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DOE/TIC-22800

TRANSURANIC ELEMENTS
IN THE ENVIRONMENT

A Summary of Environmental Research on Transuranium

Radionuclides Funded by the U. S. Department of Energy

Through Calendar Year 1979

Wayne C. Hanson, Editor
Pacific Northwest Laboratory

Prepared for the U. S. Department of Energy

Assistant Secretary for Environment
Office of Health and Environmental Research

1980

Published by
Technical Information Center/U. S. Department of Energy

NOTICE

Internationa! Copyright, © U. S. Department of Energy, 1980, under the
provisions of the Universal Copyright Convention. United States copyright is not
asserted under the United States Copyright Law, Title 17, United States Code.

Library of Congress Cataloging in Publication Data

Main entry under title:

Transuranic elements in the environment.

"DOE/TIC-22800."

Includes bibliographical references and index

1. Transuranium elements — Environmental aspects. 2. Radioecology.
3. Radioactive pollution. I. Hanson, Wayne C. II. United States. Dept.
of Energy. Office of Health and Environmental Research.
QH545.T74T73. 574.5'222 80-607069

ISBN 0-87079-119-2

Available as DOE/TlC-22800 for $18.50 from

National Technical Information Service
U. S. Department of Commerce
Springfield, Virginia 22161

DOE Distribution Category UC-11

Printed in the United States of America
April 1980

Foreword

Before 1973 environmental research into the behavior of the transuranium elements was
conducted on an ad hoc basis. It was usually prompted by some contamination event,
such as the loss of nuclear material in the military aircraft accidents at Palomares, Spain,
and Thule, Greenland, or the discovery of plutonium concentrations that exceeded
fallout levels at such locations as the Rocky Flats Plant near Golden, Colo., and the
Nevada Test Site. These research activities were usually aimed at describing the
distribution of plutonium and appraising the health hazard at the individual site. Because
this information was gathered at specific sites, it was not sufficient for generalized
statements about environmental movement. In about 1970 the Nevada Applied Ecology
Group began an integrated program at the Nevada Test Site in an attempt to provide a
broader information base on transuranium elements. This program, however, was
applicable primarily to desert environments. Some experimental studies at other locations
were concerned with the uptake of transuranium elements by vegetation, but most of
these dealt with western soils of high pH. No concerted effort was made to study
transuranic radionuclide behavior in the marine environment except for studies at Thule,
Greenland, and the Pacific Testing Grounds in the Marshall Islands.

In 1973 the U. S. Atomic Energy Commission, Division of Biomedical and
Environmental Research (BER) (now U. S. Department of Energy, Office of Health and
Environmental Research), performed an intensive study of its research efforts in support
of the development of nuclear power with special emphasis on the Liquid Metal Fast
Breeder Reactor (LMFBR). The environmental team reviewed information gathered up to
that time on transuranic cycling in various environments and concluded that a
comprehensive description of the environmental hazards of plutonium and the other
transuranium elements relative to the LMFBR could not be made with the available data
nor would it be forthcoming with the established research by BER contractors. It was
obvious that too much of the past research had been centered on studies in the western
regions, which were arid or semiarid, and essentially no studies had been made of soil
movement and plant uptake in the humid eastern regions where fuel reprocessing plants
were scheduled to operate. In addition, very little information was available on the
cycling of plutonium through aquatic food webs inclusive of the marine studies in
Greenland and the Marshall Islands. Essentially no research on the environmental
behavior of transplutonic elements was under way, and the question of the long-term
behavior and fate of the transuranium elements had not been addressed in any effective
way. Further, the question of biological modification of the transuranium elements,
which might lead to increased mobilization in the environment and possible underestima-
tions of the dose to man, was not addressed.

II!

iv FOREWORD

The conclusions of the environmental team prompted AEC to develop a research
program wliich would develop the information that was not available and which would be
as comprehensive as possible for future assessments on the impacts of transuranic
radionuclides from all stages in the nuclear fuel cycle. The program was designed to take
advantage of the high-quality research that was already under way at the Health and
Safety Laboratory, the Nevada Test Site, Lawrence Livermore Laboratory, Pacific
Northwest Laboratory, the University of California at Los Angeles, Woods Hole
Ocean ographic Institute, and the University of Washington and to complement this
research with research projects in other geographies and climates. Research activities were
selected to cover all aspects of environmental transport from soil processes to ecosystem
cycling. The objectives of the research program were to understand the cycling behavior
of the transuranium elements in our environment and to determine to what degree these
elements would be transported to us through food chains and aerial pathways. A further
objective was to develop a satisfactory description of the degree to which the
transuranium elements persist in the environment as a first step in assessing the potential
hazard of these species on a historical and geological time scale.

To achieve the broad objectives of this program, we must answer many questions. If
the transport of the transuranium elements is to be described, we must know to what
degree these elements can be mobilized in the soils and aquatic sediments where they
reside. A compendium of concentration ratios for plant uptake into food crops on various
agricultural soils must be assembled. The transport through aquatic and terrestrial food
chains must be quantified and appraised for the potential for human ingestion. Changes in
the availability of transuranic elements due to resuspension from soils also must be better
understood. Although the thrust of this work is on environmental transport, all the
research scientists are alert for unusual concentration processes that might lead to
radiological effects within environmental systems.

The areas of research just mentioned are of immediate concern, but beyond these
near-term considerations are those related to the possible long-term persistence of the
transuranic elements in available form on the scale of hundreds and thousands of years.
Such considerations are very difficult to address adequately with contemporary research.
However, two approaches are under way which may provide reasonable first approxima-
tions to the prediction of long-term behavior. One is the theoretical approach to studying
the chemical and physical processes in soil of these radionuclides with the objective of
developing good thermodynamic data. We need information on the equilibrium
concentrations of the various oxidation states in different environments, on complexation
processes, and on diffusion coefficients for various species. We can then apply this type of
information to predictive modeling. An empirical approach would be to study the
distribution and environmental behavior of naturally occurring elements that have
properties analogous to those of the transuranium elements. For instance, the availability
for plant uptake of the rare earth neodymium, which has been subjected to weathering
for thousands of years, may provide a basis for predicting the uptake of americium after
long periods of time because americium and neodymium have quite similar chemical and
physical properties. Other rare earths, uranium, and thorium are also candidates as
analogs for some of the transuranic elements in environmental studies.

The general areas of this program are outlined in Fig. 1. Research has now been
conducted for periods of time ranging from 2 yr for new work to 6 yr for work that
preceded this comprehensive program. Some of the results have been published in

FOREWORD

LMFBR PROGRAM;
PLUTONIUM AND OTHER
TRANSURANIC ELEMENTS

TERRESTRIAL

SOIL
STUDIES

ORGANIC

COMPLEXES
BIOLOGICAL

MOBILIZATION
WEATHERING

EFFECTS
COMPLEXATION
TRANSPORT
SORPTION

PLANT UPTAKE/
AVAILABILITY

VIA ROOT

ABSORPTION
VIA FOLIAR

ABSORPTION
AGRICULTURAL

DIET SURVEYS

ECOSYSTEMS AND
FOOD-CHAIN
TRANSFERS

AQUATIC

SEDIMENT
STUDIES

ARID LANDS
SHORT GRASS

PRAIRIE
ATLANTIC COASTAL

PLAIN
EASTERN DECIDUOUS

FOREST BIOME
ARCTIC-SUBARCTIC

REGIONS
TROPICAL

ENVIRONMENTS
RUMINANT FEEDING

STUDIES
MODELING

SEDIMENT

TRANSPORT
SEDIMENT-WATER

INTERFACE

BIOLOGICAL
TRANSPORT

MICROBIAL

MOBILIZATION
LITTORAL ZONE

BIOTA
PELAGIC ZONE

BIOTA
WATERFOWL

Fig. 1 United States Department of Energy, Office of Health and Environmental
Research, transuranium elements research program.

laboratory reports, government documents, and refereed journals. This volume is an
attempt to assemble the accumulated information as a synthesis document to provide an
up-to-date interpretation of the environmental behavior of the transuranium elements.

R. L. Walters

Office of Health and Environmental Research
U. S. Department of Energy

Preface

This book evolved from recommendations made at the Second Workshop on Environ-
mental Research for Transuranic Elements held at Seattle, Wash., Nov. 12—14, 1975
(proceedings available as ERDA- 76/134), under the sponsorship of the Environmental
Programs Branch of the U. S. Energy Research and Development Administration, Division
of Biomedical and Environmental Research. A sixfold expansion of research on the
environmental aspects of transuranic elements had occurred since the first plutonium
workshop was held at Estes Park, Colo., July 11-12, 1974, and a need for greater
communication of research results was identified. It was felt that investigators would be
encouraged to publish their work in open literature following the publishing of a single
publication that summarized the available information from the several Environmental
Programs Branch investigations.

The objectives of this book are to assemble the available information on the behavior
of transuranic nuclides in the environment following their release from a variety of source
terms, their translocation by physical and biological transport phenomena, and the
interpretation of the consequences of such concentrations as might be found in higher
trophic levels of food webs. All such studies require attention to sampling design to
provide accurate data, much of which may be destined as input to computer simulation
models that are capable of treating the several variables and long half-lives that are
involved in projecting the long-term environmental consequences of the nuclear fuel cycle
of the future.

The authors were asked to emphasize both similarities and differences in transuranic
nuclide behavior in various environmental settings and to identify research needs as they
perceived them. Topics were assigned to scientists who were considered to be best
qualified to address particular areas of research, often as members of a team. The
cooperation of several of the authors in making this arrangement function properly was
most gratifying and has magnified the application of tlieir research to a better
understanding of a difficult class of elements.

The outline of the volume and an initial evaluation of many of the manuscripts were
presented at the Third Workshop on Environmental Research for the Transuranic
Elements held at Woods Hole, Mass., Apr. 18-22. 1977 (proceedings available as
CONF-770429). The hope that it would provide a comprehensive review and editorial
comment of a rough draft of this book was only partially realized. That it survived at all
is due in large part to the perseverance of the session chairmen.

The importance of transuranic nuclides in terms of their long physical half-lives,
chemical toxicities, the effective linear energy transfer of their radiations, and the
appreciable public concern about their release to the environment has prompted several

Vll

via PREFACE

symposia on this subject in recent years. As a result, the reader may wonder: "Why
another publication on transuranic nuclides in the environment?" It is our intention that
this book should not be "just another pretty face" but that it should uniquely represent
an integration and synthesis of research results from a balanced program of studies with a
common funding source. How effective we have been in achieving our goal is left to the
reader to judge.

Wayne C. Hanson

Editor

Acknowledgments

The encouragement and support of R. L. Watters, W. 0. Forster, and H. M. McCammon
of the U. S. Department of Energy, Office of Health and Environmental Research, in the
planning and executing of this publication are greatly appreciated. The assistance of D. N.
Edgington, T. E. Hakonson, M. H. Smith, R. L. Watters, F. W. Whicker, and R. E.
Wildung in providing the comprehensive synthesis part of this volume and for initial
evaluation of various manuscripts is gratefully acknowledged. Many of the authors also
served as reviewers of articles in their areas of expertise. Appreciation is expressed for
additional advice on manuscripts obtained from the following persons:

A. J. Ahlquist, Los Alamos Scientific Laboratory

G. Choppin, Florida State University

J. M. Cleveland, U. S. Geological Survey, Lakewood, Colorado

P. B. Dunaway, U. S. Department of Energy, Las Vegas, Nevada

P. W. Durbin, Lawrence Berkeley Laboratory

S. W. Fowler, International Laboratory of Marine Radioactivity, Monaco

R. Fukai, International Laboratory of Marine Radioactivity, Monaco

D. W. Gillette, National Center for Atmospheric Research

J. A. Hayden, Rockwell International, Rocky Flats

J. W. Healy, Los Alamos Scientific Laboratory

P. W. Krey, U. S. Department of Energy, Environmental Measurements Laboratory

C. L. Osterberg, International Laboratory of Marine Radioactivity, Monaco

J. Pentreath, Ministry of Agriculture, Food and Fisheries, Lowestoft, England

V. Schultz, Washington State University

L. C. Schwendiman, Battelle Pacific Northwest Laboratory

F. B. Turner, University of California, Los Angeles

D. C. Wolf, University of Arkansas

D. S. Woodhead, International Laboratory of Marine Radioactivity, Monaco

Thanks are extended to many of the authors who promptly responded to the several
deadlines that came and went during the long and tedious effort to produce this volume.

We greatly appreciate the secretarial support of V. J. McCabe and M. A. Rosenthal of
the Los Alamos Scientific Laboratory and L. V. Kupinski and K. A. Tallent of Battelle
Pacific Northwest Laboratory. Financial support was supplied by DOE contracts
W-7405-ENG-36 and EY-76-C-06-1830 at Los Alamos Scientific Laboratory and Battelle
Pacific Northwest Laboratory, respectively.

M. C. Fox, Jean S. Smith, and other members of the DOE Technical Information
Center provided expert technical editing and other assistance in the publication of this

X A CKNO WLEDGMENTS

volume. R. F. Pigeon of the DOE Office of Information Services rendered iielpful advice
for achieving the editorial goals. M. L. Merritt of Sandia Laboratories gave me the benefit
of his experience in editing the DOE publication Tlie Environment of Amchitka Island,
Alaska, and otherwise provided suggestions.

Wayne C. Hanson
Editor

Contributors

D. C. Adriano

Savannah River Ecology Laboratory, Aiken, South CaroHna
T. M. Beasley

Oregon State University, Newport, Oregon
S. G. Bloom

Battelle Columbus Laboratories, Columbus, Ohio

E. A. Bondietti

Oak Ridge National Laboratory, Oak Ridge, Tennessee
A. L. Boni

E. L du Pont de Nemours and Company, Aiken, South Carolina
D. A. Cataldo

Battelle Pacific Northwest Laboratory, Richland, Washington
J. F. Cline

Battelle Pacific Northwest Laboratory. Richland, Washington
J. C. Corey

E. L du Pont de Nemours and Company, Aiken. South Carolina

F. A. Cross

National Marine Fisheries Service, Beaufort, South Carolina
R. C. Dahlman

U. S. Department of Energy, Washington, D. C.
T. J. Dobry

U. S. Department of Energy, Washington, D. C.
P. G. Doctor

Battelle Pacific Northwest Laboratory, Richland, Washington
L. L. Eberhardt

Battelle Pacific Northwest Laboratory, Richland, Washington
D. N. Edgington

University of Wisconsin, Milwaukee, Wisconsin
D. R. File

U. S. Department of Energy, Richland, Washington
R. M. Emery

Battelle Pacific Northwest Laboratory, Richland, Washington
L. D. Eyman

Oak Ridge National Laboratory, Oak Ridge, Tennessee

G. C. Facer

U. S. Department of Energy, Washington, D. C.

xii CONTRIBUTORS

R. H. Gardner

Oak Ridge National Laboratory, Oak Ridge, Tennessee
T. R. Garland

Battelle Pacific Northwest Laboratory, Richland, Washington

C. T. Garten, Jr.

Oak Ridge National Laboratory, Oak Ridge, Tennessee
R. 0. Gilbert

Battelle Pacific Northwest Laboratory, Richland, Washington
T. E. Hakonson

Los Alamos Scientific Laboratory, Los Alamos, New Mexico
W. R. Hansen

Los Alamos Scientific Laboratory, Los Alamos, New Mexico
W. C. Hanson

Battelle Pacific Northwest Laboratory, Richland, Washington

D. W. Hayes

E. L du Pont de Nemours and Company, Aiken, South Carolina
J. W. Healy

Los Alamos Scientific Laboratory, Los Alamos, New Mexico
J. H. Horton

E. L du Pont de Nemours and Company, Aiken, South CaroHna
D. C. Klopfer

Battelle Pacific Northwest Laboratory, Richland. Washington
M. R. Kreiter

Battelle Pacific Northwest Laboratory, Richland, Washington
C. A. Little

Oak Ridge National Laboratory, Oak Ridge, Tennessee
F. G. Lowman

Environmental Protection Agency. Narragansett, Rhode Island
R. W. McKee

Battelle Pacific Northwest Laboratory, Richland, Washington
M. C. McShane

Battelle Pacific Northwest Laboratory. Richland. Washington
R. P. Marshall

University of Washington. Seattle Washington
W. E. Martin

Battelle Columbus Laboratories Columbus. Ohio
J. E. Mendel

Battelle Pacific Northwest Laboratory, Richland, Washington
V. E. Noshkin

Lawrence Livermore Laboratory, Livermore, California
J. W. Nyhan

Los Alamos Scientific Laboratory, Los Alamos, New Mexico

C. R. Olsen

Lamont-Doherty Geological Observatory, Palisades, New York

D. Paine

Rockwell International, Richland, Washington
R. W. Perkins

Battelle Pacific Northwest Laboratory, Richland, Washington

CONTRIBUTORS xiii

J. E.Pinderlll

Savannah River Ecology Laboratory. Aiken, South Carolina
J. A. Robbins

University of Michigan, Ann Arbor, Michigan

E. M. Romney

University of California, Los Angeles, CaUfornia
S. M. Sanders. Jr.

E. L du Pont de Nemours and Company, Aiken, South Carolina
W. R. Schell

University of Washington, Seattle, Washington
R. G. Schreckhise

Battelle Pacific Northwest Laboratory, Richland, Washington
G. A. Sehmel

Battelle Pacific Northwest Laboratory, Richland, Washington
H. J. Simpson

Lamont-Doherty Geological Observatory, Palisades, New York
M. H. Smith

Savannah River Ecology Laboratory, Aiken, South Carolina
T. Tamura

Oak Ridge National Laboratory, Oak Ridge, Tennessee
W. L. Templeton

Battelle Pacific Northwest Laboratory, Richland, Washington
C. W. Thomas

Battelle Pacific Northwest Laboratory, Richland, Washington
R. C. Thompson

Battelle Pacific Northwest Laboratory, Richland, Washington
J. R. Trabalka

Oak Ridge National Laboratory, Oak Ridge, Tennessee
R. M. Trier

Lamont-Doherty Geological Observatory, Palisades, New York
B. E. Vaughan

Battelle Pacific Northwest Laboratory. Richland, Washington
B. W. Wachholz

U. S. Department of Energy, Washington, D. C.
M. A. Wahlgren

Argonne National Laboratory, Argonne, Illinois
A. Wallace

University of California, Los Angeles, California
R. L. Watters

U. S. Department of Energy, Washington, D. C.

F. W. Whicker

Colorado State University, Fort CoUins, Colorado
R. E. Wildung

Battelle Pacific Northwest Laboratory, Richland, Washington

Contents

FOREWORD iii

PREFACE vii

ACKNOWLEDGMENTS ix

CONTRIBUTORS xi

SYNTHESIS OF THE RESEARCH LITERATURE 1

R. L. Walters, D. .\. Edgington, T. E. Hakonson, W. C. Hanson, M. H. Smith,
F. W. Whicker, and R. E. Wildung

INTRODUCTION

Radiological Assessments, Environmental Monitoring, and Study Design 45

Wayne R. Hansen and Donald R. Elle

SOURCE TERMS

Worldwide Fallout 53

R. W. Perkins and C. W. Thomas

Transuranic Elements in Space Nuclear Power Systems 83

Thaddeus J. Dobry, Jr.

Quantities of Transuranic Elements in the Environment from Operations
Relating to Nuclear Weapons 86

Gordon Facer

Transuranic Wastes from the Commercial Light-Water-Reactor Cycle 92

M. R. Kreiter, J. E. Mendel, and R. W. McKee

INVENTORY AND DISTRIBUTION

The Detection and Study of Plutonium-Bearing Particles Following the 107

Reprocessing of Reactor Fuel

5. Marshall Sanders, Jr., and Albert L. Boni

Physicochemical Associations of Plutonium and Other Actinides in Soils 145

E. A. Bondietti and T. Tamura

Sources of Variation in Soil Plutonium Concentrations 165

John E. Finder HI and Donald Paine

xvi CONTENTS

Statistics and Sampling in Transuranic Studies 173

L. L. Eberhardt and R. O. Gilbert

Appropriate Use of Ratios in Environmental Transuranic Element Studies 187

P. G. Doctor, R. O. Gilbert, and J. E. Finder III

TERRESTRIAL ECOSYSTEMS
Experimental Studies

Review of Resuspension Models 209

/. W. Healy

Transuranic and Tracer Simulant Resuspension 236

G. A. Sehmel

Interaction of Airborne Plutonium with Plant Foliage 288

D. A. Cataldo and B. E. Vaughan

The Relationsliip of Microbial Processes to the Fate and Behavior of
Transuranic Elements in Soils, Plants, and Animals 300

R. E. Wildung and T. R. Garland

Uptake of Transuranic Nuclides from Soil by Plants Grown Under Controlled
Environmental Conditions 336

D. C. Adriano, A. Wallace, and E. M. Romney

Comparative Uptake and Distribution of Plutonium, Americium, Curium, and
Neptunium in Four Plant Species 361

R. G. Schreckhise and J. F. Cline

Field Studies

Comparative Distribution of Plutonium in Contaminated Ecosystems at
Oak Ridge, Tennessee, and Los Alamos, New Mexico 371

Roger C. Dahlman. Charles T. Garten, Jr.. and Thomas E. Hakonson

Plutonium Contents of Field Crops in the Southeastern United States 381

D. C. Adriano, J. C. Corey, and R. C Dahlman

Ecological Relationships of Plutonium in Southwest Ecosystems 403

T. E. Hakonson and J. W. Nyhan

Plutonium in a Grassland Ecosystem 420

Craig A. Little

Transuranic Elements in Arctic Tundra Ecosystems 441

Wayne C. Hanson

Models

Nevada Applied Ecology Group Model for Estimating Plutonium Transport
and Dose to Man 459

W. E. Martin and S. G. Bloom

A Model of Plutonium Dynamics in a Deciduous Forest Ecosystem 513

Charles T. Garten, Jr., Robert H. Gardner, and Roger C. Dahlman

CONTENTS xvii

AQUATIC ECOSYSTEMS

Marine Studies

A Review of Biokinetic and Biological Transport of Transuranic Radionuclides
in the Marine Environment 524

T. M. Beasley and F. A. Cross

Geochemistry of Transuranic Elements at Bikini Atoll 541

W. R. Schell F. G. Lownwn. and R. P. Marshall

Transuranium Radionuclides in Components of the Benthic Environment of
Enewetak Atoll 578

V. E. Nosh kin

Plutonium and Americium Behavior in the Savannah River Marine Environment 602

D. W. Hayes and J. H. Norton

Freshwater

Patterns of Transuranic Uptake by Aquatic Organisms: Consequences and
Implications 612

L D. Eyman and J. R. Trabalka

The Migration of Plutonium from a Freshwater Ecosystem at Hanford 625

Richard M. Emery, Donald C. Klopfer, and M. Colleen McShane

Plutonium in Rocky Flats Freshwater Systems 644

D. Paine

Plutonium in the Great Lakes 659

M A. Wahlgren, J. A. Robbins, and D. N. Edgington

Transport of Plutonium by Rivers 684

H. J. Simpson, R. M. Trier, and C. R. Olsen

BIOLOGICAL EFFECTS

Biological Effects of Transuranic Elements in the Environment: Human
Effects and Risk Estimates 691

Roy C. Thompson and Bruce W. Wachholz

Ecological Effects of Transuranics in the Terrestrial Environment 701

F. W. Whicker

Dosimetry and Ecological Effects of Transuranics in the Marine Environment 714

William L. Templeton

INDEX 722

Synthesis of the Research Literature

R. L. WAITERS, D. N. EDGINGTON, T. E. HAKONSON, W. C. HANSON,
M. H. SMITH, F. W. WHICKER, and R. E. WILDUNG

This book provides a compendium of enviromnental research related to transuranium
elements; this research has developed greatly over the last 5 yr. The individual chapters
describe studies that deal with mobility and transport in various environmental media and
physiographic provinces. The intent of this synthesis is to develop, from the information
in this book and other publications, unifying ideas and generaHzations about movement
of these elements through the environment to the human population.

Chemical, physical, and biotic processes control movement of transuranic elements
within ecosystems. As illustrated by the conceptual model in Fig. 1, transport processes
are driven by wind, water, biotic, and mechanical activity. For example, wind, water, or
mechanical resuspension of soil and sediment can result in contamination of plant and
animal surfaces, and the diet of consumer organisms may therefore contain this surficially
deposited material. Examples of biotic transport include the movement of soil
contaminants associated with a grazing animal and the subsequent redistribution of this
material through defecation and/or death. Burrowing and grooming activities, which
result in contamination of the animal, are additional examples of biotic transport.

Examples of chemical transport are the passage of soluble contaminants from soil
through plant roots or across physiological membranes (e.g., lungs or gut wall) and the
vertical leaching of soluble contaminants through the soil, although these processes may
involve biochemical and physical parameters.

To predict the behavior of transuranic elements in the environment, one must
understand (l)the ecological relationships in contaminated ecosystems, including the
content and size of compartments and the exchange of materials between compartments,
and (2) the pathways, rates, and mechanisms of transport through the ecosystems.

The behavior of transuranic elements in the environment must be described, at
present, in terms of data obtained from direct sampling of sites with different
contamination histories, sources, and ecological features. This information, together with
data from laboratory studies defining rates and mechanisms, provides the framework for
consideration of environmental fates and effects.

The results of this synthesis are, in many cases, tentative conclusions — as one would
expect in any process involving inductive reasoning. Most of the inferences must be drawn
from data pertaining to plutonium, which has been more intensively studied than other
transuranic elements now under investigation. Some of these conclusions will become
well established as more evidence accumulates; others may require modification to
emphasize exceptions.

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

ATMOSPHERE

SEDIMENT

Fig. 1 Movement of transuranic elements to man from atmosphere, aquatic, and
terrestrial components of the biosphere.

This synthesis is intended to clarify the present status of knowledge about transuranic
elements and should stimulate further research that will define more clearly the
environmental behavior of the transuranic elements and foster further analysis of new
data as they become available.

Distribution and Inventory

Sources

Atomic weapons testing has been the major source of transuranic elements in the
general environment. A portion of the debris from these tests was transferred into the
stratosphere and then slowly returned to the Earth's surface. Because the transuranic
elements were exposed to high temperatures, it was assumed that they were formed as
high-fired oxides. It was further assumed that neither plutonium nor americium would
move readily into biological systems because high-fired oxide particles would not dissolve
in natural waters or, if they did, would form insoluble polymeric hydrated oxides.
Further, it has been suggested that the behavior of transuranic elements in the
environment is a function of source and, for plutonium, isotopic composition.

If these assumptions are accepted, transuranic elements are unlike any other element
in the periodic table. However, experimental evidence relating to plutonium and
americium in a wide variety of environments does not bear out these assumptions. It is
known that a significant fraction of the plutonium deposited on the surface of the earth

SYNTHESIS OF THE RESEARCH LITERATURE 3

as fallout was produced by an (n,7) reaction with ^^^U and the subsequent decay of
^^^U through "^Np to ^^^Pu (Joseph et al., 1971).* Thus a proportion of the
plutonium in the environment was formed as single atoms long after the explosion and
was never involved in a reaction to form a high-fired oxide.

The chemical form of transuranic elements released in small quantity during nuclear
fuel reprocessing and fabrication may range from relatively insoluble oxide particles,
which are of different composition than fallout (Sanders, 1977), to relatively soluble
inorganic salts and organic complexes, which may be present in solid and liquid wastes. A
generalized representation of the chemical forms of the transuranic elements released to
the environment is given in Table 1.

TABLE 1 Major Sources and Initial Forms of
Transuranic Elements Entering the Environment

Source

Form*

Nuclear weapons testing

TuOx • MOx

(global fallout, debris)

TuOx • UjOg

Nuclear fuel reprocessing/fabrication

TuOx • nHjO

TuOx • MOx

TUNO3

Tu organic complexes

*Tu denotes plutonium and possibly americium, curium,
and neptunium. M represents metal impurities, dust, and
gaseous condensation products.

The potential movement of a transuranic pollutant from several sources, as illustrated
for plutonium in Fig. 1, can be classified according to expected initial solubility in surface
waters, interstitial waters of soils and sediments, and perhaps even on lung surfaces.
Initially, particulate oxides of transuranic elements may be largely insoluble in solution.
Ultimately, solubility will be a function of the chemical and physical properties of the
particle and the matrix in which the particle is deposited. Oxide particles of the highest
specific activity and containing the highest concentrations of impurities in the crystal
lattice will exhibit the greatest solubility. The combination of configuration and
equivalent diameter, as reflected in surface area exposed to solution, will also influence
the rate of oxide solubility. Once dissolved, transuranic elements will be subject to
chemical reactions governing dissolved salts. Hydrolyzable transuranic elements entering
the environment in solutions sufficiently acidic to maintain soluble ions and in
concentrations exceeding that of natural complexing agents will be rapidly hydrolyzed on
dilution and subsequently precipitated on particle surfaces. These include Pu(III, IV, and
likely VI), Am(III), Cm(III), and Np(IV and VI), ahhough the rates of hydrolysis will
vary between oxidation states. Conversely, chemical species of transuranic elements not
subject to marked hydrolysis, such as Pu(V) and Np(V), initially may be more soluble

*As stated in the reference, approximately 1 x 10^" atoms of ^^'Pu were generated by the
explosion of thermonuclear weapons. This amounts to 4000 kg, or about two-thirds of the total
plutonium estimated to have been deposited on the surface of the earth.

4 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

than the above species. ImmobiHzation of those chemical species may occur through
cation exchange reactions with particle surfaces or through redox reactions to
hydrolyzable forms that become insoluble.

Transuranic elements entering the environment as stable organocomplexes, as might
occur in the vicinity of a spent-fuel reprocessing facility, may be highly soluble initially
(Wildung and Garland, 1975). The duration of solubility and mobility will be a function
of the stability of the complex to substitution by major competing ions, such as Ca^ "*" and
ff*" (Lahav and Hochberg, 1976; Stevenson and Ardakani, 1972; Norvell, 1972), tlie
competition of other Hgands forming more stable compounds, and the resistance of the
organic ligand to chemical and microbial decomposition (Wildung and Garland, 1975).
Disruption of the complex may lead to marked reduction in solubility through hydrolysis
and precipitation reactions, as described for acid solutions on dilution. A portion of the
ions released may react with other, perhaps more stable, ligands. The mobility of the
intact complexes, in turn, will be principally a function of their chemical and
microbiological stability and the charge on the complex, which will govern the degree of
sorption on particles.

Initial chemical reactions and tendencies to remain soluble after release to the
environment apparently depend on the initial chemical form of transuranic elements.
However, the original source characteristics become less important as times goes on and
weathering and aging processes proceed. From a consideration of the known distribution
of transuranic elements in the environment and expected solubilities in the presence of
particle surfaces, it is clear that their behavior will be markedly influenced by their
individual chamistries and their chemical interactions in soils and sediments.

The effect of source and the immediate environment on the distribution of trans-
uranic elements can be illustrated by comparison of the concentration of plutonium
resulting from global fallout with that from more localized sources in soils, sediments,
and waters (Table 2). For nuclear weapons testing, highest concentrations of plutonium
in soils occur at the test locations. However, after stratospheric dispersion, concentrations
are relatively low [<0.1 (d/min)/g] in surface soils, fresh water, and marine sediments.
The lower concentrations of plutonium in marine sediments relative to those in soils
reflect the longer residence times in the water column. Where nuclear processing faciUties
are known to provide a source of plutonium, soil and sediment concentrations range from
fallout levels to several thousand disintegrations per minute per gram in controlled areas.
Of major significance from the standpoint of environmental behavior is the fact that
concentrations of plutonium in soils and sediments generally exceed tliose in water and
other media by many orders of magnitude.

Terrestrial Ecosystems

One way of examining the distribution of any element within an ecosystem is through tlie
use of an inventory ratio (IR). Two types of data are needed to calculate IR's: the weight
(W) of each ecosystem compartment and the concentration (C) of the element within
each compartment. The IR differs from the concentration ratio (CR) in that it takes into
account the size of the compartments. For our discussion the compartments are soil,
vegetation, htter, and animals.
The IR is calculated as

SYNTHESIS OF THE RESEARCH LITERATURE 5

where Ac is the amount of the chemical in a compartment and At is the total amount in
the ecosystem. Thus IR can be dramatically changed by the size of the soil compartment,
which is a direct function of the depth used to calculate this parameter. The amount (A)
of the chemical of interest is simply

A = WC

where W and C are given in dry-weight units. These calculations do not require
information about transfer rates or pathways within the ecosystem.

TABLE 2 Plutonium in Soils, Sediments, and Waters

Source and locations

Concentration

/2 3» '2*0py\*

Reference

Soils and sedinients,f (d/min)/g

Nuclear weapons testing
Global faUout (soU)
Debris (NTS, soil)
Bikini Atoll (soil)
Lake Michigan (sediment)

North Atlantic (sediment,
5597 to 5968 m)

Chemical processing
Savannah River, S. C. (soil)
Rocky Flats, Colo, (soil)
Hanford, Wash. (soU)
Hanford Z-9 trench, Wash, (soil,

subsurface waste site)
Maxey Flats, Ky. (soil, waste

site)
Irish Sea (sediment)

Nuclear weapons testing
Enewetak Atoll (groundwater)
Lake Michigan
North and South Atlantic
North and South Pacific

Chemical processing, runoff
Newport, S. C, Estuary
Savannah River, S. C. (freshwater

pretreatment)
Savannah River, S. C. (treated

drinking water)
Irish Sea

Hardy, 1974

Romney et al., 1976

Nevissi, Shell, and Nelson, 1976

Edgjngton, Wahlgren, and Marshall,

1976
Bowen, Livingston, and Burke, 1976

Adriano et al., 1975

Krey and Hardy, 1970

Corley, Robertson, and Brauer, 1971

Smith, 1973

Meyer, 1976

Hetherington et al., 1976

0.01-0.05
180.00-1.1 X 10'
1.10-800.00
0.20-0.90

0.00-0.017

0.05-28.00
0.01-150.00
0.01-1.50
1.50 X 10''-1.5 X 10'

0.90-9.00

0.70-105.00

Waters, (d/min)/liter

4x10"'' -1.50 Noshkinetal.,1976

5 X 10-^-7 X 10"'' Wahlgren etal., 1976

3 X lO"* -4 x 10~' Bowen, Wong, and Noshkin, 1971

1.30-9.4 X 10-" Miyake and Sugimura, 1976

5 X 10-" -5.6 X 10-' Hayes, LeRoy, and Cross, 1976

5x10-' Corey and Boni, 1976

2x 10-"

0.10-1.0

Corey and Boni, 1976
Hetherington et al., 1976

*Values should not be considered representative but rather as examples of values obtained through
various studies in specific localities. The original literature should be consulted before use of this
information for other purposes.

■[Expressed sediments on a dry -weight basis.

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Range in Inventory Ratios for Plutonium in
Major Ecosystem Compartments

Compartment

Range in
inventory ratio*

Reference

Soils

0.998 to 0.986

Romney and Wallace, 1977;

Dahlman, Garten, and Hakonson,

this volume; Little, this volume
Hanson, 1975; Romney et al., 1976;

Dahlman, Garten, and Hakonson,

this volume ; Little, this volume;

Pinderetal., 1979
3 X 10~^ to 2 X 10"" Dahlman, Garten, and Hakonson,

this volume; Little, this volume
7 X 10~^ to 6 X IC^' ' Dahlman, Garten, and Hakonson,

this volume; Little, this volume

Vegetation 2 x 10~^ to 3 x 10"

Litter

Animals

*The proportion of the total plutonium in the ecosystem that is found in
each major compartment.

More than 99% of the plutonium inventory is found in the soil compartment of most
ecosystems (Table 3), and most of the contamination occurs near the soil surface
(Francis, 1973; Little and Whicker, 1977). Notable exceptions occur in arctic systems
where lichens intercept and retain fallout for long periods of time and in ecosystems that
are still receiving aerial depositions from nuclear processing facilities. However, even in
these special cases soil will be the eventual repository after deposition ceases (Hanson,
this volume; Holm and Persson, 1975; Dahlman and McLeod, 1977).

At some sites, a considerable amount of water has percolated into the soil since
the initial deposition (e.g.. Savannah River Plant), but still the major inventory of
plutonium is in the top few centimeters of soil. The concentration of plutonium is only
occasionally higher in the subsoils below 10 cm than in the surface materials (Essington et
al., 1976). Plutonium is found at depths greater than 20 cm but usually in very low
concentrations unless soil or sediment mixing is actively in progress. Such mixing can
occur in steep canyons and delta regions of running-water ecosystems (Nyhan, Miera, and
Peters, 1975) and in terrestrial sites where natural biotic or human activities have mixed
or buried the plutonium.

Transuranic radionuclides can often be buried and remain immobile after deposition.
The exact distribution in the soil profile has an important influence on the availabiUty of
transuranic elements for resuspension and root uptake.

The proportion of plutonium associated with biotic components of the ecosystem can
be as small as 0.1% (e.g., southeastern forests) (Dahlman, Garten, and Hakonson, this
volume). This fact reflects generally lower concentrations in biota but more importantly
the small mass of biota relative to soil. Even if transuranic elements were randomly
distributed among ecosystem components, the majority would still be associated with
soil.

The amount of plutonium associated with vegetation is greater than that associated
with animals (Table 3) but ranges over five orders of magnitude. Inventory ratios for
animals range over eight orders of magnitude. Most of this variation is probably due to
the amount of surface contamination on samples and not to internal concentrations of
plutonium in plants and animals (Dahlman and McLeod, 1977).

SYNTHESIS OF THE RESEARCH LITERATURE 7

Autoradiographs of leaf tissues show that plutonium occurs primarily in discrete
particles of suspendible size on the surface (Romney and Wallace, 1977). Washing plant
materials to remove surface contamination reduces the concentration of plutonium in
subsequent analyses (Dahlman and McLeod, 1977).

In some cases (e.g., inadequately cleaned vegetables) soil can be ingested by humans.
Knowledge of the surface contamination of plant material is certainly important in
determining the amount of plutonium ingested by both animals and humans.

Analyses of animal pelts, gastrointestinal tracts, and lungs give higher concentrations
than those of tissues not exposed directly to surface contamination (Bradley, Moor, and
Naegle, 1977). Humans do not normally eat tissues exposed to direct surface
contamination to the extent that other carnivores do. Thus IR's calculated using
concentration values determined from animal samples with natural levels of surface
contamination may be more relevant in the assessment of potential environmental
problems with food chains up to, but not including, humans.

High IR values for plant Utter (Table 3) are principally due to surface contamination
because smaller soil particles are impossible to remove (Romney and Wallace, 1977). The
same is true for northern lichen-dominated communities where most dust particles are
intercepted before they reach the surface of the ground (Hanson, 1966; Holm and
Persson, 1975). Around transuranic processing facilities, aerial deposition of transuranic-
bearing particles is probably the dominant form of contamination of plants (Pinder et al.,
1979). Resuspension contributes importantly to surface deposition of contaminants and
increases plant concentration values even in relatively moist environments (Dahlman,
Garten, and Hakonson, this volume; Pinder et al., 1979).

There are few estimates of biomass of higher carnivores relative to that of vegetation.
In addition, contaminated areas are generally Umited in size and frequently include only
parts of the ranges of a few individuals. Hence reliable IR's are not available for upper
trophic levels, and animals are considered here as one compartment within the ecosystem.
Inventory ratios may have characteristic values for certain ecosystems, and identification
of ecosystem attributes allowing prediction of IR's would aid in assessing hazards.
Knowledge of the relative biomass of ecosystem components will always be useful in
modeling the long-term distribution of most contaminants, including transuranic
elements. Future research on IR's should emphasize the establishment of predictable
relationships and the identification of variables affecting IR values.

Aquatic Ecosystems

Experimental studies in the Great Lakes (Edgington and Robbins, 1975), Buzzards Bay
(Livingston and Bowen, 1976), Irish Sea (Hetherington, Jefferies, and Lovett, 1975;
Hetherington, 1978), and Trombay Harbor (Pillai and Mathew, 1976) have shown that, in
comparatively shallow bodies of water, more than 96% of the total plutonium released to
these environments is rapidly transferred to sediments. However, Bowen, Wong, and
Noshkin (1971) estimated that as of 1969 10 to 20% of the total plutonium in deep
oceans had been deposited in the sediments and that this would increase to only 30% by
1980.

In those parts of Lake Michigan and Lake Erie where sedimentation rates are greater
than 5 mm/yr, a detailed analysis of plutonium and '^''Cs profiles in sediment cores
clearly reflects the worldwide fallout maximums in 1959 and 1963 (Fig. 2). Similarly, it
has been shown that americium and plutonium profiles in sediments from the Irish Sea

8 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

12 16 20

DEPTH IN CORE, cm

Fig. 2 Histogram of the distribution o.f plutonium in a sediment core from Lake Erie.
The two peaks correspond to the material deposited during years of maximum fallout,
1963 and 1959, respectively. The dashed line represents the predicted distribution based
on a sedimentation rate of 1.15 cm/yr at the surface and a mixing depth due to
bioturbation of 4 cm.

reflect the history of releases from Windscale (Hetherington, Jefferies, and Lovett, 1975;
Hetherington, 1978). Profiles measured in sediments from the Santa Barbara channel
show a continuing input of plutonium due to erosion of California soils and direct input
from faUout (Koide, Griffin, and Goldber , 1975).

In other cores from Lake Michigan and Lake Erie, it is possible, by comparing
plutonium and cesium profiles with those for ^^'^Pb, to estimate the effects of
biotic activity on lake sediments (Robbins and Edgington, 1975) and to identify massiye
disturbances in sediments of the Crest Lakes due to large storms (Edgington and
Robbins, 1976).

Some sediment cores from Lake <)ntario have plutonium concentration profiles
exhibiting subsurface maximums similar ro those found in Lake Michigan and Lake Erie
(Bowen, 1976; Edgington and Robbins 1976). However, in other cores from Lake
Ontario and Buzzards Bay, the profiles -low no subsurface maximums (Livingston and
Bowen, 1976; Bowen, Livingston, and Burke, 1976). In these cores it is clear that there
must be mixing downward by physical or biological processes. Repeated coring in
Buzzards Bay from 1964 to 1973 showed that there was a small net loss of plutonium in

SYNTHESIS OF THE RESEARCH LITERATURE 9

the sediments. This loss was interpreted as a direct return to the water column rather than
physical redistribution of sediments.

These studies indicate that sediments will be continually reworked by physical and
biological processes. New plutonium will be continually added to the Great Lakes and
coastal waters by wind erosion and transport of sediments down river. Because of the
dynamics of the system, the inventory and distribution of plutonium will continue to
change (Edgington and Robbins, 1975).

Solubility and Chemistry

Theoretical Considerations

The transuranic elements, starting with neptunium (atomic number 93), are a subset of
the actinide series. This series is similar to the lanthanide series in that electrons are
added successively to the 5f orbitals in a manner similar to the filling of the 4f orbitals.
However, the shielding of the 5f electrons by outer electrons is less effective than that of
4f electrons; thus the chemical properties of the actinides are more complicated than
those of the lanthanides. Although the latter exist primarily in the III oxidation state and
exhibit ionic bonding, the actinides (through plutonium) can exist in multiple oxidation

TABLE 4 Comparison of Oxidation States for the
Actinide Elements in Solution*

f =

9 10

Ac Th Pa

Bk Cf

3
4
5

3
4

3 3

4
5
6

5
6

*The solid lines bound the most likely oxidation states
in aqueous solution.

States (Table 4). Because of their extreme reactivity, the II and VII oxidation states are
not likely to be encountered in the environment. The oxidation— reduction behavior of
the triad U-Np— Pu is complicated, and multiple oxidation states can coexist in solution.
Actinides with atomic numbers exceeding that of plutonium behave similarly to the
lanthanides.

The complex interactions between the various oxidation states of neptunium and
plutonium are partly governed by their total concentration in solution. When concentra-
tions are sufficient, disproportionation reactions between oxidation states are common.
However, such concentrations are unlikely to be found in the envirormient, and the stable
oxidation states will be a function of the chemical environment, e.g., the presence of

1 0 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

oxidizing or reducing agents and complexing ligands. Thus the transuranic elements
neptunium and plutonium can exist in more than one oxidation state in the environment,
whereas the transplutonium elements will be 3+ cations.

Because of their very similar electronic structures and ionic radii, transuranic elements
of a given oxidation state behave similarly chemically. Thus under most conditions
Pu(III) is similar to Am(III) or rare earths, such as La(in); Pu(IV) is similar to Th(IV);
and Pu(VI) is similar to U(VI) (Wahlgren et al, 1976). These differences in oxidation
states of the transuranic elements and the ability of the elements to form complexes with
natural ligands will greatly affect their availability for transfer in the biosphere (Dahlman,
Bondietti, and Eyman, 1976).

Standard oxidation— reduction potentials can be used to predict the possible
oxidation states of actinide elements in solution under environmental conditions. This,
however, is an equilibrium prediction of the relative thermodynamic stability of various
species and does not consider the effect of the kinetics of reaction or other factors, such
as complexation, which affect the redox couple. Because measurements have been made
of these potentials in near-neutral solution (5 < pH < 9), their magnitude must be
calculated from the known hydrolytic behavior of the various ions, their respective
formation constants, and standard potentials measured in acid solution (Kraus, 1949;
Connick, 1949). The resuhing oxidation-reduction potentials estimated for plutonium in
neutral solution are :

0.94

Pu3+ ^Ml^ Pu(0H)4 • yH2 0(s) ^^ Pu02 -^^^ Pu02(OH)2

Since the potential for the oxygen couple in neutral solution 2H2 0 ->■ Oj +
4H''"(10~^M) + 4e is +0.815 volt, the oxidized species in any oxidation— reduction couple
with a higher positive potential than this is thermodynamically unstable in water,
although in practice a considerable overpotential exists (Pourbaix, 1966) which results in
extremely slow reaction rates. The formation of complexes that drive the equilibrium
potential to more thermodynamically stable values becomes extremely important.

With values of the oxidation potential (Eh) relative to the standard hydrogen
electrode calculated for the reactions of transuranic elements in solution, it is possible to
construct Eh-pH diagrams that delineate the regions of stability of ionic and solid species
as a function of pH and soluble actinide concentrations. Earlier efforts at constructing
these diagrams and phase relationships between plutonium species have been summarized
(Bondietti and Sweeton, 1977). A comparison of Eh— pH diagrams for Pu(III) ^ Pu(IV)
and Fe(II)-> Fe(III) suggested that, under environmental conditions where ferric ion is
reduced to ferrous ion, Pu(IV) may also be reduced to Pu(III) (Bondietti and Sweeton,
1977).

An Eh-pH diagram that was constructed with published values of E° (Pourbaix,
1966) is presented in Fig. 3. The III, IV, and VI oxidation states of plutonium were
included. However, recent evidence suggests that Pu(V) can exist in aerobic environments
(vide infra). The diagram shows the effect of changes in the concentration of plutonium
in solution on the regions of stabiUty of each oxidation state. Because of the tendency to
form insoluble hydrolytic species, free Pu**"*" ions can exist principally under strongly
oxidizing acid conditions (region I). In the normal range of pH and plutonium
concentrations encountered in the environment, plutonium could be present as PuOl''",
and this form will slowly come into equilibrium with sohd Pu(0H)4 (regions II and III).

SYNTHESIS OF THE RESEARCH LITERATURE 11

Fig. 3 Eh-pH diagram of stability fields for various plutonium species. Circled numbers
represent lines of transition from one oxidation state to another and approximate the
line of equilibrium between the regions in which plutonium may be susceptible to
change by changing Eh or pH under the conditions specified. Reactions that are not
considered may also occur and contain kinetic parameters of great importance. (See also
Raiand Serne, 1977.)

Because the H2O-O2 couple is relatively insensitive to small changes in the partial
pressure of oxygen, Pu02^ is thermodynamic ally stable in solution until the concentra-
tion of dissolved oxygen is essentially zero. This diagram is different from those presented
by Polzer (1971), v^hich indicated that PUO2 is stable. Such differences reflect the
uncertainty in many of the values of the relevant equilibrium constants and the choice of
complexes considered. However, as Bondietti and Sweeton (1977) and Pourbaix (1966)
have stressed, the results of these calculations are only valid for the stated conditions. In
the formation of complexes with the 0H~ ion and other natural ligands, the effect of
insoluble compounds (such as phosphates) and the presence of natural reducing agents
must be considered.

12 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Similar Eh-pH diagrams can be constructed for americium (Pourbaix, 1966) which
indicate that the standard potential for Am(III) -^ Am(IV) is much greater than that for
Pu(III)^ Pu(IV). Hence the range of stability of Am(0H)4 moves to higher values of pH,
and it appears unlikely that Am(IV) can exist in solution under normal environmental
conditions.

The formation of complexes can also strongly affect oxidation-reduction potentials
in solution and, depending on the relative values of stability constants, can stabilize
different oxidation states in solution. Complex-ion formation in solution has been
extensively studied because of the need to understand the behavior of transuranic
elements in ion exchange, solvent extraction, and precipitation reactions. The general
tendency for complex formation depends on such factors as ionic radius and charge. The
order of stability constants is M^^ > MOa^ > M'^^ > MO2 , and for anions it is generally
COa" > oxalate^" > SOl^ for divalent ions and F~ > NO^ > Cr > ClO^ for mono-
valent ions. At relatively high concentrations of metal ions, hydrolysis reactions as acid
solutions are neutraHzed lead to the formation of low-molecular-weight hydrolytic
polymers, which can be described in some cases by simple equilibria, and higli-molecular-
weight polymers, which are not in equilibrium with the ions in solution.

Stability constants for some ligands present in natural waters have been summarized
(Rai and Seme, 1977). For oxidizing conditions at pH8, tlie conclusion was that the
dominant species are Pu02C03 0H^ and PuOj in the solution and solid phases,
respectively. Unfortunately, the values of the stability constants for many of the
hydroxo- and carbonato- complexes of plutonium, particularly Pu(IV), are not known,
and the values given in the Hterature are suspect (Cleveland, 1970; 1978).

Since the effective charges of the metal ions in UOl^ and Pu02^ ions are so similar,
the formation constants of complexes would be expected to be essentially the same for
each ligand. Woods, Mitchell, and Sullivan (1978) measured the stability constants of
complexes of PuO^^ with carbonate ions and found that at pH 8, where the HCO^ ion
predominates, there is a 1 : 1 complex (Pu : HCOf ) with a formation constant of about
4x 10^. At pH 11, where the CO^" predominates, there is a 1:3 complex, i.e.,
Pu02(C03)3". In contrast, Langmuir (1978) indicated that, for UOJ'^ in waters at pH 8
in equilibrium with partial pressures of CO2 much higher than atmospheric concentra-
tion, the predominant complex is U02(C03)3~. He also suggested that at pH <7.5
U02(HP04)2 ^ is the dominant species in natural waters with a phosphate concentration
of lO'^M.

Ahhough the effect of complexing in solution is to increase the total concentration
of metal, it is not clear if such reactions will make them more or less available for
bioaccumulation in the water column. Some of the smaller complexes may be readily
assimilated; larger ones may not. The known distribution of the transuranic elements in
the environment and their expected solubilities in the presence of particle surfaces
indicate that their biological availability also will be markedly influenced by their
chemistry/biochemistry in soils and sediments.

Terrestrial Ecosystems

Most studies indicate that plutonium is associated primarily with the solid phase in soils
and sediments (Tamura, 1976; Garland and Wildung, 1977; Edgington, Wahlgren, and
Marshall, 1976). Even in experiments where micromolar concentrations of Pu(N03)4 are
added to soil, the water-extractable and nonfilterable (<0.01 membrane filter) portion
exists principally as hydrous oxide particles with a diffusion coefficient of approximately

SYNTHESIS OF THE RESEARCH LITERATURE 13

10~^ (Garland and Wildung, 1977). Diffusion coefficients for total plutonium in soil are
on the order of 10"^ (Relyea and Brown, 1975), which indicates httle mobility "by this
mechanism. However, field studies have shown that plutonium may penetrate up to
30 cm in arid soil (Nyhan, Miera, and Neher, 1976) and that plutonium is more mobile
through biological (e.g., root uptake and transport) and physicochemical mechanisms
tlian would be predicted on the basis of diffusion alone. Furthermore, a fraction of
plutonium in •soils is readily dissolved. Studies of soils that had contained plutonium for
over 30 yr indicated that up to 13% of the plutonium was extractable with chelating
resins (Bondietti, Reynolds, and Shanks, 1976). This plutonium was probably available
for plant uptake and, in the case of perennials, may continue to be available for several
croppings, as demonstrated for clover (Romney, Mork, and Larson, 1970) and alfalfa
(Wildung et al., 1977). The chemical/biological phenomena controlling the quantity and
form of mobile plutonium are the key to predicting long-term implications of plutonium
in the environment.

Under aerobic conditions the ultimate behavior of plutonium in soils and sediments,
regardless of the chemical form of the source material, will be governed by processes that
influence hydrolysis and sorption on the solid phase and formation of soluble complexes
with organic or inorganic Ugands (Fig. 4). Initially, sorption and precipitation processes
predominate when Pu(IV) is added as the soluble nitrate and account for 98% of the total
plutonium a few hours after Pu(IV) has been added to a neutral silt loam soil (Wildung
and Garland, 1975). The addition of Pu(IV) as the DTPA complex results in nearly
100% plutonium solubility before a gradual reduction in solubility occurs by processes
described earlier. From a thermodynamic standpoint, the formation of Pu(V and VI) in
soil solution is theoretically possible. However, studies of the interactions of Pu(VI) with
organic ligands representing a range of common soil metaboUtes (Wildung, Garland, and
Cataldo, 1979), humic substances, and reducing sugars (Bondietti, Reynolds, and Shanks,
1976) suggest that Pu(VI) will be readily reduced to Pu(III) + (IV) in aerobic surface
soils. The presence of Fe(II), a reductant, may further promote the formation of reduced
plutonium under most soil conditions except highly alkaline soils.

Because Pu(IV) readily forms insoluble hydrolysis products, the interaction of these
species with mineral and organic surfaces results in the relative immobihty of plutonium
in soils and sediments. Hydrolysis products sorb on the sofid phase by mechanisms other
than ion exchange, and attempts to extract exchangeable plutonium from soils using
MgCl2 (Muller, 1978) and resins (Bondietti, Reynolds, and Shanks, 1976) resulted in the
removal of relatively small quantities (<13%) of the total plutonium. The major portion
of plutonium associated with the soUd phase in soils and sediments (Muller, 1978;
Edgington, Wahlgren, and Marshall, 1976) was extractable with citrate-dithionite, but
with citrate alone much less was extracted, which suggests the association of plutonium
with the reductant-soluble iron on the surfaces of soil/sediment particles (Wildung,
Schmidt, and Routson, 1977) or with iron in the original particles that were deposited.

The importance of hydrolysis in governing plutonium behavior extends to the soluble
fraction, at least over the short term (months). Almost all soluble and diffusible
plutonium on soil has been shown to be Pu(OH)n (Wildung et al., 1977). The
small quantity remaining in soil/sediment solutions is probably present as the Pu'*''" ion
stabilized against hydrolysis by interaction with a predominant anion (CO3" or HCO^,
depending on pH and ionic composition) and organic ligands (Wildung, Garland, and
Cataldo, 1979). Concentration of low-molecular-weight organic ligands, bicarbonate ion,
and carbonate ion are directly related to microbial metabolism and decomposition of

14 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

DIVERSE SOURCES OF PLUTONIUM

BIOLOGICAL TISSUES

SOIL

PARTICULATES

DEPOSITION

Pu(IV)L,,L2

Pu(IV)L,

TRANS-
LOCATION

ORGANIC

LIGANDS

IL,I \

Pu(IV)L,

CELL
MEMBRANE

Pu"*

TRANS-
PORT

DIFFUSION

EXUDATION

DECOMPOSTION

SOIL
SOLUTION

oh; HCO3. CO^"
ORGANIC LIGANDS

MICROBIAL ACTIVITY

SORPTION

DESORPTION

ORGANIC AND

INORGANIC SURFACE

REACTIONS

Fig. 4 Model for plutonium chemistry in the ingestion pathway. Regardless of the form
of plutonium entering soils, sediments, or water, plutonium is predominantly converted
(exceptions given in aquatic section) to Pu(IV), which is largely insoluble and associated
with the solid phase of soils and sediments. Soluble plutonium is also present primarily
as Pu(IV) stabilized by complexation with inorganic and organic ligands. The Pu(IV)
complexes are largely dissociated at the cell surface with Pu"* ^ ion transport across the
cell membrane. Mobility in biological tissues is facilitated by formation of secondary
complexes.

organic materials. There is direct evidence that plutonium forms complexes with
microbial metabolites and considerable indirect evidence supporting microbial influence
on plutonium solubility in soil (Wildung and Garland, 1977; Wildung, Gariand, and
Cataldo, 1979). The complexed Pu(IV) is probably the only plutonium that is available
for plant uptake (Fig. 4). The formation and the delivery of these complexes to roots are
the rate-limiting processes in the ingestion pathway.

Chemical properties of other transuranic elements (americium, curium, and neptu-
nium) in the environment have not been established. Laboratory studies have been
limited to studies of (1) the soil sorption of americium and neptunium, which indicate
sorption in the order Pu> Am> Np (Routson, Jansen, and Robinson, 1977); (2) the
sorption of Cm(III) and Np(V) on soil clay, which indicate sorption in the order
Cm> Np with an apparent association of neptunium with organic matter and amorphous
iron (Bondietti and Tainura, this volume); and (3) the solution behavior of ^^"^Cm in a
freshwater lake, which shows that soluble curium (50% of total) was largely anionic
(Dahlman, Bondietti, and Eyman, 1976). Field studies have been limited by relatively
low concentrations in the environment and lack of sensitive analytical methods for
certain nuclides of importance (e.g., ^'*''^'*^Am, ^^"^Cm, and ^^^Np). The aqueous
chemistry of these elements has been fairly well established (Katz and Seaborg, 1957;
Keller, 1971) and allows some predictions of behavior in soils and sediments. Major
differences in their environmental behavior as compared with plutonium would be
expected, and sorption on solid surfaces may be a function of the predominant valence
state and its tendency to hydrolyze (Dahlman, Bondietti, and Eyman, 1976). The only
stable ions of americium and curium in aqueous solutions are the trivalent cations. Their
chemistry in soUs and sediments is simOar if they are present in similar mass
concentrations. Hydrolysis reactions may be a primary factor governing the behavior of
americium and curium, but greater mobility and biological availability can be predicted
because of greater solubility of their hydroxides in comparison with Pu(0H)4. For
neptunium, Np02 is the most stable species in aqueous solution and should not be
subject to significant hydrolysis at environmental pH values (Burney and Harbour, 1974).
Of the transuranic elements, the environmental behavior of neptunium has been least
studied, but, because of its chemical characteristics, it is the most soluble in soils and may

SYNTHESIS OF THE RESEARCH LITERATURE 15

be the most available to biota. Plant uptake studies (Schreckhise and Cline, this volume)
and experimental feeding studies (Sullivan, 1979) indicate that this is likely.

Aquatic Systems

Freshwater Ecosystems. The behavior of plutonium and americium has been studied in a
wide range of freshwater systems (Table 5). Some contaminated areas have been small,
such as the U-pond on the Hanford Reservation in Washington (Emery and Klopfer,
1976), White Oak Lake in Tennessee (Dahlman, 1976), Rocky Flats ponds (Rees,
Cleveland, and Gottschall, 1978), and the ponds and canals at the Mound Laboratory in
Ohio (Bartelt et al., 1977). More extensive studies have been carried out in the Hudson
River in New York (Simpson, Trier, and Olsen, this volume) and in the Great Lakes
system (Wahlgren, Robbins, and Edgington, this volume; Bowen, 1976). The concentra-
tion of plutonium in these systems varied by more than four orders of magnitude
(Table 6). If concentrations are calculated assuming that the 239,240pjj j^ essentially all
^^^Pu, concentrations in water vary between 3x 10" ^^Af in the Great Lakes and
3 X 10~'^M in contaminated systems like U-pond on the Hanford reservation. These
concentrations are low relative to concentrations of many other trace elements. For
example, the concentration of thorium in Lake Michigan is <4 x 10~'^yif (0.1 fCi/liter)
(Wahlgren et al., 1977a).

Processes controlling the solubility of plutonium in natural waters clearly are more
complex than can be explained by a simple solubility product. For example, the
concentration measured in Lake Michigan is higher than that predicted for Pu(0H)4 and
lower than that for Pu02(0H)2 . These differences have been attributed to the formation
of hydroxyl complexes, such as Pu(OH)^ (Bondietti and Sweeton, 1977), carbonate
complexes, such as PUO2CO3 or Pu02(C03)2~ (Moskvin and Gel'man, 1958), or
PUO2CO3OH- (Rai and Serne, 1977).

A recent investigation of the reduction of Pu(lV) and (VI) by natural organic
compounds showed that up to 15% of Pu(IV) was reduced to Pu(III) at pH 4.0 and up to
75% of Pu(VI) was reduced to Pu(IV) by fulvic acid at pH 8. However, the Pu(VI) was
more stable in the presence of carbonate (Bondietti, Reynolds, and Shanks, 1976).

Wahlgren et al. (1977a) studied the behavior of plutonium in tlie water column in the
Great Lakes and other smaller freshwater lakes to determine whether differences in
chemical characteristics of lake water affect the chemical properties of plutonium
(Table 7). The concentration of plutonium in waters of these lakes varied almost
100-fold. The highest concentrations of plutonium were observed in the lakes (ELA 241
and ELA 661) with low pH, a lake with a very high concentration of sulfate (Little
Manitou), and tlie acidic southeastern United States lakes.

Using techniques to separate Pu(III) + Pu(IV) from Pu(V) + Pu(VI), Nelson and
Lovett (1978) showed that in the Irish Sea plutonium was predominantly in the
Pu(V) + Pu(VI) states. A similar observation was made in Lake Michigan waters (Wahlgren
et al., 1977b). Because Pu(V) was thought to disproportionate at lower pH than does
Pu(VI) (Pourbaix, 1966), this fraction was referred to as Pu(VI).* In all other lake waters,
Pu(lII) + (IV) apparently predominated.

*Very recent experiments at Argonne National Laboratory have shown that this assumption is not
correct (D. M. Nelson and K. A. Orlandini, Argonne National Laboratory, 1979, personal communica-
tion). Techniques have been developed after the method of Inoui and Tochiyama (1977) for
separation of Np(V) from Np(VI) to distinguish Pu(V) from Pu(VI) in water samples. Preliminary
results indicate that all the plutonium in the higher oxidation state is present as Pu(V).

1 6 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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18 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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SYNTHESIS OF THE RESEARCH LITERATURE 19

The Eh— pH diagram (Fig. 3) shows that under normal environmental conditions
Pu(III) and Pu(VI) can coexist and that the ratio of the two states will depend on the
oxidizing conditions and pH in the system. Therefore the relatively high concentrations
of plutonium in ELA lakes (other than 885) and lakes in the southeastern United States
could be due to the reduction of Pu(IV) to Pu(III) as well as to complexation of Pu(IV)
by organic ligands. The high concentration of plutonium and the very low fraction of
Pu(VT) in Little Manitou Lake, wliich contains high sulfate concentrations, could also be
due to the formation of sulfate complexes, which stabilize the Pu(IV) state.

Similar measurements have evaluated the relative concentration of Pu(III + IV) and
Pu(V + VI) in White Oak Lake water (Bondietti and Sweeton, 1977) and indicated that
Pu(IV) was the dominant oxidation state present. Plutonium(IV) rather than Pu(III) was
suggested because another study indicated that Pu(III) was unstable toward oxidation to
Pu(IV) at pH >5 (Bondietti, 1977).

Charge characteristics of plutonium in Lake Michigan water indicate that the element
is not associated with colloidal matter in the size range 0.003 < x < 0.45 /jm and that it is
almost quantitatively absorbed by anion exchange resins. In water samples from acidic
lakes, the majority of the plutonium behaves like cationic or uncharged species. These
results and the differences in oxidation state discussed earlier strongly suggest that the
solubility of plutonium is governed by different complexing agents. In waters of high pH,
the concentration of COl" and HCOi" ions is relatively high, and carbonate complexes
can form. In waters of low pH. such complexes cannot exist, and the solubility must be
due to complexing with other ligands, such as natural organic compounds.

In addition to measuring concentrations in water columns, most investigators have
measured the concentration of plutonium in surficial sediments. In a few cases
measurements have been made of plutonium in suspended particulate material. Table 6
sliows that there is some relationship between concentrations in water and concentrations
in surface sediments.

If there is mixing of the surficial sediment with the water column and a true
equilibrium between the water and particulate matter or sediment, the distribution
constant, Kp, for the reaction between filtered water (<0.45 jum) and sediment is

_ Concentration per kilogram of sediment
Concentration per Uter of water

Values of Kq vary from lO'* to 5 x 10^, but most values do not vary more than
fivefold. Considering the wide variety of sediment types involved and the differences in
source terms and sizes of aquatic environments, the small range in values strongly suggests
a commonahty in the behavior of plutonium in these systems.

Sediment characteristics affect the uptake of radionuclides, and fivefold to tenfold
variations in distribution coefficients can be explained solely in terms of differences in
distributions of particle sizes in sediments (Duursma and Bosch, 1970). Little information
is given on sediment characteristics, but sediments from small ponds and rivers probably
are generally coarser than those from deep waters of the Great Lakes.

More recent experiments have shown the distribution of plutonium between solution
and solid phases to be a true equilibrium. Sediments labeled with ^^^Pu from the Miami
River were equilibrated with Lake Michigan water. Kinetic studies indicated that
equilibrium was reached in 1 day or less. Moreover, the ratio of oxidation states in water
from this experiment is the same as that observed for ^^'Pu in Lake Michigan (D. M.

20 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Nelson and D. N. Edgington, Argonne National Laboratory, personal communication). A
few measurements have been made of the oxidation state of plutonium adsorbed on
particulate matter (Nelson and Lovett, 1978). Samples of surface sediment from the
Great Lakes and Miami River indicate that plutonium absorbed by sediment particles is
predominantly in the (III) and (IV) states. On reequilibration of sediment with water, it
has been shown that there is a conversion of Pu(III) + Pu(IV) back to Pu(V) or (VI)
(D. M. Nelson, Argonne National Laboratory, personal communication).

TABLE 8 Values of the Distribution
Coefficients (Kp ) and Concentration Ratios
(CR) for Phytoplankton for the Actinide and

Lanthanide Elements in Lake Michigan

Element

LogKD*

LogCRf

La(III)

5.2

3.0

Th(IV)

>6.5

U(VI)

-4.0

2.2

Plutonium

5.5

4.0

Pu(IV)

6.5

4.8

Pu(VI)

4.1

-4.2

Americium

>6.0

-4.2

*Values taken frc

3m Wahlgren

etal.,1976.

xpp _ pCi/kg wet tissue

pCi/kg water

Therefore the concentration of plutonium in many freshwater lakes and streams
apparently is controlled by an equilibrium between water and sediment. From the data it
is possible to calculate values of Kq for Pu(III) and Pu(IV) and Pu(VI). These values are
given in Table 8 and are compared with those for stable-element homologs, such as
La(III), Th(IV), and U(VI). As would be expected, results for ^^*Pu from U-pond and
ponds and canals in Miamisburg (Table 6) show little or no difference in behavior of the
plutonium owing to isotopic composition.

Finally, leaching experiments with sediments have shown that a major fraction of
fallout plutonium can be removed with extractants, such as dilute acids or complexing
agents (Alberts, Muller, and Orlandini, 1976). Furthermore, studies of plutonium in
natural waters have shown low but measurable concentrations of plutonium in true
solution. In Lake Michigan the plutonium concentration was essentially constant over the
whole lake (Wahlgren and Nelson, 1975). The measured concentration of 0.5 fCi/liter of
239,240pjj corresponds to a chemical concentration of 3xlO"^'^A/, or 20,000
atoms/ml. Hence molecular collision theory implies that the formation of polymeric
plutonium species in the lake (i.e., many plutonium atoms linked as — Pu— Pu or as
— Pu— 0— Pu) is unlikely. Even the possibility of the formation of dimers is vanishingly
small.

Marine Ecosystems. Studies of transuranic elements in marine and estuarine ecosystems
have encompassed a wide range of sources: worldwide fallout as a result of the testing of
nuclear weapons, lagoons in tropical atolls where many tests were performed, and direct

SYNTHESIS OF THE RESEARCH LITERATURE 21

discharges from nuclear processing plants into coastal zones. The study areas and
characteristics of the sources are given in Table 9.

Plutonium has been measured in samples of ocean waters collected since 1963,
shortly after the peak of activity in testing weapons. The concentration has decreased
from 2 to 3 fCi/hter in samples of water from the northeastern Pacific in 1964 (Pillai,
Smith, and Folsom, 1964) to about 0.2 to 0.9 fCi/liter in samples collected between 1968
and 1973 (Miyake and Sugimura, 1976). The major point in studying plutonium in the
water column of oceans is to use variations in the concentration of this element to
explain movements of water and pollutants. To this end, comparisons have been made of
the movement of plutonium relative to ^"Sr and ^^ ^Cs in the Pacific Ocean and Atlantic
Ocean (Bowen, Wong, and Noshkin, 1971; Miyake and Sugimura, 1976). The Pu/^"Sr
and Pu/^'^^Cs ratios in surface seawater are far lower tlian those found on land, which
indicates that the residence time of plutonium in the water column is less than tliat of
^^ ''Cs and ^°Sr. As early as 1968, from 10 to 20% of the total plutonium deposited over
the ocean was in deep-sea sediments at water depths of about 4000 m. The depletion of
plutonium from surface waters may be modeled in terms of settling rates of particulate
matter in the water column. The observed distribution of total plutonium in the water
column to 5000 m is explained in terms of a distribution of particles with the majority
settling at an average velocity of 195 m/yr (Bowen, Wong, and Noshkin, 1971).

Another series of water samples was collected from the Pacific Ocean in 1973 as a
part of the Geosecs Program. Analyses of these samples showed that there is a maximum
in the concentration of plutonium at a depth of 300 to 700 m across the Pacific Ocean
and that the concentration of ^^^Pu in this stratum has not changed by more than 20%
over a period of 5 yr (Bowen, 1977). This behavior can be explained by assuming a rapid
transfer of plutonium to about 400 m by biogenic debris (e.g., fecal pellets) (Beasley and
Cross, this volume) where the plutonium returns to a soluble species that can migrate
upward or downward by diffusion. Experiments in the Irish Sea have shown that
plutonium is in solution predominantly as Pu(VI) and on particles as Pu(IIl) + Pu(IV)
(Nelson and Lovett, 1978). The rapid movement of plutonium to 700 m may be
associated with particles that sink to that depth where they dissolve and the plutonium
reoxidizes to Pu(VI). The higher ^ '^ ^ Am/^ 3 9 ,2 4 op^ ^^^^^ ^^^^ ^ qqq ^ j-glative to that in
surface waters supports this hypothesis. Since the Kd for Pu(VI) is about 1000 times
lower than that for Pu(IIl), Pu(IV), or by inference Am(III), any ^^^ Am that is released
would be preferentially taken up by any remaining particles.

Plutonium in oceans occurs in solution over a wide range of concentrations
(Table 10). A more surprising result is that distribution coefficients between water and
suspended sediments are very similar to those in the Great Lakes and elsewhere.

The distribution of ^^^'^"^^Pu and ^^^Pu in waters and sediments of Enewetak Atoll
has been studied in detail (Noshkin, tliis volume). In 1976 the total inventory of
2 3 9,240p^ in water and sediments was 1.24 and 249 Ci, respectively. The Kq for
plutonium in these sediments has been independently measured in the laboratory as
1.8 X 10^. A simple model can be constructed to predict the average concentration of
plutonium in the lagoon by assuming this equilibrium constant. This model predicts the
concentration of plutonium to be 32 fCi/liter. The average concentration measured in
1976 was 16 fCi/Uter. Furthermore, there is no indication of preferential dissolution of
^^^Pu in this lagoon because the isotopic ratios of ^•^^'^'*^Pu and ^^^Pu are identical in
water and sediments. Similar Kq values for sediments from the Irish Sea and Enewetak
Atoll suggest that similar chemical reactions are occurring in all oceans. These results are

22 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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SYNTHESIS OF THE RESEARCH LITERATURE 23

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24 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

in marked contrast to the data for Bikini Atoll (Schell, Lowman, and Marshall, this
volume) where preferential mobilization of ■^^^Pu was inferred from increased ratios of
^^^Pu to ^^^'■^''Opu between nonfilterable and filterable fractions from lagoon water.
These findings were postulated to result from recoil damage from high-specific-activity
^^^Pu, which caused increased solubility. However, such an explanation must assume
that 2 3 9,2 4 0p^^j ^^^ ^^^Pu are in separate particles originally.

There is Uttle evidence to suggest that differing sources of transuranic elements affect
tlieir chemical properties when the elements are moderately well dispersed in aquatic
systems. Transuranic elements are soluble, to a limited extent, in both freshwater and
marine systems and are therefore available for transfer across biological membranes.
Plutonium apparently behaves similarly in oceans and in the Great Lakes, as shown by
values of Kq and chemical speciation. These systems can be considered oUgotrophic with
their chemical properties largely controlled by their respective carbonate cycles. Hence
the similarities in values of Kq and ratios of Pu(VI)/Pu(IV) are expected.

Because tire pH of the ocean is well buffered, plutonium apparently cannot exist
except as Pu(III) or Pu(VI) in solution in the water column or as Pu(IV) in sediments if
the relationships shown in Fig. 4 hold. However, in freshwater lakes large variations in
composition are possible, and the pH can be relatively low (about 4). Under these
conditions dramatic changes in concentrations of plutonium are observed and can be
explained by the presence of Pu(in) or Pu(IV) as complexes.

Environmental studies show the danger of using the results of laboratory experiments
with moderately concentrated solutions (\0~^M) to predict the behavior of plutonium in
tlie environment, where the maximum observed concentration has not exceeded
10"' ■^M. Somewhere within the concentration range of 10"'^ to 10" ^M, plutonium
ceases to exhibit the properties of simple ions, and tlie possible formation of polymeric
species must be considered.

Transport

Terrestrial

Most environmental plutonium exists in a strongly adsorbed state on surface soils. Hence
most investigators have concluded that the transport of this element, at least over the last
30 yr, has been governed by processes regulating the distribution and transport of soil
(Essington et al., 1976; Hakonson, 1975; Hakonson, Nyhan, and Purtymun, 1976;
Hakonson and Nyhan, this volume; Hayes and Horton, this volume; Romney and Wallace,
1977; and Sprugel and Bartelt, 1978). In natural systems soil-erosion processes are mainly
driven by wind and water.

Wind Erosion. Wind transport of plutonium in soil can be documented anywhere that
appropriate soil, vegetation, and climatic conditions exist. These conditions exist when
soil is loose, dry, and of optimum particle size; the soil surface is relatively smooth;
vegetation cover is sparse; and winds are sufficiently strong to initiate soil movement
(Beasley, 1972).

Wind redistributes plutonium in soil, as inferred from samphng of contaminated sites
(Little, 1976; this volume; Markham, Puphal, and Filer, 1978; Romney and Wallace,
1977) and from studies focused specifically on wind transport of plutonium (Gallegos,
1978; Sehmel, 1978; Anspaugh, Shinn, and Wilson, 1974; Anspaugli, Shinn, and Phelps,
1974; 1975). These observations and field studies, primarily in arid regions, imply that

SYNTHtJJS OF THE RESEARCH LITERATURE 25

wind transport of soil is highly seasonal and is relatively more important in dry, sparsely
vegetated areas than in mesic, heavily vegetated areas.

In the arid western United States, wind erosion of soil occurs primarily in the spring
and late summer months, coinciding with periods of high wind and low surface soil
moisture. Studies in the humid soutlieast United States suggest that wind is a minor cause
of transport of plutonium in soil (Dahlman, Bondietti, and Eyman, 1976) because of the
low incidence of high winds and the heavy cover of vegetation.

Soil particle sizes and plutonium concentrations in soil affect the importance of wind
as a plutonium transport vector. Plutonium concentrations of various soil size fractions
can differ by several orders of magnitude and, depending on source characteristics, are
generally highest in the smaller size fractions (Nyhan, Miera, and Neher, 1976; Nyhan,
Miera. and Peters, 1976; Tamura, 1975; Little and Whicker, 1977). Furthermore, wind
preferentially moves certain sizes of soil particles, depending on the physical characteris-
tics of soil, the wind speed, and the soil moisture (Beasley, 1972). The relationship of
some of these factors to plutonium transport by wind is illustrated for a 1-month
sampling period at two locations in the fallout zone at Trinity Site, New Mexico, in
Table 1 1. Within 1 km of ground zero, very Uttle of the plutonium activity was present in

TABLE 1 1 Mass and Plutonium Content of Dust and Soil

Samples from Two Locations in the Trinity Fallout

Zone for the Period 7-14-76 to 8-10-76

Sample
mass, g

Dust*

Soil

Distance from
crater, km

% mass
<53 jum

% Pu in
<53Mm

% mass
<53 \i.m

%Pu in

<53 iim

1
45

4.02
2.10

7
54

2
45

9
36

0.8
73

*Saltated dust collected in the zone 0 to 15 cm above the ground
surface with accumulative Bagnold dust sampler.

the silt— clay (<53 /jm) size fraction of dust or soil samples. However, about 45 km from
the crater along the fallout pathway, a much higher percentage of the plutonium in dust
and soil samples was present in this size fraction. These differences demonstrate the
potential importance of the relationship of soil particle sizes to plutonium concentration
in understanding plutonium transport within ecosystems. Sih-clay particles may be
transported farther and are more likely to remain attached to biological surfaces than are
larger size particles (Romney and Wallace, 1977; Romney et al., 1963; Little, 1976; this
volume).

Plutonium suspended by wind can be redeposited on soil or intercepted by biological
surfaces. Redeposition of plutonium on soils can lead to major changes in the distribution
of the element within an ecosystem, as shown by work at the Nevada Test Site (Romney
et al., 1963; Essington et al., 1976). These studies showed that plutonium associated with
blown sand accumulates around the bases of shrubs where many of the desert life
processes function (Romney and Wallace, 1977). Our understanding of soil plutonium in
otlier areas and cUmates is Umited. However, the accumulation of plutonium around
vegetation clumps (or other natural or man-made obstacles) may be common to all
regions where wind is a major soil-erosion agent.

26 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

The deposition of plutonium on biological surfaces can be inferred directly from
concentration ratios

_ Transuranic concentration in receptor
Transuranic concentration in donor

based on field data (Hakonson, 1975; Little, 1976; this volume; Romney and Wallace,
1977; Dalilman, Bondietti, and Eyman, 1976). Such ratios are much higher than those
derived from greenliouse studies (Francis, 1973; Price, 1973; Schulz, 1977) and imply
that root uptake cannot account for concentrations measured in field samples. Physical
processes are evidently more important than chemical processes in transporting
plutonium to vegetation.

Wind is apparently more important in contaminating vegetation in dry regions than it
is in humid regions, as shown by plutonium CR's. These ratios decrease from about 10~*
in United States deserts (Hakonson, 1975; Little, 1976; Romney and Wallace, 1977) to
10" ■^ in mesic ecosystems of the southeast (Dalilman, Bondietti, and Eyman, 1976).

Additional observations implicate wind-driven processes in contaminating vegetation
with plutonium. For example, plutonium concentrations are inversely correlated with
height of plants above tlie ground (Hakonson and Johnson, 1974; Dahlm an, Garten, and
Hakonson, this volume). Tlius low-growth forms, such as grasses, forbs, lichens, and
mosses, generally' exhibit higlier plutonium concentrations than shrubs or trees. This
pattern is consistent with soil flux— heiglit relationships, which show that most of the soil
mass transported by wind is within 1 m of the ground surface (Selimel, 1978; Gillette,
Blifford, and Fenster, 1972; Phelps and Anspauglr, 1977).

Water Erosion. Physical transport of transuranic elements by raindrop splash or surface
runoff has received Uttle attention in terrestrial ecosystems, althougli these processes
certainly occur (Romney and Wallace, 1977; Hakonson, Nyhan, and Purtymun, 1976;
Sprugel and Bartelt, 1978; Muller, Sprugel, and Kohn, 1978). For example, Beasley
(1972) has shown that a 5-cm rainstorm causes disaggregation of 200 metric tons of soil
per hectare by raindrop splash and surface-water runoff. The importance of soil splash up
from raindrops in contaminating vegetation is unknown, although the process certainly
occurs.

In certain cases (e.g., intermittent streams) water movement of sediments may be the
dominant mechanism of plutonium transport (Hakonson, Nyhan, and Purtymun, 1976).
The process is primarily the physical transport of plutonium sorbed on soil particles
rather than movement of dissolved plutonium (Hakonson, Nyhan, and Purtymun, 1976;
MuUer, Sprugel, and Kohn, 1978).

The relationship of plutonium concentration to soil particle size is also important in
assessing transport because water movement preferentially sorts soil according to particle
size (Hakonson, Nyhan, and Purtymun, 1976; Muller, Sprugel, and Kohn, 1978). For
example, as water velocity decreases, successively smaller soil size fractions remain in
suspension. Hence silt— clay fractions, wltich usually contain higher concentrations of
plutonium, are probably carried greater distances than larger size fractions.

Water transport of soil across landscapes redeposits plutonium within local watershed
soils and stream channel sediments (Hakonson, Nyhan, and Purtymun, 1976), down-
stream ponds (Muller, Sprugel, and Kohn, 1978), rivers (Hayes and Horton, this volume;
Sprugel and Bartelt, 1978), lakes, and oceans. Studies of intermittent streams at Los

SYNTHESIS OF THE RESEARCH LITERATURE 21

Alamos showed that stream-bank soils are a repository of effluent plutonium and serve as
a source of the element to stream-bank biota (Hakonson et al., 1979).

It has been estimated that rivers contribute about 150 to 500 Ci/yr of ^^^Pu to
oceans (Simpson, Trier, and Olsen, this volume).

Studies of the Savannah River showed tliat about 0.005% of the total plutonium in
the watershed is lost to the coastal zone annually (Hayes and Horton, this volume). In
contrast, the annual loss for a typical midwestern river, the Miami (Muller, Sprugel, and
Kohn, 1978; Sprugel and Bartelt, 1978), and for the Hudson (Simpson, Trier, and Olsen,
this volume) is an order of magnitude greater. These results indicate a residence time of
between 10^ and 2 x 10"* yr. In each river the majority of the plutonium is transported
with suspended matter. The differences in loss rates are probably related to differences in
watershed morphology. A major fraction of the plutonium transported in the Savannah
River is probably held up in impoundments at the upper reaches.

Biotic Activity and Mechanical Disturbance. Plutonium concentrations in animal tissues
demonstrate the dominance of physical processes in transporting plutonium to animals in
natural ecosystems. In addition to the gastrointestinal tract, highest concentrations of
plutonium are associated with the pelt and, to a lesser degree, lung tissue as a result of
interaction with plutonium on soil particles (Hakonson, 1975; Little, 1976; this volume;
Hakonson and Nyhan, this volume; Bradley, Moor, and Naegle, 1977).

Work at the Nevada Test Site with cattle (Smith, 1977) shows that considerable
amounts of soil are routinely ingested by grazing herbivores. Cattle ingest several hundred
grams of soil daily under normal range conditions. Transport of plutonium occurs when
these animals move to other areas with subsequent deposition through defecation and/or
death of the animal (Arthur and Alldredge, 1979). The amount of plutonium transported
in this manner is considered small. Foraging by herbivores, such as cattle, deer, rodents,
and insects, may subject a substantial amount of the plutonium in soil to digestion
processes over prolonged grazing histories. Whether the chemical form of ingested
plutonium is altered as it passes through the gastrointestinal tracts is not known, but in
vitro studies indicate changes in solubility in an artificial rumen, simulated abomasal, and
intestinal fluid procedure (Earth, 1977).

Mechanical disturbances, such as soil tilling and construction, can transport large
amounts of plutonium on a local scale. Plowing enhances mixing of plutonium with the
soil profile and also can cause large increases in airborne soil particles. Soil tilling activities
at the Savannali River Plant increased local air concentration of plutonium 100-fold
(Milham et al., 1976). Mechanical harvesting of agricultural crops also results in surface
contamination of edible grains (Adriano et al., 1975).

Existing data pertaining to plutonium distribution in natural ecosystems suggest that
physical processes driven by wind will become less important as plutonium migrates into
the soil profile. Contemporary data from fallout areas contaminated in 1945 (Hakonson
and Nyhan, this volume) show tliat less than 50% of the soil column inventories of
plutonium occurs in the surface 2.5 cm of soil. Similar relationships have been observed
in a Los Alamos intermittent stream initially contaminated in 1963 (Hakonson et al.,
1979) and in the grassland study site at Rocky Flats (Little, this volume). A change in
physical transport of plutonium would probably change the relative importance of
chemical and biological transport processes. For example, long-term cropping studies of
Romney and Davis (1972) and Schreckhise and Chne (this volume) suggest that migration
of plutonium into soil may create conditions more favorable for uptake by deeper rooted
plants.

28 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Plant Uptake. Transuranic elements in terrestrial environments can enter plants by foliar
absoiption and root uptake. The route o^ entry into plants will depend on the nature of
the source; climatic conditions affecting ueposition, retention, and chemistry of particles
on leaf surfaces; the foliar surface area exposed; and soil conditions affecting resuspension
and solubility.

The root is the major ion-absorbing organ of the plant. Somatic cells in the leaf
possess the same potential for absorption; however, they are protected by a waxy cuticle.
Foliar absorption is an efficient route of entry for nutrients (Bukovac and Wittwer, 1957;
Wittwer, Bukovac, and Tukey, 1963), fission products, and activation products (Tukey,
Wittwer, and Bukovac, 1961; Athalye and Mistiy, 1972). Foliar absorption of ^^^Pu and
^^ ^ Am can occur and is dependent on chemical form and environmental conditions with
up to 10^^ and 10^^ of the foliar deposits absorbed and translocated to seeds and roots
(Cataldo and Vaughan, this volume; Cataldo, Garland, and Wildung, 1978). About one
one-millionth of the plutonium applied as oxide was absorbed by leaves; availability was
dependent on particle size. The availability of americium applied as the oxide was two to
five times as great as that of the less soluble plutonium oxide at comparable particle size.
Thus foHar uptake appeared to be related to transuranic solubihty.

Ion uptake by plant roots is apparently a metaboUcally mediated process in which
ions are transported across the cell membrane. The process is concentration dependent
over a broad range (10"^ to \0~^M), exhibits a degree of ion selectivity, and may allow
for accumulation of ions against a concentration gradient (Nissen, 1973). Although the
transport process is selective, plants accumulate nonnutrient ions. Processes leading to the
deliveiy of soluble transuranic species to root membranes have been described. It is
critical to determine if discrimination occurs at the membrane level because this would
limit transuranic uptake by plants and incorporation into food chains.

Because of the relatively low uptake of plutonium and americium from soil by plants
(CR values of 10"^ to 10~^), it has been generally assumed that marked discrimination
occurs. Evidence is increasing that solubility in soil rather than plant discrimination at die
membrane level limits transuranic uptake by plants. As expected from their respective
aqueous chemistries, transuranic elements are sorbed by soil in the order
Pu > Am ~ Cm > Np (Table 12). Uptake of these elements is apparently inversely related
to soil sorption. The addition of complexing agents, wliich markedly increases transuranic
solubility in soil (Wildung and Garland, 1975), also increases plant uptake (10- to
10,000-fold) (Energy Research and Development Administration, 1976). Thus indirect
evidence supports soil solubility as the primaiy factor governing transuranic availability to
plants.

Experiments with plants grown in hydroponic solutions containing plutonium aid in
distinguishing soil sorption and plant root discrimination when uptake is compared with
uptake by plants grown in soils. Wlien hydroponically grown soybeans (Glycine max)
were placed in /.iM^^^Pu-DTPA solutions and permitted to accumulate plutonium for up
to 49 hr (Wildung et al., 1977), CR's [(^iCi/g diy plant) per (juCi/ml nutrient solution)]
for shoot tissues were 6 X 10"^ and 3 x 10"' after 1 and 24 hr, respectively. The
Pu— DTPA complex supplied in the growth medium was not detected in the exudates.
Similarly, leaves of bush beans (Phaseolus vulgaris) exhibited CR's in nutrient solution of
0.8 and 5.1 for Pu(lV) and Pu(VI), respectively. Thus plants can" accumulate soluble
plutonium effectively; much of the apparent discrimination found in soil-plant studies
resuhs from the effect of soil sorption in reducing the quantity of soluble plutonium
available to the plant.

SYNTHESIS OF THE RESEARCH LITERATURE 29

TABLE 1 2 Distribution Coefficients (Kd ) for Soil

Sorption and Relative Plant Uptake of the

Transuranic Elements

Relative
Element Log Kj)* uptakef

Plutonium(IV)

4.0

Americium(III)

Curium(III)

Neptunium(V)

3x lO'*

*Plutonium (Prout, 195 8); americium and neptunium
(Routson, Jansen, and Robinson, 1977); curium (Routson,
1978, personal communication).

t^Ci/g barley per mCi/container (~0.1 mCi/container);
transuranic element may not have been uniformly
distributed in container (Schreckhise and Cline, this
volume).

Plutonium is probably transported across biological membranes in the Pu(IV) state,
particularly in plant roots (Fig. 4). Plutonium(IV) has been identified in the plant xylem
of plants grown in a solution containing predominantly Pu(VI) (Delaney and Francis,
1978).

Once in the root plutonium is probably translocated downward in the root and
upward in the xylem stream to shoot tissues as Pu(IV). Simple organic acids typical of
microbial and plant metabolites quantitatively reduce Pu(VI) to Pu(IV) (Wildung et al.,
1977). Plutonium(IV) dominates in the plant xylem regardless of the oxidation state
supplied in nutrient solution (Delaney and Francis, 1978). The low solubility of Pu(IV)
limits translocation in plants unless complexed, and several anionic and cationic
complexes of plutonium have been identified in the xylem stream of plants supphed
Pu(IV) and Pu(VI) (Wildung et al, 1977). During growth a fraction of tlie plutonium is
lost from the root with other inorganic and organic exudates and by decomposition of
slouglied cells (Fig. 4). The plutonium associated with this material in the rhizosphere
may be subject torecychng into the plant, subsequent modification, leaching, and
diffusion.

Translocation in plants can also serve as a primary factor governing entrance of
transuranic nuclides into foodstuffs. Plutonium was mobile in barley and soybeans but
was not unifomily distributed in the plant (Garland et al., 1974). In general,
concentrations of plutonium in the leaves of soybeans were 5 to 10 times as high as those
in stems. The lowest plutonium concentrations were observed in barley and soybean
seeds, which minimized the amount of plutonium ingested with these edible tissues.

Animal Uptake. The gut absorption of plutonium by mammals requires the presence of
soluble forms, and solubility is governed by chemical reactions similar to those previously
discussed (hydrolysis and complexation). When large amounts of plutonium (>1 mg) are
introduced into the gut as Pu(VI) in the absence of foodstuffs (starved animals) and in
the presence of large excesses of a holding oxidant, Pu(VI) is absorbed in significant
quantities (Weeks et al., 1956). In Chicago plutonium is present largely as Pu(VI) in
chlorinated drinking water (Larsen and Oldham, 1978). However, the reducing potential
in the gut seems to be sufficient to reduce very low concentrations of Pu(VI) to Pu(lV),
which would limit uptake in the absence of the holding oxidant. This conversion would

30 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

be particularly pronounced if reducing substances, such as food residues, were present
(Sullivan et al., 1979). In contrast, when plutonium is present as Pu(IV) complexes, such
as in microbial and plant tissues, preliminary studies indicate that gut absorption is
somewhat increased (Sullivan and Garland, 1977; Ballou et al., 1978). This probably
occurs because of increased Pu(IV) solubility in the gut and transport across the gut wall
as Pu(IV) ion or a low-molecular-weight plutonium complex. For elements that are not
readily hydrolyzed in the pH range of the digestive tract, e.g., Np(V), incorporation in
plant tissues may reduce gut transport relative to direct absorption from administered
solutions (Sullivan, 1979).

The absorption of transuranic elements other than plutonium, under similar
conditions in the gut, may be related to solubility following principles outlined in the
previous section. If this is true, gut absorption in the presence of foodstuffs will follow
the order Np(V) > Cm(III) ~ Am(III) > Pu(IV) ^ Pu(VI).

Gut absorption has not been studied in terrestrial invertebrates; thus comparisons
with the observations reported for marked uptake in marine invertebrates (vide infra)
cannot be made.

Terrestrial Food Webs. The CR is used to assess the degree of bioaccumulation. Extensive
reviews of CR data based on greenhouse and field studies are presented elsewhere
(Francis, 1973; Schulz, 1977; Energy Research and Development Administration, 1976;
Price, 1973). Table 13 summarizes transuranic-element CR's based on laboratory studies,
and Table 14 summarizes those based on field studies. These transuranic elements are

TABLE 13 Transuranic-Element Concentration Ratios
Based on Experimental Studies

Element Agricultural crops Native plants Reference

238,239,24opy ^q_,o_^q-3 ^q-s_^q-a FraHcls, 1 97 3 ; Schulz, 1977 i

Brown, 1976; Price, 1973
^'"Am lO-'-lO' 10-= -10-' Francis, 1973; Schulz, 1977;

Price, 1973
'""•Cm lO-^-lO-' Schieckhise and Cline, this

volume; Price, 1972; 1973
'^'Np 10-' -10-' Schieckhise and Cline, this

volume; Price, 1972; 1973

TABLE 14 Transuranic-Element Concentration Ratios
Based on Field Studies

Element Agricultural crops Native plants Native animals* Reference

'^*Pu 10-''-10° Hakonson and Nyhan, this volume;

Dalilman, Bondietti, and Eyman,

1976
2 3 9,24opy 10"= -IQ-' 10-" -10° 10-" -10-' Hakonson and Nyhan, this volume;

Dalilman, Bondietti, and Eyman,

1976; Little, 1976; Durbin, 1975
'"'Am 10-^-10" 10-^ -lO*" Durbin, 1975

*Based on whole-body burdens.

SYNTHESIS OF THE RESEARCH LITERATURE 31

generally not concentrated (i.e., CR < 1) by terrestrial plants and animals. On the basis of
laboratory studies, neptunium may be an exception (Schreckhise and Cline, this volume;
Price, 1972). However, there are no data from which to judge the behavior of this
element under field conditions, particularly in acid soils where Fe(II) would reduce
Np(V).

Tables 13 and 14 show tliat CR's based on greenliouse studies are much lower than
those derived from field data. The higher CR's based on field data are likely due to
surficial contamination of plants with small soil particles, whereas CR's based on
greenhouse studies generally reflect only root uptake.

Agricultural plant species accumulate transuranic elements to about the same degree
as native plants. The concentrations of transuranic elements in fruits and grains are 10"^
to 10~ ^ times lower than those in vegetative parts (Schulz, 1977).

Field data from a number of study sites containing up to several hundred picocuries
per gram of soil show that plutonium transfer to native and domestic animals is also very
small (Little, 1976; this volume; Hakonson and Nyhan, tliis volume; Bradley, Moor, and
Naegle, 1977; Smith, 1977). Concentration ratios in internal tissues of rodents are
comparable to those observed in internal tissues of vegetation. Concentrations of
plutonium in internal tissues (i.e., liver, muscle, and bone) can seldom be measured owing
to the low gut availability of this element (Durbin, 1975).

Aquatic Food Webs. Transuranic elements can enter aquatic environments at a number
of points in complicated food chains encompassing all trophic levels from microbes to
vertebrates. Summaries of trophic-level studies in freshwater and marine environments
(Dahlman, Bondietti, and Eyman, 1976; Hetherington at al., 1976) indicate that
plutonium CR's relative to water generally decrease at higher trophic levels.

Marine benthic invertebrates and invertebrate predators feeding on them exhibit tlie
highest levels of plutonium in coastal fauna (Noshkin, 1972; Pillai and Mathew, 1976).
Although these observations generally correlated with the high fraction of the plutonium
inventory found in sediments, experimental studies show that marine invertebrates have
remarkably hi^ assimilation efficiencies relative to terrestrial mammals (Beasley and
Cross, 1979).

No clear correlation between sources of transuranic elements and marine fish
concentrations can be made at this time because of limited data. The evidence from both
field and experimental studies shows variations in uptake which can be attributed to tlie
element under study, the chemical species, and the type of fish. Studies with ^^ ''Pu show
that plaice can absorb plutonium as Pu(VI) by direct uptake from seawater, but
absorption across the gut from labeled food or sediment is very low (Pentreath, 1978a).
Elasmobranch fish, such as the thornback ray, however, do appear to absorb plutonium
across the gut wall relatively easily (Pentreath, 1978b). Environmental observations
indicate that americium is relatively more available to plaice than is plutonium (Pentreath
and Lovett, 1978).

Except for high CR's for plutonium in phytoplankton relative to water, which appear
to result from a surface-adsorption phenomenon (Beasley and Cross, tliis volume), and
observations of a fourfold increase in the concentrations of plutonium in starfish relative
to those in the mussels on which they feed (Noshkin et al., 1971), no apparent
biomagnification has been observed in aquatic systems (Dahlman, Bondietti, and Eyman,
1976).

52 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Further research should be conducted to estabhsh the possibihty of biomagnification
of plutonium in invertebrate food chains and to determine the magnitude of the uptake
of other transuranic nuclides.

In freshwater systems, such as the Great Lakes, food webs are moderately simple
compared to those in oceans. Studies in the Great Lakes (Edgington, Wahlgren, and
Marshall, 1976; Bowen, 1976) indicate that, although conspicuous biomagnification of
plutonium occurs between water and phytoplankton, there is a net decrease of an order
oi magnitude for each higlier trophic level in the food chain. Results from studies at
Hanford U-pond, Rocky Flats, and the Miami River in Ohio are comparable to those from
Lake Michigan. The CR's (Table 15) reflect not only biological variation but also
variations in the concentration or chemical form of plutonium or americium in the water
column.

TABLE 15 Accumulation of Plutonium by Aquatic Organisms Leading to Man
(for Fish and Shellfish Given for Muscle Only)

Freshwater (concentration ratio)

Marine
(concentration ratio)

Aquatic
organism

Great

Lakes*

Miami
Riveit

Rocky

Flatst U-pond §

White Oak
Lakef

Atlantic/

Pacific** Irish Sea

Phytoplankton
Mi.ved
Qadophora

Macrophytes

5700
3800

220-2900
3000
1600

7000
2000

500-5000

260-3500

Zooplankton

350

1400-1700

Benthic

organisms
My sis

Ron toporeia
Mytilus
Worms

760-1600

260-490 2000tt
4000

Crustacea
Crayfish
Crab

600-1300

Fish
Benthic
Plank tivores
Piscivores

250
14-37
1-7

600

ND§§

3
0.04

1-13

Aquatic birds

1.0

*Edgington, Wahlgren, and Marshall, 1976.

jWayman, Bartelt, and Alberts, 1977.

ifPaine, this volume.

§ Emery and Klopfer, 1976.

HEyman and Trabalka, tliis volume.

**Noshkin, 1972.

tfHetherington et al., 1976.

§§ND, not detected in flesh.

SYNTHESIS OF THE RESEARCH LITERATURE 33

The transfer to humans seems Umited because the transuranic elements are not
significantly enriched in fresh edible fish (Edgington, Wahlgren, and Marshall, 1976;
Dahlman, Bondietti, and Eyman, 1976; Eyman and Trabalka, 1977; Pentreath and
Lovett, 1976; 1978; Pentreath et al., 1979).

Prediction of Long-Term Behavior

The long half-lives of several isotopes of the transuranic elements necessitate tlie
estimation of their behavior and effects over thousands of years. The behavior of
transuranic elements over a 30-yr interval may not properly represent behavior over more
extended periods. Uncertainties arise principally from effects of physical and biogeo-
chemical processes on the redistribution and form of transuranic elements in the
environment and from effects of these changes on biological availability.

Several research approaches have been taken to estimate the long-term behavior of
transuranic elements. These include (1) basic studies of environmental influences and
mechanisms that may alter distribution and biological availability over time; (2) investiga-
tions of the behavior of naturally occurring elements that have been in the environment
over geologic time and may exhibit analogous behavior; and (3) investigations of the
distribution and behavior of transuranic elements presently in the environment as a result
of defense activities. These approaches have developed information highly useful for
predictive purposes, but considerable research remains to be done before a reliable model
can be developed. It is essential to understand factors influencing the chemical speciation
of plutonium in the vicinity of biological membranes prior to uptake and how these
chemical changes influence the transfer within organisms and between trophic levels. This
will require more refined mechanistic studies (approach 1) using studies of analog
elements and plutonium distribution from fallout (approaches 2 and 3) to verify
predictions. For example, the behavior of plutonium is largely governed by the chemistry
of its lower oxidation states, Pu(III) and Pu(IV). However, Pu(V and VI) may be present
in highly oligotrophic lakes. Thus, under conditions in wliich the valence state controls
plutonium chemistry, the behavior of naturally occurring Th(IV) and U(VI) may serve as
analogs of Pudll + JV) and PufVI) in tests of predictions with respect to matrix and
environmental factors (e.g., pH, Eh, and ionic composition). The results of investigations
to define the distribution of plutonium from defense activities (approach 3) can be used
in a similar manner.

The complexity of the environmental chemistry of plutonium has required a major
basic research effort. Unless the mechanisms responsible for the behavior of plutonium
are known, it is not possible to develop or validate predictive models. For example, to
determine the validity of certain analog elements of plutonium, one must first determine
the predominant plutonium valence states and the conditions under which these valences
exist. Only then can comparisons be made with naturally occurring elements with similar
valences. Americium and neptunium, which have more than one oxidation state, must be
studied in this manner. The chemistry of these elements is less complex than that of
plutonium, and more rapid progress can be expected. Curium has only one oxidation
state, Cu(III), and analog chemistry should be straightforward.

Perhaps the most important factor Umiting our ability to predict the transport of
transuranic elements in ecosystems is our knowledge of ecosystem structure and function.
Prediction of the behavior of transuranic elements in the environment requires
infoirnation as to the concentrations of these elements in important ecosystem

34 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

compartments, the physical size of the compartments, and the processes and rates
controlling movements of materials between compartments. Knowledge of ecosystem
structure and function is useful in predicting the transport of any insoluble element
where physical and biotic transport processes dominate. Quantification of physical
and biotic processes controlling soil and sediment transport provides a basis for predicting
the behavior of elements tiglitly bound to them.

Ecological Effects of Transuranic Elements in the Environment

The main purpose of the studies described in this volume is to provide information that
can be used to help predict the consequences of transuranium elements in the
environment. Such consequences include possible harmful effects on man and other
species from current and potential levels of these elements in the biosphere. The
prediction of consequences requires detailed knowledge of source terms, environmental
transport, biological uptake, and biological effects expected from uptake. This review
emphasizes effects that might be expressed in species populations, such as mortality and
natality, and resulting perturbations in population density and community composition.

It is clear that effects induced by transuranic elements at the population or
community level have not been measured directly because environmental levels have not
been sufficient to produce obvious changes. Subtle changes in populations or communi-
ties are readily masked by natural variations, and ecologists are ordinarily unable to
measure small perturbations and identify their causes. However, indirect calculations and
extrapolations to low doses can be used to infer ecological consequences of transuranic
elements presently in the environment. The task of predicting ecological impacts of a
given level of transuranic elements in a particular environment is not simple. How the
contaminant will behave; i.e., how it will be distributed among the various ecosystem
components; how this distribution will change with time; and what physical, chemical,
and biological factors will affect the distribution, must be understood. There is also tlie
question of doses to critical biological tissues. Most transuranic elements are alpha
emitters that exliibit generally heterogeneous distributions in tissues, and this makes the
calculation of effective doses difficult. Finally, the relation of effective doses to biologic
effects must be understood.

The bank of data from which ecological impacts of the transuranic elements can be
predicted is limited. For instance, the relationships of tissue concentrations of transuranic
elements to concentrations in soil, air, or water are accurately known for only a few
ecosystems. Current data pertain to plutonium and americium; research on other
transuranic elements only recently has been initiated. Our experience with transuranic
elements in the environment has been too brief to allow us to predict long-term behavior
confidently. Another problem is tliat the microdosimetry of transuranic elements has
been well studied in only a few laboratory animals. Finally, predictable dose— effect
relationships exist mainly for plutonium in laboratory animals. Thus there is considerable
uncertainty as to dose— effect relationships for all the transuranic elements in plants and
aquatic organisms.

To measure direct relationships between amounts of transuranic elements and effects
would require purposely contaminating ecosystems at levels permitting direct observa-
tions of biological effects. In practice, however, this approach is not feasible for
ecosystem-scale investigations, and such studies have not been done. Ecosystems have
been contaminated with transuranic elements through mishaps or experiments for other

SYNTHESIS OF THE RESEARCH LITERATURE 35

purposes, but the levels have generally been orders of magnitude below those presumably
required to cause detectable ecological changes. Aquatic and terrestrial organisms exposed
to locally high levels of transuranic elements have been studied, but no evidence of
transuranic-related effects deleterious to a population has been reported (Bradley, Moor,
andNaegle. 1977).

Numerous investigators have directly assayed plutonium and a few other transuranic
elements in tissues of a variety of environmentally exposed aquatic and terrestrial
organisms, including humans. In these experiments the tissue burdens and resulting
radiation dose rates have generally been less than dose rates experienced during
evolutionary time from natural sources of radiation. This is due to the low levels of
transuranic elements in the environment and also to their low solubihty and biological
mobility. At such low levels of radiation exposure, ecological changes would be
undetectable. Laboratory studies of a variety of aquatic and terrestrial species have shown
that radiation dose rates several orders of magnitude higher than those resulting from
natural background sources are necessary to produce gross changes in mortality or
natality. This is true even for tlte more radiosensitive stages of comparatively sensitive
organisms.

Altliough gross ecological effects from transuranic elements are not Ukely to be
demonstrated at tlie levels hitlierto experienced in the environment, there is reason to
expect a statistically determined incidence of biological effects, such as tumors and
genetic alterations. In the absence of sufficient data to the contrary, a linear dose— effect
relationship is generally assumed for cancer induction and genetic mutations at low doses.
If this assumption is correct, then any dose, however low, imposes some risk. Since
concern for most plants and animals is generally for populations rather than for
individuals, modest increases in genetic or somatic effects are not expected to have
measurable consequences. A different attitude prevails for humans, however, where there
is concern for individual organisms.

Dose rates can be calculated and compared with natural background or with literature
on dose— effect relationships. Table 16 lists dose rates calculated from measured tissue
concentration of plutonium in a variety of organisms exposed to elevated environmental
levels and in humans exposed to fallout. The data from Windscale, Rocky Flats, and the
Nevada Test Site apparently represent some of the organisms exposed to the higliest doses
of transuranic elements studied. Even in those higlily localized cases, calculated dose rates
are about tlie same as or less than those for natural background, and measurable
population-level changes are not expected (National Academy of Sciences-National
Research Council, 1972). Doses to humans exposed to fallout plutonium have been so
low that specific biological effects cannot be demonstrated (Thompson and Wacliholz,
this volume).

Levels and distributions of transuranium nuclides in water, sediments, and selected
biota, particularly in locally contaminated freshwater and marine sites, have been
examined extensively. However, there are few data pertaining to biologic effects. In fact,
there have been no good opportunities to observe effects of transuranic elements in
natural aquatic ecosystems. Reported water concentrations of plutonium and other
transuranic elements in natural environments have been 1 pCi/liter or less, and dose rates
appear to be three to eight orders of magnitude less than dose rates required to produce
detectable effects (Templeton. this volume; Till. Kaye, and Trabalka, 1976). Present data
suggest that aquatic systems can receive several orders of magnitude more transuranic
activity than experienced in the past before ecological changes will be detectable.

36 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 16 Examples of Dose Rates Calculated from Measured Tissue Concentrations
of Plutonium in Environmentally Exposed Organisms

Dose rate.

Organism

Tissue

Environ men t/ Location

mrad/day

Reference

Mussels

Viscera

Windscale/Irish Sea

1.6*

Hetherington et al., 1976

Crab

Gill

Windscale/Irish Sea

3.1*

Hetherington et al., 1976

Plaice

Bone

Windscale/Irish Sea

0.04*

Hetherington etal., 1976

Liver

Windscale/Irish Sea

0.17*

Hetherington etal., 1976

Kidney

Windscale/Irish Sea

0.15*

Hetherington et al., 1976

Fish

Embryos

Windscale, White Oak Lake,
U-pond, Enewetak lagoon,
Lake Michigan

10-* to 10-'

Till, Kaye, and Trabalka,
1976; TiU and Franks,
1977

Small mammals

Whole

Rocky Flats

1.7

Little, 1976

Carcass

Nevada Test Site

3.3

Bradley, Moor, and Naegle,
1977

Arthropods

Whole

Rocky Flats

0.9

Little, 1976

Cotton rats

Carcass

Savannah River Plant

0.007

McLendon etal., 1976

Humans

Bone

United States

10-"

Mclnroy et al., 1977

Lymph nodes

United States

10-'

Mclnroy etal., 1977

* Includes dose from ^" ' Am.

Opportunities to observe and quantify ecological changes resulting from transuranic
contamination of terrestrial environments have also been extremely limited. Terrestrial
ecosystems contaminated with plutonium at levels of 10 to 1000 ixC'ilm^ have been
examined carefully but without demonstrable effects (Whicker, this volume). Assays of
plutonium in plants and animal tissues from such contaminated areas reveal levels of
plutonium generally less than 10 pCi/g. Dose rates from such plutonium concentrations
are a few millirad per day. Chronic dose rates of at least a few rad per day are generally
required to cause detectable ecological changes (Whicker and Fraley, 1974; Turner,
1975). Calculations based on a substantial body of information suggest that man could
occupy and derive sustenance from land containing 20 to 200 juCi ^^^Pu/m" without
exceeding the nonoccupational maximum permissible dose to the lung or other critical
organs (Healy, 1974; Martin and Bloom, 1976). Other calculations suggest that ^^^Pu
levels of 1 to 1000 mCi/m'^ would be required to cause significant mortality in plant and
animal populations. Mammals would probably show mortality at lower levels than plants
(Whicker, this volume).

Summary

The preceding discussion leads to a number of generalizations that can be summarized as
follows:

l.The nature of the source for release to the environment is important in the
initial deposition and distribution of transuranic elements. However, as environmental
factors, such as erosion, chemical weathering, and biological processes, proceed, tlie
original chemical and physical properties are altered and source influence diminishes.

2. The major repositories of plutonium and americium are soils and sediments.

3. Suspended particles in air and water act as vectors for the physical movement of
plutonium and americium, and erosional processes are the principal means of translational
movement in the environment.

SYNTHESIS OF THE RESEARCH LITERATURE 3 7

4. In spite of the large fraction of plutonium and americium residing in soils and
sediments, chemical and biological processes produce a veiy small fraction of soluble
species in terrestrial and aquatic environments. These species are incorporated in
biological tissue, but the concentrations in biota have not produced demonstrable
deleterious radiation effects.

5. An increase or decrease in the soluble fraction of plutonium over long weathering
times cannot be demonstrated at tliis time. However, preliminary observations of
naturally occurring analog elements indicate that plant uptake and transfer of plutonium
and americium througli food chains would not be expected to change appreciably over
time.

6. Concentrations of plutonium do not increase from one trophic level to the next in
natural food webs except for sorption by phytoplankton and one observation of starfish
feeding on mussels.

7. The environmental chemistry of transuranic elements in marine and in oligotrophic
freshwater systems is similar in a number of ways. However, significant differences in
chemical species exist in many lakes where chemical conditions, such as pH and ligand
concentration (botli organic and inorganic), may be different.

8. Present levels of transuranium elements in our environment have not produced
discernible ecological effects.

Important reservations are implicit in the above generalizations mainly because of
insufficient information on fundamental processes and lack of data pertaining to
transuranic elements other than plutonium. Three lines of investigation are necessary in
future studies:

1. Develop process and dose models as a framework to identify specific research areas
where important data are lacking.

2. Expand research related to neptunium, americium, and curium to provide a
broader base of information about the environmental behavior of the transuranic
elements.

3. Investigate the kinetics of the Pu(IV)^Pu(V and VI) oxidation and the factors
controlling tliis valence distribution since increasing evidence suggests that oxidation
mechanisms occur that make plutonium more soluble than predicted in some environ-
mental media.

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42 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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SYNTHESIS OF THE RESEARCH LITERATURE 43

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44 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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Radiological Assessments, Environmental
Monitoring, and Study Design

WAYNE R. HANSEN and DONALD R. ELLE

Studies of the behavior of transuranic elements in the environment form the basic data
for applied programs in radiological assessment, environmental monitoring, derivation of
radiation-protection standards, and environmental impact statements. This chapter
introduces some of the major information requirements of these applications of
transuranic research data. Giaracteristics of the source terms from nuclear activities
usually are needed for an analysis of environmental pathways or deployment of
monitoring systems. Major inhalation and ingestion pathways are considered in
assessments of hazards from transuranics and are discussed from the viewpoint of research
needed.

In conducting radiological assessments, writing environmental impact statements,
attempting to derive standards, or designing monitoring programs for transuranic
elements, one must rely on data from existing studies of transuranic elements in the
environment. In each of these types of assessments, man is usually the major receptor
considered for a variety of pathways. The data used to estimate the radiological impact of
transuranics on man derive from the results of research carried out with a variety of
objectives. The objectives may have been hmited to the assessment of a specific pathway
at a specific site. Data obtained for a particular pathway or portion of a pathway and
geographical area often are applied, with modifying assumptions, to other geographical
areas for lack of data specific to the area of interest.

The design of environmental monitoring programs for estimating radiological effects
on man must include consideration of a large number of factors. The following discussion
reiterates some of these factors in study design and analysis for use in radiological
assessments. A brief discussion of the nontechnical influences on radiological assessments
and thus study design is also included. Statistical considerations and modeling are
considered elsewhere but will be referred to as necessary. Hopefully, a statistician will
always be included in the design phase of any study or monitoring system.

Prior to the projected growth of the nuclear-power industry, efforts to study the
environmental behavior of transuranic elements were centered around dispersal by
nuclear weapons testing programs. Early radiological assessments of the behavior of
transuranium elements in the environment relied on conservative assumptions owing to
the lack of empirical data and concentrated on plutonium, which was the major
transuranic element in weapons manufacture and testing. Weapons plutonium is still the
major transuranic element available for study. Data are less available or nonexistent for
curium, americium, and neptunium, but assessments should still include these elements.

46 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

The need for additional data on plutonium dispersal and behavior in the event of an
accident with weapons components was recognized early. Examples of early studies
designed to provide data for radiological assessments are Operation Plumbbob in 1957 at
the Nevada Test Site and Operation Roller Coaster in 1963 at the Tonopah Test Range.
Jordan (1971) described the objectives of these tests as being primarily concerned with
obtaining data about the dispersion characteristics and biomedical impacts on animals
exposed to the airborne plutonium and with evaluating instrumentation and decontam-
ination methods. During the design of the experiments, the personnel involved decided
not to attempt measurements of resuspension because of the complex nature of the
process. Analysis of the data from these tests provided the experience and guidance
required to deal with the dispersion of plutonium from nuclear weapons accidents in
Palomares, Spain, and Thule. Greenland. In addition, the test areas in Nevada provided
research areas for the ongoing study and evaluation of the long-term environmental
behavior of the residual plutonium in a desert ecosystem.

Environmental Assessments, Impact Statements

With the growth in the number of light water reactors (LWR), the projected recycle of
plutonium in LWR, and the projected liquid-metal fast breeder reactor (LMFBR), the
detailed radiological assessments of transuranics increased to keep pace with planned fuel
fabrication and fuel reprocessing facilities. Beginning in 1970 the National Environmental
Policy Act (NEPA) of 1969 (U. S. Congress, 1970) required that prior to implementing
"major Federal actions significantly affecting the quality of the human environment, a
detailed statement" must be prepared which addresses the environmental impact, any
adverse environmental effects that cannot be avoided, alternatives, relationships between
short-term uses and long-term productivity of the environment, and any irreversible and
irretrievable commitments of resources. The Council on Environmental Quality (1976)
issued guidelines for the preparation of environmental impact statements in 1976.
Actions taken in the past by the U. S. Atomic Energy Commission and presently by the
U. S. Nuclear Regulatory Commission and the U. S. Department of Energy have been
considered major actions that significantly affect the human environment. Both the
environmental impact statement on the LMFBR program by the U. S. Energy Research
and Development Administration (1975) and the environmental statement by the U. S.
Nuclear Regulatory Commission (1976) on the use of recycle plutonium in mixed-oxide
fuel in light-water-cooled reactors (GESMO) contain radiological assessments that
estimate the radiation doses from transuranics. In each case esfimates of the radiological
impacts on man were made from data available at the time of the preparation of the
environmental impact statement. In most cases the limited data available required that
conservative assumptions and extrapolations be used in predicting the impacts as required
by NEPA and the CEQ guidelines. These estimates become decision-making tools and are
the subject of debate in hearings or litigation for licensing of facilities.

Generic environmental statements and modeling efforts, such as those carried out by
Bloom and Martin (1976), have been based on hypothetical individuals who obtain air,
food, and water from the area of maximum transuranic concentration. The estimated
radiation dose for such broad studies is usually for transuranics from a postulated source.
Existing facilities usually have accumulated some data that describe the source of
transuranics. Existing facilities carry out environmental monitoring programs that are
designed to detect changes in transuranics and other radionuclides in environmental

ENVIRONMENTAL ASSESSMENTS, MONITORING, AND STUDY DESIGN 47

media, such as air, water, soil, vegetation, and animals, and thus verify the results of
emission and effluent monitoring programs from which the radiological impacts from
existing facilities are estimated. The design of the surveillance programs and the
radiological assessments performed, however, are still heavily dependent on data provided
by the studies of transuranics in different ecosystems.

General Aspects of Environmental Monitoring

The general design of networks and of programs for the measurement of radioactive
materials in the environment has been described by the International Commission on
Radiological Protection (1965) and by the World Health Organization (1968). The
International Atomic Energy Agency (1966; 1975) has published two guides for
environmental monitoring, and more recently, the National Council on Radiation
Protection and Measurements (1976) published a report Qntitled Environmental Radia- ■
tion Measurements. In addition to the recommendations of scientific bodies such as ICRP
and NCRP, more specific guidance for environmental monitoring is provided by
government agencies (U.S. Atomic Energy Comimission, 1974). The Nuclear Regulatory
Commission issues general guidance as regulatory guides. Regulatory Guide 4.5, issued in
1974, for example, deals with the sampling and analytical procedures for plutonium in
soil. The regulatory guides are not regulations but represent methods that are acceptable
for licensing actions or compliance with regulations for operating facilities. The
Department of Energy relies on surveillance programs tailored to specific sites and
problems. Its contractors issue annual reports of the methods and results of the
surveillance programs. The DOE follows A Guide for Environmental Radiological
Surveillance at ERDA Installations (Corley etal., 1977), which is based on a
state-of-the-art review of environmental monitoring practices. The specific objective of
the guide was to develop guidance for achieving comparable, high-quality, environmental
monitoring and reporting programs at DOE installations, which encompass a wide variety
of nuclear activities, i.e., plutonium production, reactor operation, and research studies.
Environmental monitoring systems for all types of facilities have many similarities.
Although somewhat specific to reactors, the Environmental Radioactivity Surveillance
Guide issued by the U. S. Environmental Protection Agency (1972) contains general
information on samphng methods and frequencies that can be applied.

Monitoring systems are usually designed to verify that a facility is operating within
limits specified as safe by a scientific body. These limits become law where incorporated
into state or federal codes, such as Qjde of Federal Regulations, Title 10, Part 20
{Federal Register, 1976). Each monitoring program should, prior to deployment, identify
the pathways to man for transuranic elements in addition to the normal pathways
requiring monitoring by regulation. Figures 1 and 2 are examples of simplified pathway
diagrams for the movement of radionuclides to man. Not all the pathways in the diagrams
will be present for a given site. For a specific facihty and location, however, all pathways
should be identified and analyzed for their contribution of transuranic elements to the
total radionucHde uptake by man or biota. As pathways are identified and analyzed, the
number of pathways requiring routine monitoring will be reduced. In the analysis of
pathways, the short-term and long-term aspects of accumulation and movement of
transuranics should be kept in mind.

The considerations that are included in the idenfification and analysis of the
pathways are the many aspects of environmental studies described in other chapters of

48 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

this book, starting with the source term. Characteristics of the source term are important
for the selection of monitoring methods and instrumentation. Some of the general
characteristics of the transuranics to be considered are

1. Quantity.

2. Rate of release.

3. Chemical form.

4. Physical characteristics, such as particle size distribution or ionic state.

5. The presence of radionuclides.

6. The presence of nonradioactive chemicals.

For routine operation of such facilities as plutonium fabrication, reactor-fuel fabrication,
and reactor-fuel reprocessing, the emission and effluent monitoring systems will provide
the infomiation about quantities and rates of release. The chemical fomi may be inferred
from the operations being carried out by the facility. The physical characteristics of the
emissions or effluents can be identified by a specific study or inferred from the operating
parameters for the waste-treatment system. The transuranic isotopic composition is
usually well defined by criticality and safeguards calculations or by analysis of actual
effluent samples. The presence of other radionuclides, such as fission products, may
interfere in monitoring transuranics or may be helpful by serving as tracers from which
ratios with respect to transuranics can be determined. The presence of nonradioactive
chemicals combined with the transuranics in emission or effluents may alter the original
chemical forms introduced into the waste-treatment systems. For normal operation of
nuclear facilities handHng transuranics, the quantities released usually are small.
Monitoring for the transuranics in the environment from nonnal operations of facilities
fabricating plutonium metal, heat sources, or reactor fuel or reprocessing reactor fuel
become oriented to long-term buildups. The source-tenn parameters for unexpected or

DIRECT RADIATION

INGESTION

DEPOSITION

CROPS

AND

PLANTS

DIRECT

AIR

—

-r

DEPOSITION

SOIL

MAN

RADIATION
INGESTION

RADIOACTIVE
MATERIALS

ANIMALS

INHALATION

INH

^LAl

ION

Fig. 1 Simplified pathways between radioactive materials released to atmosphere and
man. [From International Commission on Radiological Protection (1965).]

ENVIRONMENTAL ASSESSMENTS, MONITORING, AND STUDY DESIGN 49

DIRECT RADIATION

RADIOACTIVE
MATERIALS

AQUATIC
PLANTS

FISHING

AND SPORTS

GEAR

SOIL

SAND AND
SEDIMENT

> . _.

UIHbtl HAUIAIIUiM

■

SURFACE
WATER OR
GROUND-
WATER

AQUATIC
ANIMALS

MAN

IRRIGATION
WATER

RADIOACTIVE
MATERIALS

LAND
PLANTS

SOIL

LAND
ANIMALS

INGESTION

Fig. 2 Simplified pathways between radioactive materials released to groundwater or
surface water (including oceans) and man. [From International Commission on
Radiological Protection (1965),]

sudden releases, such as may occur from accidents, are identified after an occurrence.
Although facilities incorporate engineered protection against such releases, environmental
surveillance programs must be prepared to trace the movement of transuranics released to
the environment after the initial assessments of the emission or effluent monitoring
results. Existing areas of transuranics in the environment also present a challenge to the
design of an adequate monitoring system. Once the transuranics are deposited in the
environment, however, the pathway considerations for routine and accidental sources are
much the same at a given site. To predict the radiological impact from areas of potential
future contamination requires a knowledge of the parameters used to estimate the dose
from transuranic pathways.

The methods for calculating dose use ICRP models that predict the metaboUc fate of
radionuchdes on the basis of the chemical and physical characteristics of the mode of
intake. The parameters needed for dose calculations should be considered during the
analysis of pathways for the design of monitoring systems so that the number of
assumptions needed for the dose estimations can be minimized. Details of the
dose-estimation methods and consequences are discussed in many other pubhcations. At
the risk of being redundant, some of the methods are discussed below in relation to
pathway information.

50 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Pathways and Inhalation

The inhalation pathways for a number of exposure modes for man have been identified
by ICRP. In addition to the direct inhalation of airborne transuranics released from an
operating facility, other secondary pathways are possible and have been discussed by
Healy (1974), These secondary pathways are the primary inhalation pathways from
existing areas of transuranics in the environment. Resuspension by wind is discussed in
another chapter. Other means of generating airborne transuranics, such as agricultural
activities (including home gardening), carriage into homes on clothes and pets, children's
play, and movement by vehicular activity, need further quantification.

Estimates of the doses from inhalation of the different airborne transuranics use the
ICRP II lung model (International Commission on Radiological Protection, 1960), which
categorizes the inhaled material as soluble or insoluble. The Task Group on Lung
Dynamics model (International Commission on Radiological Protection, 1966) has three
categories- of solubiUty. Additionally, the deposition of particulates in three respiratory
regions is dependent on particle size in the Task Group model. Residence time in the lung
is dependent on solubihty classification in both models. After transuranics have resided in
the environment for months or years, the solubiUty and particle sizes may change. Data
are needed which evaluate the changes of solubility and particle-size distribution with
time and weathering. Although some work has been started for plutonium, more
long-term studies of the solubihty and particle-size changes are needed. Data for
neptunium, americium, and curium is sparse to nonexistent. The derivation or collection
of some of this information could be incorporated into the design of present
environmental monitoring programs. More detail about the association of the transuranics
with soil particles, particle-size distribution, and chemical changes with time and
weathering is gathered in separate projects that are usually beyond the scope and budget
of monitoring programs.

Pathways and Ingestion

Although inhalation is the primary pathway for human exposure, ingestion also plays an
important role. Transuranics can be ingested through numerous pathways. These
pathways are strongly influenced by local water-use practices, agricultural systems, sport
fisheries and wildlife use, and estuarine and marine fisheries, and by the amount of soil
directly ingested on contaminated plants and from the hands. The deliberate ingestion of
soil by children is a special pathway for evaluation identified by Healy (1974). Dose
assessments for transuranics have generally indicated that inhalation is the dominant
pathway. The hmited data available for such transuranics as neptunium, americium, and
curium, however, indicate that these elements are possibly more available than plutonium
for plant uptake, as indicated by an examination of the data available for plant uptake of
transuranics. Thomas and Healy (1976) concluded that current information is inadequate
for accurate dose assessments. The research needs discussed in the following chapters on
physical and biological transport mechanisms are many of those which would reduce the
great number of uncertainties contained in present dose assessments for transuranics from
ingestion. Pathways in particular that need further quantification are the contamination
of food plants by resuspension and the consequences of incorporation of transuranics
into organic molecules.

The identification of local food webs is one of the more important aspects of dose
assessment and design of monitoring programs. Regional diets throughout the United

ENVIRONMENTAL ASSESSMENTS, MONITORING, AND STUDY DESIGN 51

States and the world vary widely according to the local agricultural, sport fisheries and
wildlife, commercial fisheries, and water-use practices. Food imported or exported may
serve to reduce or increase the total dose of transuranics. Lx)cal gardens or small farms
may lead to maximization of dose to an individual through the food pathway. Dose
assessments have usually been calculated for a "maximum individual," i.e., one who
obtains all sustenance at a facility site boundary. For an actual facility where food chains
to man could be identified, the use of the "fence-post man" would appear to contain
numerous conservatisms.

Sampling for transuranics in soils, plants, or other media in areas with existing levels
of radionuclides must be designed according to the particular needs of the study and must
include consideration of existing levels of transuranics in the environment. Methods for
inventory sampling may not always be adequate for defining resuspendible material
leading to inhalation or plant contamination. Studies must be carried out which provide
information about inventory mobility and the consequential effect on human health.
General comments throughout this section have referred to the general considerations of
monitoring program design with Httle reference to specific guidance. The selection of
sampling methods and measurement techniques for environmental media, including those
used to calculate radiation dose to people but also commonly used as trend indicators,
have been discussed in several review articles and publications; NCRP 50, the DOE guide,
ICRP-7, and the NRC regulatory guides are examples. As already discussed, the proper
selection of media samples, based on the detailed pathway analysis, is important in
assessing the dose to people. It would not be possible to detail these considerations here
for all media; therefore discussion of soil contamination is presented as an example.

As guidance for ingestion and inhalation of transuranics in the environment was being
developed, consideration of soil concentration limits and resulting problems in appli-
cability of a soil Hmit led to the recognition that dose limits to the lung and bone are
most important. The calculation of this dose can be based on air measurements, in which
case more information on particle size and the physical parameters of the transuranics is
necessary than is normally developed in environmental surveillance programs. Thus the
need for research input and cooperation with environmental programs becomes
important.

Although the air pathway is of primary concern, the environmental measurements
eventually must be or will be translated into soil concentrations. Thus the usual inventor>'
measurements and sample techniques will not provide adequate information on the
resuspendible and respirable fraction of transuranics in the environment. Definitions of
such things as resuspendible surface, sample collection methods, sample preparation, and
particle-size determinations are all factors in the radiological assessment of transuranic-
contaminated soils. Much of the necessary data is yet to be determined. SampHng
techniques have been reviewed by several researchers, including Bernhardt (1976).
Differences in techniques exist which necessitate evaluation and verification of
methodology used in assessing radiological impacts of transuranics in the environment.

References

Bernhardt, D. E., 1976, Evaluation of Sample Collection and Analysis Techniques for Environmental
Plutonium, Technical Note ORP/lv-76-5, U. S. Envlionmental Protection Agency, NTIS.

Bloom, S. G., and W. E. Martin, 1916, A Model to Predict the Environmental Impact of the Release of
Long-Lived Radionuclides, Final Report to the U. S. Environmental Protection Agency, Battelle,
Columbus Laboratories, NTIS.

52 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Corley, J. P., D. H, Denham, D.E. Michels, A. R. Olsen, and D. L. White, 1977, Guide for

Environmental Radiological Surveillance at ERDA Installations, ERDA Report ERDA-77-24,

Battelle, Pacific Northwest Laboratories, NTIS.
Council on Environmental Quality, 1976, Environmental Quality, The Fifth Annual Report of the

Council on Environmental Quality, Appendix D, Preparation of Environmental Impact State-
ments: Guidelines, pp. 506-522, GPO.
Federal Register, 1976, Special Edition, Code of Federal Regulations, Title 10, Energy, GPO.
Healy, J. W., 1914, A Proposed Interim Standard for Plutonium in Soils, USAEC Report LA-548 3-MS,

Los Alamos Scientific Laboratory, NTIS.
International Atomic Energy Agency, 1966, Manual on Environmental Monitoring in Normal

Operation, Safety Series No. 16, STI/PUB/98, International Atomic Energy Agency, Vienna.
, 1975, Objectives and Design of Environmental Monitoring Programmes for Radioactive

Contaminants, Safety Series No. 41, STI/PUB/385, International Atomic Energy Agency, Vienna.
International Commission on Radiological Protection, 1960, Report of Committee II on Permissible

Dose for Internal Radiation, Health Phys. , 3: 27-39.
, 1965, Principles of Environmental Monitoring Related to the Handling of Radioactive Material,

ICRP Publication 7, Pergamon Press, Inc., New York.
, 1966, Task Groups on Lung Dynamics. Deposition and Retention Models for Internal Dosimetry

of the Human Respiratory Tract, Health Phys., 12: 173-207.
Jordan, H. S., 1971, Distribution of Plutonium from Accidents and Field Experiments, in Proceedings

of Environmental Plutonium Symposium, Los Alamos, N. Mex., Aug. 4-5, 1971, USAEC Report

LA-475 6, pp. 21-24, NTIS.
National Council on Radiation Protection and Measurements, 1976, Environmental Radiation

Measurements, NCRP Report No. 50, NCRP Publications, Washington, D. C.
Thomas, R., and J. W. Healy, 1976, An Appraisal of Available Information on Uptake by Plants of

Transplutonium Elements and Neptunium, ERDA Report LA-6460-MS, Los Alamos Scientific

Laboratory, NTIS.
U. S. Atomic Energy Commission, 1974, Measurements of Radionuclides in the Environment:

Sampling and Analysis of Plutonium in Soil, Office of Standards Development, NTIS.
U. S. Congress, 1970, Public Law 91-100, National Environmental Policy Act of 1969, 9\st Congress,

S. 1075.
U. S. Energy Research and Development Administration, 1915, Final Environmental Statement on the

liquid Metal Fast Breeder Reactor Program.
U.S. Environmental Protection Agency, 1972, Environmental Radioactivity Surveillance Guide,

Report ORP/SID-72-2, Office of Radiation Programs.
U. S. Nuclear Regulatory Commission, 1976, Final Generic Environmental Statement on the Use of

Recycle Plutonium in Mixed Oxide Fuel in Light Water Cooled Reactors, Health Safety and

Environment , Office of Nuclear Material Safety and Safeguards, NTIS.
World Health Organization, 1968, Routine Surveillance for Radionuclides in Air and Water.

Worldwide Fallout

R. W. PERKINS and C. W. THOMAS

Since the first nuclear weapons test at Alamogordo, N. Mex., on July 16, 1945,
approximately 360,000 G (360 kCi) of '^^-^'^'^Pu lias been injected into the atmo-
sphere. In addition, 1 7,000 G (17 kG) of^^^Pu entered the atmosphere in April 1964 as
a result of the high-altitude burnup of a SNAP-9 satellite power source. Since most of the
plutonium from nuclear weapons testing, as well as tliat from the SNAP-9 burnup,
entered the stratosphere, fallout lias been worldwide. The deposition is influenced by
meteorological conditions and also by topographical features of the earth s surface.
Residence time in the stratosphere is about 10 to 11 months: however, because of the
high-altitude burnup of the SNAP-9 device, it was 2 yr before significant amounts of this
debris reached the earth s surface.

In addition to plutonium, substantial amounts of '^^^ Am are formed from the decay
of the weak beta emitter ^"^^ Pu and are an important constituent of fallout.

Tlie majority of radioactivity entering the stratosphere during this past decade has
been a result of the Giinese nuclear weapons testing. Tlie ratio of plutonium to ^^'^ Cs has
been relatively constant throughout the nuclear weapons period, and thus a measurement
of ^^"^Cs permits a reasonable estimate of the plutonium deposition. The ratio of
transuranic elements in fallout is substantially different from tliat in power reactor
wastes, which contain far more americium and curium relative to plutonium. Fresh
fallout from thermal nuclear weapons contains large amounts of short-lived ^^'^U and
^^^Np, and these may contribute substantially to the radiation exposure at the earth's
surface.

The first significant injection of transuranium elements into the atmosphere occurred as
the result of the nuclear weapons testing in Alamogordo, N.Mex., on July 16, 1945.
Between then and 1952 further nuclear detonations resulted in additional injections to
the atmosphere; however, because of their relatively low yield, most of this debris was
confined to the troposphere. On Nov. 1, 1952, the first thermonuclear device was
detonated. This 14-Mt explosion injected large amounts of debris into the stratosphere.
The relatively high energy yield of this fusion device, together with a much higher
integrated neutron flux, greatly increased the production of the transuranium elements.
The majority of the transuranium elements and other nuclear debris which has been
injected into the atmosphere was produced during the 1961 and 1962 United States
(U. S.) and Union of Soviet Socialist Republics (U.S.S.R.) nuclear testing programs. A
nuclear weapons test-ban agreement between the United States, United Kingdom, and
Soviet Union in early 1963 suspended atmospheric testing. However, in late 1964 the
Chinese exploded their first atmospheric nuclear test, and since that time they have
continued testing in the northern hemisphere. France was not a member of the test-ban
agreement, and in mid-1966 they began atmospheric testing in the southern hemisphere.
The test-ban agreement in 1963 did not rule out underground. tests, which do not vent to

34 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

the atmosphere. Since 1963 major underground testing, which included several hundred
underground nuclear devices, has been conducted but has in only a few cases released
radioactivity to the atmosphere.

Most of the studies of transuranium elements from nuclear weapons testing have been
concerned with the measurement of 239,240pjj isotopes. However, the complete decay of
the accompanying short-lived ^'^^Pu (15 yr) results in the formation of a quantity of
^^^Am which approaches that of ^■^^'^'^^Pu. A major injection of ^^^Pu into the
atmosphere occurred in April 1964 when a navigational satellite failed to achieve a stable
orbit and disintegrated on reentry into the atmosphere. The 17 kCi of ^^^Pu that was
added to the atmosphere was relatively small compared with the 360 kCi of 239,240p^
that has been added by nuclear weapons testing. However, it greatly increased the
worldwide ^^^Pu deposition, and its point-source injection has been useful in developing
models describing global atmospheric mixing.

Other incidents have added to the environmental distribution of transuranic elements
but not on a worldwide scale. An aerial refueling explosion involving a B-52 bomber
carrying four plutonium-bearing nuclear weapons and a KC-135 tanker occurred on
Jan. 16, 1966, 28,000 ft above the Mediterranean coastline near the Spanish village of
Palomares, Spain. The high-explosive component part of two weapons exploded on
impact, releasing the weapons plutonium inventory over the hillside outside the village.
On Jan. 21, 1968, a B-52 with four plutonium-bearing nuclear weapons on board
attempted an emergency landing at Thule Air Force Base. At 9000 ft over the base, the
crew bailed out, and the abandoned plane crashed on the ice of North Star Bay. The
high-explosive components of all weapons detonated, and the plutonium inventory was

scattered over the ice.

The operation of nuclear reactors also results in the production of transuranium

elements, and the potential exists for release of some of these to the atmosphere during
reactor operation and subsequent fuel processing. The modern nuclear power plants,
which are designed for the generation of electric energy, use very long fuel exposure
periods and may in the future recycle the fuel to burn the resulting ^^^Pu. This results in
successive neutron capture of the transuranium elements and production of very
substantial quantities of higher mass elements. It has been estimated that approximately
2 X 10^ Ci of transuranium elements may be produced as radioactive waste through the
year 2000. Whereas 2 3 9,2 4 0pjj ^^^ ^'^^ Am are the main transuranium alpha activities
from nuclear weapons testing, ^^^Pu, ^'^^Am, and the curium isotopes will be the
principal alpha activities from nuclear reactor operations. Accidental releases of
transuranium elements to the atmosphere have occurred both from nuclear plant
operation and from the transport of nuclear weapons. The total amounts released to the
atmosphere by these processes have been relatively minor; however, such accidents may
have rather significant local effects.

Distribution of Transuranium Elements from Nuclear Explosions

The amounts of transuranium elements from nuclear testing distributed over the
world surfaces have been estimated on the basis of the nuclear tests of all nations (Hardy,
1964; United Kingdom Atomic Energy Authority, 1972; 1973; 1974; 1975; Nakahara
etal., 1975). Tables 1 to 5* are summaries of the individual tests performed by each

*Publication of this book does not constitute a DOE endorsement of the accuracy or completeness

of the list of alleged tests contained in these tables.

(Text continues on page 59.)

WORLDWIDE FALLOUT 55

TABLE 1 United States Nuclear Detonations

Height of

Cloud top,

Date

Name

burst, ft

Type

Yield

Location

Trinity

July 16, 1945

Trinity

100

Tower

19 kt

35,000

Alamogordo,
N. Mex.

World War 11

Aug. 5, 1945

World War II

-1,850

Air

20 kt

Hiroshima, Japan

Aug. 9, 1945

World War II

-1,850

Air

20 kt

Nagasaki, Japan

Crossroads

June 30, 1946

Able

520

Air

20 kt

35,000

Bikini Atoll

July 24, 1946

Baker

-90

Underwater

20 kt

8,000

Bikini Atoll

Sandstone

Apr. 14, 1948

X-ray

200

Tower

37 kt

56,000

Enewetak Atoll

Apr. 30, 1948

Yoke

200

Tower

49 kt

55,000

Enewetak Atoll

May 14, 1948

Zebra

200

Tower

18kt

28,000

Enewetak Atoll

Ranger

Jan. 27, 1951

Able

1,060

Air

1 kt

17,000

Nevada Test Site

Jan. 28, 1951

Baker

1,080

Air

8kt

35,000

Nevada Test Site

Feb. 1, 1951.

Easy

1,080

Air

1 kt

12,000

Nevada Test Site

Feb. 2, 1951

Baker-2

1,100

Air

8kt

36,000

Nevada Test Site

Feb. 6, 1951

Fox

1,435

Air

22 kt

42,000

Nevada Test Site

Greenhouse

Apr. 7, 1951

Dog

300

Tower

Enewetak Atoll

Apr. 20, 1951

Easy

300

Tower

47 kt

40,000

Enewetak AtoU

May 8, 1951

George

200

Tower

Enewetak Atoll

May 24, 1951

Item

200

Tower

Enewetak Atoll

Buster-Jangle

Oct. 22, 1951

Able

100

Tower

<0.1 kt

8,000

Nevada Test Site

Oct. 28, 1951

Baker

1,118

Air

3.5 kt

29,000

Nevada Test Site

Oct. 30, 1951

Charlie

1,132

Air

14 kt

40,000

Nevada Test Site

Nov. 1, 1951

Dog

1,417

Air

21 kt

40,000

Nevada Test Site

Nov. 5, 1951

Easy

1,314

Air

31 kt

45,000

Nevada Test Site

Nov. 19, 1951

Sugar

Surface

1.2 kt

16,000

Nevada Test Site

Nov. 29, 1951

Uncle

-17

Underground

1.2 kt

11,000

Nevada Test Site

Tumbler-Snapper

Apr. 1, 1952

Able

793

Air

1 kt

16,000

Nevada Test Site

Apr. 15, 1952

Baker

1,050

Air

1 kt

16,000

Nevada Test Site

Apr. 22, 1952

Charlie •

3,447

Air

31 kt

42,000

Nevada Test Site

May 1, 1952

Dog

1,040

Air

19 kt

42,000

Nevada Test Site

May 7, 1952

Easy

300

Tower

12 kt

34,000

Nevada Test Site

May 25, 1952

Fox

300

Tower

llkt

41,000

Nevada Test Site

June 1, 1952

George

300

Tower

15 kt

37,000

Nevada Test Site

June 5, 1952

How

300

Tower

14 kt

41,000

Nevada Test Site

Ivy

Oct. 31, 1952

Mike

Surface

14 Mt

-100,000

Enewetak AtoU

Nov. 15, 1952

King

1,480

Air

High yield

-70,000

Enewetak AtoU

Upshot-Knothole

Mar. 17, 1953

Annie

300

Tower

16 kt

41,000

Nevada Test Site

Mar. 24, 1953

Nancy

300

Tower

24 kt

42,000

Nevada Test Site

Mar. 31, 1953

Ruth

300

Tower

0.2 kt

14,000

Nevada Test Site

Apr. 6, 1953

Dixie

6,020

Air

11 kt

43,000

Nevada Test Site

Apr. 11,1953

Ray

100

Tower

0.2 kt

13,000

Nevada Test Site

Apr. 18, 1953

Badger

300

Tower

23 kt

35,000

Nevada Test Site

Apr. 25, 1953

Simon

300

Tower

43 kt

45,000

Nevada Test Site

May 8, 1953

Encore

2,425

Air

27 kt

41,000

Nevada Test Site

May 19, 1953

Harry

300

Tower

32 kt

43,000

Nevada Test Site

May 25, 1953

Grable

524

Gun

15 kt

38,000

Nevada Test Site

June 4, 1953

Climax

1,334

Air

61 kt

43,000

Nevada Test Site

(Table continues on the next page.)

56 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 (Continued)

Height of

Ck)ud top,

Date

Name

burst, ft

Type

Yield

Location

CasUe

Feb. 28, 1954

Bravo

Surface

15 Mt

114,000

Bikini Atoll

Mar. 26, 1954

Romeo

Barge

Bikini Atoll

Apr. 6, 1954

Koon

Surface

100 kt

Bikini Atoll

Apr. 25, 1954

Union

Barge

Bikini Atoll

May 4, 1954

Yankee

Barge

Bikini Atoll

May 13, 1954

Nectar

Barge

Enewetak Atoll

Teapot

Feb. 18, 1955

Wasp

762

Air

1 kt

22,000

Nevada Test Site

Feb. 22, 1955

Moth

300

Tower

2kt

25,000

Nevada Test Site

Mar. 1, 1955

Tesla

300

Tower

7kt

30,000

Nevada Test Site

Mar. 7, 1955

Turk

500

Tower

43 kt

44,000

Nevada Test Site

Mar. 12, 1955

Hornet

300

Tower

4kt

35,000

Nevada Test Site

Mar. 22, 1955

Bee

500

Tower

8kt

40,000

Nevada Test Site

Mar. 23, 1955

Ess

-67

Underground

1 kt

12,000

Nevada Test Site

Mar. 29, 1955

Apple I

500

Tower

14 kt

32,000

Nevada Test Site

Mar. 29, 1955

Wasp Prime

740

Air

3kt

32,000

Nevada Test Site

Apr. 6, 1955

36,620
(mean sea
level)

Air

3kt

55,000

Nevada Test Site

Apr. 9, 1955

Post

300

Tower

2kt

16,000

Nevada Test Site

Apr. 15, 1955

Met

400

Tower

22 kt

40,000

Nevada Test Site

May 5, 1955

Apple II

500

Tower

29 kt

43,000

Nevada Test Site

May 15, 1955

Zucchini

500

Tower

28 kt

35,000

Nevada Test Site

Wigwam

May 14, 1955

Wigwam

-2,000

Underwater

30 kt

29°N126°W

Safety experiment

Jan. 18, 1956

Surface

Nevada Test Site

Redwing

May 4, 1956

Lacrosse

Surface

20 kt

Enewetak Atoll

May 20, 1956

Cherokee

4,320

Air

Several
megatons

Bikini Atoll

May 27, 1956

Zuni

Surface

Bikini Atoll

May 30, 1956

Erie

300

Tower

Enewetak Atoll

June 6, 1956

Seminole

Surface

Enewetak Atoll

June 11, 1956

Flathead

Barge

Bikini Atoll

June 11, 1956

Blackfoot

200

Tower

Enewetak Atoll

June 16, 1956

Osage

680

Air

Enewetak Atoll

June 25, 1956

Dakota

Barge

Bikini Atoll

July 8, 1956

Apache

Barge

Enewetak Atoll

July 10, 1956

Navajo

Barge

Bikini Atoll

July 20. 1956

Tewa

Barge

Bikini Atoll

July 21, 1956

Huron

Barge

Enewetak Atoll

Plumbbob

May 28, 1957

Boltzmann

500

Tower

12kt

33,000

Nevada Test Site

June 2, 1957

Franklin

300

Tower

140 kt

17,000

Nevada Test Site

June 5, 1957

Lassen

500

Balloon

0.5 kt

7,000

Nevada Test Site

June 18, 1957

Wilson

500

Balloon

10 kt

35,000

Nevada Test Site

June 24, 1957

Priscilla

700

Balloon

37 kt

43,000

Nevada Test Site

Julys, 1957

Hood

1,500

Balloon

74 kt

48,000

Nevada Test Site

July 15, 1957

Diablo

500

Tower

17kt

32,000

Nevada Test Site

July 19, 1957

John

20,000
(mean sea
level)

Rocket

~2kt

44,000

Nevada Test Site

July 24, 1957

Kepler

500

Tower

10 kt

28,000

Nevada Test Site

July 25, 1957

Owens

500

Balloon

9.7 kt

35,000

Nevada Test Site

July 26, 1957

Pascal A

Underground

Slight

6,000

Nevada Test Site

Aug. 7, 195 7

Stokes

1,500

Balloon

19 kt

37,000

Nevada Test Site

Aug. 18, 1957

Shasta

500

Tower

17 kt

32.000

Nevada Test Site

WORLDWIDE FALLOUT 5 7

TABLE 1 (Continued)

Height of

Cloud top.

Date

Name

burst, ft

Type

Yield

Location

Plumbbob (Continued)

Aug. 23, 1957

Doppler

1,500

Balloon

11 kt

38,000

Nevada Test Site

Aug. 30, 1957

Franklin Prime

750

Balloon

4.7 kt

32,000

Nevada Test Site

Aug. 31, 1957

Smoky

700

Tower

44 kt

38,000

Nevada Test Site

Sept. 2, 1957

Galileo

500

Tower

11 kt

37,000

Nevada Test Site

Sept. 6, 1957

Wheeler

500

Balloon

197 tons

17,000

Nevada Test Site

Sept. 6, 1957

Coulomb B

Surface

0.3 kt

18,000

Nevada Test Site

Sept. 8, 1957

Laplace

750

Balloon

1 kt

20,000

Nevada Test Site

Sept. 14,1957

lizeau

500

Tower

11 kt

40,000

Nevada Test Site

Sept. 16, 1957

Newton

1,500

Balloon

12 kt

32,000

Nevada Test Site

Sept. 23, 1957

Whitney

500

Tower

19 kt

30,000

Nevada Test Site

Sept. 28, 1957

Charleston

1,500

Balloon

12 kt

32,000

Nevada Test Site

Oct. 7, 195 7

Morgan

500

Balloon

8kt

40,000

Nevada Test Site

Safety experiment

Dec. 9, 1957

Coulomb C

Surface

0.5 kt

Nevada Test Site

Hardtack— Phase I

Apr. 28, 1958

Yucca

86,000

Balloon

12°37'N

May 5, 1958

Cactus

Surface

163°orE

May 11, 1958

Fii

Barge

Lnewetak Atoll

May 11, 1958

Butternut

Barge

Bikini Atoll

May 12, 1958

Koa

Surface

Enewetak .Atoll

May 16, 1958

Wahoo

-500

Underwater

Enewetak Atoll

May 20, 1958

Holly

Barge

Enewetak Atoll

May 21, 1958

Nutmeg

Barge

Bikini Atoll

May 26. 1958

Yellowwood

Barge

Enewetak Atoll

May 26, 1958

Magnolia

Barge

Enewetak Atoll

May 30, 1958

Tobacco

Barge

Enewetak Atoll

May 31, 1958

Sycamore

Barge

Bikini Atoll

June 2, 1958

Rose

Barge

Enewetak Atoll

Junes, 1958

Umbrella

-150

Underwater

Enewetak Atoll

June 10, 1958

Maple

Barge

Bikini Atoll

June 14, 1958

Aspen

Barge

Bikini Atoll

June 14, 1958

Walnut

Barge

Enewetak Atoll

June 18, 1958

Linden

Barge

Enewetak Atoll

June 27, 1958

Redwood

Barge

Bikini Atoll

June 27, 1958

Elder

Barge

Enewetak AtoU

June 28, 1958

Oak

Barge

Enewetak Atoll

June 29, 1958

Hickory .

Barge

Bikini Atoll

July 1, 1958

Sequoia

Barge

Enewetak Atoll

July 2, 1958

Cedar

Barge

Bikini Atoll

July 5, 1958

Dogwood

Barge

Enewetak Atoll

July 12, 1958

Poplar

Barge

Bikini Atoll

July 22, 1958

Olive

Barge

Enewetak Atoll

July 26, 1958

Pine

Barge

Enewetak Atoll

Aug. 1, 1958

Teak

25 2,000

Rocket

Megaton
range

Johnston Island

Aug. 12, 1958

Orange

141,000

Rocket

Megaton
range

Johnston Island

Hardtack — Phase 11

Sept. 12, 1958

Otero

-480

Underground

38 tons

9,000

Nevada Test Site

Sept. 17, 1958

Bernalillo

-456

Underground

15 tons

7,500

Nevada Test Site

Sept. 19, 1958

Kddy

500

Balloon

83 tons

11,000

Nevada Test Site

Sept. 21,1958

Luna

-484

Underground

1.5 tons

Low diffuse cloud

Nevada Test Site

Sept. 26, 1958

Valencia

-484

Underground

2 tons

5,500

Nevada Test Site

Sept. 28, 1958

Mars

Underground

13 tons

Low diffuse cloud

Nevada Test Site

Sept. 29, 1958

Mora

1,500

Balloon

2kt

18,500

Nevada Test Site

Oct. 5, 1958

Hidalgo

377

Balloon

7 7 tons

12,000

Nevada Test Site

Oct. 5, 1958

Colfax

-350

Underground

5.5 tons

5,500

Nevada Test Site

Oct. 8, 1958

Tamalpais

-330

Underground

72 tons

Low diffuse cloud

Nevada Test Site

(Table continues on the next page.)

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 (Continued)

Height of

Cloud to(

Date

Name

burst, ft

Type

Yield

Location

Hardtack — Phase II (Continued)

Oct. 10. 1958

Quay

100

Tower

79 tons

10 000

Nevada Test Site

Oct. 13, 1958

Lea

1,500

Balloon

1.4 kt

17,000

Nevada Test Site

Oct. 14, 1958

Neptune

-98.5

Underground

115 tons

1 1 ,000

Nevada Test Site

Oct. 15, 1958

Hamilton

Tower

1.2 tons

6,000

Nevada Test Site

Oct. 16, 1958

Dona Ana

450

Balloon

37 tons

11,000

Nevada Test Site

Oct. 17, 1958

Vesta

Surface

24 tons

10,000

Nevada Test Site

Oct. 18, 1958

Rio Arriba

72.5

Tower

90 tons

13,500

Nevada Test Site

Oct. 22, 1958

Socorro

1,450

Balloon

6 kt

26,000

Nevada Test Site

Oct. 22, 1958

WrangeU

1,500

Balloon

115 tons

10,000

Nevada Test Site

Oct. 22, 1958

Rushmore

500

Balloon

188 tons

11,500

Nevada Test Site

Oct. 24, 1958

Catron

72.5

Tower

21 tons

8,500

Nevada Test Site

Oct. 24, 1958

Juno

Surface

1.7 tons

5,500

Nevada Test Site

Oct. 26, 1958

Ceres

Tower

0.7 tons

6,000

Nevada Test Site

Oct. 26, 1958

San ford

1,500

Balloon

4.9 tons

26,000

Nevada Test Site

Oct. 26, 1958

De Baca

1,500

Balloon

2.2 kt

17,500

Nevada Test Site

Oct. 27, 1958

Chavez

5 2.5

Tower

0.6 tons

6,500

Nevada Test Site

Oct. 29, 1958

tvans

-848

Underground

55 tons

Nevada Test Site

Oct. 29, 1958

Humboldt

Tower

7.8 tons

7,500

Nevada Test Site

Oct. 30, 1958

Santa 1 e

1,500

Balloon

1.3 kt

18,000

Nevada Test Site

Oct. 30, 1958

Blanca

-835

Underground

19kt

7,700

Nevada Test Site

Oct. 30, 1958

Titania

Tower

0.2 tons

6,000

Nevada Test Site

Argus

Aug. 27, 1958

Argus- 1

-300 miles

Rocket

1-2 kt

38°S 12°W

Aug. 30, 1958

Argus-2

-300 miles

Rocket

1-2 kt

50°S 8°W

Sept. 6, 1958

Argus-3

-300 mUes

Rocket

1-2 kt

50'S 10°W

Continental (1)

Sept. 15, 1961

Antler

-1,319

Underground

2.4 kt

Low diffuse cloud

Nevada Test Site

Sept. 16, 1961

Shrew

Underground

Low

diffuse cloud

Nevada Test Site

Oct. 10, 1961

Chena

Underground

Low

diffuse ch

3Ud

Nevada Test Site

Oct. 29, 1961

Mink

Underground

Low

diffuse cloud

Nevada Test Site

Dec. 3, 1961

lisher

-1,193

Underground

13.5 kt

Low

diffuse cloud

Nevada Test Site

Dec. 10, 1961

Gnome

-1,184

Underground

3 ± 1 kt

Low

diffuse cl(

DUd

Nevada Test Site

Dec. 13, 1961

Mad

-594

Underground

430 tons

Low

diffuse cl<

jud

Nevada Test Site

Dec. 17, 1961

Ringtail

Underground

Low

diffuse cloud

Nevada Test Site

Dec. 22, 1961

leather

Underground

Low

diffuse cloud

Nevada Test Site

Continental (II)

Jan. 9, 1962

Stoat

Underground

4.5 kt

Low diffuse ch

jud

Nevada Test Site

Jan. 30, 1962

Doormouse

Underground

Low

diffuse cl(

jud

Nevada Test Site

Feb. 9, 1962

Armadillo

-786

Underground

6.6 kt

Low

diffuse cli

jud

Nevada Test Site

Feb. 15, 1962

Hardhat

-950

Underground

5.9 kt

Low diffuse ch

3Ud

Nevada Test Site

leb. 19, 1962

Chinchilla

-504

Underground

1.8 kt

Low

diffuse cl(

3Ud

Nevada Test Site

Feb. 24, 1962

Platypus

Underground

Low

diffuse ch

jud

Nevada Test Site

Mar. 5, 1962

Danny Boy

-110

Underground

430 tons

Low

diffuse ch

3Ud

Nevada Test Site

Mar. 6, 1962

Frmine

Underground

Low

diffuse ch

jud

Nevada Test Site

Mar. 8, 1962

Brazos

Underground

7.8 kt

Low

diffuse cl(

3Ud

Nevada Test Site

Mar. 31, 1962

Chinchilla 11

Underground

Low

diffuse cloud

Nevada Test Site

Apr. 14, 1962

Platte

Underground

1.7kt

Low

diffuse ch

jud

Nevada Test Site

May 12, 1962

Aardvark

-1,444

Underground

37kt

Low

diffuse ch

aud

Nevada Test Site

May 19, 1962

tel

Underground

Low

diffuse ch

aud

Nevada Test Site

June 6, 1962

Packrat

Underground

Low

Low diffuse ch

jud

Nevada Test Site

June 13, 1962

Des Moines

Underground

Low

Low diffuse ch

jud

Nevada Test Site

June 21, 1962

Daman 1

Underground

Low

diffuse ch

oud

Nevada Test Site

June 27, 1962

Haymaker

Underground

56 kt

Low diffuse cl(

jud

Nevada Test Site

June 28, 1962

MarshmaUow

Underground

Low

Low diffuse ch

oud

Nevada Test Site

July 6, 1962

Sedan

-635

Underground

100 kt

12,000

Nevada Test Site

July 7, 1962

Little I eller I

SlighUy
above
ground

Surface

Low

8,000

Nevada Test Site

July 11, 1962

Johnie Boy

Shallow
depth

Underground

500 tons

11,000

Nevada Test Site

WORLDWIDE FALLOUT 39

TABLE 2 United Kingdom Nuclear Detonations

Date

Name

Type

Yield

Location

Oct. 3, 1952

Hurricane

Ship

Kiloton range

Monte Bello Islands

Oct. 14, 1953

Totem

Tower

Kiloton range

Tests held at Emu Field,

Oct. 26, 1953

Totem

Tower

Kiloton range

300 miles northwest
ofWoomera

May 16, 1956

Mosaic

Tower

Kiloton range

Monte Bello Islands

June 19, 1956

Mosaic

Tower

Kiloton range

Monte Bello Islands

Sept. 27, 1956

Buffalo

Tower

Kiloton range

Marahnga

Oct. 4, 1956

Buffalo

Surface

Low yield

MaraUnga

Oct. 11, 1956

Buffalo

Air drop

Low yield

Marahnga

Oct. 22, 1956

Buffalo

Tower

Kiloton range

MaraUnga

May 15, 1957

Grapple

Air drop

Megaton range

Christmas Island Area

May 31, 1957

Grapple

Air drop

Megaton range

Christmas Island Area

June 19, 1957

Grapple

Air drop

Megaton range

Christmas Island Area

Sept. 14, 1957

Antler

Tower

Low yield

MaraUnga

Sept. 25, 1957

Antler

Tower

Kiloton range

MaraUnga

Oct. 9, 195 7

Antler

Balloon

Kiloton range

Maralinga

Nov. 8, 195 7

Grapple

Air drop

Megaton range

Christmas Island Area

Apr. 28, 1958

Grapple

Air drop

Megaton range

Christmas Island Area

Aug. 22, 1958

Grapple

Balloon

Kiloton range

Christmas Island Area

Sept. 2, 1958

Grapple

Air drop

Megaton range

Christmas Island Area

Sept. 11,1958

Grapple

Air drop

Megaton range

Christmas Island Area

Sept. 23, 1958

Grapple

Balloon

Kiloton range

Christmas Island Area

Mar. 1, 1962

Pampas

Underground

Low

Nevada Test Site

July 17, 1964

Underground

Low

Nevada Test Site

Sept. 10, 1965

Underground

Low to intermediate

Nevada Test Site

nation and include the yield of each device. During the course of nuclear weapons testing
from 1945 through 1976, it has been estimated by Harley (1975) and updated by using
announced nuclear tests that approximately 230 Mt of fission yield were introduced into
the atmosphere, which produced approximately 360 kCi of 239,240py ^^ lesser
amounts of other transuranic elements.

Prior to the detonation of the first thermonuclear device (Mike) in 1952, atmospheric
injections were confined mainly to the troposphere, and the mass of most of the
transuranic isotopes was lower than about 243. In debris from the Mike, which was
detonated at the Enewetak Atoll on Nov. 1, 1952, transuranium elements with masses
through 255 were observed. The much higher neutron yield of the Mike and subsequent
fusion devices than that of earlier fission devices permitted very substantial multiple
neutron capture by uranium, which allowed production of the very heavy elements. In
the detonation process, multiple neutron capture by ^^*U results in the production of
extremely neutron-rich products, the beta decay of which produces nuclides along the
line of greatest stability. Table 6 shows the relative abundance of the transuranium
isotopes that were formed in the Mike test (Diamond etal., 1961) as well as those
measured in fallout debris. The isobars of significant half-Ufe are shown together with

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Union of Soviet Socialist Republics Nuclear Detonations

Date

Type

Yield

Cloud
top, ft

Location

Aug. 29, 1949

Oct. 3, 1951*
Oct. 22, 1951*

Aug. 12, 1953
Aug. 23, 1953

Oct. 26, 1954*

Aug. 4, 1955*
Sept. 24, 1955*
Nov. 10, 1955*
Nov. 23, 1955* Air

Mar. 21, 1956*
Apr. 2, 1956*
Aug. 24, 1956
Aug. 30, 1956
Sept. 2, 1956
Sept. 10, 1956
Nov. 17, 1956

Jan. 19, 1957
Mar. 8, 1957
Apr. 3, 1957
Apr. 6, 1957
Apr. 10, 1957
Apr. 12, 1957
Apr. 16, 1957
Aug. 22, 1957
Sept. 9, 1957*
Sept. 24, 195 7
Oct. 6, 1957
Oct. 10, 1957
Dec. 28, 1957

Feb. 23, 1958
Feb. 27, 1958
Feb. 27, 1958
Mar. 14, 1958
Mar. 14, 1958
Mar. 15, 1958
Mar. 20, 1958
Mar. 21, 1958
Mar. 22, 1958
Sept. 30, 1958
Sept. 30, 1958
Oct. 2, 1958
Oct. 5, 1958
Oct. 10, 1958
Oct. 12, 1958
Oct. 15, 1958
Oct. 18, 1958
Oct. 19,1958

Thermonuclear
Fission

Megaton range

<1 Mt

Large

Substantial size
Moderate intensity
Megaton range
Thermonuclear
Small explosion

Megaton range

Large

Below megaton range

Small range

Medium range
Moderate to high
Moderate to high
Moderate

Relatively largef

Largef

Small

U.S.S.R.

U.S.S.R.
U.S.S.R.

U.S.S.R.

U.S.S.R.
U.S.S.R.
U.S.S.R.
U.S.S.R.

U.S.S.R.

Siberia

U.S.S.R.

Siberia

Arctic

U.S.S.R.

Arctic

Siberia

Arctic

Siberia

Arctic

WORLDWIDE FALLOUT

TABLE 3 (Continued)

Cloud

Date

Type

Yield

top, ft Location

Oct. 20, 1958

Largef

Arctic

Oct. 22, 1958

Largej

Arctic

Oct. 24, 1958

Largef

Arctic

Oct. 25, 1958

Relatively large

Arctic

Nov. 1, 1958

Relatively low

Siberia

Nov. 3, 1958

Relatively low

Siberia

Sept. 1, 1961

Atmospheric

Intermediate

Semipalarinsk

Sept. 4, 1961

Atmospheric

Low

Semipalatinsk

Sept. 5, 1961

Atmospheric

Low to intermediate

Semipalatinsk

Sept. 6, 1961

Atmospheric

Low to intermediate

East of StaUngrad

Sept. 10, 1961

Atmospheric

Several megatons

Novaya Zemlya

Sept. 10, 1961

Atmospheric

Low to intermediate

Novaya Zemlya

Sept. 12, 1961

Atmospheric

Several megatons

Novaya Zemlya

Sept. 13, 1961

Atmospheric

Low to intermediate

Semipalarinsk

Sept. 13, 1961

Atmospheric

Low to intermediate

Novaya Zemlya

Sept. 14, 1961

Atmospheric

Several megatons

Novaya Zemlya

Sept. 16, 1961

Atmospheric

Order of a megaton

Novaya Zemlya

Sept. 17, 1961

Atmospheric

Intermediate

Semipalatinsk

Sept. 18, 1961

Atmospheric

Order of a megaton

Novaya Zemlya

Sept. 20, 1961

Atmospheric

Order of a megaton

Novaya Zemlya

Sept. 22, 1961

Atmospheric

Order of a megaton

Novaya Zemlya

Oct. 2, 1961

Atmospheric

Order of a megaton

Novaya Zemlya

Oct. 4, 1961

Atmospheric

Several megatons

Novaya Zemlya

Oct. 6, 1961

Atmospheric

Several megatons

Novaya Zemlya

Oct. 8, 1961

Atmospheric

Low

Novaya Zemlya

Oct. 12, 1961

Atmospheric

Low to intermediate

Semipalatinsk

Oct. 20, 1961

Atmospheric

Several megatons

> 12,000 Novaya Zemlya

Oct. 23, 1961

Atmospheric

About 25 Mt

Novaya Zemlya

Oct. 23, 1961

Underwater

Low

South of Novaya Zemlya

Oct. 25, 1961

Atmospheric

Intermediate to high

Novaya Zemlya

Oct. 27, 1961

Atmospheric

Low to intermediate

Novaya Zemlya

Oct. 30, 1961

Atmospheric

55 to60Mt

>12,000 Novaya Zemlya

Oct. 31, 1961

Atmospheric

Several megatons

Novaya Zemlya

Oct. 31, 1961

Atmospheric

Intermediate to high

Novaya Zemlya

Nov. 2, 1961

Atmospheric

Low to intermediate

Novaya Zemlya

Nov. 2, 1961

Atmospheric

Low to intermediate

Novaya Zemlya

Nov. 4, 1961

Atmospheric

Several megatons

Novaya Zemlya

Feb. 2, 1962

Underground

Semipalatinsk

Aug. 5, 1962

Atmospheric

30 Mt

Novaya Zemlya

Aug. 7, 1962

Atmospheric

Low

Central Siberia

Aug. 10, 1962

Atmospheric

<1 Mt

Novaya Zemlya

Aug. 20, 1962

Atmospheric

Order of several megatons

Novaya Zemlya

Aug. 22, 1962

Atmospheric

Low megaton

Novaya Zemlya

Aug. 25, 1962

Atmospheric

Order of several megatons

Novaya Zemlya

Aug. 25, 1962

Atmospheric

Low

Semipalatinsk

Aug. 27, 1962

Atmospheric

Several megatons

Novaya Zemlya

Sept. 2, 1962

Atmospheric

Intermediate

Novaya Zemlya

Sept. 8, 1962

Atmospheric

Megaton

Novaya Zemlya

Sept. 15, 1962

Atmospheric

Several megatons

Novaya Zemlya

Sept. 16, 1962

Atmospheric

Several megatons

Novaya Zemlya
(Table continues on the next page.)

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLES (Continued)

Cloud

Date

Type

Yield top, ft Location

Sept. 18, 1962

Atmospheric

Few megatons

Novaya Zemlya

Sept 19, 1962

Atmospheric

Multimegatons

Novaya Zemlya

Sept. 21, 1962

Atmospheric

Few megatons

Novaya Zemlya

Sept. 25, 1962

Atmospheric

Multimegatons

Novaya Zemlya

Sept. 27, 1962

Atmospheric

<30Mt

Novaya Zemlya

Oct. 7, 1962

Atmospheric

Intermediate

Novaya Zemlya

Oct. 14, 1962

Atmospheric

Low

Semipalatinsk

Oct. 22, 1962

High altitude

Few hundred kilotons

Central Asia

Oct. 22, 1962

Atmospheric

Several megatons

Novaya Zemlya

Oct. 27, 1962

Atmospheric

Intermediate

Novaya Zemlya

Oct. 28, 1962

High altitude

Intermediate

Central Asia

Oct. 28, 1962

Atmospheric

Low

Semipalatinsk

Oct. 29, 1962

Atmospheric

Intermediate

Novaya Zemlya

Oct. 30, 1962

Atmospheric

Intermediate

Novaya Zemlya

Nov. 1, 1962

High altitude

Intermediate

Central Asia

Nov. 1, 1962

Atmospheric

Intermediate

Novaya Zemlya

Nov. 3, 1962

Atmospheric

Intermediate

Novaya Zemlya

Nov. 3, 1962

Atmospheric

Intermediate

Novaya Zemlya

Nov. 4, 1962

Atmospheric

Intermediate

Semipalatinsk

Nov. 17, 1962

Atmospheric

Low

Semipalatinsk

Dec. 18, 1962

Atmospheric

Intermediate

Novaya Zemlya

Dec. 18, 1962

Atmospheric

Intermediate

Novaya Zemlya

Dec. 20, 1962

Atmospheric

Low

Novaya Zemlya

Dec. 22, 1962

Atmospheric

Intermediate

Novaya Zemlya

Dec. 24, 1962

Atmospheric

About 20 Mt

Novaya Zemlya

Dec. 25, 1962

Atmospheric

Few megatons

Novaya Zemlya

*Date of announcement not necessarily shot date.

|Mr. McCone on Oct. 24, 1958, announced that these seven tests had a high yield, meaning that
each had an explosive power equal to millions of tons of TNT.

their decay properties and half-lives. Their total activities are normalized to ^^^Pu to
permit comparison of their relative production rates. The radioisotope ^"^^Pu (a
beta-decay isotope with a 14.7-yr half-life), which decays to ^"^^Am, is the most
abundant activity.

The distribution on the earth's surface of transuranic elements produced during
nuclear weapons testing depends on whether the debris is contained in the stratosphere or
troposphere. Such partitioning is dependent on many things, including yield of the
device, the "burst" height, and the height of the troposphere. Figure 1 shows the percent
of debris in the troposphere as a function of the yield of a nuclear device (Ferber, 1964).
From these data it can readily be seen that devices in the low-kiloton range place most of
the debris in the troposphere, whereas weapons in the megaton range inject most of the
debris into the stratosphere. Prior to 1952 all the nuclear explosions were in the
low-kiloton range; the residence time for this debris is about 20 to 40 days (Stewart,
Crooks, and Fisher, 1955; United Nations, 1964; Krey and Krajewski, 1970a).
After 1952 numerous multimegaton tests took place in which most of the debris was
injected into the lower stratosphere where the residence half-time is about 1 yr (Thomas
etal., 1970).

WORLDWIDE FALLOUT

TABLE 4 Republic of France Nuclear Detonations, 1960 to 1971

Date of detonation

Type

Yield

Location

Feb. 13, 1960

Tower

60 to 70 kt

Reggan, Algeria

Apr. 1, 1960

Surface

Small

Reggan, Algeria

Dec. 27, 1960

Tower

Small

Reggan, Algeria

Apr. 25, 1961

Tower

Small

Reggan, Algeria

Nov. 7, 1961

Underground

Weak

Sahara Desert

May 1, 1962

Underground

Middle

Sahara Desert

Mar. 18, 1963

Underground

Weak

Sahara Desert

Mar. 30, 1963

Underground

Weak

Sahara Desert

Oct. 20, 1963

Underground

Middle

Sahara Desert

Feb. 14, 1964

Underground

Weak

Sahara Desert

June 15, 1964

Underground

Weak

Sahara Desert

Nov. 28, 1964

Underground

Weak

Sahara Desert

Feb. 27, 1965

Underground

Middle

Sahara Desert

May 30, 1965

Underground

Weak

Sahara Desert

Oct. 1, 1965

Underground

Weak

Sahara Desert

Dec. 1, 1965

Underground

Weak

Sahara Desert

Feb. 16, 1966

Underground

Weak

Sahara Desert

July 2, 1966

Barge

Small

Mururoa Island

July 19, 1966

Air

Small

Mururoa Island

Sept. 11, 1966

Balloon

Small

Mururoa Island

Sept. 24, 1966

Barge

Small

Fangataufa Island

Oct. 4, 1966

Barge

200 to 300 kt

Mururoa Island

June 5, 1967

Balloon

Small

Mururoa Island

June 27, 1967

Balloon

Small

Mururoa Island

July 2, 1967

Balloon

Small

Mururoa Island

July 7, 1968

Balloon

Small

Mururoa Island

July 15. 1968

Balloon

0.5 Mt

Mururoa Island

Aug. 3, 1968

Balloon

Low to intermediate

Mururoa Island

Aug. 24, 1968

Balloon

Low megaton (first H bomb)

Fangataufa Island

Sept. 8, 1968

Balloon

Low megaton

Mururoa Island

May 15, 1970

Balloon

Low

Mururoa Island

May 22, 1970

Balloon

Intermediate

Mururoa Island

May 30, 1970

Balloon

Intermediate (megaton range)

Fangataufa Island

June 24, 1970

Balloon

Low

Mururoa Island

July 3, 1970

Balloon

Intermediate (1 Mt)

Mururoa Island

July 27, 1970

Balloon

Low

Mururoa Island

Aug. 2, 1970

Balloon

Low to intermediate

Fangataufa Island

Aug. 6, 1970

Balloon

Intermediate

Mururoa Island

Junes, 1971

Balloon

Low

Mururoa Island

June 12, 1971

Balloon

Intermediate

Mururoa Island

July 4, 1971

Balloon

Low

Mururoa Island

Aug. 8, 1971

Balloon

Low

Mururoa Island

Aug. 14, 1971

Balloon

Intermediate

Mururoa Island

June 25, 1972

Low

South Pacific

July 1, 1972

Low

South Pacific

July 29, 1972

Low

South Pacific

(Table continues on the next page.)

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 (Continued)

Date of detonation

Type

Yield

Location

July 21, 1973
July 28, 1973
Aug. 18, 1973
Aug. 24,1973
Aug. 28, 1973

June 16, 1974
July 7, 1974
July 17, 1974
July 25, 1974
Aug. 15, 1974
Aug. 24, 1974
Sept. 15, 1974

Low
Low
Low
Low
Low

Low

Unknown

Low

South Pacific
South Pacific
South Pacific
South Pacific
South Pacific

South
South
South
South
South
South
South

Pacific
Pacific
Pacific
Pacific
Pacific
Pacific
Pacific

TABLE 5 People's Republic of China Nuclear Detonations

Date of detonation

Type

Yield Location

Oct. 16, 1964

Tower

-20 kt Lop Nor

May 14, 1965

Air drop

>20kt

May 9, 1966

Air drop

200 to 500 kt

Oct. 27, 1966

MissUe,notHH§*

<20kt

Dec. 28, 1966

Tower

300 kt

June 17, 1967

Air drop

3 Mt

Dec. 24, 1967

Air drop

15 to 25 kt

Dec. 27, 1968

Air drop

3 Mt

Sept. 22, 1969

Underground

-25 kt

Sept. 29, 1969

Air drop

3 Mt

Oct. 14, 1970

Air drop

3 Mt

Nov. 18, 1971

Tower

-20 kt

Jan. 7, 1972

Atmospheric

<20kt

Mar. 18, 1972

Atmospheric

20 to 200 kt

June 27, 1973

Missile?

1 to 3 Mt?

June 17, 1974

Atmospheric

1 Mt

Oct. 27, 1975

Underground

20 kt

Jan. 23, 1976

Atmospheric

<20kt

Sept. 25, 1976

Atmospheric

200 kt

Oct. 27, 1976

Underground

200 kt

Nov. 17, 1976

Atmospheric

4 Mt

*HH§ stands for launching by missile to high altitude.

WORLDWIDE FALLOUT

TABLE 6 Relative Abundance of Heavy Elements Produced During Mike Test
Compared with That Measured in Worldwide Fallout (mass abundances at time = 0)

Isobar

Type
decay

t.^. yr

Relative abundance, atoms

Relative abund:
Mike

mce, activity

Mass No.

Mike

Fallout

239

Plutonium

2.44 X 10*

1.0

240

Plutonium

6.54 X 10*

0.363

0.18

1.35

0.669

241

Plutonium

-0

0.039

0.013

242

Plutonium

3.87 X 10'

1.9x10-'

0.004

1.2 X 10-'

2.53 X 10-*

243

Americium

7.37 X 10'

2.1 X 10-'

6.9x10-'

244

Plutonium

8.3 X 10'

1.2 X 10-'

3.5 X 10-'

245

Curium

8.5 X 10'

1.2 X 10-*

3.6 X 10-*

246

Curium

4.76 X 10'

4.8 X 10-'

2.4x10-*

247

Curium

1.54x10'

3.9 X 10-*

6.2x10-'

248

Curium

3.5 X 10»

1.2 X 10-*

8.4 X 10-*

249

Berkelium

-(3

0.852

1.1 X 10-'

3.2 X 10-'

250

Curium

SF*

1.13 X 10*

~3x ll-«

-6.5 xlO-«

251

Californium

9.0 X 10'

-1.4 X 10-'

-3.8 X 10-'

252

Californium

2.63

1.0 X 10-'

9.3 X 10-*

253

Californium

-13

0.049

5. Ox 10-'"

2.5 X 10-*

254

Californium

-0

0.164

5.0 X 10-"

7.4 X 10-*

255

Einsteinium

-0

0.107

4.0 X 10-"

9.1 X 10-*

*SF, spontaneous fission.

102

101

I
a.
</)

O
a.
O
a:

IQO

10-

o
cc

10-2

T — I I I 1 1 ny —■ i_ 1^1 1 1 IM|^^^ 1 I I ""I ' — ' ' I ' I

Typical

Probable
maximum

J ' I I I iiri

I I I

UiA

50 100
— Kilotons-

500 1

TOTAL YIELD

5 10
-Megatons -

50 100

Fig. I Percent of total activity initially injected in the troposphere as a function of total
yield for air bursts in a tropical atmosphere.

66 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Figure 2 shows the distribution of radioactive fallout (in millicuries per 100 square
miles) between 2 and 35 days following the Mike detonation explosion in the Marshall
Islands (Machta, 1964). Tlrese results represent only that fractional amount of the debris
which was contained in the troposphere. Since the residence time in the troposphere is on
the order of 20 to 40 days, the deposition rate was quite rapid and was confined mainly
to the hemisphere where the test took place; the higher concentration was near the
latitude where the explosion occurred.

The distribution of debris of stratospheric origin is considerably different from that
of tropospheric origin. Most of the debris leaving the stratosphere does so through the
tropopause discontinuity, which occurs near the midlatitudes and is almost independent
of the latitude of the detonation. Empirical box models (Krey and Krajewski, 1970b),
which describe the movement of radioactivity from the upper to the lower stratosphere,
between hemispheres, from the stratosphere to the troposphere, and the deposition rate
on the earth's surface, have been developed and appear to be reasonably satisfactory.

The movement of radioactive debris in the troposphere is influenced by all the forces
of the weather. Rain and snow will scavenge radioactive particles, which will cause the
debris to be distributed unevenly on the earth's surface. Recently, measurements have
shown that the scavenging of radionuclides by cirrus cloud ice particles resulted in major
depletion of radionuclides from atmospheric layers of 1 .3 to 2.8 km thick at about 10 Vr.-
(Young, Wendell, and Wogman, 1975). The mixing of air masses as they move west to
east across the United States and are orographically Ufted over mountain ranges can
increase the ground-level concentration of radionuclides on the downwind side. This
effect is presumably due to the downwind mixing of high-level air, which contains higher
concentrations of both cosmogenic and nuclear-weapons-produced radionuclides. This
effect is shown in Fig. 3 where the atmospheric concentrations of ^Be and ' "'^Cs for a
period of iVj yr for Quillayute, Wash. (48°N, 125°W), Richland, Wash. (46°N, 119°W),
and Rocky Flats, Colo. (40°N, 106°W), are compared. Storm systems originating in the
Aleutians move air masses over the Quillayute sampling site which are orographically
lifted several thousand feet by the Cascade Mountain range before they descend to the
Ricliland site. The air mass is again lifted by the Rocky Mountain range before it descends
to the Rocky Flats sampling site. The average annual air concentrations of ^Be and ' ^^Cs
during 1973 through early 1975 were 2.1 to 2.4 times as great at Richland and 2.9 to 3.1
times as great at Rocky Flats as those at Quillayute (Thomas, 1972).

Production and Characteristics of Individual Transuranic Elements

Because of their methods of production, the relative abundances of the transuranium
elements are considerably different in nuclear detonations than in reactor operations. In
nuclear detonations neutron capture occurs in extremely rapid succession, producing
uranium or plutonium isotopes of very liigh mass which rapidly decay to form a spectrum
of transuranium elements. In this case there is no opportunity for the decay of the
various uranium isotopes, which could break the chain of successive neutron capture. In
the reactor production of radionuclides, the neutrons are captured only one at a time,
and the resulting product may decay before additional neutron capture. In Table 7 the
amounts of the various transuranium elements resulting from the Mike nuclear test are
compared with those which result from nuclear power generation. It is immediately
evident that the transuranium elements resulting from nuclear energy generation are
much higher relative to ^^^Pu, particularly in the region just below and above 239 than

WORLDWIDE FALLOUT

O 4*

u. -

^ 2

^ 5

>
O
Z

« E

J5 o

c c
.2 .2
o o
"o. "S-

X X

0 « ^

C >> i/^

01 ^3 ^

2 -a
ago.

E "^ w

■^ c r-

1 ^!^

r. -^ ^

C '^ ..

o ± ^

"^ 'e -

3 01 (J"

o S i

oj o ^

•B ^ .2

cj '^ 'S

.2 o) V

•^ '^

Cd W5 CJ

OS .Si 'e

w C

3 O

ir E o

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

QUILLAYUTE
1.0

CASCADE

MOUNTAINS

QUILLAYUTE,
1.0

ROCKY
MOUNTAINS

RICHLAND
2.3

ROCKY FLATS
3.0

Fig. 3 Relative concentrations of ^Be and ' ''Cs in air from late 1973 tlirough early
1975 normalized to 1 for Quillayute.

those from nuclear weapons testing. Comparable data for the production of isotopes of
mass greater than 246 by the nuclear industry were not available, but higlier
concentrations relative to ^^^Pu would be expected through perhaps mass 252.

Reported releases of transuranium elements from nuclear plant operations indicate
that, in general, these have been very small. Loss of material around the Rocky Flats
plant (Krey and Hardy, 1970) has resulted in some environmental contamination, and
elevated atmospheric concentrations have been observed through resuspension. Accidents
involving aircraft carrying nuclear weapons (Langliam, 1970) have resulted in the spread
of plutonium over limited areas; however, these appear to have resulted in rather minor
injections of plutonium into the atmosphere.

Since the testing of the first theriTionuclear device in 1952, substantial amounts of
fission products, as well as transuranium elements, have entered the stratosphere. These
injections result in the long-term, relatively slow deposition of radioactivity over the
entire surface of the earth. As a first approximation, the transuranium elements appear to
behave in their atmospheric transport in essentially the same manner as other fission
products. Figure 4 shows, for example, that the ratio of ' ^^Cs to 2 3 9,2 4 0pj^ ^^^ ^^^^

WORLDWIDE FALLOUT

TABLE 7 Relative Compositions of Transuranium

Elements from Power Reactors and Mike Shot

(values normalized to ^^^Pu)

Mass No.

Isotope

Power reactors*

Mike shot

238

Plutonium

44.1

0.015

239

Plutonium

240

Plutonium

1.65

1.35

241

Plutonium

306

Americium

242

Plutonium

Americium

4.3 X 10-'
5.64

1.2

xlO-'

243

Plutonium
Americium

11.2

6.9

X 10-'

244

Curium

15.11

3.5

X 10-'

245

Curium

0.22

3.6

xl0-'»

246

Curium

4.5 xlO-^

2.4

X lO-''

247

Curium

6.2

X 10"'

248

Curium

8.4

X 10-«

249

Berkelium

3.2

X 10-'

250

Curium

-6.5

X 10-«

251

Californium

-3.8

X 10-'

252

Californium

9.3

X 10-«

253

Californium

2.5

X 10""

254

Californium

7.4

X 10"*

255

Einsteinium

9.1

X 10-«

*Assuming 30,000 Mwd/ton exposure (Schneider, 1974).

0.05

0.01 —

0.005

Fig. 4 Activity ratio of ^^ ' '^ ""Pu/' ^ ''Cs in tropospheric air. •, Richland, Wash. A,
New York, N. Y. ■, Harwell, England.

70 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

reasonably constant in fallout since the U.S. and U.S.S.R. nuclear tests of 1961 and
1962. These data are based on measurements beginning in 1962 at Richland, Wash.
(46°N), and in later periods in New York, N. Y. (4rN), and Harwell, England (52°N).
The ^^^'^'*^Pu/^ ^^Cs ratio appears to be constant in most cases within the accuracy of
the measurements and thus tends to indicate a general constancy of the transuranium
production and atmospheric behavior relative to that of the fission products.

100

E 10

■D

>10 Mt

1 to 10 Mt

0.1 to 1 Mt

0.02 to 0.1 Mt

Harwel

Richland, Wash., Measurements -

1 —

I I I

YEAR

Fig. 5 Concentrations of '^^Cs in surface air at 46°N latitude since 1953. The
concentrations prior to 1962 were estimated by normalizing concentrations measured at
Harwell, En^and.

On the basis of observed '^^Cs concentrations at 46°N latitude since 1962 and an
extrapolation back to 1953 by normalizing Harwell, England, to Richland, Wash., ' ^^Cs
air concentrations during the period 1962 to 1964 (as indicated in Fig. 5), it should be
possible to obtain a good estimate of the airborne plutonium concentrations during this
entire period. Such extrapolations are, of course, subject to some uncertainty.

An atmospheric sampling program using high altitude aircraft has been conducted
since 1959 (Hardy, 1973). Sampling aircraft normally operate at four latitudes — 70°N,
35°N, 10°N, and 40°S. Sampling altitudes normally range from 15.000 to 70,00011.
Figure 6 shows the ratios of ^"^^Pu to^-^'^Pu, ^'"Pu to ^^'^Pu,and ^''^Pu to "^Pumair
at 70°N latitude as a function of time. It is evident that there is considerable variation in
these ratios which is undoubtedly associated with the type and energy of the weapon
responsible for the plutonium isotope production. There is a substantial increase in the
heavy-to4ight plutonium isotopes immediately following the 1961 and 1962 U.S.-
U.S.S.R. test series.

Figure 7 shows the concentrations of ^^^Pu and 239,240p|j j^om 1962 to the
present. These measurements, which were made near Richland, Wash., show that seasonal
variations in the ^^^ ,2 3 9,2 4 0pu ^^^.^ gjp^jigp ^q those of other nuclear-weapons-produced
radionuclides of stratospheric origin; maximum concentrations occur in the late spring.

WORLDWIDE FALLOUT 71

0.26

0.20

0.16

0.020 —

0.015

I-

K 0.010

240pu/239pu -

241PU/239PU _

0006

—

242pu/239pu -

0.005

—

0 004

/ V-,

N^-V

0.003

- \

^^^x -

0.002

1 1 1 1 1 1

1 1 1 1 1

I960

1962

1964
YEAR

1966

1968

1970

Fig. 6 Atom ratios of plutonium isotopes from an air column (15,000 to 17,000 ft
high) at 70° N latitude.

and minimum concentrations occur in the winter. The rate of decrease in the 2 3 9,2 40pjj
concentrations from 1963 through 1967 corresponded to a stratospheric half-residence
time of 10 to 11 months, which is similar to the half-residence times calculated from
measurements of other radionuclides of stratospheric origin. The 2 39,240pjj Qoncentra-
tions remained fairly constant from 1967 to 1972, primarily because of yearly injections
of plutonium by thermonuclear tests conducted by the Chinese at Lop Nor (44°N);the
contribution from the French tests in the South Pacific (23°S) may also have significance.
From 1962 througli 1965 the ^^^Puand '^^ '^'^^Pu in surface air at Richland, Wash.,
came primarily from the 1961 and 1962 U.S. and U.S.S.R. series. The ^^^Pu/^^^-^'^^Pu
activity ratio averaged about 0.020 in 1964. The activity ratio stayed almost the same in
1965, but, by the spring of 1966, it had increased to 0.042, which suggests that ^^^Pu
from the SNAP-9A burnup was present. The amount of SNAP-9A '^^^Pu present was
determined from the '■'^Pu concentrations and the ^^^pxi/^^^ •^'^^Pu activity ratios;the
activity ratio in debris from nuclear weapons tests was assumed to be 0.020. These
considerations indicate that the ^^^Pu in Richland air t>om 1967 to 1971 came largely
from SNAP-9A. From 1967 through 1969, the concentrations of SNAP-9A plutonium at
Richland remained fairly constant, which indicates that the '"'^Pu was being transferred
into the northern hemispheric lower stratosphere at a rate comparable to the rate at
which ^^^Pu was being deposited on the earth's surface. This suggests that a substantial
amount of ■^ '^^Pu was retained in the upper stratosphere, and its slow movement into the
lower stratosphere maintained a nearly constant level for about 2 yr. The fact that the
^''^Pu concentrations showed the usual seasonal variations typical of radionuclides of

12 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

1.0

0.1

0.01

£ 0.001

n
O

c
E

0.0001

A. A ,

^■■\(-^ " ^^■

<\ r\i

J \ I L

1 \ I \ r

238pu

\ I ■

•- A

U*^ '*' l\

J I L

J\J:

J I L

I-

o
z
o
o

1.0

0.1

0.01

0.001

iJ^

■V..A A. .A.

I I r

239,240pu

\ i

J L

lOr

1.0 r-

1 r

< 0.1

0.01

0.001

SNAP
burnup

I I \ r

238pu/239, 240 py

J L

1962 1964 1966 1968 1970 1972 1974 1976

YEAR

Fig. 7 Concentrations of ^^*Pu and ^^''^""Pu in surface air at Richland, Wash.

WORLDWIDE FALLOUT 73

5.0

1.0

a.
o

Q.
CO

0.1

0.01

O , Southern hemisphere, HASL
A , Northern hemisphere, HASL
hemisphere, Battelle

1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974 1975

YEAR

Fig. 8 Average yearly activity ^ ^ ^ Pu/

2 3 9.240

Pu ratios in surface air.

Stratospheric origin indicates that the transfer involved movement into the northern
stratosphere and then to the troposphere. The hemispheric yearly averages compared with
the yearly average at Richland, Wash., are shown in Fig. 8. Concentrations of SNAP-9A
^^^Pu in the northern hemisphere and at Ricliland have decreased rapidly since 1968 and
1969, respectively.

Similar changes in ground-level air concentrations were observed at other locations in
the northern and southern hemispheres (Hardy, 1976). These results indicate that the
stratospheric debris injected into the high stratosphere may not produce high concentra-
tions of the debris in ground-level air until 2 yr later. These ground-level concentrations
may in some regions remain nearly constant for about 2 yr before they begin to decrease.

Although not formed direcdy in the nuclear weapons detonation, considerable
amounts of ■^'' 'Am are present in fallout debris. This, of course, results from the decay of

24 1

2 39 ,240i

Pu. On the basis of the amount of ^•^'''^'♦"Pu in the atmosphere and the ratio of
^^^Pu to '"^'Pu observed in the Mike test, one can calculate the ^"^ 'Am as a function of
time in the atmosphere. The ^^' Pu and the '^'''Am ratios are plotted in Fig. 9 together
with the observed concentrations of ^'^'Am as measured from samples taken at a
monitoring station in Richland, Wash. It is evident that the airborne concentrations are in
reasonably good agreement with those calculated. Also, the ratio of ^"^ 'Am to 2 3 9,2 4 Op^
does increase, as would be expected, as the debris ages. On the basis of the yields of
transuranium elements, which were observed in the Mike tests, and the 2 3 9,2 4 Op^
updated inventory established by the Environmental Measurements Laboratory (EML),
the total amounts of the other transuranium elements that have entered the atmosphere
can be estimated. These values are shown in Table 8. For isotopes of mass greater than
244, the total atmospheric injections are in the range of hundredths to tens of curies, and
the total alpha-decay activity of all the transuranium elements of mass greater than 241 is
only about 1% of the ^^''•'^"Pu.

74 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

100

n
O

c
E
2 0.1

0.01

0.001

■^-llpu (based on 239,240pu using 240pu/239pu activity ratio
of 1.35 and 24lpij/239pu activity ratio of 63)

2''lPu measurements made
at Richland, Wash.

241 Pu (based on 239,240pu
using 240pu/239pu activity ratio
of 0.67 and 24lpu/239pu activity
ratio of 21

241 Am calculated from Mike 241 Pu
and assuming residence half-time of 1 yr

•241 Am calculated from
fallout 241 Pu assuming
residence half-time of 1 yr

^.i:

241 Am measured yearly
averages at Richland, Wash

I "

1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974 1975

YEAR

Fig. 9 Estimated concentrations of ^ " ' Pu and ^^ ' Am as compared with the measured
concentration of ^ "' Am in surface air since 1962 at Richland, Wash.

TABLE 8 Relative Abimdance and Estimated
Amounts of Transuranium Elements That
Have Been Injected into the Atmosphere

Activity abundance

Total injection,* kCi

^"Pu

154

^""Pu

1.35

209

24Ipy

9720

'*' Am

336t

242 p^j

1.2 X

10-'

0.19

^"'Am

6.9 X

10-'

1.07

244p^,

3.5 X

10-'

5.2 X 10-'

'*'Cm

3.6 X

10-*

5.6 X 10-'

^"♦Cm

2.4 X

10-*

3.7 X 10-'

'"'Cm

6.2 X

10-'

8.8 X 10-'

'"'Cm

8.4 X

10-'

1.3 X 10-'

'"'Bk

3.2 X

10-'

0.49

''"Cm

6.5 X

10-'

8.8 X 10-*

' " Cf

3.8 X

10-'

5.8 X 10-'

2 5 2^,-

9.3 X

10-*

1.4 X 10-'

2 5 3^,-

2.5 X

10-*

3.9 X 10-'

254cf

7.4 X

10-'

1.1 X 10-'

^'^Es

9.1 X

10-*

1.4 X 10-'

♦Assumes 360,000 Ci (360 kCi) ^ ' » -^ " o p^ atmo-
spheric injection with Mike ratios.

t Americium-241 fomied on total decay of

Pu.

>10 Mt -

1 to 10 Mt -

0.1 to 1 Mt -

0.02 to 0.1 Mt

WORLDWIDE FALLOUT 75

241pu/239,240pu

• , Monthly samples
O , Yearly averages

Denotations

United States

Union of Soviet Socialist Republics

— Chinese

I I I
I II
i JJ

I I
1 .1

I —

1961 1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974

YEAR

Fig. 10 Concentration ratio of ^^ ' Pu and ^ ^ 'Pu in air at Richland, Wash.

The most abundant plutonium isotope produced during nuclear detonation is the
weak beta-emitting '^^^Pu. The atmospheric concentrations observed in air at Richland,
Wash., from 1963 to 1972 ranged from 20 (d/min)/10^ m^ at standard temperature and
pressure in 1963 to 0.7 (d/minVlO^ m^ in 1972, whereas the ratio of ^"^^ Pu/^^^'^'^^Pu
was about 15 (Thomas and Perkins, 1974). These data are summarized in Fig. 10.

Americium-241 , which is the daughter of ■^'^'Pu, in global fallout can be estimated
from the plutonium isotopic composition data. Americium-241 is an alpha emitter with a
half-life of 433 yr. It is a bone seeker, and, on the basis of the International Commission
on Radiological Protection (ICRP) maximum permissible concentrations in air and water,
its toxicity is comparable to that of ^^^Pu. In the nuclear power industry, ^'^^Am is
particularly important because it is a relatively large contributor to the total alpha
activity of higli-burnup nuclear fuel (Thomas and Perkins, 1974). Like ^'*°Pu, most of
the ^'^'Pu in large nuclear weapons test debris is produced in the detonation. By making
appropriate weighting and radioactive decay corrections, including in-growth from
parent— daughter relationships, the ^"^ ^Am/'^^^ '■^'*°Pu activity ratio of integrated global
fallout in February 1974 was estimated to be 0.22 (Krey et al., 1976). Two soil samples
analyzed for ^"^^Am fallout (Krey et al., 1976) gave ^^lAm/^^^'^^^Pu ratios of 0.25
and 0.22.

Further calculations of the parent— daughter relationships indicate that the ^"^^Am
content of present integrated fallout in soil will peak in 2037 and will represent 42% of
the '"^''^•^Pu activity.

The distribution of the ^^^Pu/^^^Pu atom ratio in soil is shown in Fig. 11. Similar
patterns emerge from the ^^^ Pu and ^^■^Pu data. There is a marked reduction in the ratio
in the southwestern United States and along the west coast of South America owing to

76 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

U.f

1 I 1 1

I 1

•

.1 •»

• •
•

•

•
•

••• v- •'

•

•
•

•

0.1

•

0.05

. •

1 1 1 1

1 1 1

40
NORTH

0
EQUATOR

40
SOUTH

Fig. 1 1 Atom ratio of ^ '*" Pu/^ ' Pu in worldwide fallout soil sample.

the relatively higher deposition of atypical debris from the Nevada Test Site and from the
French testing site at Mururoa Atoll, respectively.

Figure 11 also shows a slightly reduced ^^^Pu/^^^Pu mass ratio in the equatorial
region between 30°N and 30°S. This observation can be explained by the following
considerations. The neutrons generated in a nuclear detonation increase with the yield.
Therefore it seems reasonable that the ^'*°Pu/^^^Pu atom ratio will also increase with the
size of the nuclear detonation. Most of the nuclear test sites are located within the 30°N
to 30°S region where the tropopause height is at its maximum and only the debris clouds
from the larger yield tests in tliis region had sufficient momentum to penetrate the
stratosphere. Debris from these tests, which have an elevated ■^^^Pu/^^^Pu ratio, which
entered the stratosphere is largely deposited in the middle latitudes because of the greater
transfer rates from the stratosphere to the troposphere at these latitudes. By contrast,
debris from the smaller yield tests, which have lower ■^'*"Pu/^^^Pu mass ratios, remained
within the troposphere and are deposited on the earth's surface predominantly at the
latitude of the detonation. Therefore a relatively greater amount of fallout plutonium
from low-yield detonations is deposited in the equatorial regions, which gives rise to the
reduced ^''^Pu/^^^Pu mass ratio (Hardy, Krey, and Volchok, 1973). The average atom
ratios for ^^^Pu/^^^Pu, ^""'Puj^^^Pu, and ^^^Pu/^^^Pu corrected to January 1971
were 0.179 ± 0.014, 0.0083 ± 0.0017, and 0.0036 ± 0.0011, respectively. Samples con-
taining obvious contributions from the Nevada Test Site were excluded in the
determining of these ratios.

The concentration of ^^^Pu from the SNAP-9 A device had been measured in surface
monthly air samples from northern to southern latitudes (Hardy, 1977). The yearly
average of these data is shown in Fig. 1 2. At most latitudes the ^^^Pu from the SNAP-9A
incident was barely detectable in 1966 but built up to a maximum in 1967 and 1968. The

WORLDWIDE FALLOUT 77

60 —
50 —
40
30
20 —
10
0

Guayaquil (0

J L

1 — \ 1 \ \ [

New York (40^N) —I 60

1 \ — \ \ \ r-

Lima (10°S)

— I — r

Antarctica (70°S) _

1 — \ — \ — \ — I r

Point Arenas

1965 1966 1967 1968 1969 1970 1971 1972 1973
YEAR

50 S) _

1965 1966 1967 1968 1969 1970 1971 1972 1973
YEAR

Fig. 12 SNAP-9A Pu tropospheric air concentration.

IS TRANSURANIC ELEMENTS IN THE ENVIRONMENT

notable exception to this trend was the Antarctica sample (70°S), where 1966
represented the maximum concentration. It has been theorized that air movement above
21 km involves an ascent of air over the summer pole, a mesospheric meridional flow
from the summer to the winter hemisphere, and a descent of air over the winter pole.
This, coupled with the residence half-time at this altitude of 6 months (Thomas et al.,
1970) and reentry burnup of the SNAP-9A generator occurring at 46 km in the southern
hemisphere over the Indian Ocean, produced a condition where the meridional flow from
the northern hemisphere to the southern hemisphere had either begun or was about to
begin. All these factors resulted in a disproportionation of ^■^^Pu deposition, 73% in the
southern hemisphere and 27% in the northern hemisphere, and could have transferred
debris over the south pole, resulting in a more rapid movement downward at the pole
than at other southern or northern latitudes. Whatever the mechanism is for distribution
of SNAP-9A debris througliout the hemisphere, these data show that debris injected in
April 1964 at 46 km altitude reached a maximum in ground-level air about 2 or 3 yr later,
depending on the latitude, and by 1971 it was largely depleted from the atmosphere.

From October 1970 until January 1971, soils were collected by EML (Thomas and
Perkins, 1974) from undisturbed areas at 65 sites around the world to determine the
total deposition of plutonium. Each sample consisted of ten 8.9-cm-diameter cores taken
to 30-cm depth, which represented a surface area of 622 cm^. The measured ^^*Pu
included that from weapons testing plus the SNAP -9A contribution. The ■^^^'^'*°Pu was
assumed to be entirely derived from weapons testing. The EML estimated the weapons
^^^Pu contribution by multiplying the 239,240p|j y^^gg ^y 0.024, their average weapons
238p^/239,240py ratio fouud for six soils collected before fallout from the SNAP-9A
(Krey et al., 1976). These soils were selected to cover a range of latitudes from 71°N to
35°S. The SNAP-9A ^^^Pu was simply the difference between the total measured ^^^Pu
and the weapons ^•^^Pu. The EML deposition sites were grouped into ten-degree latitude
bands and the deposition values averaged as shown in Table 9. The average activities of
^^*Pu per square kilometer, ^^^ '^"^^Pu per square kilometer, and the ^^^Pu/^^^'^''°Pu
ratios in each ten-degree latitude band are shown in Figs. 13 to 15, respectively.

The distribution pattern for weapons plutonium shows heaviest deposition in the
northern hemisphere temperate latitudes and a minimum in the equatorial region. The
rise in the southern hemisphere temperate zone is, at its peak, about one-fifth of that in
the northern hemisphere maximum. The SNAP-9A ^^*Pu has an entirely different
distribution pattern. Most of the SNAP debris was deposited in the southern hemisphere
where the total fallout is 2.5 times as great as that in the northern hemisphere.

Short-Lived Transuranic Radionuclides in Fallout from Nuclear Weapons Testing

Nuclear debris from the past several Chinese tests has been examined to estimate the
radiation exposure resulting from individual short-Hved radioisotopes (Thomas, 1979a,
1979b, 1979c; Thomas and Jenkins, 1974; Thomas, Jenkins, and Perkins, 1976; Thomas
etal., 1976a, 1976b). Table 10 shows the ratios of the concentrations of '^^^Np and
^^''U relative to ''^^Ba. It is evident that the ratios of each of these transuranium
elements to the major fission product, ''*°Ba, are rather high. This fact becomes
important when one calculates the radiation exposure from a submersion dose or from
ground shine. In fallout debris from such a test, the radiation exposure from these
short-lived transuranic radionuclides makes up a significant portion of the total exposure
of fresh fallout debris.

WORLDWIDE FALLOUT 79

TABLE 9 Average Latitudinal Distributions of
Cumulative ^ 3 9 ,2 4 o p^ ^^^ 2 3 s p^ p^iiout *

Latitude

2 3 i

'Pu

band,
degrees

2 3 9.240 p„

2 38py

Hemisphere

Weapons

SNAP-9A

2 3 9.240 py

Millicuries per square kilometer

Northern

90-80

(0.10 ± 0.04)

(0.002 ± 0.001)

«0.001)

0.020

80-70

0.36 ± 0.05

0.009 ± 0.001

<0.001

0.025

70-60

1.6 ± 1.0

0.038 ± 0.025

0.026 ± 0.015

0.040

60-50

1.3 ± 0.2

0.031 ± 0.004

0.013 ±0.004

0.034

50-40

2.2 ± 0.5

0.053 ± 0.011

0.026 ± 0.011

0.036

40-30

1.8 ±0.6

0.042 ± 0.014

0.025 ± 0.015

0.037

30-20

0.96 ± 0.07

0.023 ± 0.002

0.011 ± 0.004

0.035

20-10

0.24 ± 0.10

0.006 ± 0.002

0.003 ± 0.002

0.038

10-0

0.13 ±0.06

0.003 ± 0.001

<0.001

0.023
x = 0.032 ± 0.0073

Southern

0-10

0.30 ± 0.20

0.007 ± 0.005

0.010 ± 0.007

0.057

10-20

0.18 ±0.05

0.004 ± 0.001

0.036 ± 0.021

0.222

20-30

0.39 ±0.16

0.009 ± 0.004

0.070 ± 0.042

0.203

30-40

0.40 ± 0.12

0.009 ± 0.003

0.061 ± 0.020

0.175

40-50

0.35 ± 0.21

0.008 ± 0.005

0.069 ± 0.038

0.220

50-60

(0.20 ± 0.09)

(0.005 ± 0.002)

(0.044 ± 0.023)

0.245

60-70

(0.10 ±0.04)

(0.002 ± 0.001)

(0.022 ± 0.012)

0.240

70-80

(0.03 ± 0.01)

(0.001 ± 0.001)

(0.008 ± 0.005)

0.300

80-90

(0.01 ± 0.004)

«0.001)

(0.004 ± 0.002)

0.400

Northern
Southern
Global

x= 0.229 ± 0.092

Kilocuries deposited (through 1971)

256 ±33

6.1 ± 0.8

3.1 ± 0.8

0.036

69 ± 14

1.6 ±0.3

10.8 ± 2.1

0.180

325 ± 36

7.7 ± 0.9

13.9 ± 2.2

0.066

♦Results in parentheses were derived by extrapolation; error terms are standard deviations.

TABLE 1 0 Ratio of" ^ ^ U and ' ^ ^ Np Activities to

' '^^ Ba Chinese Test Debris Collected

in Surface Air Samples at Richland, Wash.

Date

2 3 7y/i 4 0

"'Np/'^°Ba

May 9, 1966

4.52

31.3

Dec. 27, 1968

3.97

29.4

Sept. 29, 1969

4.76

34.8

Oct. 15, 1970

3.70

27.2

June 26, 1973

7.55

46.5

June 17, 1974

9.71

44.2

80 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

0.1

— Southern

I-
<

o
o

0.01

0.001

20 40 60

DEGREES LATITUDE

Fig. 1 3 Concentration of ^ ^ * Pu as a function of latitude in integrated soil sample.

1.0

o
E

o 0.1 —

o
u

0.01 I 1 \ L

20 40 60

DEGREES LATITUDE

Fig. 14 Concentration of ^ ^' '^^'Pu as a function of latitude in integrated soil sample.

WORLDWIDE FALLOUT 81

1.0 r-

0.01

Southern

hern >»

Northern

■.—■■—.--./ \

20 40 60 80

DEGREES LATITUDE

Fig. 15 Activity ratio of 23 8py^239 ,24 op^ ^ 3 function of latitude in integrated soil
sample.

References

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119(6): 2000-2004.
Ferber, Gilbert J., 1964, Distribution of Radioactivity with Heights in Nuclear Clouds, in Proceedings

of the Second Conference on Radioactive Fallout from Nuclear Weapons Tests, Germantown, Md.,

Nov. 3-6, 1964, USAEC Report CONF-765, pp. 629-645, NTIS.
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1964. USAEC Report HASL-142, p. 225, Health and Safety Laboratory, NTIS.
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Report HASL-273, pp. 111-112, Health and Safety Laboratory, NTIS.
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USAEC Report HASL-302, pp. Bl 1 1-119, Health and Safety Laboratory, NTIS.
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USAEC Report HASL-315, p. B-1 18, Health and Safety Laboratory, NTIS.
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Society Winter Meeting 1974, ERDA Report HASL-291, pp. 1/105-109, Health and Safety

Laboratory, NTIS.
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HASL-235, Health and Salety Laboratory, NTIS.

82 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

, and B. Krajewski, 1970a, Tropospheric Scavenging of '"Srand ^H, in Precipitation Scavenging

(1970). DOE Symposium Series, Richland, Wash., June 1-5, 1970, R. J. Engelmann and W. G. N.

Slinn (Coordinators), pp. 447-463, CONF-700601, NTIS.
, and B. Krajewski, 1970b, Comparison of Atmospheric Transport Model Calculations with

Observations of Radioactive Debris,/. Geophys. Res., 75(15): 2901-2908.
, E. P. Hardy, C. Pachucki, F. Rourke, J. Coluzza, and W. K. Benson, 1976, Mass Isotopic

Composition of Global Fallout Plutonium in Soil, in Transuranium Nuclides in the Environment,

Symposium Proceedings, San Francisco, 1975, pp. 671-678, STI/PUB/410, International Atomic

Energy Agency, Vienna.
Langham, W., 1970, USAF, A^mc/. Safety, 65: 36.
Machta, L., 1964, Status of Global Radioactive Fallout Predictions, in Proceedings of the Second

Conference on Radioactive Fallout from Nuclear Weapons Tests, Germantown, Md., Nov. 3-6,

1964, USAEC Report CONF-765, NTIS.
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the 15th Chinese Nuclear Test, //ea/r/iP/? vs., 29(2): 291-300.
Schneider, K. J. (Ed.), 1974, High-level Radioactive Waste Management Alternatives, ERDA Report

BNWL-1900 (Vol. 1), Battelle, Pacific Northwest Laboratory, NTIS.
Stewart, N.G., R. N. Crooks, and E. M. R. Fisher, 1955, The Radiological Dose to Persons in the U. K.

due to Debris from Nuclear Test E.xplosions, in Report for the M.R.P. Committee on the Medical

Aspects of Nuclear Radiation, Report AERE-HP/R-1701, Atomic Energy Research Establishment.
Thomas, C. W., 1972, Comparison of Atmospheric Radionuclide Concentrations at Neah Bay and

Richland, Washington, USAEC Report BNWL-1651(Pt. 2), Battelle Pacific Northwest Lalc'?-

tories, NTIS.
, 1979a, Radioactive Fallout from Oiinese Nuclear Test, December 14, 1978, DOE Report

PNL-2912, Battelle, Pacific Northwest Laboratory, NTIS.
, 1979b, Radioactive Fallout from Chinese Nuclear Weapons Test, March 15. 1978, DOE Report

PNL-2913, Battelle, Pacific Northwest Laboratory, NTIS.
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PNL-2914, Battelle, Pacific Northwest Laboratory, NTIS.
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ERDA Report BNWL-B-367, Battelle, Pacific Northwest Laboratory, NTIS.
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Winter Meeting of the American Nuclear Society on Environmental Levels of the Transuranium

Elements, Washington, D.C., Oct. 30, 1974, American Nuclear Society.
, C. E. Jenkins, and R. W. Perkins, 1976, Behavior and Characteristics of Radioactive Debris from

the Chinese Nuclear Weapons Test of June 26. 1973. ERDA Report BNWL-2160, Battelle, Pacific

Northwest Laboratory, NTIS.
, W. B. Silker, C. E. Jenkins, and R. W. Perkins, 1976a, Behavior and Characteristics of Radioactive

Debris from the Oiinese Nuclear Weapons Tests of January 7. 1972, and March 18, 1972, ERDA

Report BNWL-2161, Battelle, Pacific Northwest Laboratory, NTIS.
, J. K. Soldat, W. B. Silker, and R. W. Perkins, 1976b, Radioactive Fallout from Chinese Nuclear

Weapons Test, September 26, 1976, ERDA Report BNWL-SA-6054, Battelle, Pacific Northwest

Laboratory, NTIS.
, J, A. Young, N. A. Wogman, and R. W. Perkins, 1970, The Measurement and Behavior of

Airborne Radionuclides Since 1962, in Radionuclides in the Environment, Advances in Chemistry

Series, No. 93, American Chemical Society.
United Kingdom Atomic Energy Authority, \912, Radioactive Fallout in Air and Rain, Results to the

Middle of 1972, Report AERE-R-7245.

, 1973, Radioactive Fallout in Air and Rain, Results to the Middle of 1973, Report AERE-R-7540.

, 1 974, Radioactive Fallout in Air and Rain, Results to the Middle of 1974. Report AERE-R-7832.

, 191 S, Radioactive Fallout in Air and Rain, Results to the End of 1975, Report AERE-R -8267.

United Nations, New York, 1964, Report of the United Nations Scientific Committee on the Effects

of Atomic Radiation, General Assembly, Supplement No. 14 (A/5814).
Young, J. A., L. Wendell, and N. A. Wogman, 1975, Qrrus Scavenging as a Mechanism for the

Production of Radionuclide Concentration Minimums Between 6 arid 9km, USAEC Report

BNWL-1950(Pt. 3), p. 136, BatteUe, Pacific Northwest Laboratories, NTIS.

Transuranic Elements in Space
Nuclear Power Systems

THADDEUS J. DOBRY, JR.

In the 20 years of the space age, the U. S. Department of Energy and its predecessors, the
U. S. Energy Research and Development Administration and the U. S. Atomic Energy
Commission, have liad a growing role in our country 's exploration and exploitation of
space. From a few early earth orbital missions through lunar landings to long-term
outer-planetary journeys, the safety, compactness, reliability, and life of nuclear isotope
power supplies fiave been essential to mission success. Technology improvements are
continuing to virtually eliminate the release of radioactive fuel during normal operations
and accident situations.

Between June 1961 and December 1976 the United States launched 19 spacecraft
designed with electrical systems powered by the transuranic element plutonium, which
contained approximately 80% ^^^Pu and 17% ^^^Pu by weight. Of these 19 systems, 7
were U. S. Department of Defense (DOD) satelUtes and 12 were National Aeronautics and
Space Administration (NASA) scientific spacecraft (2 weather satelUtes, 6 Apollo lunar
experiments, 2 Pioneer interplanetary probes, and 2 Viking Mars landing vehicles).
Table 1 lists these launchings and gives the status of the systems.

Of the 941,600 Ci of "^Pu (700 Ci of ^^^Pu) launched to date, 379,200 Ci (282 Ci
of ^^^Pu), approximately 40%o, is in long-term earth orbit. There is 222,500 Ci of ^^*Pu
(165 Ci of ^^^Pu), or approximately 24% of the total, on the lunar surface; 160,000 Ci
(119 Ci of ^^^Pu), approximately 17%o, has been ejected from our solar system, and
84,000 Ci (63 Ci of^^^Pu), 9% of the total, is on the surface of the planet Mars. The
remaining 10%, 95,900 Ci (71 Ci of ^^^Pu) was involved in three in-flight vehicle aborts.

These aborts did not result in nuclear accidents and were of the following three types:

1. The DOD Transit satellite 5 BN-3 launched in April 1964 from the Western Test
Range reentered the atmosphere, and the 17,000 Ci of ^^^Pu (13 Ci of ^^^Pu) promptly
burned up at high altitude over the Mozambique channel. Prior to 1967 plutonium metal
was used as a fuel, and burnup with subsequent atmospheric dilution was a design and
safety requirement. Since 1967 progress has been made in virtually eliminating reentry
burnup of the fuel in the event of an in-flight abort and in minimizing the probability of
releases of radioactive fuels from launch aborts. An intact reentry— intact impact
pliilosophy has been invoked to counter aborts leading to uncontrolled random
worldwide land or sea impacts.

2. The NASA Nimbus B-1 weather satellite launched in May 1968 from the Western
Test Range was a range safety destruct action that resulted in the intact impact of 34,400
Ci of ^^^Pu (25 Ci of ^^^Pu) in the Santa Barbara channel. The fuel remained intact in
two containers, which were subsequently recovered.

84 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

3. After being launched in April 1970, the Apollo 13 was aborted in flight, and
44,500 Ci of ^^*Pu (33 Ci of ^^^Pu) attached to the lunar landing vehicle was
deliberately disposed of in the Pacific Ocean near the Tonga Trench as a preplanned flight
contingency.

All the above aborts were identified and their probabilities and consequences were
analyzed in the risk assessments submitted for the pre flight presidential approval actions.

The space program has handled approximately 1 MCi of plutonium, and only 17,000
Ci has been released to the environment as worldwide fallout. The activity disposed of in
the deep ocean is contained in corrosion-resistant materials that should prevent the
release of the fuel for a period of time equivalent to 10 half-lives of the ^^^Pu fuel.
Safety design requirements have been established which limit the release of fuel to the
environment. Since accidents are not expected to occur each time a system is launched

TABLE 1 Summary of Launched Space Nuclear Power Systems

Launch
date

Fuel form

Activity

,Ci

Mission

2 3 spy

239pu

Disposition

Transit 4-A

6/61

Plutonium metal

1,800

1.3

In >1000-yr
earth orbit

Transit 4-B

11/61

Plutonium metal

1,800

1.3

In >1000-yr
earth orbit

Transit 5-BN-l

9/63

Plutonium metal

17,000

In >1000-yr
earth orbit

Transit 5-BN-2

12/63

Plutonium metal

17,000

In>1000-yr
earth orbit

Transit 5-BN-3

4/64

Plutonium metal

17,000

Aborted, burned up
on reentry

Nimbus B-1

5/68

PuOj microspheres

34,400

Aborted, containers
recovered

Nimbus 111

4/69

PuOj microspheres

37,600

In ~3000-yr
earth orbit

Apollo 12

11/69

PuOj microspheres

44,500

On lunar surface

Apollo 13

4/70

PuOj microspheres

44,500

Aborted, intact in
Pacific Ocean

Apollo 14

1/71

PuOj microspheres

44,500

On lunar surface

ApoUo 15

7/71

PuOj microspheres

44,500

On lunar surface

Apollo 16

1/72

PuOj microspheres

44,500

On lunar surface

Pioneer F

3/72

Plutonium molybdenum cermet

80,000

5 9.5

Ejected from
solar system

Transit

9/72

Plutonium molybdenum cermet

24,000

In<1000-yr
earth orbit

Apollo 1 7

12/72

PuO^ microspheres

44,500

On lunar surface

Pioneer G

4/73

Plutonium molybdenum cermet

80,000

5 9.5

Ejected from
solar system

Viking-1

8/75

Plutonium molybdenum cermet

42,000

On Mars surface

Viking-2

9/75

Plutonium molybdenum cermet

42,000

On Mars surface

Les8/9

3/76

Pressed PuO^

280,000

208.4

In >100,000-yr
earth orbit

Total

941,600

700

TRANSURANIC ELEMENTS IN SPACE NUCLEAR POWER SYSTEMS 85

and since the fuel inventory required in the system could be contained in multiple
structures, the extent of release per container is controlled within the heat-source design
by a probabiUstic scaling factor, where the release probability, including the occurrence
probability of the accident, is inversely proportional to the quantity of fuel released. Risk
assessments of current space systems have indicated that a source term to the biosphere
of 1 to 10 Ci of respirable fuel might be expected with a probability of from 10"^ to
10~^. This source term could be either an atmosphere release at altitude or a ground
point-source release. Releases to the hydrosphere are controlled by the integrity of the
fuel containers and the dissolution rate of the fuel form. On the basis of release-rate
experiments, plutonium concentrations in seawater are generally in the picocurie per
millihter range at distances of 10 m or more from the point source.

In summary, progress has been made in the space program to virtually eliminate the
release of radioactive fuel during normal operations and launch aborts. As future trends
require larger systems with liigher electrical power levels and larger fuel inventories,
more-stringent system safety requirements and more-sophisticated analytical and test
methods to improve the quality or risk assessments and source-term evaluations are being
developed and enforced.

Quantities of Transuranic Elements
in the Environment from Operations
Relating to Nuclear Weapons

GORDON FACER

Only nuclear explosions near or above the earth 's surface or under water have contributed
substantial amounts of transuranic materials to the world bioenvironment. The amounts
of transuranics placed in the environment through underground test ventings, accidents
involving U. S. nuclear weapons, and releases during weapon production operations have
been negligible in comparison with those from atmospheric testing of nuclear explosives.
On the order of 10^ Ci of plutonium has been dispersed within our environment from
about 400 nuclear explosive tests, including those by the United States, Great Britain,
and Russia, between 1945 and 1963, plus more recent nuclear explosive tests in the
atmosphere by China, India, and France.

The main source of transuranic material, particularly plutonium, presently in the human
environment, other than that which occurs in nature (Meyers and Lindner, 1971), is
nuclear weapons.* Weapons testing in the atmosphere since 1945 has distributed by far
the largest part of the existing transuranic inventory throughout the world. However,
smaller amounts of transuranic materials have reached the environment as the result of
accidents, both real and simulated, with nuclear weapons and of releases of transuranic
materials during weapon development and fabrication operations. It must be assumed
that other countries have had releases of transuranic materials comparable to those for
which the United States was responsible.

All U. S. weapons explosions in atmospheric or near-surface (ground or water)
environments took place between 1945 and September 1963. The United States, Great
Britain, and Russia joined in terminating atmospheric testing when the Limited Test Ban
Treaty was established in September 1963. Only China, India, and France (not parties to
the Limited Test Ban Treaty) have continued testing nuclear explosives in the atmosphere
since the 1963 date.

The quantities of transuranics released to the environment from nuclear testing are
somewhat uncertain. First, the amounts of transuranics that have been placed within test
devices and the numbers of such devices that have been tested are topics that have been
closely held by the respective testing countries. Second, even if we knew the amount of
materials in each specific test device, there would be no accurate means for determining
the amount oi material that may have reached the environment from the detonation of
those devices. Some undisclosed amount of the transuranic material was expended in the

*F()r this discussion, the term "nuclear weapons" is used to mean all nuclear explosives, including
some designed for peaceful applications.

TRANS URANICS FROM NUCLEAR WEAPONS OPERATIONS 8 7

fission process; in addition, an unknown amount became environmentally inaccessible
because of the circumstances under which the tests were done. However, even with these
limitations, certain approximations can be made.

About 195 (Glasstone, 1962) U. S. nuclear tests have been conducted in emplacement
locations from which transuranic materials might have reached the environment,
including all atmospheric and most underwater and cratering tests.* Allowing for as many
nuclear tests by other countries as the United States has conducted brings us to an
approximate worldwide total of about 400 tests. [Carter and Moghissi (1977) reported
389 such tests through June 1975.]

Some of the devices tested have been of a pure fission design. Many others, however,
probably reflected a variety of designs involving combinations of fission and fusion
processes. The transuranic material released to the environment per test certainly has
varied considerably through the range of tests that have been done. For this discussion I
have assumed that, as an order of magnitude, more than 100 Ci and less than 1000 Ci of
transuranic source material was residual to each test. Hence a residual of between 4x10"*
and 4x 10^ Ci (1 Ciof^^^Pu= 16 g) of transuranic materials might remain environmen-
tally available from worldwide nuclear weapons testing in the atmosphere; however, only
a small part of the real total may remain accessible to the human environment today.

The nature of individual test emplacements has a considerable influence on the
amount of the residual transuranic material that actually becomes environmentally
available. Of the 195 U. S. tests (Glasstone, 1962), 60 were fired at or near the earth's
surface and 45 were fired atop steel towers. About 90 others were detonated in a way
that would make most residual material environmentally available, e.g., devices emplaced
on tethered balloons and those positioned by airdrop, rocket, or gun.

The fraction of transuranic material released to the world environment from nuclear
tests fired very close to the surface has probably been relatively small (Glasstone and
Dolan, 1977). Plutonium particles in particular have a strong tendency to attach to other
materials; hence most of the residual plutonium from a near-surface explosion would
become attached to the enormous amount of earthen material disturbed by the
explosion. Most of these plutonium-laden earthen materials, which were in the form of
large particles, remained fairly close to the detonation point after the test explosion. (A
different but somewhat comparable situation exists with the near-surface explosions on
moored barges in the shallow lagoons of Bikini and Enewetak.) Part of the work that has
been done by the Nevada Applied Ecology Groupt (NAEG) (Dunaway and White, 1974;
Wliite et al., 1975; 1976; 1977) at Nevada on the behavior of this earth-entrained
plutonium has been aimed at defining the nature of these distributions.

The nuclear explosives detonated on steel towers represent an intermediate situation
wherein relatively little surface material was disturbed and the fraction of transuranic
residue that became associated with earthen materials was much smaller. However, the
towers used in these tests furnished materials that probably influenced the behavior of
those transuranics. The typical tower was made of heavy structural steel with an open

*ror this discussion, there would be no point in including the contained nuclear tests from which
plutonium does not become environmentally available.

fThe Nevada Applied Ecology Group was established in 1970 within the Nevada Operations Office
of the U. S. Atomic Energy Commission to design a comprehensive studies program looking into
specific environmental problems that might already exist or that might arise in connection with
nuclear weapons test activities.

88 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

square framework about 22 by 22 ft (7 by 7 m) in cross section and extending high
enough that the explosion fireball did not reach the ground surface. This meant that most
towers were between 300 and 500 ft (90 and 150 m) high (Glasstone, 1962); the tallest
was 700 ft (213 m). Steel guy wires gave these towers lateral support. When a nuclear test
was fired, the great heat vaporized much of the tower steel and carried it upward v^thin
the rapidly rising cloud. As the vaporized tower material cooled, it condensed, and most
of it tended to fall nearby, carrying along with it some of the residual transuranic
materials from the tested device which were condensing at the same time.

Balloon-suspended nuclear tests and airdropped nuclear tests, which were detonated
well above the ground, contributed very little material to which the residual transuranic
materials might become attached. Hence probably a fairly large fraction of the
transuranic material from those tests became widely dispersed throughout the world.

Much of the material that was released into the environment from nuclear testing has
by today become relatively inaccessible to the environment. That which fell on the lakes,
oceans, and seas is in the process of sinking to the bottom or already has reached the
bottom sediments (Edgington, Wahlgren, and Marshall, 1976), and much of that which fell
on land areas will soon be beneath the immediate surface layer (Essington and Fowler,
1976). Because of these factors and because of the fairly large uncertainties as to the
amounts of transuranic materials that originally had been injected into the world
environment from nuclear weapons testing, quantitative (source-term) estimates based on
weapons testing history may not be as useful in handling specific localized problems as
are regional estimates of the materials present based on local samplings.

On the other hand, regional estimates may not extrapolate well as a means of
determining the world source term. Environmental releases of transuranics from nuclear
weapons have varied considerably by latitude and have been far greater in the northern
hemisphere than in the southern hemisphere. If we should assume that the U. S.
plutonium distribution level (Harley, 1971) in the surface soils of about 1 mCi/km^
persists worldwide (about 4.8 x 10^ km^), there would be 480,000 Ci, or about
8,000 kg. Hardy, Krey, and Volchok (1973) have estimated that worldwide there was
325,000 Ci, or about 5200 kg, of ^^^Pu and ^"^^Pu in weapons-fallout debris. Although
quantities of transuranics in this range are entirely credible, my review of atmospheric
nuclear testing indicates that those estimates may be somewhat high. All things
considered, a 1 x 10^ Ci source term for environmentally available ^^^Pu, although very
approximate, appears conservatively suitable.

As mentioned, there is some transuranic material in the environment as a result of
accidents with nuclear weapons. A certain amount of this transuranic dispersal took place
as the result of deliberate tests of the behavior of weapons under accident conditions.*
Several such accident-simulation tests were done at and near the Nevada Test Site and at
the Tonopah Test Range near Tonopah, Nev. Although there was on the order of 10 to
10^ Ci of plutonium (total) involved in those tests, some of the material was recovered
and removed from the environment by personnel manually searching for and picking up
the scattered metal pieces. On the basis of site -in tensive inventories at the locations of
these safety tests (White and Dunaway, 1975) conducted under the NAEG studies
program, about 160Ci of ^^^Pu and ^"^^Pu remain environmentally available (in the

*U. S. nuclear weapons are designed so that no accident can create the nuclear circumstances
necessary to deUver nuclear yield. However, certain accident conditions could cause the materials
associated with a nuclear device to burn or detonate and thus disperse transuranic materials.

TRANSURANICS FROM NUCLEAR WEAPONS OPERATIONS 89

top 5 cm of soil) in the immediate vicinity of those accident-simulation test sites;
probably there are on the order of a few curies from the same tests dispersed outside the
immediate vicinity but within a few miles of those locations. No such inventory yet has
been feasible for the sites of U. S. atmospheric tests.

Very few accidents with U. S. plutonium weapons have placed transuranic materials
in the environment. The following weapons accidents are noted (U. S. Atomic Energy
Commission, 1974). One in 1960 involved the burning of a missile on its launcher at
Maguire Air Force Base, N. J. That fire completely melted plutonium in the nuclear
warhead but apparently did not lead to any appreciable dispersal. A second accident in
January 1966 resulted from the collision of a B-52 bomber with a tanker aircraft over
Palomares, Spain. Two nuclear weapons on board the bomber fell to the ground; the
impact detonated the high explosive in the weapons, and the contained plutonium was
dispersed nearby, mainly within an area totaling about 500 acres (200 ha). As the result
of extensive cleanup by the U. S. Air Force, the amount of residual plutonium that is
environmentally available from that accident has been estimated to be quite small. A
third accident (U. S. Air Force, 1970) involved the in-flight fire and crash of a B-52
bomber on the ice of an Arctic bay at Thule, Greenland, in January 1968. When the plane
impacted, there was a large explosion and intense fire. After that accident considerable
plutonium was found associated with the black crustation of burned jet fuel distributed
over about 30 acres (12 ha) of the snow-covered surface of the bay ice. More of the
material was found on the aircraft wreckage. Almost all this plutonium was removed
through cleanup operations. An estimated 25 Ci probably went to nearby soils and
bottom sediments; of the remainder, only a small fraction of the total plutonium in the
accident appears to have been dispersed via the atmosphere away from the crash location.

During Operation Hardtack II at Johnston Island in 1962, four THOR missiles
being used in connection with the atmospheric nuclear tests failed to perform properly.*
Three of these missiles had to be destroyed in flight; transuranic materials from the
attached nuclear explosives were scattered over nearby ocean areas. A fourth THOR
caught fire on the launch pad on July 25, 1962, and high explosive, associated with the
nuclear explosive, was detonated as a protective measure. Plutonium metal and plutonium
oxide were scattered, as a result. There was extensive decontamination of the launch pad
following this specific incident; the more accessible plutonium was removed, and some
less-accessible material was painted over or paved over.

From the record of these accidents, it appears that from 10 to 100 Ci of plutonium
may remain available in the environment from that type happening involving U. S.
nuclear weapons.

Finally, transuranic materials have been released in the course of operations at the
laboratories and plants where nuclear weapons are designed and built. An amount
estimated at between 10 and 100 Ci of plutonium was released to the soil through leakage
from stored waste at the Rocky Flats Plant over a period of several years. About 300 Ci
of ^^^Pu was released at Mound Laboratory in 1969 owing to a break in a waste transfer
line (U. S. Atomic Energy Commission, 1975); this material, however, was not weapons
related. The total from all such releases at all weapons laboratories and facilities probably

*This information is based on personal communications with Layton O'Neill, Nevada Operations
Office. Although the individual incidents were made public through press releases in 1962, no
comprehensive account has been located in published literature.

90 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

does not exceed 400 Ci, of which a small percentage may have reached the general off-site
environments.

Something should be said v^th regard to the ingrowth of americium as a result of the
decay of plutonium. During the production of weapons-grade plutonium in a reactor, a
small fraction of the material produced is ^'*^Pu, which has a 13-yr half-life for decay
by beta emission to ^"^^Am. As this decay progresses, the americium activity grows to
approach the long-term level. Through this ingrowth mechanism the inventory of
environmentally available transuranic material has substantially increased. Because the
percent ^'*^Pu produced depends on the conditions under which the material was
produced in the reactor, the amount of ^"^^ Am that will result from the ingrowth process
is not a fixed fraction that can be predicted. For plutonium more than 20 or 30 yr
after its production, however, americium activity at several percent of the curie level of
the host plutonium is to be expected.

In summary, on the order of 10^ Ci of weapons plutonium has probably been broadly
distributed wdthin the world environment from all sources. Of that amount, between 10^
and 10** Ci is probably concentrated in surface soils around the U. S. test sites. Plutonium
that remains dispersed and environmentally available from actual weapon accidents may
be on the order of 10^ Ci with perhaps an additional 10^ Ci dispersed and environmen-
tally available as a result of weapon accident tests. On the order of 10^ Ci of plutonium is
probably accessible in the environment owing to spills and releases at laboratories and
plants.

References

Carter, M. W., and A. A. Moghissi, 1977, Three Decades of Nuclear Testing, Health Phys., 33: 55-71.

Dunaway, P. B., and M. G. White, 1974, The Dynamics of Plutonium in Desert Environments, Nevada
Applied Ecology Group Progress Report as of January 1974, USAEC Report NVO-142, Nevada
Operations Office, NTIS.

Edgington, D. N., M. A. Wahlgren, and J. S. Marshall, 1976, The Behavior of Plutonium in Aquatic
Ecosystem: A Summary of Studies on the Great Lakes, in Environmental Toxicity of Aquatic
Radionuclides: Models and Mechanisms, Proceedings of the 8th Rochester International
Conference on Environmental Toxicology, Rochester, N. Y., June 2— 4, 1975, M. W. Miller and
J. N. Stannard (Eds.), pp. 45-79, Ann Arbor Science Pubhshers, Inc., Ann Arbor, Mich.

Essington, E. H., and E. B. Fowler, 1976, Distribution of Transuranic Radionuclides in Soils, in
Transuranics in Natural Environments, Symposium Proceedings, Gathnburg, Tenn., Oct. 5-7,
1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178, p. 41, Nevada Operations
Office, NTIS.

Glasstone, Samuel, 1962, The Effects of Nuclear Weapons, Appendix B, U.S. Atomic Energy
Commission, GPO.

, and Philip J. Dolan, 1977, The Effects of Nuclear Weapons, Sec. 9.50, U.S. Department of

Defense and U. S. Department of Energy, GPO.

Hardy, E. P., P. W. Krey, and H. L. Volchok, 1973, Global Inventories and Distribution of Fallout
?\\iXom\xm, Nature, 241: 444.

Harley, J. H., 1971, Worldwide Plutonium Fallout from Weapons Tests, in Environmental Plutonium
Symposium Proceedings, Los Alamos, New Mexico, USAEC Report LA-4756, Los Alamos
Scientific Laboratory, NTIS.

Meyers, Wm, A., and M. Lindner, 1971, Precise Determination of the Natural Abundance of ^^''Np
and ^^'Pu in Katanga Pitchblende,/. Inorg. Nucl. Chem.. 33: 3233-3238.

U. S. Air Force, 1970, USAFNucl. Saf. 65(2), No. 1, Special Edition, Crested Ice.

U. S. Atomic Energy Commission, 1974, Plutonium and Other Transuranium Elements, USAEC
Report WASH-1 359, NTIS.

TRANSURANICS FROM NUCLEAR WEAPONS OPERATIONS 91

, 1975, Investigation of the Circumstances Associated with the Appearance of Plutonium-238

Contamination in Waterways Adjacent to Mound Laboratory, Albuquerque Operations Office,

unnumbered report.
White, M. G., and P. B. Dunaway (Eds.), 1975, The Radioecology of Plutonium and Other

Transuranics in Desert Environments, Nevada Applied Ecology Group Progress Report, ERDA

Report NVO-153, Nevada Operations Office, NTIS.
, and P. B. Dunaway (Eds.), 1976, Studies of Environmental Plutonium and Other Transuranics in

Desert Ecosystems, Nevada Applied Ecology Group Progress Report, ERDA Report NVO-159,

Nevada Operations Office, NTIS.
, P. B. Dunaway, and W. A. Howard (Eds.), 1977, Environmental Plutonium on the Nevada Test

Site and Environs, ERDA Report NVO-1 7 1 , Nevada Operations Office, NTIS.

Transuranic Wastes from the Commercial
Light-Water-Reactor Cycle

M. R. KREITER, J. E. MENDEL, and R. W. McKEE

Airborne and transuranic-contaminated wastes generated in postfission activities are
identified by quantity and radioactivity for the case in which spent fuel is declared waste
(once-through cycle) and that in which spent fuel is reprocessed and the recovered
uranium and plutonium are recycled. Because no standard defining transuranic wastes is
available at this time, in this chapter the waste source is used as the basis for such a
definition.

Radioactive wastes are generally treated to reduce their volume andjor mobility. For
convenience the radioactive wastes discussed are categorized according to the treatment
they require. A selected treatment process as well as the final treated volume is presented
for each of the seven categories of waste.

In addition to wastes generated during the operation of fuel reprocessing and
mixed-oxide-fuel fabrication plants, transuranic wastes resulting from activities associated
with decommissioning postfission fuel-cycle facilities are identified. Dismantlement is the
mode assumed for decommissioning the facilities.

A listing of projected nuclear power growth is presented both for the Organization for
Economic Cooperation and Development nations and for the United States to provide
perspective regarding the quantities of waste generated.

Radioactive wastes result from the fissioning of nuclear fuels used in producing energy at
nuclear power plants. In this chapter radioactive wastes are defined as all materials
actually or potentially contaminated with radioactivity and subsequently disposed of
when worn out, defective, or of no further use.

Radioactive wastes can be categorized as transuranic contaminated or nontransuranic
contaminated. Currently, there is no standard or criterion defining a commercially
generated transuranium waste. A proposed rule making would consign to licensed burial
grounds wastes that have been contaminated with no more than 10 nCi of transuranic
elements per gram of waste [Fed. Regist. (Washington, D.C.j, 39: 32922 (Sept. 12,
1974)] . Ten nanocuries per gram was chosen as representing the upper range of
concentration of radium in the earth's crust. This proposal would imply that wastes
exceeding 10 nCi/g can be classed as transuranic. Studies are in progress to assess the
numerical validity of the transuranic limit in this proposed rule making (Adam and
Rogers, 1978). For this chapter we will assume that reactor operations and spent-fuel
storage-basin operations do not normally produce transuranic wastes [Fed. Regist.
(Washington, D.C.), 39: 32922 (Sept. 12, 1974)]. Transuranic wastes would include
spent fuel if it is declared waste, high-level waste, cladding hulls, and others that will be
identified later in this chapter.

TRANSURANIC WASTES FROM LWR CYCLE 93

Both as-generated (untreated) and treated wastes are, in most cases, addressed in the
following text. Untreated wastes are generally exposed to some form of treatment to
reduce activity levels, to decrease the volume to be handled, and/or to decrease the
mobility of the waste. Treatment processes can be as uncomplicated as a simple
compaction scheme or technically quite sophisticated, as with high-level waste vitrifica-
tion. The technology for a variety of alternative waste-treatment processes is at this time
commercially available or under active development (Energy Research and Development
Administration, 1976; U.S. Department of Energy, 1979).

Attention is focused here on the quantities and radioactivities of those transuranic-
contaminated wastes generated in postfission activities involved in the hght-water-reactor
(LWR) fuel cycle for commercial power production only. A description is given of wastes
resulting from operating and decommissioning of the fuel-cycle facilities. The projected
waste characteristics are often necessarily conjectural owing to lack of hard data from
plant operating experience.

Nuclear Power Growth

A nuclear power generation forecast to the year 2000 is presented for the OECD nations*
to provide a frame of reference for the magnitude of the worldwide generation of wastes
(excluding those nations with centrally planned economies). Forecasts of installed nuclear
capacity have been decUning as a result of a downward trend in projected electric power
requirements from all energy sources. For example, in the 3-yr period between 1973 and
1976, the high and low nuclear power growth estimates for OECD nations decreased 20%
and 33%, respectively (Muda, Haussermann, and Mankin, 1977). Such reductions
generally reflect uncertainties due to lower than expected grov^h in energy use and
greater than anticipated delays because of concerns about safety and the environment.

Table 1 shows the results of a 1976 estimate of the nuclear power growth for OECD
nations. Since 1976 an additional downward revision of from 5 to 10% in the low growth
estimate has been proposed (letter from R. Gene Clark, Chief, Nuclear Energy Analysis
Division of DOE to M. W. Shupe, DOE, Richland Operations Office, July 12, 1978). This
table also provides a recent projection of the nuclear power growth for the United States
(U.S. Department of Energy, 1978).

Waste Descriptions and Classifications

For the LWR fuel cycle, there are two generic operating modes to be considered, that
with and that without spent-fuel reprocessing. In the nonreprocessing mode, currently
referred to as the once-through cycle, energy values contained in irradiated fuel removed
from a nuclear power plant are not recovered by reprocessing and recycHng. Irradiated
fuel is considered a radioactive waste and after storage for some period is sent to disposal.
The reprocessing mode includes different alternatives, for example, the recycle of
uranium only or the recycle of uranium and plutonium. In this chapter we are discussing
the transuranium wastes resulting from the recycling of uranium and plutonium. In this
recycle uranium and plutonium are chemically recovered from the irradiated fuel and
then purified and formed into fresh fuel for generating electricity in a nuclear power
plant. Recycling uranium and plutonium requires a plant for processing irradiated fuels

*The OECD (Organization for Economic Cooperation and Development) nations are Australia,
New Zealand, Canada, Japan, United States, and Western Europe.

94 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Nuclear Power Growth Estimates, GW(e)

Year

1980

1985

1990

1995

2000

OECD

Minimum

147

293

467

681

829

Maximum

172

389

680

1090

1640

USA only

127

195

283

380

and a plant for fabricating fresh fuel elements containing the recycled plutonium. The
plants are commonly referred to as fuels reprocessing plants and mixed-oxide fuels
fabrication plants, respectively. In this chapter these two plants are considered to be the
only sources of transuranic-contaminated wastes for the reprocessing fuel-cycle mode. It
is assumed that the recovered uranium and plutonium are converted to UF^ and PUO2,
respectively, at the fuels reprocessing plant. The wastes from these conversion processes
are included with the fuels reprocessing plant wastes.

Primary wastes are untreated initial wastes issuing from a fuel-cycle facility. The
primary wastes are processed to form treated wastes. Only treated wastes are allowed to
leave the confines of the originating plant and are always under careful control. Treated
wastes are of two types: (1) those which are treated to reduce their activity levels so that
they can be released to the environment without harm to man and (2) those which are
conditioned for long-term containment so that their radioactivity will remain confined
and out of contact with man's environment. The latter are covered in this chapter.

Secondary wastes are generated in treating primary wastes and in the subsequent
handling of treated wastes. Thus secondary wastes are generated not only from the initial
waste-processing steps but also from the storage, transportation, and disposal steps. The
amount of secondary wastes is generally small compared with that of primary wastes.
Nevertheless, an assessment of waste management is not complete until the effects of
secondary wastes are included. Secondary wastes are of the same general classifications as
primary wastes and require the same treatments. Most can be recycled to incoming
primary-waste streams for treatment. In remote locations without primary-waste-
treatment facihties (for instance, isolation sites), special facilities must be provided for
treating the secondary wastes.

Many methods of classifying radioactive wastes are in use, e.g., the kind of
radioactivity contained, the amount of radioactivity contained, the untreated physical
form, and the treated physical form. It is convenient to classify the primary and
secondary wastes into categories according to the treatment they require; i.e., all wastes
requiring a similar treatment are included in one category. The categories and a brief
generic description of each are given in Table 2. The first three waste categories listed in
Table 2 are generated to some degree in almost any facility in which radioactive materials
are processed, treated, or handled. Thus both primary and secondary wastes in these
categories are found throughout the postfission LWR fuel cycles. The remaining four
waste categories are specific to certain fuel cycles. Spent fuel as a waste is specific only to
the once-through cycle, and high-level liquid waste and hulls, only to the fuel cycles that
use fuels reprocessing.

TRANSURANIC WASTES FROM LWR CYCLE 95

TABLE 2 Classification of Primary Transuranic Wastes
from the Postfission LWR Fuel Cycle

Waste category

General description

Gaseous

Compactible trash and
combustible wastes

Concentrated liquids, wet
wastes, and particulate
solids

Failed equipment and non-
com pactible, noncombustible
wastes

Spent UO2 fuel

High-level Uquid waste

Hulls and assembly hardware

Mainly two types: (1) large volumes of ventilation air, potentially
containing particulate activity, and (2) smaller volumes of vessel
vent and process off gas, potentially containing volatile radio-
isotopes in addition to particulate activity.

Miscellaneous wastes, including paper, cloth, plastic, rubber,
and filters. Wide range of activity levels dependent
on source of waste.

Miscellaneous wastes, including evaporator bottoms, filter sludges,
resins, etc. Wide range of activity levels dependent
on source of waste.

Miscellaneous metal or glass wastes, including massive process
vessels. Wide range of activity levels dependent on
source of waste.

Irradiated PWR and BWR fuel assemblies containing fission products
and actinides in ceramic UOj pellets sealed in Zircaloy
tubes.

Concentrated solution containing over 99% of the fission products
and actinides, except uranium and plutonium, in the
spent fuel. Contains about 0.5% of the uranium and
plutonium in the spent fuel.

Residue remaining after UO^ has been dissolved out of spent fuel.
Includes short segment of Zircaloy tubing (hulls) and
stainless-steel assembly hardware. Activity levels are
next highest to high-level liquid wastes.

A large plant can have many sources of wastes belonging to the same category. Thus
the generation rate (waste volume/time) and radioactive content of the wastes from each
source must be analyzed and summed to obtain the overall description of wastes in each
category. This has been done for the main generic plant components of the postfission
LWR fuel cycle in Tables 3 to 7. The waste volumes and activities are given per GW(e)-yr;
1 GW(e)-yr corresponds to the annual electric power needs of about 500,000 people in
the United States. For the generic LWR fuel cycle, on which this chapter is based,
38 metric tons of UO2 fuel must pass through the cycle to generate 1 GW(e)-yr. The
radioactivity contributions of important isotopes present in the wastes are given
individually in the tables along with total radioactivities.

Besides the wastes generated from operation of the main plant components of the
postfission LWR fuel cycle, wastes will be generated from the decommissioning of these
facilities (see Table 8), and miscellaneous secondary wastes will be generated in the
ancillary activities of the postfission LWR fuel cycle, such as transportation and geologic
isolation. The volume of these secondary wastes is minor, and their radioactivity content
is insignificant compared with that of the primary wastes. These ancillary secondary
wastes will be treated at the main plant waste-treatment facilities where possible. Special
waste-treatment facilities, which are scaled-down versions of the large-plant waste-
treatment facilities, will be installed where required, e.g., at geologic repository receiving
stations.

96 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Spent Fuel as a Primary Waste (TRU)

Weight, MTHM/GW(e)-yr*

Radionuclide content,! Ci/GW(e)-

■yr

Important activation products

1 4 /-

4.2

= = le

1.6 X 10 =

*"Co

1.6 X 10=

'^Zr

3.1 X 10'

Total activation products

4.8 X 10=

Important fission products

1.6 X 10"

«5Kr

3.4 X 10=

'"Sr

2.5 X 10*

•°«Ru

6.5 X 10'

1 2 9 j

1.3

'^^Cs

3.5 X 10*

Total fission products

5.3 X 10'

Important actinides

5.4 X 10'

4.3 X 10*

1.5 X 10"

1.9 X 10=

Total actinides

4.6 X 10*

*Weight is in metric tons heavy metal.
tContent 1.5 yr after removal from reactor.

Nuclear-Power-Plant Wastes

For this chapter the general wastes issuing from nuclear power plants are not considered
to be transuranic contaminated. However, for the once-through cycle, irradiated fuel
issuing from a nuclear power plant would be considered a transuranic (TRU) waste. The
characteristics of spent fuel as a waste are given in Table 3. For the purposes of the
generic treatment used here, a reference LWR fuel assembly has been defined as a
composite with properties characterized as between those of PWR and BWR fuel
assemblies. Each assembly weiglis about 430 kg (SOwt.v? core and 20 wt.% Zircaloy
cladding and stainless-steel hardware) and is slightly over 5 m long.

After a period of storage in water basins, the spent fuel, if declared waste, will be
placed in a container, which will be subsequently filled with helium or a metal with high
thermal conductivity and sealed for ultimate disposal.

Fuel Reprocessing Plant Wastes

The fuel reprocessing plant TRU primary wastes are described in Tables 4 to 6. The
characteristics of each waste after treatment are also shown in the tables.

The waste treatments shown in the tables are those defined as references for this
chapter. Other waste treatments could result not only in differing characteristics in the
treated wastes but also in differing amounts and types of secondary wastes. The effects of
secondary-waste management are included in the tables. The volumes of primary waste
shown have been increased appropriately to refiect the recycle of secondary wastes.

TRANSURANIC WASTES FROM LWR CYCLE 97

The general-operations wastes from the fuel reprocessing plant are described in
Table 4. The three categories of solid and Uquid general-operations waste include all the
miscellaneous wastes; the volumes and activities can vary widely depending on plant
operation. The wastes are all packaged and transported to off-site geologic isolation after
treatment. Combustible wastes are incinerated to ensure that the treated wastes sent to
the geologic repository are nonflammable. High-efficiency particulate aerosol (HEPA)
filters are included in the "compactible trash and combustible wastes" category. The
filter cartridges are punched out and packaged with compaction; the combustible filter
frames are incinerated. The incinerator ashes are immobilized cement, as are all
concentrated Uquids, wet wastes, and particulate soUds. The remaining failed equipment
and noncompactible, noncombustible wastes are disassembled and packaged. Large
equipment is disassembled to fit into boxes 1.2 by 1.8 by 1.8 m. All remaining
general-operations wastes are packaged in 55- or 80-gal drums.

High-level liquid waste is described in Table 5. Although of relatively small volume,
particularly after treatment, high-level liquid waste initially contains over 100 times more
radioactivity than the rest of the wastes combined. The first solvent-extraction battery in
the Purex process separates plutonium and uranium from the remaining radionuclides.
The plutonium and uranium are extracted into an immiscible organic fluid and separated
from the starting aqueous solution, which still contains over 99% of the nonvolatile
fission products and actinides other than plutonium and uranium. After concentration
this aqueous solution becomes high-level liquid waste. Since the extraction of plutonium
and uranium is not perfect, some is left behind as "waste losses." The amount of
plutonium and uranium present in high-level liquid waste can vary owing to waste loss. It
is assumed that 0.5% of the plutonium and uranium in the spent fuel ends up in the
high-level liquid waste.

The reference treatment for high-level liquid waste used for Table 5 calculations is
vitrification. This treatment encapsulates the waste in a durable, temperature- and
radiation-resistant glass that is cast in stainless-steel canisters. The canisters are
hermetically sealed before they leave the fuel reprocessing plant.

The second most radioactive waste from the fuel reprocessing plant includes the hulls
and assembly hardware (Table 6). This soUd waste results from the feed-preparation step
in the fuel reprocessing plant. In the feed-preparation step, the fuel assemblies are
mechanically chopped to expose the fuel so that it can be dissolved in nitric acid. The
hulls are the chopped segments of Zircaloy tubing from which the UO2 fuel has been
dissolved. The assembly hardware consists mainly of the fuel-assembly end fittings, which
are removed before the fuel is dissolved.

The activity in hulls and assembly hardware wastes is from (1) neutron activation of
Zircaloy and stainless steel; (2) neutron activation of trace materials, such as uranium in
the cladding metal; and (3) residual fission products and actinides, which were either not
dissolved or had diffused into the metal surfaces so that they could not be dissolved.

Gaseous streams originating in a fuels reprocessing plant must be treated to remove
airborne radioactive materials before they are discharged to the atmosphere. The principal
gas streams include dissolver offgas, process vessel off-gases, ventilation air, vaporized
excess water, and off-gases from the uranium fluorination process. The transuranic
elements contained in the untreated gaseous streams are a small fraction of the airborne
radioactive materials, and most are transported by entrainment.

Estimated transuranic activities discharged to the atmosphere following two different
treatments are given in Table 7. The treatment in case 1 (Fig. 1) consists of directing the

98 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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TRANSURANIC WASTES FROM LWR CYCLE 99

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100 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 5 High-Level Liquid Waste (TRU)

Primary waste

Volume,* mVGW(e)-yr

Radionuclide content,! Ci/GW(e)-yr

Important fission products

1.3 X 10^

"Kr

'"Sr

2.3 X 10«

'"'Ru

7.3 X 10*

1 2 9 J

6.7 X 10-'

'^'Cs

3.6x10'

Total fission products

5.3x10''

Important actinides

1.8 X 10="

3.5 X 10"

3.2 xlO"

6.5 xlO'

Total actinides

7.2 X 10'

Treated waste

Treatment

Vitrification

Volume, mVGW(e)-yr

2.8

Containers/GW(e)-yr

(stainless-steel canister, 30 by

300 cm)

12.6

*38 metric tons heavy metal must pass through
the cycle to generate 1 GW(e)-yr.

IContent 1.5 yr after removal from reactor.

dissolver and vessel off-gases and the ventilation air through an atmospheric protection
system (APS). This system is comprised of a Group III prefilter followed by a single bank
of HEPA filters. After treatment the gas is released througli a stack to the atmosphere.
Vaporized excess water and the off-gases arising from UF^ conversion are sent directly to
the stack without treatment.

The treatment in case 2 (Fig. 1) directs the dissolver off-gas through iodine, carbon,
and krypton removal systems and the vessel off-gas through a filter, an iodine recovery
system, and a nitrogen oxide removal system before entry into the APS. The remaining
streams are handled in the same manner as in case 1. In the reference processes, iodine is
removed by reaction with a silver-loaded adsorbent; carbon is captured in the form of
CO2 on a zeolite (molecular sieve) and converted to calcium carbonate; and krypton is
cryogenically liquefied along with xenon and argon, followed by fractionation. Removing
iodine, carbon, and krypton in conjunction with the APS results in about twice the
reduction in transuranic activity discharged as with the APS alone.

Mixed-Oxide-Fuel Fabrication Wastes

Tlie wastes from mixed-oxide-fuel fabrication are shown in Table 8. It is assumed that
about 20% of the fuel is mixed oxide; thus 7.6 metric tons of mixed oxide is equivalent
to 1 GW(e)-yr for the reference system.

There are no unique waste streams from the mixed-oxide-fuel fabrication plant such
as those which occur at the fuel reprocessing plant. All are general-operations wastes and

TRANSURANIC WASTES FROM LWR CYCLE 101

TABLE 6 Hulls and Assembly Hardware (TRU)

Primary waste

Vi)lume.* niVt;W(e)-yr

12.3

Radionuclide content ,t Ci/GW(e)-yr

Important activation piroducts

'^C

4.6

5 S j..^

1.5 X 10=

•^To

1.5 X 10'

''Zr

3.1 X 10'

Total activation products

4.8 X 10=

Important fission products

2.4 X 10'

'"Sr

1.2 X 10'

'"'Ru

3.6 X 10'

"'Cs

1.8 X 10'

Total fission products

2.9 X 10"

Important actinides

9 X 10-'

3.5 X 10'

1.6 X 10'

3.2 X 10^

Total actinides

3.9x10'

Treated waste

Treatment

Package without

compaction

Volume, mVGW(e)-yr

12.8

Containers/GW(e)-yr

9.2

(canister, 76 by 300 cm)

*38 metric tons heavy metal must pass through the
cycle to generate 1 GW(e)-yr.

fContent 1.5 yr after removal from reactor.

TABLE 7 Release of Airborne Transuranics
from a Fuels Reprocessing Plant

Activity,* Ci/GW(e)-yr

Into

atmospheric

protection

system

Into stack

Element

Case 1 Case 2

Neptunium

3.6 X 10^"

3.6x10"* 2x10"*

Plutonium

1.4

3x10-" 2x10-''

Americium

6.5 X 10-'

6.5 X 10-' 3 x 10-'

Curium

1.3 X 10"'

1.3x10-= 6.5x10-*

* Activity 1.5 yr after removal from reactor; 38 metric
tons heavy metal must pass through the cycle to generate
1 GW(e)-yr.

102 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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TRANSURANIC WASTES FROM LWR CYCLE 103

TABLE 8 Mixed-Oxide-Fuel Fabrication Wastes (TRU)

Compactible trash

Concentrated liquids.

Failed equipment and

and combustible

wet wastes, and par-

noncompactible, non-

Gaseous wastes

wastes

ticulate solids

combustible wastes

Primary waste

Volume,* mVGW(e)-yr

8.4 X 10'

Radionuclide content

(important actinides),t

Ci/GW(e)-yr

3.7 X 10-'°

1.2 X 10-'

1.3 X 10"'

1.3 X 10-''

1.4

4.6 X 10'

5 X 10'

5x10'

2.2 X 10"'

7.3

1.5 X 10'

0.8

Total actinides

1.4

4.6 X 10'

6.6 X 10'

5.1

Treated waste

Treatment

Filtration of

Incineration, with

Immobilized with

Package with mini-

particulates

ash and blowdown
immobilized with
cement. Filters
compacted
separately.

cement

mum treatment

Volume, m^/GW(e)-yr

0.8 (filters)t

4.2

2.3

Containers/GW(e)-yr

21 (55-gal drums)

1 1 (55-gal drums)

7.5 (55-gal drums)
0.4 (boxes, each
1.2 by 1.8 by 1.8 m)

*7.6 metric tons of mixed-oxide fuel produced per 1 GW(e)-yr.

t Based on fuel fabrication 1 yr after reprocessing (2.5 yr out of reactor).

t Included as primary waste in compactible-trash and combustible-wastes column.

as such can vary considerably in volume and radioactive content, depending on the
day-to-day variables of plant operation. Since there are no volatiles in the gaseous wastes,
filtration to remove particulates is sufficient treatment. As in the fuel reprocessing plant,
excess wastewater is vaporized with the gaseous waste; thus there are no releases of liquid
waste to the environment. The highly contaminated liquid wastes are immobilized in
concrete. The combustible wastes are incinerated and immobilized with cement.

After immobilization and packaging, the mixed-oxide-fuel fabrication wastes are sent
to geologic disposal in either 55- or 80-gal drums or in 1.2- by 1.8- by 1.8-m boxes.

A filtration system comprised of a prefilter and two banks of HEPA filters was used
as the reference for removing contaminated airborne particulates from mixed-oxide-fuel
fabrication-plant ventilation air discharged to the atmosphere. Vaporized excess water
from processing is not filtered but is sent directly to the stack. Estimated activities of the
major transuranic elements directed into the filtration system and subsequently
discharged to the atmosphere are given in Table 9.

Decommissioning Wastes

Nuclear power plants and postfission fuel-cycle facilities become contaminated during
power-production, fuel-cycle, and waste-treatment operations. On retirement these
facilities become a waste that requires management, commonly termed decommissioning.
Various alternatives are available for decommissioning these retired facilities. Three basic
decommissioning modes are considered.

1 04 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 9 Release of Airborne Transuranics from a
Mixed-Oxide-Fuel Fabrication-Plant Center

Activity,*

Ci/GW(e)-yr

Into filter

Into

Element

system

stack

Neptunium

3.7 X 10-'°

2 X 10-'^

Plutonium

1.4

5 X 10"*

Americium

2.2 X 10"'

8 X 10""

* Activity 2.5 yr after discharge; 7.6 metric tons of
mixed-oxide fuel produced per 1 GW(e)-yr.

Protective storage. At shutdown the facOity is prepared to be \ef{ in place tor an
extended period. Temporary physical barriers are constructed between the environment
and radioactive contamination in the facility. Continuing surveillance is required after the
facility has been placed in protective storage. Surveillance continues until all radioactivity
in the facility has decayed or further decommissioning activities are carried out.

Entombment. At shutdown the facility is prepared to be left in place until all
radioactivity has decayed to nonhazardous levels. Permanent physical barriers are
constructed between the environment and the radioactive contamination in the facility.
Minimal surveillance is required at an entombed facility.

Dismantlement. At shutdown all potentially hazardous amounts of radioactive
contamination are removed from the facility to an approved disposal site. Plans for future
use of the site dictate which noncontaminated portions of the facility remaining after
dismantlement will be demolished and removed.

Combinations of these basic modes can also be used to decommission a retired
faciUty. For example, a facility can be placed in protective storage at shutdown and
dismantled after radioactive decay has reduced radiation levels in the facOity.

The decommissioning mode assumed here for each fuel-cycle facility is dismantle-
ment. Decommissioning by dismantlement requires that all potentially hazardous
amounts of radioactivity be packaged and removed from the site to an approved disposal
location. Uncontaminated portions of the facOity can be reclaimed for other uses or
demolished and removed. In either case there would be no restrictions on subsequent use
of the site; no residual from its use in the nuclear fuel cycle would remain.

For mixed-oxide-fuel fabrication plants, immediate dismantlement after a 30-yr
useful hfe is assumed. For nuclear power plants and fuel reprocessing plants, a 30-yr
useful life is assumed, but dismantlement is preceded by 50- and 30-yr periods of
protective storage, respectively, to allow short-Uved activity to decay. During protective
storage of nuclear power plants and fuel reprocessing plants, the radioactivity is
consolidated in portions of the facility with relatively high contamination levels.
Appropriate security measures are established, and a surveillance and monitoring program
is maintained. Because most wastes generated in preparing for protective storage will be
stored on site, major shipments of wastes will occur only at the final dismantlement.

Wastes generated during decommissioning are listed in Table 10. The wastes shown in
Table 10 must be packaged and shipped from the site by truck or rail to an approved
disposal location. There will also be atmospheric releases of gases and releases of water to

TRANSURANIC WASTES FROM LWR CYCLE 105

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106 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

the ground. The gases will be subjected to the standard off-gas treatments normally used
in the plants, and the atmospheric releases will be controlled to below those when the
plants were operating. Water releases will also be below established limits. Water
inventories from cooling basins, for instance, will be released to the ground in the
location of the site but only after suitable analyses are made to ensure that the activity
contained in the basin water is below established limits.

References

Adam, J. A., and V. L. Rogers, 191%, A Classification System for Radioactive Waste Disposal — What

Waste Goes Where? Report NUREG-0456 FBDU-224-10, Office of Nuclear Material Safety and

Safeguards, U.S. Nuclear Regulatory Commission.
Energy Research and Development Administration, 1976, Alternatives for Managing Wastes from

Reactors and Post-Fission Operations in the LWR Fuel Cycle, ERDA Report ERDA-76-43, NTIS.
Miida,J.,W. Haiussermann, and S. Mankin, 1977, Nuclear Power Programmes and Medium-Term

Projections in the OECD Area, in Nuclear Power and Its Fuel Cycle, Symposium Proceedings,

Salzburg, Austria, 1977, pp. 213-232, STI/PUB/465, Vol. 1, International Atomic Energy Agency,

Vienna.
U. S. Department of Energy, 1978, Report of Task Force for Review of Nuclear Waste Management:

Draft, DOE/ER-0004/D.
, 1979, Technology for Commercial Radioactive Waste Management, Report DOE/ET-0028, NTIS.

The Detection and Study of Plutonium-
Bearing Particles Following the Reprocessing
of Reactor Fuel

S. MARSHALL SANDERS, JR., and ALBERT L. BONI

A method has been developed to identify and study individual airborne particles
containing ^^^Pu from fission-fragment and alpha-particle tracks produced by them in a
polycarbonate film with a nuclear-track-emulsion coating. Membrane filters, used to
collect the particles from atmospheric effluents, are cast into films composed of a
polycarbonate matrix containing the particles. When a particle is located, the amount of
^^^Pu in it is determined by counting the tracks, a small portion of the film containing
the particle is isolated, the emulsion removed, the polycarbonate dissolved, the track
replicas oxidized, and the elemental composition of the ^^^Pu-bearing particle
determined by electron-microprobe analysis. The elemental compositions, sizes, struc-
tures, and ^^^Pu contents were determined for 558 plutonium-bearing particles isolated
from various locations in the exhaust from a nuclear processing facility at the Savannah
River Plant. These data were compared with data from natural aerosol particles.

Nuclear fuel reprocessing facilities at the Savanriah River Plant release to the atmosphere
minute quantities (<1 mCi/yr) of ^^^Pu in particulate form. These particles have been
isolated and studied as to size, elemental composition, and radioactive properties with
autoradiographic techniques.

Leary (1951) first used an autoradiographic technique to measure particle size-
frequency distributions of radioactive aerosols in 1950. With his procedure, aerosols fed
to and discharged from a decontamination pilot plant at Los Alamos Scientific
Laboratory were collected on filter paper. A sample of the filter paper was then placed in
contact with nuclear track emulsion for various exposure times. Assuming that the
aerosol particles contained no nonradioactive material and that the isotopic composition
of the radioactive compounds was known, Leary determined the size of each radioactive
particle by counting the number of alpha-ray tracks produced by it in the emulsion for a
given exposure time. This method distinguished between the radioactive and inert
particles and thus was particularly useful for aerosols in which the abundance of these
radioactive particles was low relative to atmospheric dust. It had, however, two serious
deficiencies: (1) the actual particles were never observed and (2) plutonium could not be
distinguished from uranium. In spite of these deficiencies, this method was the basis for
other techniques for more than a decade.

Quan (1959) overcame the first of these deficiencies in 1959 by permanently bonding
the aerosol particles between the nuclear track emulsion and the Millipore filter used in
their collection so that the particles were not separated from the tracks they produced in
the emulsion. With Quan's method, the Millipore filter with the contaminated surface
upward was cemented with collodion to a stainless-steel frame. The upper surface was

107

1 08 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

then covered with a thin collodion membrane and a gelatin bonding medium. A square of
Kodak autoradiographic permeable base film was then floated onto the surface of the
gelatin with the sensitive emulsion upward. After exposure and development, the
Millipore filter was made transparent with a collodion solution.

About the same time Moss, Hyatt, and Schulte (1961) developed a simpler method of
relating the particles with the tracks they produce. These investigators collected airborne
plutonium samples on membrane filters in a processing plant at Los Alamos Scientific
Laboratory wdiere plutonium metal was handled in glove boxes. The membrane filters
were pressed on a glass slide covered with nuclear track emulsion with the contaminated
side in contact with the emulsion. When excess water was blotted from the slide, the
emulsion, which had been softened by submerging it in water for several minutes before
using it, became tacky. The filter was then separated from the emulsion, and the particles
were left embedded in the emulsion.

To calculate the particle size. Moss, Hyatt, and Schulte (1961) assumed that the
particles were spheres of pure ^^^PuOa. However, some particles appeared larger than
calculated, which led the group to speculate that, "when plutonium dust settles on
surfaces and is resuspended during cleaning, the resuspended particles are much larger and
are only partly composed of plutonium."

Andersen (1964) made a particle size study ot plutonium aerosols in employee work
areas at Hanford Laboratories in 1963. Here samples were collected on membrane filters
and HoUingsworth and Vose type 70 filter paper. The filters were contact exposed to
nuclear track emulsions by a method similar to that of Leary (1951). The filters were
mounted in X-ray exposure holders altered so that the nuclear track film could be
positioned reproducibly. These and the use of a microscope-stage micrometer permitted
sufficient reproducibility in film location that a particle in question could be readily
relocated. He found the plutonium particles to be small with a geometric mean diameter
less than 0.04 to 0.1 jum. He assumed that plutonium was not attached to dust particles
since copious quantities of plutonium were generally involved during particle formation
and the effect of foreign particles was not extensive.

An autoradiographic technique for the location and examination of alpha active dust
particles collected on glass-fiber paper, developed by Stevens (1963), was used by
Sherwood and Stevens (1963; 1965) to analyze laboratory air samples taken at the
Atomic Energy Research Establishment at Harwell, England, in 1963. Each filter was
mounted in an Araldite (epoxy resin) mixture, which renders the filter transparent, and
was covered with autoradiographic stripping film. After exposure and development, the
sample was viewed with a high-powered optical microscope. The particles that were
identified as radioactive by the alpha-particle tracks emanating from them were sized, and
their radioactivity was determined by counting the number of tracks. Stevens and
Sherwood found that relatively few of the particles collected were pure plutonium or
plutonium compounds. Most of the particles were large inert particles contaminated with
plutonium.

As late as 1965 Kirchner (1966) used the contact-exposed method of Leary (1951) to
analyze air samples obtained from work areas in plutonium chemistry and fabrication
plants at Rocky Flats. Although this procedure did not permit examination of inert
particles, Kirchner beUeved that the autoradiographs indicated that agglomeration with
inert or other active particles was rare. He also believed that, despite good agreement in
the activity median diameters reported at Harwell (Sherwood and Stevens, 1965), lung

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 109

retention patterns from accidental exposures at Rocky Flats support the premise that
aerosols there are primarily pure compounds of plutonium.

Ettinger, Moss, and Johnson (1971 ) developed a technique in 1971 to measure ^^^Pu
particles, which was a modification of the one used earlier by Moss, Hyatt, and Schulte
(1961), in which the particles were embedded in the emulsion. The samples were
collected from several ^^^Pu02 glove-box areas at Los Alamos Scientific Laboratory by
two-stage (cyclone-filter) air samplers and on gross filter samplers using Millipore filters.
Their modified technique permitted multiple exposures of aerosols to additional nuclear
track plates. Additional plates were placed against the plate containing the embedded
aerosols for shorter periods of time and thus increased the range of sensitivity by
providing both long and short exposures to the same particles. Ettinger felt that the alpha
tracks in the emulsion being symmetrical suggested that the plutonium was not attached
to inert aerosols.

Hayden (1976), at Rocky Flats, was the first to combine alpha-particle and
fission-fragment tracks to isolate '^^^Pu-bearing particles from other fissile material in
1974. The sample was placed in intimate contact with a 10-jum -thick polycarbonate film.
A cellulose nitrate film was then placed on the polycarbonate film. The package was
allowed to set for a predetermined exposure time. The cellulose nitrate was removed, and
the remaining package was irradiated in a reactor for a desired neutron fluence. Both
films were then etched and examined for tracks. The fission-fragment tracks appeared in
the polycarbonate film and the alpha-particle tracks in the cellulose nitrate film. The
presence of fission-fragment and alpha-particle tracks indicated that the particle
contained ^^^Pu. The presence of fission-fragment tracks alone indicated that the particle
contained uranium. The Murri (Hayden, Murri, and Baker, 1972) equation was used to
calculate particle size from the fission-fragment track count, and the Leary (Leary, 1951)
equation was used to calculate particle size from the alpha-particle track count. However,
the sample could not be mounted with precise positioning of reference symbols so that a
specific particle could be evaluated. Thus the actual size of the particle or the presence or
composition of inert material could not be determined.

Only two previous studies of plant effluents have been made, neither of which has
used autoradiographic techniques. The first was of effluent aerosols downstream from
high efficiency particulate air (HEPA) filters undertaken by Mishima and Schwendiman
(1972) in 1971 at the plutonium finishing plant at the Hanford site. Filter and cascade
impactor samples were taken of the stack gases and various exhaust systems of the plant
to determine the aerodynamic characteristics and distribution of plutonium-bearing
particles with their associated radioactivity. They found that the overall efficiency of the
exhaust system was high, that little, if any, of the alpha radioactivity leaving the stack
was being recycled back into the ventilation system, and that the plutonium present
appeared to be attached to large, nonradioactive particles.

Systematic measurements and analyses of plutonium-bearing parficles in off-gas were
also made by Elder, Gonzales, and Ettinger (1974) and Seefeldt, Mecham, and Steindler
(1976) at Los Alamos Scientific Laboratory in 1972. They collected samples from five
operations — two research and development, two fabrication, and one recovery — where
isotopes of plutonium were handled at Rocky Flats, Mound Laboratory, and Los Alamos
Scientific Laboratory. The sampling stations were upstream from the HEPA filters, where
aerosol concentrations were adequately high. Particle size characteristics were determined
by radiometric analyses of the material deposited on each of the eight stages of Andersen
impactors and material deposited on a backup membrane filter that collected particles

1 1 0 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

that passed through the impactor. The material at each impactor stage had been
characterized as being within a certain particle size range. Since the stage at which a
particle is deposited is a complex function of actual particle size, shape, and density, the
unit of size measurement used by them was the activity median aerodynamic diameter
(AMAD), which is the diameter of a unit-density sphere with the same settling velocity as
a plutonium-bearing particle in a population divided so that the radioactivity of all the
larger particles equals that of the smaller ones.

They found that the two fabrication facilities produced the largest AMAD (4.0 and
2.7 ^m) and the recovery facihty produced the smallest AMAD (0.3 fim). The two
research and development facilities produced intermediate size particles.

In 1975 Sanders (1976: 1977; 1978; 1979) began a study of plutonium-bearing
particles in various parts of the chemical separations process exhaust system at the
Savannah River Plant using autoradiographic tecliniques to record both fission-fragment
and alpha-particle tracks.

Methods and Materials

Particle Collection

Particles are collected by drawing a fraction of exhaust air through membrane filters.
These filters are polycarbonate films that are 47 mm in diameter and 5 jum thick with
3 X 10^ 0.1-/.im-diameter pores per square centimeter, which gives a filter porosity of
0.024. The filters are supported in a polycarbonate aerosol holder.* Air is drawn through
the holder by a small diaphragm pump with a Vitonf diaphragm at a rate of 4 liters/min
to give a face velocity at the filter of 3.8 cm/sec. At this flow the total efficiency for
particle collection by the processes of impaction, diffusion, and interception, calculated
according to Spurny et al. (1969) is 100% for all particles with diameters of 0.001 )um
(the diameter of gas molecules) or larger.

Arrangement of the air-sampUng system is shown in Fig. 1. As particles accumulate on
the membrane filters, membrane porosity and airflow are reduced. Integrated airflow is
measured with a dry-type test meter J in series with the diaphragm pump to determine the
fraction of the exhaust sampled. When nitrogen dioxide is present, exhaust gas is passed
through two gas-drying towers between the filter and the pump. The first tower contains
indicating Drierite§ to remove moisture from the air and save the Ascarite^ in the second
tower. The self-indicating Ascarite, in turn, absorbs nitrogen dioxide to protect the pump
and the dry test meter. A small flowmeter is mounted on the exhaust side of the dry test
meter to give an indication of the instantaneous flow rate through the system. Air from
the meter is fed back into the exhaust system to prevent its release to the service area.

Film Preparation

Figure 2 shows the procedure for converting the particle-containing filter membrane to a
polycarbonate film. After air has been sampled, the radioactivity retained on each filter is

*The aerosol holders and membrane filters were produced by Nuclepore Corporation, Pleasanton,
Calif., and obtained from them or Bio-Rad Laboratories, Richmond, Calif.
fTrademark of E. I. du Pont de Nemours & Company, Inc.
rfManufactured by the American Meter Division of Singer.
§Trademark of W. S. Hammond Drierite Company,
f Trademark of Arthur H. Thomas Company.

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 11 1

SAMPLE LINE

FILTER

EXHAUST

DRIERITE
TOWER

ASCARITE
TOWER

LD.

i — bLI

DIAPHRAGM
PUMP

FLOWMETER

DRY TEST
METER

Fig. 1 Arrangement of sample collection equipment.

47-mm MEMBRANE
HOLDER

ALPHA COUNTER

POLYCARBONATE

MEMBRANE

FILTER

ACRYLIC SUPPORT

POLYCARBONATE
FILM ON ACRYLIC
SUPPORT

DICHLOROETHANE
SOLUTION

2-in. BY 2-in.
GLASS PLATE

ACRYLIC GLASS
SANDWICH

Fig. 2 Procedure for preparing polycarbonate films.

/ / 2 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

measured before it is liandled in the laboratory. Each fiUer is then dissolved in a 40%
(vol. /vol.) solution of 1 ,2-dichloroethane in dichloromethane. The filters are folded, and
each is placed in a 1-ml volumetric flask. A second clean, unused filter is placed in the
same flask to give sufficient polycarbonate to form a 50-mm^ film. Volume of the
dichloroethane solution in the flask is adjusted to about % ml. This mixture is stirred
until the polycarbonate filters dissolve. The flasks are stoppered and allowed to stand for
30 min to allow trapped air bubbles to rise to the surface.

The clear polycarbonate solution containing the particles is poured onto a clean,
50-mm (2-in.) square glass plate (see Fig. 2). One edge of a second 50-mm^ glass plate is
used to spread the solution evenly over the surface of the first plate. The solution is
stirred continuously with the second plate for about half a minute while the solution
thickens. A 50-mm^ 1.6-mm-thick acrylic support with a 45-mm-diameter hole is placed
on top of the wet film. The support and plate combinafions are placed in covered petri
dishes for 16 hr while the films continue to dry.

The glass plates are then removed by dipping the support and plate combinations in
distilled water and prying the supports from the glass with tweezers.

Film Irradiation

The cast film is irradiated in a thermal neutron fluence of about 9 X 10'^ neutrons/ciu^
to produce fission -fragment tracks in the polycarbonate film by which particles
containing fissionable material can be identified. Films are arranged for irradiation by
stacking the supports on top of each other, thus sandwiching each film between two
supports. Included in the stack are blank films that are prepared in the same way as the
sample films from clean unused filters. The assembled stack is wrapped with cellophane
tape. Wrapped with each stack are preweighed 25.4-mm-diameter 0.25-mm-thick type
302 stainless-steel disks. The induced radioactivity from 27-day ^ ' Cr in these disks is
later measured to determine the thermal neutron tluence to which the particles are
exposed.

The packaged stacks are irradiated in a 3-in.-diameter hole in a light-water-cooled
enriched-uranium-fueled standard pile with graphite reflectors (Axtmann et al., 1953).
Following irradiation, the induced radioactivity of the stacks is allowed to decay several
days before the packaged stacks are returned to the laboratory.

Film Etching

The polycarbonate film is etched for 10 min in 6A'NaOH at 52 to 55°C to make the
fission -fragment tracks visible with an optical microscope. During this etching process, a
portion of all polycarbonate surfaces is dissolved: the outer surface of the cast film, the
surface around the particle, and especially that along the fission-fragment tracks.

Emulsion Coating

For the identificafion of the fissionable material in each particle, the alpha-particle
emission rate is measured by coating the polycarbonate film with a photographic
emulsion, which is developed after a predetermined exposure time.

Kodak type NTB nuclear track emulsion is used to coat irradiated films. Under
darkroom lighting (No. 2 Wratten-filtered) a 4-oz. jar of emulsion is partly immersed in a
water bath that is maintained at 40°C until the emulsion melts (between 15 and 20 min).
Slightly over half the molten emulsion is carefully poured into a narrow polyethylene

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 113

container in the water bath. The molten emulsion is tested by dipping a clean glass
microscope sUde into it and exarnining the coat on the glass under a safeHght to
determine whether bubbles are present. If bubbles are present, they are scooped from the
surface of the molten emulsion with a porcelain spoon.

The polycarbonate films are coated with emulsion by holding the supports containing
the films vertically by one corner and dipping them into the clear molten emulsion for
about 1 sec. The films are kept vertical until the excess emulsion has drained off. The
coated films are then placed horizontally in a TH-Junior temperature-humidity test
chamber* that is maintained at 28°C and about 80% relative humidity until the emulsion
cools and gels (about 30 min).

Exposure

The polycarbonate films are exposed for 1 week before being developed to determine the
alpha-particle emission rate for each aerosol particle. The films are stored in spun
aluminum Desicoolersf containing 60 g of indicating Drierite during this exposure of the
emulsion to the particles. The Desicoolers are sealed with black adhesive tape and stored
in a refrigerator between 4 and 5°C for the duration of the exposure.

Emulsion Processing

At the end of the exposure period, the alpha-parficle tracks in the emulsion are
developed, and all substances other than tracks are removed from the emulsion. The
emulsion is developed in a 1:1 solution of DektolJ developer for 3 min at 17 C
(Eastman Kodak Company, 1976; Boyd, 1955; Kopriwa and Leblond, 1962).

Immediately following development the film is rinsed in 28% (vol. /vol.) acetic acid
for 10 sec. The high acid concentrafion is used to prevent reticulation of the emulsion
and its separation from the supporting polycarbonate film.

The rinsed emulsion is fixed by placing it for 5 min in a 1 : 3 dilution of Kodak rapid
fixer concentrate § containing 2.8% (vol. /vol.) hardener concentrate.

A batch process is used to wash all chemicals except the metallic silver from the
emulsion. The emulsion-covered film is placed in distilled water, and the chemicals in the
emulsion and wash water are allowed to approach equilibrium for 2 min. The emulsion is
then placed in a second container of distilled water while the water in the first container
is being changed. This process is repeated a total of eight times. After the water wash, the
emulsion-coated polycarbonate films are placed in racks and allowed to dry in a dust-free
atmosphere.

Track Counting

The film is prepared for track counting by placing the acrylic support on a 50-mm-square,
1.0-mm-thick polycarbonate block.

Particles with tracks are located under a Bausch and Lomb zoom stereomicroscope by
using transmitted light and a magnificafion of 105 X. Each particle with tracks is circled
with a felt-tip marking pen. After a particle has been marked, the support and block
holding the film are moved to a Zeiss photomicroscope where the fission-fragment and

*Manufactured by Tenney Engineering, Inc., Union, New Jersey.

■fTrademark of Fisher Scientific Company.

if Trademark of Eastman Kodak Company.

§ Kodak Photographic Products catalog numbers 146 4106 or 146 4114.

1 14 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

alpha-particle tracks are counted by using transmitted light and a magnification of 1 000 X .
Epiplan, flat-field objectives are used because they are corrected for uncovered specimens
and do not require cover glasses. The numbers of alpha-particle tracks in the lower and
upper emulsions are added to give the total number of alpha tracks observed.

Three Polaroid pictures of tracks from a single particle — one with the focal plane in
the lower emulsion, one in the polycarbonate film, and one in the upper emulsion — are
shown in Fig. 3.

Identification of Fissionable Materials

Table 1 gives the theoretical ratios of alpha-particle to fission-fragment tracks which
would be produced from particles irradiated with a fluence of 8.64 x 10''* thermal
neutrons/cm^ when there is a 7-day interval between film casting and etching and during
exposure to nuclear track emulsion. The stipulation that etching follows film casting by 7
days is included because spontaneous fissions will add to the number of fission-fragment
tracks during this period.

The atom percents in the uranium mixtures in Table 1 are given in Table 2, and those
in the plutonium mixtures are given in Table 3.

The isotopic mixtures of plutonium contain ^"^'Pu. All but 0.0023% of this nuclidp
decays by beta emission to ^'*'Am, which has a 433-yr half-life. The amount of mis
americium nuclide in a mixture will reach a maximum of 0.887 of the initial ^"^ ' Pu atom
percent in 74.6 yr. Americium-241 will add additional alpha tracks to those from
plutonium. Thus two ratios are given in Table 1 for each plutonium mixture, one for
freshly purified plutonium and one for 75-yr-old plutonium containing the maximum
^^ ^ Am activity. In two of the three mixtures, this caused an increase in alpha-particle-to-
fission-fragment ratios. However, with heat-source plutonium, the decrease in *^^Pu
activity was not compensated for by the increase in "'^ ' Am activity.

This identification procedure can be used to distinguish particle-bound plutonium
from uranium. Table 2 shows that, of the six isotopic mixtures of uranium, only the
higWy enriched uranium mixture will give a number of fission-fragment tracks
comparable to that of the plutonium mixtures. Even if there should be enough uranium
to produce fission-fragment tracks, mixtures of these isotopes would not produce
alpha-particle tracks.

This procedure can be used not only to identify plutonium but also to identify the
plutonium isotopic composition in a particle. For example, a particle having 10
fission-fragment tracks would also have 5 alpha-particle tracks if the mixture were
low-irradiation plutonium, 640 alpha-particle tracks if it were higli-irradiation plutonium,
and 5080 alpha-particle tracks if it were heat-source plutonium.

Table 1 includes (in addition to uranium and plutonium track data) a number of
curium and californium nuclides that could mimic the plutonium mixtures. Some of these
nuclides decay by spontaneous fission. The polycarbonate film should be allowed to
stand several weeks after casting and then be etched both before and after thermal
neutron irradiation to detect spontaneous fissioning. Under these conditions tracks due to
spontaneous fissioning will appear in unirradiated films.

Measurement of Plutonium and Uranium Ratios

For a demonstration of the effectiveness of this identification method in distinguishing
between plutonium and uranium, samples of particles were obtained from two sources of

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 115

O
a

E
o

¥^

•

'^y

4M»

•/

e
o

4>
«

"2 -^
c u
n ^-^

4> O

?■ s

■< w

tb a.

1 1 6 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Theoretical Number of Fission-Fragment and Alpha-Particle Tracks from
10' ° Atoms and the Ratio of Alpha-Particle to Fission-Fragment Tracks

Ratio of alpha-

Fission-fragment

Alpha-particle

particle tracks to

tracks

per

tracks per

fission-fragment

Nuclides

10' " atoms

10' ° atoms

tracks

Power reactor fuel

(4% " 5 U)

3.99 X

10^

2.14 X 10"'

5.36 X 10"*

Low-burnup uranium

(2.5% ^^^U)

2.50 X

10^

1.50 X 10"'

6.00 X 10""

Highly enriched uranium

(90% ^ ^ = U)

8.98 X

10^

6.54

7.29 X 10""

Natural uranium

(99% ^ ^ « U)

7.18X

10'

5.74 X 10""

7.99 X 10""

Depleted uranium

(-100% ^^*U)

2.49 X

10'

3.60 X 10"^

1.44 X IC

High-burnup uranium

(1%^^ = U)

8.88 X

10'

1.44 X 10~'

1.62 X 10"^

^"'Cm

1.25 X

10^

8.10

6.48 X 10"'

242/«^j^

1.31 X

10^

4.20 X 10="

3.19X 10"^

233U

9.06 X

10^

8.20 X 10^

9.05 X 10"'

2'»5Cm

3.73 X

10"

1.56 X 10"

4.17 X 10"'

Low-irradiation plutonium

(94% ^ ^ ' Pu)

1.20 X

10"

(6.48 to 7.52) X 10'

(5.40 to 6.27) X 10"'

^''^Cm(SF)

<6.64

8.10

>1.22

25 1 Cf

8.30 X

10"

1.48 X 10'

1.78

^''«Cm(SF)

5.78 X

10'

3.37 X 10^

5.84

High-irradiation plutonium

(40% ^^'Pu)

8.02 X

10^

(5.13 to 9.31) X 10"

(6.40 to 1.16) X 10'

249(2 J-

2.88 X

10"

3.79 X 10=

1.32x 10'

^ = ^Cf(SF)

3.13X

10*

5.02 X 10'

1.61 X 10'

^"^Cm

1.19x

10"

4.43 X 10"

3.71x10^

Heat-source plutonium

(80%^^«Pu)

2.40 X

10^

(1.22 to 1.07) X 10*

(5.08 to 4.46) X 10^

^ = °Cf(SF)

1.56 X

10"

1.02 X 10'

6.50 X 10^

^''*Cf(SF)

1.56 X

10'

2.76 X 10"

1.76 X 10'

^"'Am

5.43 X

10'

3.07 X 10=

5.66 X 10'

252Cf

5.53 X

10^

5.02 X 10'

9.08 X 10"

^^^Cm(SF)

2.03 X

10'

7.34 X 10*

3.62 X 10 =

known nuclide mixtures, one of low-irradiation plutonium and one of highly enriched
uranium. Polycarbonate films were prepared containing particles from either one or the
other source. The films were irradiated and coated with emulsion, and the emulsion was
exposed and developed by this procedure. The number of alpha-particle and fission-
fragment tracks with each particle were counted.

The data from 315 particles containing low-irradiation plutonium are given in Table 4
and those from 350 particles containing higlily enriched uranium are given in Table 5.
The data were ranked according to the number of observed fission-fragment tracks per
particle to determine whether the number of tracks influenced the measured ratios. The
mean and standard deviation of the ratios in each track interval are also given.

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING Ul
TABLE 2 Atom Percents in Six Typical Mixtures of Uranium

Power

Low-

High-

Highly

Natural

reactor

bumup

enriched

Depleted

Nuclide

uranium

fuel

uranium

uranmm

234,J

0.005

0.030

0.018

0.017

1.160

0.001

235U

0.720

4.000

2.506

0.890

90.000

0.250

236^

0.320

0.363

0.393

2.770

0.014

238U

99.275

95.650

97.113

98.700

6.070

99.735

TABLE 3 Atom Percents in Three Typical
Mixtures of Plutonium

Nuclide

■Pu 5 X 10"" 1 X 10-

'Pu 0.0115 2.9 80.3

'Pu 93.6 39.6 15.87

Low-

High-

Heat-

irradiation

source

Plutonium

24 0py

5.9

25.6

3.00

24. Pu

0.4

16.8

0.72

242p^;

0.013

15.0

24 4 p^,

0.02

From Tables 4 and 5 the mean ratio (alpha-particle-to-fission-fragment tracks) for
low-irradiation plutonium is 9.1 X 10"^ and that t\)r highly enriched uranium is
1.8 X 10~^ Thus '^^Pu can clearly be distinguished from ^^^U by this procedure if
there is a sufficient number o{ tracks. However, these ratios are 1.7 and 2.5 times the
theoretical ratios given in Table 1. For highly enriched uranium, all alpha-particle tracks
were observed as single tracks only, some of which may have been due to background
radiation; this would explain the higlier mean ratio for uranium. With plutonium the
higlier observed ratios are probably due to the geometry of the media in which the tracks
are formed.

Quantitative Radiographic Analysis

Alpha-particle and fission-fragment track counts will provide not only a ratio from which
the fissionable material carried on the particles can be identified but also an estimate of
the quantity of the radioactive nuclides present. One femtocurie (1 fCi = 10' '^ Ci) of
^^^Pu will produce about 22 alpha particles in a week, and, when irradiated with a
fluence of 8.64 x 10'"* thermal neutrons/cm', it will produce about 40 fission fragments.
In a mixture of low-irradiation plutonium, the number of fission fragments produced will
be increased to 53 with between 28 and 33 alpha particles, depending on the age of the
mixture. Only about half of these particles will produce tracks; yet this radiographic
technique is much more sensitive than electron-microprobe analysis, which is not sensitive
to less than 10 fCi (Sanders. 1976).

It 8 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 Analyses of Particles Containing Pu

Total

Ratio of alpha-

Fission-

Number

fission-

alpha-

particle tracks to

fragment

particle

fission-fragment

Standard

tracks

particles

tracks

deviation

3-4

125

108

0.86

0.68

5-9

528

358

0.68

0.49

10-14

783

569

0.73

0.44

15-19

607

489

0.81

0.45

20-24

570

432

0.76

0.38

25-29

294

275

0.94

0.50

30-34

408

417

1.02

0.38

35-39

262

287

1.10

0.32

40-44

672

819

1.22

0.44

45-49

423

398

0.94

0.45

50-54

212

302

1.42

0.72

55-59

0.71

60-64

371

362

0.98

0.30

65-69

136

124

0.91

0.08

80-84

160

172

1.08

0.46

85-89

1.00

90-94

182

147

0.81

0.18

100-104

1
315

104

0.54
0.91

Total

5981

5442

TABLE 5 Analyses of Particles Containing U

Total

Ratio of alpha-

Fission-

Number

fission-

alpha-

particle tracks to

fragment

particle

fission-fragment

tracks

particles

tracks

3-4

124

435

0.0069

5-9

146

935

0.0021

10-14

460

0.0000

15-19

293

0.0000

20-24

214

0.0000

25-29

0.0000

30-34

126

0.0000

35-39

110

0.0000

40-44

3
350

125

0.0000

Total

2780

0.0018

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 119

Particle Isolation

After a particle has been identified and photographed and the tracks have been counted,
it is excised from the fihn in a polycarbonate square. For this the support and block
holding the film are returned to the stereomicroscope. hi transmitted illumination and at
a magnification of 105 x. two parallel cuts are made through the emulsion-coated film on
either side of the particle with an ultra microlance. The film is then rotated through 90°,
and two more cuts are made; the cut film then forms a square [Fig. 4(a)]. The cut
square is then probed in one corner by a 15-mm-long. electrolytically sharpened tungsten
needle (made by placing a pair of 0.52-mm-diameter tungsten wires in a 37VNaOH
solution and applying a 60-Hz 10-volt poiential between them for 10 to 15 min). With
this needle the cut square containing the particle is lifted from the film and placed on a

POLYCARBONATE SQUARE
ON GLASS SLIDE

(d)

MOTION OF COVER GLASS
(e)

BERYLLIUM
SAMPLE BLOCK

GLASS
ICROBRUSH

POLYCARBONATE
SOUARE

REMOVAL OF POLYCARBONATE FROM
PARTICLE ON BERYLLUM BLOCK

Fig. 4 Procedure for mounting particle for fissionable material identification. See text
for explanation of (a) etc.

120 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

glass microscope slide [Figs. 4(b), 4(c), and 4(d)]. The polycarbonate square is freed
from the needle by rotating it so that the corner of the square opposite that stuck by the
needle strikes the slide and causes the square to rotate and fall.

The emulsion layers are then removed from the polycarbonate square by placing a
cover glass on top of the square. Water is introduced between the cover glass and slide
with a glass microbrush made from a 20-|Ul glass disposable pipet. (A 0.025-mm-diameter
tungsten wire doubled and threaded through the lumen forms a loop at one end. A small
amount of glass wool is placed through this loop and is then drawn into the end of the
pipet. The glass fibers are then cut off about 2 mm from the end of the pipet.) The
microbrush is dipped in water, and the glass fibers are touched to the edge of the cover
glass to allow the water to flow from the brush to between the slide and the cover glass.

The emulsion is then removed by gently moving the cover glass a few millimeters
from side to side [Fig. 4(e)] ; this rolls the swollen emulsion from the surface of the film
but not from the fission-fragment tracks themselves. The cover glass is carefully lifted
from the glass microscope slide, taking care not to lose the polycarbonate square
containing the particle.

Particle Mounting

To mount a particle, the polycarbonate square is placed in a selected grid location on a
beryllium sample mounting block* [Fig. 4(f)] . These sample mounting blocks are 25 mm
in diameter and 13 mm thick and fit the standard electron-microprobe sample holders,
which grip the sides and provide the necessary electrical contact. The top surface of the
block is highly polished and contains a grid network of 1-mm squares inscribed on the
surface. The squares are numbered in mirror-image fashion both vertically and
horizontally through the center.

With coaxial (reflected light) illumination and 15 X magnification under a stereo-
microscope, the polycarbonate squares are moved from the microscope slide to the
beryllium block with an electrolytically sharpened tungsten needle.

The polycarbonate square is then dissolved and washed back from the particle with
dichloroethane, leaving the particle usually connected to the main body of polycarbonate
by a thin isthmus of plastic. This connection does not seriously affect the microprobe
analysis and aids in later locating the particles and holding them on the beryllium block.
A glass microbrush is rinsed in dichloroethane to remove any foreign material and is filled
by immersing the bristled end in a second beaker of dichloroethane. The magnification
was increased to 105 x. Dichloroethane from the brush is dispensed on the beryllium
block just in front of the polycarbonate square until the square is engulfed in the
solution. The microbrush is then used to push the solution back from the particle. Gelatin
replicas of the fission-fragment tracks remained with the particles.

The beryllium block is returned to the photomicroscope where a second Polaroid
picture of each particle is made at a magnification of 556 x to identify the particles after
the gelatin has been removed.

The gelatin with each particle is oxidized by exposure to an oxygen plasma for 3 hr in
a low-temperature asher.t In this asher a gas plasma is generated in oxygen with the
energy of electrons in the gas. Power is supplied to electrons at 13.56 MHz by a

*Walter C. McCrone Associates, Inc., catalog number XIlI-403-3.
t Manufactured by international Plasma Corporation.

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 121

radiofrequency generator. Since the energy to do this with a low-temperature asher is
provided through the electrons instead of heat energy, higli-temperature degradation,
volatiUzation, or fusion of the inorganic constituents of the particles is eliminated.

Figure 5 illustrates the last three stages in the preparation of one particle. Part a is the
particle in the polycarbonate film with emulsion stripped off. Part b is the same particle
with the polycarbonate removed showing the gelatin replicas of the fission-fragment
tracks. Part c is a scanning electron micrograph of the particle after oxidation of the
gelatin. In this photograph traces of the gelatin replicas and silver grains can be seen. Here
what had appeared to be a single particle is actually a conglomeration of at least five and
possibly ten smaller particles.

Particle Sizing

For control of particles after the gelatin track replicas have been oxidized, the beryllium
sample block is returned to the photomicroscope where each particle is located and
photographed again under reflected Hglit by using Polaroid film and a magnification of
556 X. An arrow pointing to the particle is marked on the film so that there will be no
mistake in what is intended for analysis.

The size of each particle is estimated from these Polaroid pictures taken after
oxidation of the completely denuded particles. An average of the smallest and largest
dimensions of the photographed particle is measured in micrometers and divided by the
magnification.

Elemen tal A nalvsis

For the detemiination of elemental composition of the particles, the particles are
analyzed on a Cameca MS46 electron microprobe, equipped with four crystal,
wave-length-dispersive spectrometers (take-off angle, 18°) and an EDAX
701/MICROEDIT* energy-dispersive analyzer. X-ray intensities resulting from the
electron bombardment of the particles and particle sizes and shapes are estimated. These
estimates, along with estimated average densities, are used in the FRAME program
(Yakowitz. Myklebust, and Heinrich, 1973) as modified for particles work by Armstrong
and Buseck (1975) oh a UNIVACt 1110 computer. This calculation gives the particle
composition in elemental weiglit percents. Ratios of the elemental weight percents are
used to calculate enrichment factors, explained in the appendix, which are used to
compare the composition of these particles with that of other aerosols.

Sampling Locations

Particles were collected from air in both exhaust s>stems in nuclear fuel reprocessing
facilities at the Savannah River Plant. A schematic diagram of these systems is given in
Fig. 6. System I takes room air from inside wet cabinets (where plutonium is in solution)
and tVom work areas and exhausts it via the JB-Line stack (Sanders, 1976). System II
takes air from the mechanical line (where plutonium is handled in metallic form) and
exhausts it via the 291-F stack. In System I samples were taken of unfiltered cabinet air
from the fifth and sixth levels (sampling points 29 and 30, respectively), of filtered air
from both locations (sampling point 27), unfiltered room air from the fifth level

*Trademark of EDAX Internation, Inc.
fTrademark of Sperry Rand Corporation.

122 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Lf)

«4.-l

c«

■4^

•a

k«

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Vrt

*1M

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_4>

"a;

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£3

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r^ •

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 123

SAMPLING POINTS
O SYSTEM I
<C> SYSTEM 11

JB-LINE 156-ft
STACK

6th LEVEL
WET CABINET

HEPA

MECHANICAL LINE
AIR SAMPLE EXHAUST

FURNACE OFF-GAS
VESSEL VENT

PROCESS VACUUM
SYSTEM

6th LEVEL
ROOM AIR

',24;

AIR DRYER
SYSTEM

MECHAN
CAL LINE

5th LEVEL
WET CABINET

5th LEVEL
ROOM AIR

SYSTE

A-LINE

772-F OFF-GAS
SYSTEM

PROCESS
VESSEL VENT

WARM CANYON

HOT CANYON

RECYCLE
VESSEL VENT

SAND FILTER

re^

OLD |_|

B-LINE

CENTRAL
CANYON AIR

lAl

SYSTEM 11

Fig. 6 Savannah River Nuclear Fuel Reprocessing Facility exhaust systems.

(sampling pt)int23). and o\ an at the 1 56-t't level of the JB-Lhie stack (sampling
pomt 28). In S\stem II samples were taken o\ mechanical line air from just beyond the
first HHPA filters located m back o{ the cabinets (sampling point A or 31); o'i the
combined air from the mechanical line, air sample exhaust, furnace off-gas vessel vent,
process vacuum system, and au-dryer system after the second HEPA filter (sampling
pt>int B or lb): of the air leaving the sand filter, which also contained air t>om the
support laborator\ i)ff-gas system o\ building 772-F. the fuel dissolving and extraction
process vessel vent system, and building 221-F canyons containing the process vessels
(sampling point C);and of an from the 50-ft level in the 291-F stack where air from the

i24 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

sand filter mingles with that from the uranium recovery A-line and other sources
(sampHng point D).

A total of 121 particles were analyzed from System 1(16 from sampling point 23, 67
from point 29, and 38 from point 30) and 417 from System II (125 from sampling
point A, 107 from point B, 1 14 from point C, and 71 from point D). These figures do not
include 20 particles that contained no elements with atomic numbers greater than 9 and
were assumed to be organic.

Grouping of Data by Enrichment Factors

The results were expressed in terms of "enrichment factors" (dimensionless ratios of
elemental concentrations), which enabled the intercomparison of the compositions of
plutonium-bearing particles with other atmospheric aerosols and the intracomparison
among particles collected from different sampling points. A definition of enrichment
factors and an explanation of their development and application in this work are given in
the appendix.

For a comparison of the chemical composition of the particles collected from
Systems I and II with each other and with the average for global crustal aerosol, the
particle analyses were grouped according to the level of the enrichment factors. Four
groups were established for each element by using the elemental concentration data in
Table A.l of the appendix. The first group contained particles with no detectable
amounts of the element sought. The second group contained detectable amounts with
enrichment factors less than one standard deviation below the geometric mean
enrichment factor, EFg/Sg. The third group contained particles with enrichment factors
between the lower and upper limits of one standard deviation from the geometric mean
enrichment factor, EFg/Sg and EFg X Sg, respectively. The fourth group contained
enrichment factors greater than one standard geometric mean enrichment factor,
EFg X Sg. The third column of Table 6 gives the percent of the particles analyzed which
gave positive analyses for each element. The fourth, fifth, and sixth columns of Table 6
contain the percent of those having positive analyses which had enrichment factors less
than, between, and more than the lower and upper limits of the geometric standard
deviation.

For a comparison of the chemical composition of particles collected at the various
sample points in System II with each other and with global crustal aerosol (Table A.l),
this process was repeated, and the results are listed in Table 7.

Particles with no detectable amounts of an element were not counted with those with
enrichment factors less than the lower limit for the geometric standard deviation (s^,)
because there can be no zero or negative concentration of enrichment-factor values in
log-normal frequency distributions. Thus the size of the three groups is expressed as the
percent of the particles giving positive analyses rather than the percent of the total
number of particles.

Particle Evaluation by Size

In this study particles were selected for analysis on the basis of the number of observed
fission -fragment tracks. Since there were many more particles than could be analyzed,
those having 3 or 4 tracks were generally passed over in favor of those surrounded by 50
or more tracks. The selection of particles for analysis, however, was not biased by

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 125

TABLE 6 Comparison of Analyses of Particles
from Systems I and II

% of positive analyses*

Positive

Less

Greater

Element

System

analyses, %

thant

Within t

than§

Silicon

100

Aluminum

100

Iron

Calcium

Sodium

Potassium

Magnesium

Titanium

Phosphorus

Manganese

100

Barium

0.5

100

Sulfur

Chlorine

Chromium

Nickel

•

Zinc

Cobalt

100

Scandium

0.2

100

Copper

Tungsten

100

0.5

100

Cadmium

0.2

100

*The percent of the positive analyses less than, within, and
greater than one geometric standard deviation of the global
geometric mean enrichment factor.

t EF < EFg/sg.

tEFg/sg< EF < EFgSg.

§EF> EFgSg.

126 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Comparison of Analyses of Particles from
Sampling Points A, B, C, and D of System II

% of positive ar

lalyses*

Sampling

Positive

Less

Greater

Element

point

analyses, %

thant

Within t

than§

Silicon

100

Aluminum

100

Iron

100

Calcium

Sodium

Potassium

Magnesium

Titanium

Phosphorus

100

Manganese

100

Barium

100

Sulfur

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING \21

TABLE 7 (Continued)

% of positive analyses*

Sampling

Positive

Less

Greater

Element

point

analyses, %

thant

Withint

than§

Chlorine

Chromium

100

Nickel

100

Zinc

Cobalt

100

Scandium

100

Copper

100

Tungsten

100

Cadmium

100

*The percent of the positive analyses less than, within, and
greater, than one geometric standard deviation of the global

geometric mean enrichment factor.
tEF < EFg/sg.

|EFg/sg < EF < EFg Sg.

§EF >EFgSg.

physical size. The size of the particles was not measured until after the particles had been
mounted and the polycarbonate film containing the tracks dissolved. Thus the size
distribution of the analyzed particles is indicative of the size distribution of particles in
the aerosol carrying most of the plutonium.

Cumulative frequency plots were constructed for particles from Systems I and II.
Particles in each system were first ranked in order of their approximate diameter (in
micrometers) from the smallest to the largest. A list of the number of particles with
successively larger diameters was made. A cumulative total of the number of particles at
increasing diameter segments was calculated and then normalized by dividing by the total
number of particles from each system. This gave the fraction of the particles having a
diameter equal to or smaller than any particular diameter. Table 8 lists the particle

128 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Comparison of Size Distributions of Particles from
Systems I and II with Natural Aerosols*

Fraction with diameter <

Diameter

System

Sampling

Natural

(D). Mm

point A

point B

point C

point D

aerosol

0.4

0.03

0.01

0.5

0.04

0.02

0.9

0.07

0.04

0.06

1.1

0.09

0.08

0.25

1.2

0.10

0.42

1.4

0.11

0.64

1.7

0.14

0.11

1.8

0.28

0.02

0.17

0.83

2.2

0.30

0.32

0.21

0.91

2.5

0.34

0.23

2.7

0.35

0.04

0.01

0.37

0.24

0.949

3.0

0.42

0.31

3.3

0.43

0.32

3.6

0.53

0.10

0.06

0.46

0.34

0.979

3.9

0.11

0.983

4.0

0.07

0.51

0.38

4.4

0.52

0.41

4.5

0.54

0.13

0.10

0.989

5.0

0.11

0.55

0.48

5.4

0.62

0.20

0.18

0.60

0.994

5.8

0.20

0.59

6.1

0.22

0.61

0.56

6.3

0.23

0.26

0.996

6.7

0.27

0.64

0.58

7.0

0.67

7.2

0.67

0.33

0.35

0.68

0.997

7.4

0.68

7.8

0.36

0.70

8.0

0.68

0.34

0.37

0.71

0.59

0.998

8.6

0.39

0.63

9.0

0.75

0.40

0.46

0.74

0.69

0.999

10.0

0.41

0.48

0.78

0.75

10.8

0.79

0.44

0.60

0.81

11.7

0.46

0.61

0.82

0.79

12.6

0.83

0.54

0.62

0.83

0.80

1.000

13.5

0.55

0.66

0.82

14.4

0.84

0.58

0.89

14.9

0.67

0.83

16.2

0.88

0.61

0.71

17.1

0.62

0.89

18.0

0.92

0.64

0.75

0.90

20.7

0.70

0.81

0.91

0.92

21.6

0.93

0.71

0.82

23.4

0.74

0.85

0.930

24.3

0.75

0.86

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 129

TABLE 8 (Continued)

Fraction with diameter < D

Diameter System Sampling Sampling Sampling Sampling Natural
(D), Mm I point A point B point C point D aerosol

0.972
0.986

25.2

0.94

0.78

0.87

26.9

0.95

0.81

0.91

0.939

27.9

0.83

0.92

28.8

0.93

0.956

30.6

0.86

0.935

31.5

0.87

0.974

32.4

0.88

0.944

33.5

0.89

0.963

34.2

0.959

0.94

0.972

0.982

35.1

1.000

36.0

0.975

0.944

0.981

39.6

0.968

0.991

41.4

0.983

1.000

50.4

0.992

53.9

0.992

59.4

1.000

62.9

1.000

*The percent ot the positive analyses less than, within, and greater than
one geometric standard deviation of the global geometric mean enrichment
factor.

diameters (in micrometers); and, in columns 2, 3, 4, 5, and 6, the fraction of the particles
having diameters equal to or less than each diameter measured in System I and sampling
points A, B, C, and D in System II, respectively. These fractions are also plotted on the
logarithmic probability graph given in Fig. 7.

For comparison a cumulative frequency plot was also made of the size distribution of
particles in natural atmospheric aerosols. A very simple function that has been used
extensively in atmospheric research to express particle size distribution in both natural
and polluted atmospheres is

where N is the number concentration or total number of particles per unit volume having
diameters from the lower limit of definition of aerosols up to diameter D (in
micrometers). From the relationships

dD=Dd(lnD) (2)

and

In D= In lOlouD (3)

130 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

<
Q

atural aerosol
particles with
diameters > 1 Mm
near the surface
in continental air

10 20 30 40 50 60 70 80 90 95 98 99
PARTICLES WITH DIAMETER < D, %

Fig. 7 Size distribution plots for natural and collected partijles.

the more useful expression

ddog D)

= (ln 10)aD"

(4)

is obtained, where c= b — 1 and dN/d (log D) is the number distribution. Junge (1965)
found c to be about 3 over the size range — 0.7 < log D < 1.5 or 0.2 < D < 32 //m.
Integrating the first equation between Do and D (Dq < D) gives

^, aD-^l^° a/1 1\

(5)

Instead of expressing the distribution as the number of particles per unit volume, it can
be expressed as a fraction, F, of the total number of particles, or

(6)

where Nj is the total number of particles when D = °° and Nj - a/3Do. For a reasonable
distribution, only those particles which could be easily seen with an optical microscope
were included. Thus Do was assumed to be 1 jum, and Eq. 6 can be expressed as

i
D^

F= 1 --^

(7)

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 131

The frequency distribution for natural aerosols with particle diameters between 1
fxm and D, calculated from this expression, is given in column 7 of Table 8 and plotted in
Fig. 7.

To see how closely the distribution of particle diameters resembles a log-normal
distribution, we assumed that the observed diameters represented a sample of a
population having a log-normal distribution. The geometric mean diameter, Dg, and
geometric standard deviation, Sg, were calculated from these data by using equations
similar to those given earlier for the geometric mean enrichment factor and geometric
standard deviation. These values are given in Table 9. Values for the upper 68.27% limit
for the diameters were calculated from the product of D^ and Sg. The best-fit log-normal
probability curves were plotted on the logarithmic probability graph in Fig. 7 by drawing
straiglit lines through coordinates for Dg and Dg X Sg on the 50.00 and 84.14*
cumulative percent abscissae, respectively.

TABLE 9 Distribution of Particle Diameters in
Systems I and II

Geometric

Data

mean

standard

Sample

points.

diameter

deviation

Skewness

System

location

(Dg)

(Sg)

(SK)

121

4.64

2.92

0.71

125

12.27

2.24

0.04

107

10.82

1.93

0.34

114

4.48

2.75

0.37

5.43

2.69

0.23

To detennine the degree of asymmetry, we calculated the skewness (SK) of these
frequency distributions by using the relationship

^ 3 /InDg-lnD^ed") (g)

where Dmed is the median diameter. A perfect log-normal distribution has a skewness of
zero. If a distribution has a higher tail to the right than to the left, it is positively skewed.
Most of the distributions encountered here are negatively skewed, i.e., have higher
left-hand tails. Calculated skewness values are given in Table 9.

Particle Evaluation by Plutonium Content

Another characteristic studied was the distribution of plutonium among the particles as
indicated by the observed number of fission-fragment tracks in the surrounding
polycarbonate.

The track distribution among particles from both systems was evaluated in the same
way as the particle diameters. The fraction of the particles with the number of tracks

*50.00 + ^^

132 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

equal to or less than a selected number, T, are given for sampling points A, B, C, and D in
Table 10. Figure 8 is a logarithmic probability plot of cumulative percent of particles
from each of these sampling points. Figure 9 is a similar plot for particles from four
locations in System I. The calculated geometric mean for the number of fission-fragment
tracks per particle, the geometric standard deviation, and the skewness for particles from
each sampling point are given in Table 1 1. Best-fit log-normal probability curves for each
distribution are plotted in Figs. 8 and 9. For comparison of the track distributions for
particles from the various sampling points in System I with those from System II, the
probabiUty curve for the track distribution for particles from sampling point A in
System II is plotted with the distributions from System I in Fig. 9.

Table 10 Distribution of Fission-Fragment Tracks Among Plutonium-Bearing
Particles Collected from Sampling Points A, B, C, and D

Fraction with tracks < T

Number

Number
of

Sampling

Sampling Sampling

tracks

point A

point B

point C

point D

tracks

point A

point B

point C point D

0.04

0.86

0.60

0.956

0.05

0.61

0.965 0.958

0.09

0.87

0.62

0.13

0.01

0.88

0.15

0.06

0.03

0.89

0.63

0.972

0.19

0.01

0.09

0.65

0.21

0.11

0.90

0.66

0.26

0.13

0.06

0.68

0.31

0.03

0.15

0.10

0.91

0.69

0.34

0.04

0.20

0.13

0.91

0.36

0.24

0.21

0.73

0.38

0.26

0.25

0.92

0.75

0.40

0.05

0.32

0.28

0.93

0.76

0.44

0.07

0.38

0.32

0.945

0.77

0.986

0.45

0.07

0.44

0.42

0.950

0.47

0.08

0.48

0.46

0.78

0.50

0.11

0.49

0.58

0.955

0.991

0.51

0.15

0.54

0.63

0.80

0.54

0.17

0.56

0.69

0.81

0.59

0.21

0.59

0.70

0.84

0.60

0.22

0.63

0.72

0.85

0.63

0.26

0.66

0.75

0.960

0.89

0.64

0.31

0.70

0.77

0.90

0.68

0.37

0.72

0.82

1.000

0.70

0.75

0.91

0.72

0.39

0.78

0.92

0.74

0.40

0.82

0.965

0.75

0.44

0.87

0.83

0.93

0.76

0.45

0.89

0.85

0.970

0.78

0.48

0.90

0.86

0.980

0.93

1.000

0.79

0.50

0.92

0.89

0.953

0.81

0.52

0.93

0.972

0.82

0.56

0.947

0.92

0.981

0.84

0.58

0.93

100

0.990

0.991

0.85

0.59

0.944

150
200

0.995
1.000

1.000

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 133

5 10 20 30 40 50 60 70 80 90 95
PARTICLES HAVING ^T TRACKS, %

98 99

Fig. 8 Distribution of the number of tracks per particle for particles collected from
sampling points A, B, C, and D in System II.

Discussion

The most abundant elements in average crustal rock (and soil) are oxygen (46.60%),
silicon (27.72%), aluminum (8.13%), iron (5.00%), calcium (3.63%), sodium (2.83%),
potassium (2.59%), magnesium (2.09%), and titanium (0.44%) (Mason. 1966). Except for
oxygen, which was not detected by electron-microprobe analyses, these elements are also
found in most inorganic particles (Tables 6 and A.l). This supports the idea that most
plutonium-bearing particles are airborne crustal material to which minute quantities of
plutonium have become attached.

Of particular interest is the quantity of "'^''Pu contained on these particles. One
femtocurie o\' ""'''Pu irradiated under the conditions described here should produce 41
fission-fragment tracks. The mmmuim detection limit tor electron-microprobe analysis of
plutt)nium is abt)ut 0.2 pg. or about 10 fCi, of ^^^Pu (Sanders, 1976), which is equivalent
to 410 fission-fragment tracks. Because of this relatively low sensitivity of electron-
microprobe analysis, plutonium could be detected by this method in only 1 of the 558
particles selected for analysis, even though all the particles produced tlssion-fragment
tracks. This single particle was a small, l-/;m-diameter particle collected from unfiltered
wel-cabinel exhaust. It contained 73% PUO2 by weight (equivalent to 170 fCi of ^^^Pu)
in ct)mbination with Fe203 and mica.

134 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

20 30 40 50 60 70 80 90 95
PARTICLES HAVING >T TRACKS,

98 99

99.8

Fig. 9 Distribution of the number of tracks per particle for particles collected from the
mechanical line (sampling point A), wet cabinets, and room air.

TABLE 1 1 Distribution of Fission-Fragment Tracks Among Plutonium-Bearing
Particles from Various Sources in Systems I and II

Geometric mean

Geometric

Data

of number of

standard

Geometric mean

points,

fission tracks

deviation

Skewness

activity particle,

Source

(Tg)

(Sg)

(SK)

fCi

System 1

Unfiltered fifth-level

wet-cabinet air

15,987

3.76

2.56

-0.20

0.09

Unfiltered sixth-level

wet-cabinet air

7,042

3.32

2.99

-0.51

0.08

Fifth-level room air*

1.00

8.40

-0.98

0.02

Filtered wet-cabinet

air'

0.87

4.14

-0.29

System II

Sampling point A air
Samphng point B air
Sampling point C air
Sampling point D air

0.02

200

14.74

2.69

-0.43

0.36

107

32.38

1.78

0.23

0.79

114

16.50

1.75

-0.22

0.40

17.01

1.65

0.24

0.41

*Values determined graphically.

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 135

Of the major crustal elements listed in Table 6, silicon and iron were the most
ubiquitous, being found in most particles. The enrichment-factor distribution for these
elements, however, does not fall within the log-normal distribution for crustal material.
For the enrichment factors of an element to match the log-normal distribution of crustal
material in aerosols, there should be about 16% of the enrichment factors of less than one
geometric standard deviation, 68% within one geometric standard deviation of the mean,
and another 16% above one geometric standard deviation. This lack of conformity may
result from the low values for the geometric standard deviations of the enrichment factors
for these elements in aerosols.

Only the enrichment factors for sodium and chlorine fall within the log-normal
distribution for crustal material. This may be due to the relatively high solubiUty of
compounds of these elements and, in the case of chlorine, the high value for the
geometric standard deviation.

Particles from System I contain a greater variety of elements than those from
System II, and thus all but four elements are contained on a higher proportion of
particles from System I than from System II. The most striking example was nickel.
Although 56% of the particles from System I contained nickel, only 9% of those in
System II did. The major crustal elements (those in Table A.l comprising 0.4% or more
of crustal material) are contained on over half the particles from System I and, except for
magnesium in particles from sampling points C and D and titanium, are also contained on
over half the particles from System II. Some of the minor elements (those comprising
0.1% or less of crustal material) are present in over half the particles, viz, nickel,
chromium, and zinc in particles from System I and sulfur, chromium, and zinc in particles
from samphng point B of System II. The chromium and nickel may have come from the
304L stainless-steel alloy of cabinets and exhaust ducts or the Hastelloy*-C alloy in the
wet cabinets. However, few of the particles contained the proper ratio of chromium to
nickel found in either alloy. Also, \{ Hastelloy-C contributed the nickel in the particles,
some molybdenum should also have been detected.

Of the elements that are present on less than 10% of the particles, all but copper on
particles from System II have high enrichment factors. This indicates that the minor
constituents of crustal material are not uniformly distributed among particles but are
concentrated on a few particles where they represent a major constituent.

The plutonium-bearing particles were larger than natural aerosol particles collected at
relatively low altitude (<2.3 km), as shown in Fig. 7. Particles collected from sampling
points A and B of System II were larger than those from System I, with geometric mean
diameters two or three times as great as those of particles from other locations.

The size of about 95% of the plutonium-bearing particles ranges between 0.4 and 37
jum in diameter. Morrow (1964) estimated that with normal respiration all particles in a
monodispersed aerosol of unit-density spheres 37 /im in diameter will be deposited in the
nasopharyngeal region of the respiratory tract. [With larger (>37 jum) particles the
fraction deposited rapidly decreases.] As the diameter decreases, the fraction deposited in
the respiratory tract decreases until a minimum of 20% deposition is reached for particles
that are around 0.1 to 0.2 iin\ in diameter, where the particles tend to remain airborne.
As the diameters decrease below 37 jum, a larger fraction is deposited in the
tracheobronchial region until 70% of the particles 5 jum in diameter are deposited in the
tracheobronchial region and only 5% in the nasopharyngeal and 5% in the alveolar

*Trademark of Cabot Corporation, Boston, Mass.

136 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

regions. With still smaller particles, the fraction deposited in the tracheobronchial region
decreases until at 0.2 /um diameter only 10% is deposited in the tracheobronchial and 10%
in the alveolar regions. For dust particles with a density of around 2.5, this distribution
will be shifted toward smaller diameters so that 100% deposition in the nasopharyngeal
region occurs around 5 nm.

Particles from all parts of System II also contained, on the average, more plutonium
per particle than those from System I. As shown in Table 1 1 , the geometric mean number
of tracks per particle from unfiltered wet-cabinet air was just over three for both fifth-
and sixth-level cabinets (averaging about 0.08 fCi per particle), whereas that for filtered
wet-cabinet air was about one-third of this, or almost the same for room air (averaging
about 0.02 fCi/particle).

A comparison of the mean diameters of particles collected from different sampling
points, given in Table 9, with the mean number of fission-fragment tracks for particles
from the same location, given in Table 11, indicates a possible relationship between
particle size and plutonium content. Correlation coefficients between the cube of the
particle diameter and the number of fission-fragment tracks from each particle from
sampHng points B, C, and D were calculated. These are given in Table 12. These
coefficients differ significantly from that expected from a random sample from a
population of paired variables having a correlation coefficient of zero. Thus, even though

TABLE 12 Correlation and Coefficient of Alienation for the Cube of the

Diameter and the Number of Fission-Fragment Tracks for

Particles from Sampling Points B, C, and D of System II

Sampling Number of Correlation
point particles coefficient

107

0.69

114

0.29

0.36

the points on a plot of particle diameter cubed vs. number of fission-fragment tracks
appear scattered, there is a significant correlation between the quantity of plutonium in
particles collected from sampling points B. C, and D in System 11 and the particle volume.
(Tracks with particles collected at other sampling points, where only ^"'^Pu could be
found, were counted but not recorded for each particle. Only where a ratio of
alpha-particle to fission-fragment tracks was needed to distinguish plutonium-bearing
particles from those having other fissionable materials were the track counts recorded.)

Summary and Conclusions

The elemental compositions, sizes, structures, and ^"'^Pu contents were determined tor
558 plutonium-bearing particles collected from various locations in the exhaust t rom a
reactor fuel reprocessing facility. Airborne particles were collected on polycarbonate
membrane filters. Particles containing ^^^Pu were identified by fission-fragment and
alpha-particle tracks produced by them in a polycarbonate film with a nuclear-track-
emulsion coating. When located, the amount of "^^^Pu in each particle was determined by

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 137

counting the tracks, a small portion of the film containing the particle was isolated, the
emulsion removed, the polycarbonate dissolved, the track replicas oxidized, and the
elemental composition of the ^^^Pu-bearing particle determined by electron-microprobe
analysis. These data were compared with data from natural aerosol particles.

Most of the collected particles were composed of aggregates of crustal materials. Of
the particles, 3.6% was organic and 1.7% was metallic, viz., iron, chromium, and nickel.
High enrichment factors for titanium, manganese, chromium, nickel, zinc, and copper
were evidence of the anthropogenic nature of some of the particles. The amount of
plutonium in most particles was very small (less than 1 fCi of ^^^Pu). Thus plutonium
concentrations had to be determined by the fission-track counting method. Only one
particle contained sufficient plutonium for detection by electron-microprobe analysis.
This was a l-jum-diameter particle containing 73% PUO2 by weight (estimated to be 170
fCi of ^^^Pu) in combination with FejOa and mica. The plutonium-bearing particles
were generally larger than natural aerosols. The geometric mean diameter of those
collected from the mechanical line exhaust was larger than that of particles collected
from the wet-cabinet exhaust (12.3 /jm vs. 4.6 /im). Particles from the mechanical hne
also contained more plutonium per particle than those from the wet cabinets. The
amount of plutonium per particle decreased with the distance of each sampling point
from the mechanical Une.

The size and ^^^Pu content distribution among particles collected from the sand
filter effluent and at the 50-ft level of the 291-F stack were almost the same. The
geometric mean and standard deviation of the diameter of ^^^Pu-bearing parricles at the
50-ft level was 5.43 ± 2.69 jum. The relatively large size of these particles is believed to be
due to coagulation of submicrometer particles by thermal and turbulent mechanisms to
fomi larger agglomerates. The elemental composition of these particles, which contain
very small amounts of plutonium in combination with crustal elements not used in the
recovery process, supports this assumption. Scanning electron micrographs, such as
Fig. 5(c), also show these particles to be agglomerates of smaller dissimilar particles.

Fleischer and Raabe (1977) have observed alpha-decay-induced fragmentation of
^^^Pu02 particles probably caused by the heavy recoiling nuclei. When suspended in
water, these particles produce fragments, or subparticles, which contain from 50 to
10,000 ^^^Pu atoms, the abundance of which follows a power-law relation v^th the
largest particles being the least abundant. The possibiHty exists that PUO2 particles, large
enough to be trapped on HEPA filters, fragment owing to alpha decay. The small
fragments then pass through the filters where they coagulate with dust composed of
crustal elements. The larger dust particles may not have passed through the HEPA filters
but entered the exhaust system through leaks in the ducts, as illustrated in Fig. 10. Such
leaks might remain undetected as long as the exhaust system remained under negative
pressure with respect to the atmosphere.

The geometric mean and standard deviation of the number of fission-fragment tracks
per ^^^Pu-bearing particle collected from the 50-ft level during July, August, and
September 1977 was 17.01 ± 1.65 tracks. One femtocurie of ^^^Pu in a mixture of
low-irradiation plutonium will produce 52.6 fission fragments when irradiated with a
fluence of 8.64 x 10^ '^ thermal neutrons/cm^. Only about half, or 26.3, of the fragments
will produce tracks in the polycarbonate film. Thus the calculated geometric mean
radioactivity on the ^^^Pu-bearing parricles leaving the stack is 0.65 fCi/parricle. During
these 3 months a total of 82 juCi of ^^^Pu was discharged to the environment. This

138 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

amounts to an average of 0.62 nCi/min. Thus during this period about 10^ ^^^Pu-bearing
particles per minute were discharged from the 291-F stack to the environment. With a
flow rate in the stack of 2 x 10^ cfm, the average ^^^Pu-bearing particle concentration in
stack air was 5 particles/ft^ .

EXHAUST

SUBMICROMETER
PARTICLES

ATMOSPHERIC
DUST

SOURCE

LEAK

SAMPLE

Fig. 10 Formation of plutonium-bearing particles in exhaust systems by the coagulation
of submicrometer plutonium particles with atmospheric dust.

Acknowledgment

I gratefully acknowledge the assistance of E. F. Holdsworth and J. T. Armstrong of the
Chemistry Department of the Arizona State University, Tempe, Ariz., who performed the
electron-microprobe analyses.

Appendix: Use of Elemental Enrichment Factors to Express Particle Compositions

Background

Two recent developments in aerosol studies have provided valuable tools for the analysis
of particle composition data. The first is the use of ratios of elemental concentrations
called "enrichment factors" to compare aerosol compositions. Begun in the early
seventies, this technique has gained wdde acceptance in the last few years (Rahn, 1971;
1976; Zoller, Gladney, and Duce, 1974; Duce, Hoffman, and ZoUer, 1975; Neustadter,
Fordyce, and King, 1976). The second development is the availability of data on the
composition of natural aerosols. In the last few years, Rahn (1976) published a
compilation of 104 data sets of trace elements in aerosols along with the geometric mean
and geometric standard deviation of the enrichment factors for each of the elements.
These data sets were from sampling sites ranging from highly industriahzed temperature
zones to the tropics and poles and represent all continents except South America as well

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 139

as various marine locations. As a framework from which to view much of the order in
atmospheric aerosols, Rahn used the concept of aerosol-crust enrichment factors for the
elements. This concept has been applied to analyzing data collected in this study to
provide for (l)the intercomparison of the compositions of plutonium-bearing particles
with atmospheric aerosols compiled by Rahn and (2) the intracomparison among particles
collected from different sampling points.

Microprobe Analyses of Particles

For comparison, results of microprobe analyses must be expressed as elemental ratios
because not all elements that may be present in an aerosol are detected by microprobe
analysis. The microprobe used in this study is quantitative only for elements with atomic
numbers greater than 10. It is only semiquantitative for oxygen (the most-abundant
element in crustal material) as well as for other major elements of low atomic number,
such as hydrogen, fluorine, and carbon. Atmospheric aerosols are known to contain, in
addition to elements and oxides, carbonaceous material, such as sooty carbon and
organics, and water-soluble ionic material, such as sulfate, nitrate, and ammonium ions.
Thus elemental weight percents, normalized to 100 on the basis of the elements detected,
cannot be compared. Even the addition of a hypothetical oxygen concentration,
calculated on the supposition that all elements are present as oxides of known valence,
will still not account for the organic fraction of particles. However, a ratio of the
concentrations of one element to another will normally be relatively unaffected by the
concentrations of other elements that may be present and thus can be used for
comparisons even when a complete analysis of all the elements in an aerosol or single
particle is not available.

Enrichment Factors

A dimensionless ratio of elemental concentrations, called the enrichment factor, has been
defined as

EF(X) = ^-^^^^^iH£l£l (A.l)

where EF(X) is the enrichment factor of element X in an aerosol relative to some source
material and X/Ref is the ratio of the concentration of element X to the concentration of
the reference element, Ref, in both the aerosol and the source material.

Source Material

Elemental ratios in aerosols or in single particles are nonnalized by dividing them by
ratios of the same elements in a standard source material to obtain the enrichment
factors. If a particle is composed of the same material as the source, the enrichment
factor will be 1.00 for all elements. If the ratio of an element to the reference element is
greater or less than the same ratio in the source material, the enrichment factors will be
greater or less than 1.00, and the particle is said to be either enriched or depleted,
respectively, in that element.

The most commonly used crustal source material for continental enrichment-factor
calculations is globally averaged crustal rock. (For marine enrichment-factor calculations,
sea salt is used.) The selection of rock may seem strange because there is little doubt that

140 TRANS UR AM C ELEMENTS IN THE ENVIRONMENT

soil rather than rock is the precursor to the crustal aerosol. Some 93% of the earth's
continental surface is covered by soils (Kothny, 1973). Many of these soils are in states of
loose aggregation which can easily be made airborne by the wind. Chemically, however,
Rahn (1976) has found that the composition of the crustal aerosol is not unambiguously
that of soil. Elements in natural aerosols with rock-like enrichment factors include silicon,
iron, calcium, potassium, and chromium; those with soil-like enrichment factors are
titanium and barium. One would expect natural aerosols to be, hke soil, depleted in the
more-soluble elements. Except for glacial activity and to a lesser extent in deserts,
physical weathering processes, which ultimately produce small particles from boulders,
are very slow and are accompanied at all stages by intense chemical weathering. Thus
large masses of physically pulverized rock which have not been chemically weathered are
not available for aerosol production.

Rahn (1976) speculates that remote continental aerosols are never as depleted in the
soluble elements (e.g., sodium, potassium, calcium, and magnesium) as they should be
relative to rock (if natural aerosols were purely soil derived) because of the presence of
small amounts of marine aerosol. Soluble elements, especially sodium and magnesium, are
abundant in the marine aerosol; thus only small amounts of this aerosol in remote
continental areas would noticeably raise the proportions of soluble elements in an aerosol
collected there.

In addition to the similarity in the elemental composition of aerosol and crustal rock,
available analytical data are much less numerous and less reliable for soils, especially for
several interesting trace elements that are enriched in aerosols.

For these reasons the majority of investigators who calculate aerosol-crust enrichment
factors have chosen one of the several available tables of elemental abundances in average
crustal rock. Because the composition of plutonium-bearing particles are compared with
data reported by Rahn (1976), the same crustal-rock composition used by him [that
reported by Mason (1966)] was selected as the source material composition for this
work. Column 2 of Table A.l gives the elemental concentrations in globally averaged
crustal rock for those elements found in plutonium-bearing particles.

Reference Elemen t

Of the various elements that seem to be reliably crust derived in aerosols, akuiiinum,
silicon, and iron are generally considered to be the most suitable reference elements.
(When sea salt is the source material, the nearly universal choice is sodium.) An
acceptable crustal reference element should have higli concentrations in rock and soil,
very low pollution potential, ease of detemiination by a number of analytical techniques,
and tYeedom from contamination during sampling. Iron has markedly higher pollution
potential than aluminum and so is less suited tor use with urban or rural aerosols. Silicon
is probably the most unambiguous elemental indicator of crustal material. Unfortunately,
silicon has been detennined in so few aerosol samples that it cann(.)t be used as the
reference element where comparisons are to be made. Aluminum is a major element
(81,300 ppm in rock), well determined by a variety of analytical techniques, and has a
minimum of specific pollution sources.

Thus f\M- this work enrichment factors for element X in most particles were calculated
using

(X/AI)rock

PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 141

TABLE A.l Elemental Concentrations in Average Crustal Rock and
Geometric Mean Enrichment Factors of Various Aerosols

Geometric mean

enrichment factors

Remote

Concentration,
ppm

Global

marine
EFg

continental

Urban

Element

EFg/Sg

EFg

EFgSg

Silicon

277,200

0.62

0.79

1.01

0.7

0.79

Aluminum

81,300

1.00

1.0

1.00

Iron

50,000

1.05

2.06

4.06

2.5

1.5

2.2

Calcium

36,000

1.15

2.84

7.04

1.5

2.9

Sodium

28,300

0.64

4.44

30.8

10^-10^

0.4

1.81

Potassium

25,900

0.99

1.98

3.98

1.5

1.63

Magnesium

20,900

0.64

2.38

8.90

10' -10'

0.7

2.0

Titanium

4,400

1.01

1.39

1.92

1.2

1.63

Phosphorus

1,050

0.79

2.63

8.71

2.6

Manganese

950

1.45

3.91

10.5

3.2

Barium

425

2.61

5.50

11.6

4.8

Sulfur

260

228

608

1620

490

Chlorine

130

100

740

5470

lo-'-io^

300

Chromium

100

2.50

8.11

26.3

6.2

Nickel

8.74

31.9

116

100

10.8

Zinc

79.7

257

832

400

300

Copper

34.0

102

304

150

149

Tungsten

1.5

4.89

19.1

74.3

11.0

Cadmium

0.2

274

1920

13400

5000

2000

940

with aluminum as the reference element and average crustal rock as the source material.
However, 18 particles from System I and 37 from System II contained no aluminum.
Thus the enrichment factors had to be based on silicon rather than on aluminum, where

FF^Y'l = ^^/^Oparticle (Si/Al)g aerosol _ ^ -,g (^/Si)particle

(X/Si)rock (Si/Al)rock

(X/Si)rock

(A.3)

(The second set of ratios is the geometric mean of the global aerosol-crust enrichment
factor explained in the next section.)

With the use of these two relationships, the enrichment factors were calculated from
the elemental weight percents obtained for 115 particles in System I and 156 particles in
System II. Six small (0.5 to 3.6 jum in diameter) iron particles in System I and two
particles [~15 nm in diameter and containing potassium, chromium, and iron (1 : 3 : 3)]
from sample point A of System II contained neither aluminum nor silicon and were thus
not included in the study.

Comparative Aerosol Data

For a comparison of the elemental composition of plutonium-bearing particles with that
of atmospheric aerosols, enrichment factors calculated for elements in these particles
were giouped according to data supplied by Rahn (1976) for aerosols. In his report

142 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

trace-element concentrations in aerosols from 104 published and unpublished data sets
were used to calculate enrichment factors. From the enrichment factors in each data set,
the geometric mean enrichment factor (EFg) and geometric standard deviation (Sg) of the
logarithmic frequency distributions of enrichment factors were calculated for each
element by the following formulas:

EFg = exp(^^lnEFi) (A.4)

and

Sg = exp

where N is the number of data points and EFj is the enrichment factor of the \th point.

The geometric mean enrichment factors obtained by Rahn (1976) for 19 elements are
given in Table A.l for global, remote marine, remote continental, and urban aerosols. The
geometric means of the global aerosol enrichment factors include data from all points and
may be weighted too heavily toward cities, but they can serve as a useful first
approximation to a general aerosol. The urban enrichment factors are geometric means
for 29 cities. The enrichment factors for remote continental and remote marine areas
were read from the enrichment-factor plots and are therefore somewhat subjective.

Values for EFg/sg and EFg X Sg, respectively, were calculated with the use of global
values to obtain the lower and upper limits for 68.27% of the enrichment factors closest
to the geometric mean. (When describing concentrations at selected statistical levels
remote from a mean, the Sg is a multiplier or divider of the EFg, whereas its counterpart,
Gaussian standard deviation, functions as an increment to the arithmetic mean. This is a
consequence of the fact that multiplying and dividing values are equivalent to adding and
subtracting their logarithms.) The results from these calculations are also given in
Table A.l.

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PLUTONIUM-BEARING PARTICLES FROM FUEL REPROCESSING 143

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144 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

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Physicochemical Associations of Plutonium
and Other Actinides in Soils

E. A. BONDIETTI and T. TAMURA

Soil physicochemical behavior of plutomum and other actinides is discussed with primary
emphasis on the behavior ofplutonium in acnial contaminated soil and the importance of
actinide speciation in interpreting laboratory results. The behavior of actinides in soil is
strongly influenced by physical form and/or oxidation state. The chemistry ofplutonium,
americium, curium, and neptunium is reviewed, particularly with respect to the oxidation
states likely to control their behavior in most soils. Several aspects of sorption to soils are
discussed, particularly those for plutonium. Tlie comparative behavior of plutoniwn,
thorium, and uranium in soil is also illustrated to provide a perspective for evaluating
long-term environmental behavior. The relative hazard associated with plutonium-
contaminated soil is evaluated and the importance of both physicochemical form of the
plutonium and the soil particle-size association is emphasized.

The exposure of internal organs to ionizing radiation is tiie major potential hazard
associated with the production and release of actinide alpha emitters to the environment.
Two predominant pathways of exposure from this environmental contamination are
inhalation and ingestion.

Inhalation of discrete radioactive particles (i.e.. PUO2) and carriers (i.e., contaminated
soil particles) can result in exposure of the lung and of other body organs following
transport of particles or ions through the lung to other organs. Ingestion of radionuclides
present in biologically assimilated fomis or as surface contamination also serves as a
source of possible exposure to critical organs.

The relative importance of inhalation and ingestion as pathways for human exposure
depends on many environmental parameters exclusive of the physicochemical associations
of the radioelement in soils. However, the physicochemical properties of the radioelement
in soils strongly influence the pathway and magnitude of transport. This is illustrated by
the ~10^ greater assimilation by plants of plutonium added in monomeric fomi as
compared with PuOt microspheres (Adams et al., 1*^75).

This chapter examines two aspects of the physicochemical associations ofplutonium
and other actinides in contaminated soils: (l)the case in which plutonium is
monomerically distributed in soil (i.e., the potentially most biologically available form)
and (2) the case in which plutonium exists or its origin is traced to a discrete particulate
source. However, it is also important to evaluate general principles of actinide— soil
interactions. For this purpose we also explore several aspects of sorption behavior to soil
colloids.

145

146 TRANSURANIC ELEMENTS IN THE ENVIRONMENT
General

Chemical Characteristics of the Transuranium Elements (Np, Pu, Am, and Cm)
and Related Ac tinides (U and Th)

The alpha-emitting elements of significance in the uranium and thorium fuel cycles
include Th, U, Np, Pu, Am, and Cm. These radioelements can be released to the biosphere
by fuel reprocessing, radioactive waste handUng and disposal, and fuel fabrication. Many
of these elements have already been dispersed through defense-related activities.

Although the geochemistries of U and Th are reasonably well understood, the largely
man-made elements (Pu, Np, Am, and Cm) are only now under study. Fortunately the
chemical characteristics of the transuranic elements are very similar to the naturally
occurring rare earths (oxidation state III), Th (oxidation state IV), and U (oxidation
states IV and VI). Indeed, early studies on the solution and solid-phase chemistry of the
transuranic elements used these analogies (Hindman, 1954; Connick, 1954; Cunningham
and Hindman, 1954; Thompson et al., 1949).

The complexities of the chemistry of any element usually depend on the number of
oxidation states that the element can exhibit. For the actinide elements under discussion,
these oxidation states are represented by the M^"*", M'*''", MOt(V), and M02^(VI) species.
Uranium is a classic example of tlie influence of oxidation state on environmental
chemistry. Both U and Th in the tetravalent state are extremely resistant to leaching.
However, the oxidation of U(IV) to U(VI) results in much higher U mobilities in the
environment (Adams, Osmond, and Rogers, 1959). Likewise, altliough uranyl ion is
relatively stable in seawater as the uranyl carbonate complex, ^^^Th formed from the
radiodecay of ^^"^U rapidly becomes depleted with respect to ^^'*U (and ^^*U) (Cherry
and Shannon, 1974).

Environmentally Important Oxidation States of the Transuranium Elements

Curium is trivalent in solution and probably will be present in air-ignited oxides as
Cm2 03. Americium is also likely to be trivalent in environmental solutions, although the
IV, V, and VI oxidation states are known in the laboratory. The dioxide, Am02, can be
formed on ignition. Plutonium can exhibit valences of III, IV, V, and VI in solution and
the IV state as the dioxide. Neptunium is probably tetravalent or pentavalent in
environmental solutions and tetravalent in tlie oxide. Since limited experimental
information is available, these valence assignments for environmental systems are subject
to revision.

Environmental conditions, such as pH and Eh, will control the oxidation-state
distribution, although the kinetics of the redox reactions are unknown. Experimental
determinations of Pu and Np oxidation states in an environmental context have been
undertaken (Bondietti and Reynolds, 1976; Bondietti and Sweeton, 1977: Bondietti,
Reynolds, and Shanks, 1976; Bondietti, 1976). A number of speciation diagrams have
been constructed (Polzer, 1971; Andelman and Rozzell, 1970; Rai and Serne, 1977)
which attempt to evaluate the oxidation-state distribution of Pu in environmental
solutions. These investigators have recognized that understanding the environmental
speciation of Pu is critical to evaluating its biogeochemistry. The oxidation states of Pu in
natural water have actually been determined (Bondietti and Sweeton, 1977).

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 147

TABLE 1 Effect of Clay Treatment on Adsorption of
Actinide Elements to Miami Silt Loam Clay

Cation e.xchange
capacity, meq/ 100 g

Percent adsorbedf

Treatment*

^^''Th(IV)

''"'Cm(in)

^^'U(IV)

^^^Np(V)

Intact clay

Organic matter removed
Fe and organic
matter removed

17
11

9.9

99.7
99.8

99.7

98.6(±0.2)t
99.6(±0.2)t

95.6(±0.5)t

95.6(±0.2)t
96.4(±0.4)t

99.1(±0.5)t

61.8
49.7

18.2

*Organic matter removed with NaOCl; Fe removed with sodium dithionite.

tpH 6.5; 5mM CaCNOj)^; solution/clay ratio of 400/1; 48 hr equilibration; U and Np at
<microgram per gram levels; Th and Cm at <nanogram per gram levels.
jMean ± standard deviation.

Interactions with Environmental Colloids

Effect of Oxidation State on Sorption The partitioning of the actinides between solid
and solution phases may be dependent on the charge characteristics of the element, the
physicochemical characteristics of the solid, and the composition of the solution.
Complexation by OH (hydrolysis) and other ligands affects sorption because all four
common oxidation states (III, IV, V, and VI) form complexes of varying stabilities. For
example, the competition between hydrolysis and complexation by carbonate dominates
the sorption behavior of uranyl ion in natural solutions. Above pH 7.5 (and in
equilibrium with atmospheric CO2) soluble uranyl carbonate complexes can predominate;
below this pH sorption to particulates readily occurs (Starik and Kolyadin, 1957).
Another example is Pu(IV), which is extensively hydrolyzed in near-neutral solutions and
is probably not adsorbed by normal ion exchange mechanisms (Tamura, 1972).
Oxidation— reduction reactions, since they affect oxidation state, also influence sorption.
Thus NpOt shows poor adsorption to soil, but reduction to Np(IV) increases sorption
(Bondietti, 1976).

The different actinide oxidation states with respect to sorption are compared in
Table 1. The distribution of Th(IV), U(VI), Cm(III), and Np(V) between a soil-clay
fraction and a 5mAf Ca(N03)2 solution showed that relative sorption followed the
oxidation-state order IV > III > VI > V under the specified conditions. The organic
matter and free iron oxide (Fe) coatings were removed to evaluate the effects of colloid
surface constituents [and thus cation exchange capacity (CEC)] . Only Np(V) sorption
was strongly influenced by these treatments. The removal of organic matter decreased the
CEC by 35% and Np sorption by 20%.

Removal of organic matter and Fe did not further affect the CEC, but the Np(V)
sorption value decreased to 29% of the intact clay value. This observation of an apparent
surface-dependent sorption mechanism suggests that, even for the MO2 oxidation-state
species, which is largely unhydrolyzed at environmental pH's, mass-action relationships
may not describe adsorption equilibria.

Mass-action expressions have been used to describe ion-exchange equilibria. The
exchange of Np02 on a sodium-saturated clay can be expressed as

NpOt + NaC ^ NPO2C + Na""

(1)

148 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

and the equilibrium expression as

j^ ^ (NpO^CHNa)
(NpOt)(NaC)

where K is the equilibrium constant, C is clay, and the parentheses denote activity.

In a system in which neptunium is present in trace quantities and the clay is
essentially sodium saturated, the activity coefficients of tlie sodium clay, neptunium clay,
and neptunium and sodium ions can be considered to be a constant. Equation 2 can be
rewritten in the form

_ [NpO.C] [Na]
[NpOt] [NaC] ^ ^

where K' incorporates the constant activity coefficients and the brackets denote
concentrations. By definition, the distribution coefficient (Kj) for neptunium is

[Np02 J

and Eqs. 3 and 4 can be combined to form

> [NaC]

Since the clay remains essentially sodium saturated, NaC is a constant, and a log Kj vs.
log [Na] plot should yield a straight line. The slope of the line is a function of the
exponent of Na ; in this case the slope is --1. If the clay is calcium saturated, it can be
written as Cao.sC in Eq. 1. The final expression of Eq. 5 would contain the exponent of
0.5, and thus the slope would be -0.5.

Data reported by Routson, Jansen, and Robinson (1975) on neptunium sorption by
two soils at different sodium and calcium ion concentrations are shown in Fig. 1. The
notable feature in the plot is the absence of any effect of sodium on neptunium sorption.
Thus the slope of the Np02 Kj vs. Na concentration plot approaches 0 rather than —1.
Calcium exerted a more pronounced effect on Np sorption, but even here the Np Kj vs.
Ca^ concentration plot had a slope of about —0.3 rather than -0.5.

The data of Routson, Jansen, and Robinson (1975) and the effect of clay surface
treatment on Np02 sorption (Fig. 1) indicate that electrostatic interactions alone do not
explain NpOa sorption. One surface component, organic matter, appears to have an
influence. In general, the stability of Np(V) chelates is comparable with divalent cation
chelates (Zn^"^, Ca^"*", etc.). The interactions of Np(V) with soil humic acids reflect this.
Figure 2 represents the observed distribution of Zn^"^, Cd^"^, Ca^"^, Sr^"*", and NpOt
between complexed and free forms in the presence of soil humic acids. As is illustrated in
the figure, Zn and Cd form stronger complexes than Ca or Sr, which is expected. The
Np02 cation forms complexes that are slightly stronger than Ca. It should be apparent
from the above samples that even the least hydrolytic actinide oxidation state interacts
with soil constituents in complicated ways.

ASSOCIATIONS OF Pii AND OTHER ACTINIDES IN SOILS 149

100
8 0

6.0

0 001

D
CD

^ 0.4

Ca2* CONCENTRATION {M)
001 01

1.0

"1 1 — I I I I

I I r

KjjNp Burhank sand, f(Na)

KjNp South Carolina subsoil, f(Na

Sodium
Calcium

J I I I I

001

0 1 10

Na* CONCENTRATION (M)

10.0

Fig. I Distribution coefficients of neptunium in selected soils. (Modified after Routson,
Jansen, and Robinson. 1975.)

Plutonium Sorption. Plutonium in oxidation state IV is very insoluble in water in the
absence of soluble complexers. Given a solubility product (Ksp) tbr Pu(0H)4 of ~10^^^
(Coleman, 1965), soluble monomeric Pu(IV) species should be difficult to assay in
near-neutral solutions. Considering the various hydrolytic species [Pu(0H)3, Pu(0H)2 ,
etc.] , concentrations of soluble Pu in equilibrium with crystalline PUO2 might approach
those depicted in Fig.-3. This figure was plotted using the hydrolysis constants evaluated
by Baes and Mesmer (1976). By analogy to U(IV), a negatively charged pentahydroxy
species, Pu(OH)^, was postulated to exist in their analysis of hydrolytic constants.

Studies on the effect of pH on Pu(IV) sorption by soils have shown that, in the pH
range 2 to 8, 99+% of the added Pu is lost from solution (Rhodes, 1957; Rogers, 1975).
Rogers (1975) showed that maximum sorption was at about pH 5.5; sorption was less at
lower and at higher pH's. Above pH 8, Rliodes (1957), Rogers (1975), Prout (1958), and
Nishita (1978) obsei^ved substantial increases in the concentration of Pu in the
supernatant (Fig. 4). Rogers (1975) attributed this behavior to dispersed soil colloids that
failed to sediment during centrifugation. Rliodes (1957) and Rogers (1975) observed that
this decrease in sorption might also have been due to the dispersal of Pu polymer or
hydroxy species.

Prout (1958) observed a decrease in adsorption for three Pu oxidation states (III, IV,
and VI) above pH 7 to 8 which might argue against polymer dispersion. In addition, he
found that radiostrontium and radiocesium in low-ionic-strength solutions also showed a
decrease in adsorption above pH 7 to 8. In higher ionic-strength solutions, the adsorption
increased with increasing pH. Figure 5 illustrates this ionic-strength effect using selected
adsorption curves reported by Prout (1958). The concentration of strontium used was

150 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10-5

10-8

5 10-7 2

FREE METAL (M)

10"6

Fig. 2 Comparison of the relative complexing of Zn, Cd, Np(V), Ca, and Sr by soil
humic acids (pH 7.0).

5 X 10~^yif; if no sodium nitrate were added, strontium adsorption increased to a
maximum around pH 7 and then decreased. When 1% NaNOa was added, no decrease was
observed with increasing pH. Cesium showed the same effect: tracer Cs (5 x \Q~^M)
displayed maximum adsorption around pH 7.7, but increasing the Cs concentration to
5 X lO'^^M removed this effect. These resuUs suggest that the decrease above pH 7 to 8
observed with ~10~^MPu was due to dispersion of clay-size particles containing sorbed
ions. As the ionic strength was increased in the Sr and Cs studies, the clay remained
flocculated and the observed K^ increased.

Plutonium represents an element in which the simultaneous presence of more than
one oxidation-state species in solution can influence the observed adsorption behavior.
Determination of the extent of this problem has unfortunately been ignored in most Pu
adsorption experiments reported in the literature. Bondietti and Reynolds (1976),

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 131

9 10

Fig. 3 Estimated concentrations of hydrolyzed Pu(lV) species in solutions saturated
with crystalline PuO^. (Adapted from C. F. Baes and R. E. Mesmer, The Hydrolysis of
Cations, John Wiley & Sons, Inc., New York, 1976.)

however, reported the presence (and problem) of multiple oxidation states in Pu— clay
equilibrations. These data are summarized in Table 2.

When ^^^Pu(IV) (as the nitrate) was added to treated clays, the organic-matter
removal treatment showed the least Pu sorption at 3 weeks (61.5%). With time, sorption
values approached the other two treatments (i.e., 99+%). The reduced adsorption was not
directly caused by the removal of organic matter but was influenced by disproportiona-
tion of Pu(IV), yielding Pu(III) and Pu(V) and/or Pu(VI) during the tracer addition. The
low initial adsorption in the organic-matter removal treatment was apparently due to the
fact that the clays were "oxidized" from the NaOCl treatment used to remove organic
matter and the PuOt and/or PuOj , resulting from disproportionation, became
stabilized. For the Fe removal treatment, the clays were in a reduced state, which
minimized the presence of Pu(V + VI). The memory of these treatments is observable
2 yr after treatment (Table 2). For the intact clay, 79% of the soluble Pu was present as
Pu(III + IV) (Bondietti and Reynolds, 1976). The largest amount of oxidized Pu
[Pu(V) + Pu(VI)] was found in the organic removal treatment (35%o). Very little oxidized
Pu was found in the iron treatment (7%). The presence of more than one oxidation state
in the control solutions provides support for an initial disproportionation reaction since
oxidized or reduced clay surfaces were not present.

152 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Fig. 4 Literature observations on the effect of pH on Pu sorption to soils. A,
soil/solution ratio. 1/12.5 (Nishita, 1976». •.soil/solution ratio. 1/10 (Proiil. 1958). □,
soil/solution ratio. l/10(Prout. 1958). .soil/solution ratio. 1/20 (Rhodes, 1957).

It is significant to note that, in the results described above, the initial ^^^Pu
concentration was about 3 ng/ml (1.3 x lO^^M). Experiments described by Jacobson
and Overstreet (1948) and Prout (1958) indicated that Pudll) was adsorbed more readily
than Pu(IV), which was itself adsorbed more readily than Pu(VI). However, the molar
concentrations of Pu used by these investigators [Jacobson and Overstreet (1948),
7 X 10" "^M; Prout (1958), 10~^M] were greater than those used here. Consequently the
reason the Pu(IV) sorption was interrnediate between Pu(III) and Pu(VI) was probably
due to an initial disproportionation of the added Pu(IV). The true order is probably
IV > III > VI, as shown in Tables 1 and 2 (intact clay). Additional support for this
concept is provided in data by Bondietti, Reynolds, and Shanks ( 1976), where the Kj for

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 153

Fig. 5 Effect of ionic strength on sorption of radiostrontium and radiocesium by soil.
[Modified from E. E. Prout, Adsorption of Radioactive Wastes by Savannah River Plant
Soil Soil Science, 86: p. 15 (1958).]

TABLE 2 Adsorption and Solution-Phase Characterization of
2 3 8pu(iV) Added to Miami SUt Loam Clay (pH 4.0)

Percent adsorbedf

'oUowing

indicated

Soluble phase

equilibration
3 18

time (weeks)
52 104

characterization

$(104 weeks)

Treatment*

Percent Pu(III + IV)

Percent Pu(V + VI)

Intact clay

99.9

99.8

99.9

Organic matter removed

50.0

61.5

99.8

99.4

Organic matter and

Fe removed

99.8

99.9

99.8

99.9

Control §

71.0

71.7

78.6

82.1

*Organic matter removed with NaOCl; Fe removed with sodium dithionite.
tpH 4.0;5mA/Ca(NO3)2 ; solution/clay ratio of400/l.
XSee Bondietti and Reynolds (1976) for methodology.
§No clay present.

i54 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Th was 20 times as high as that for Pu(IV) when both elements were equilibrated with
montmorillonite clay. No valences were determined, but the concentration of ^^^Pu used
undoubtedly resulted in some initial disproportionation.

This discussion is meant to illustrate that the adsorption of Pu to soils is more
complex than the simple distribution of a radioelement between a solid phase and a
solution. For Pu the problem of mixed-oxidation-state species is significant. Attempts to
delineate the soil/sediment chemistry of Pu must consider that more than one oxidation
state may be present in stock solutions or may be formed during the experiment.
Attempts to correlate Pu adsorption with soil type may be confounded by the complex
interplay between soil components and the stability of various Pu oxidation-state species.
Thus Glover, Miner, and Polzer (1976) and Polzer and Miner (1976) observed that the
adsorption of Am(III) by various soils was as great as or greater than that of Pu(IV). They
also noted that the variability in Am(III) sorption was much less than that in Pu. At the
Pu concentrations used (10~^ to \Q~^M), disproportionation may have been the cause
of this variability. A small proportion of Pu(V) or Pu(VI) with their correspondingly
lower sorption tendencies would provide erroneous sorption values for "Pu(IV)."

Plutonium at Contaminated Sites

General

Contaminated field sites provide the best situations for studies to understand the behavior
of plutonium in environmental systems. Full appreciation of the behavior of Pu in these
sites requires knowledge of the initial character of the contaminating event(s). From this
information conclusions can be drawn regarding Pu behavior up to the time of samphng
and potential behavior extrapolated for different source terms.

Nevada Test Site. One of the largest contaminated areas in the United States is the
Nevada Test Site (NTS), which serves as the test area for nuclear detonations. Within the
NTS several sites were used for safety shot evaluation; these sites, which have been
declassified, contain dispersed plutonium from a series of liigh-explosive detonations
simulating an accidental detonation of a subcritical atomic device. The detonation would
be expected to produce a wide range of particles. Since plutonium metal is relatively
reactive, the oxide form would be expected to be produced (Cunningham, 1954).
Although a size distribution as a function of distance from ground zero (GZ) might be
expected, this relationship is difficult to establish since considerable cleanup took place
after the test. Some indication of decreasing size with increasing distance from GZ has
been reported by Tamura (1975). Samples taken from 500 to 6700 ft from GZ showed
that at 500 ft the 125- to 50-jum soil-size fraction contained 25% of the activity; at 6700
ft this fraction contributed less than 2%.

The solubility of plutonium oxides decreases with increasing ignition of the oxide
(Cunningham, 1954). The plutonium in the safety shot sites was not subjected to fission
temperatures but to explosion temperatures. The lower solubility of plutonium at NTS
has been reported by Tamura (1976), who subjected contaminated soils from NTS to 8Af
nitric acid extraction at room temperatures. Compared with samples from Oak Ridge
Nafional Laboratory (ORNL) and Mound Laboratory (ML), the NTS samples were only
one-fifth to one-eighth as soluble.

Rocky Flats. The contamination at the Rocky Flats (RF) plant in Colorado was caused
by leaking barrels of Pu-contaminated cutfing oil (Krey and Hardy, 1971). Before

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 155

containment the plutonium-bearing oil was filtered to remove the large particles (Navratil
and Baldwin, 1977). The plutonium would not oxidize in the oil; however, once the
plutonium was in the soil and the protective oil was leached, the oxide form would
predominate. With room-temperature extraction using 8M nitric acid, the RF samples
showed 15 to 20% dissolution; this can be compared with the 10 to 15% observed in the
NTS samples. Tliis suggests that both RF and NTS plutonium are similar in character; the
higher solubility of the RF vs. the NTS samples in the mineral acid may be ascribed to the
smaller size of the plutonium or to the lower temperature of ignition of the RF samples.

Mound Laboratory. Contamination in the canal at ML occurred in 1969 through
sedimentation of eroded soil particles contaminated with ^^*Pu. Initially, the plutonium
was in an acidic solution as plutonium nitrate. During transfer the pipeline ruptured and
the plutonium was sorbed on soil particles. The sorbed plutonium was eroded during
cleanup operations by intense rains (Rogers, 1975). Thus, unlike the metallic origin of the
plutonium at the NTS and RF locations, the ML contamination was originally in a soluble
form. That the character of the plutonium differed from the NTS and RF samples is
exemphfied by the higher solubility (80 to 85%) in cold 8M nitric acid. The
contamination in the canal should be differentiated from the soil contamination reported
by Mullerand Sprugel (1977). They reported that the source of the ^^*Pu in the soil 1
mile east of the Laboratory was aerial emissions from stacks. These emissions were not
characterized physically or chemically (MuUer and Sprugel, 1977).

Oak Ridge National Laboratory^ The ORNL site involves two different contaminating
situations. One of the sites was formerly a holdup pond of wastewater for radionuclide
retention. After the pond was drained in 1944, the bottom sediment was exposed, and a
young forest developed on the floodplain. The depth distribution of the plutonium
suggests that the plutonium was sorbed on particles that settled to the bottom. The 8M
nitric acid extraction revealed that 60 to 75% was soluble at room temperature and 1-hr
extraction. This relatively high extraction suggests a monomeric hydrolyzed form of
plutonium.

The second site of contamination was the bottom sediment of a pond that served
initially as a waste-receiving pond for ORNL liquid waste. With an improved
waste-management system, the pond then served as a secondary settling pond for effluent
from a low-level wastewater treatment plant (Tamura, Sealand, and Duguid, 1977). The
higher activity level in the pond and its close proximity to ORNL suggest that the pond
served as the primary waste retention system before overflowing into the White Oak
Creek. The 8M nitric acid extraction revealed that 90% of the Pu in this sediment was
soluble.

Citric acid extractability of the NTS, ML, and ORNL (tloodplain) samples has been
pubhshed (Tamura, 1976). The increasing order of extraction by the citrate was: NTS,
1%; ORNL, 25%; and ML, 50%. The time of contact of the citric acid with the sohds was
30 min at room temperature. Unpublished data by these investigators show that citric
acid treatment of the RF sample extracted approximately 1 0% of the plutonium.

The percentage range in Pu extractability by the mineral acid and citric acid from the
different site samples reveals the differences in the plutonium at these sites. Although not
established quantitatively, these differences should also be reflected in the uptake
coefficients of vegetation grown at the sites. The extraction data presented further
suggest that plutonium derived from metallic sources, such as at NTS and RF, is less
soluble than that derived from an initially solubilized form, such as at ML and ORNL.

156 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Chemical Characterization

Plutonium Behavior in ORNL Soil

We have conducted a partial characterization of Pu behavior in an alluvial soil
contaminated in 1944 at ORNL. The nature of the Pu source material appears to have
been low-level waste solutions in which the Pu was probably present in monomeric forms.
This soil probably represents the oldest contamination event since the tlrst milligram and
gram amounts of Pu were extracted from the Oak Ridge graphite reactor in January and
February 1944. The study area was used for low-level radioactive-waste retention from
March 1944 to September 1944 and was subsequently abandoned.

Because the mode of release was soluble Pu rather than calcined Pu02, the site is
proving invaluable for developing concepts on the biogeochemistry of Pu. Plutonium
distributions in the contaminated floodplain soil and biota are available (Dahlman,
Gartin, and Hakonson, this volume). Total soil Pu has been determined by hot 8M HNO3
leaching. These values are the same as those for HF— HNO3 leaches.

Comparative Behavior of Pu, Tli, and U. Comparative extractions have been conducted
to compare the extractabilities of Pu, U, and Th. Two extractants are used, \M HNO3
and 10% sodium carbonate and 5% sodium bicarbonate. The former represents a weak
acid extraction, and the latter is used for U extraction in ore processing. Table 3

Table 3 Extraction of U, Th, and Pu from
Floodplain Soil Using Mild Extractants

Percent extracted*

Extractant

Th Pu

IM HNO3
10%Na2CO3-5%NaHCO,

73
71

7.9 7.7
45 54

♦Based on Ul UNO, extractable LI (8.18 Mg/g).
Th (16.8 /jg/g), and Pu (135 dpm/g).

illustrates that Pu and Th are extracted similarly, whereas U is much more readily
extracted. Althougli U extraction changes little between the acid and basic systems, Th
and Pu are extracted more readily with carbonate. In the case of carbonate extraction,
considerable organic matter (humic material) was solubilized; dialysis studies in carbonate
indicate that the carbonate-extracted Pu was not bound to the solubilized organic matter
but was present as a diffusible carbonate complex. As is discussed later, however, some of
the Pu was associated with humic acids in the soil.

Thorium-234 and plutonium-236 were added to replicate 10-g samples containing
20 ml of 0.5M nitric acid to further characterize the relative behavior of Pu and Th in the
floodplain soil. The samples were equilibrated for 0.5. 1, 4, and 24 hr. The amount of
indigenous 239,240pjj ^^^ ^^^Th in solution and the amount of added isotopes in
solution were assayed. The results are plotted in Fig. 6. The hot 8M nitric acid extractable
^^ Th and '''''Pu were used as tlie basis for inventorying the total indigenous elements;
thus, witliin 0.5 hr for Th and 1 hr tor Pu. both indigenous isotopes achieved the same
soil-to-solution distribution as the added isotopes. The sluggishness of the ^'''"Pu toward

ASSOCIATIONS OF Pu AND OTHER ACT IN IDES IN SOILS 157

Q
Q
<

_l
<

o
cc

2 5

EQUILIBRATION TIME, hr

Fig. 6 Comparative behavior of added ^^*'Pu, ^^^Th, and indigenous 2 3 9.24opu ^j^j
^ ^ ^ Th during equilibration with O.SM HNO3 . Soil was contaminated with ^ ^ ' "^ "" Pu in
1944.

equilibration during the first 0.5 hr is not yet understood. It may have been due to the
presence of a small amount of Pu(V and VI) at the start of the equilibration. It is
apparent, however, that the SM nitric acid soluble isotopes rapidly redistribute in the
same manner as the added isotopes and that the solution-phase indigenous isotopes are in
equilibrium with soil-bound elements. This behavior strongly suggests that both Pu and
&M nitric acid soluble Th are surface sorbed rather than occluded or trapped in mineral
matrices. The strong carbonate-extraction results further support this concept since only
a few elements are soluble in this extractant. Iron, for example, is not solubilized.

Other Observations on Physicochemical Associations. Bondietti, Reynolds, and Shanks
(1976) discussed the probable association of part of the Pu with soil organic matter. This
conclusion was based on the solubilization of part of the soil humic acids using chelating
resin (Na form). The resin was prebuffered to the soil pH (6.5); the subsequent
decalcification of soil solubilized 15% of the soil organic C. These soluble humates
contained 5% of the soil Pu. In addition to the humic-associated Pu, 13% of the Pu was
associated with the resin itself. Assuming equal distribution of Pu with soil organic C
(which is not likely) and assuming that the resin-associated Pu was organically bound
initially, 55% of the Pu at most was associated with organic matter. In reality, the
fraction was probably substantially less. Repeated treatments of the soil with NaOCl to
destroy organic matter did, however, remove 82% of the soil Pu. The bleach treatment,
which was conducted at pH 9.5, minimized inorganic mineral destruction. Removal of
most of the Pu with bleach also suggests that the Pu was surface sorbed.

t58 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

4 8 12

EXTRACTION TIME, weeks

Fig. 7 Extraction of Pu from contaminated soil by chelating and cation-exchange resins.

Mild Extractants. Bondietti, Reynolds, and Shanks (1976) reported on the transfer rates
of this soil-bound Pu to chelating resin. The objective was to evaluate what fraction of the
soil Pu would transfer from the soil sohd phase to a resin solid phase as an approximation
of soil— root interactions. For three soil samples, the transfer of Pu was characterized by
an initial period of rapid Pu transfer followed by much slower rates (Fig. 7). Between 9
and 1 1% of the Pu desorbed from the soil in 14 weeks. The soil-bound Pu did not transfer
to a cation-exchange resin (Dowex-50) (Fig. 7). These observations indicate that there is a
form of Pu present that will transfer from the soil to resin, but it is not bound by cation
exchange. Also, the mobile species are exhausted rather quickly, which suggests that only
Pu on the surface of soil peds is involved in the redistribution.

Tamura (1976) reported that 7% of the Pu desorbed when equilibrated with sodium
citrate (pH 7.3), which suggests a fraction similar to the resin-extractable Pu. Extraction
of the soil with citric acid (pH 3) removed 23% of the Pu in that study. Bondietti,
Reynolds, and Shanks (1976), using DTPA (pH 6.5) to extract Pu from this soil, found
that 28% was removed with no difference between 4 and 20 hr of equilibration. In
addition to the extractions using organic chelating agents, the soil was equOibrated for
18 hr with 10 ^M sodium bicarbonate. At a 4/1 solution/ soil ratio, 0.08% of the Pu was

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 159

TABLE 4 Concentration Ratios for U, Th, and Pu
of Plants Grown on Contaminated Soil*

Concentration ratio (x

IQ-^)

Plant

238U

232-j^

2 3 9,240 p„

Leaves and stems

Snapbean (2)

9.3

2.1

1.9 ± 0.7

Soybean (3)

17 ± 1.7t

4.6 ±0.35t

2.0 ± 0.04t

Millet

17 ± 0.7t

0.1 ± 0.5t

0.2 ± 0.07t

Tomato (2)

23 ± 1.4t

5.8 ± 0.4t

6.0 ± 0.5t

Garden beet

8.3

2.5

3.4

Fruit, seed, and storage organ

Soybean (2)

0.55 ± 0.12t

0.1 ± 0.05t

<0.05 ± O.Olt

Squash

Whole

1.9

0.45

0.081

Peeled

<3.0

<0.18

<0.15

Irish potato

Whole (4)

9.9 ± O.Olt

2.0 ± 0.03t

1.4 ± 0.08f

Peeled (3)

0.9 ± 0.03i

<0.12 ±0.02t

0.18 ± 0.02t

Snapbean (pod)

1.0

0.3

0.1

MUlet

1.0

2.0

0.2

Tomato (4)

1.5 ± 0.07t

Not detemiined

0.1 ± 0.03t

Tomato (1)

0.7

0.17

Not determined

Beet (peeled)

0.3

0.6

*Values are ± standard errors of replicate analysis; for single samples, U and Th
analyses are typically ±10%; Pu analysis is ±50% (counting error). Data set represents
plants cleaned by washing. Edible tissues were peeled or cleaned as if being prepared
for cooking; several analyses (not included) showed high and similar CR's for three
elements, indicating soil contamination; in no case, however, was Th and/or Pu
significantly higher than LI.

tMean ± standard deviation.

found in the aqueous phase after uhracentrifugation (9000 x g for 1 hr). This
corresponds to a desorption Kj of 5 X 10^ ml/g.

Extraction by Growing Plants. Like chemical extractants, plants grown on this
floodplain soil reveal that U is extracted more readily than Pu or Th. Table 4 illustrates
this for both whole plants and reproductive and storage organs. Both Th and Pu appear in
vegetative tissue in similar concentrations relative to 871/ nitric acid extractable soil values.
This observation suggests that the Pu is probably present in the III or IV oxidation state
rather than in the V or VI. The similarity to Th suggests that the IV state dominates.

Conclusions on Chemical Associations of Pu in an Alluvial Soil

Tamura (1976) concluded that the Pu in the tloodplain soil was most likely present in
monomer rather than polymer form. This conclusion was reached because the Pu was
highly extractable in citrate and leached in cold 8Af HNO3. The extraction of the Pu by
strong carbonate solution and the extractability with diethylenetriaminepentaacetic acid
also suggest that monomeric forms are present as do the isotopic-exchange results.

The Pu associated with this soil appears to represent surface-sorbed associations, with
humic materials representing one, but not the only, site. About 7 to 11% of the Pu is

160 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

readily mobilized by citrate or resin equilibrations where the pH is maintained at natural
values. A small fraction (0.08%) is soluble in dilute sodium bicarbonate. On the basis of
chemical extractions with dilute nitric acid and strong carbonate solutions and by
plant-extraction data, the Th and Pu appear to behave similarly, which suggests similar
chemical associations in the soil. Since the soil itself is largely alluvium, the extractable
Th is probably associated with colloidal surfaces of secondary minerals. The intrusion of
Pu may have resulted in similar associations.

For this soil and for the type of contamination that was believed to have occurred, it
appears reasonable to conclude that 30+ yr after deposition in soil Pu is not assimilated
by vegetation to a greater extent than natural Th. To the extent that Th is recognized as
an element which does not readily transfer in biological food chains, a similar behavior
for Pu should be observed many centuries after its release to the biosphere.

Physical Characterization

Observations of Particle-Size Association ofPu in Contaminated Sites

As noted in an earlier section, plutonium can actually be released into the environment
by several modes. In addition, the size and character of the plutonium as it is being
introduced are subject to change as it interacts with environmental material. After the
interaction, some of the properties of the plutonium can be controlled by the matrix
properties.

Studies of the association of plutonium with soil and sediment particles have been
reported by Tamura (1976) for several contaminated sites. He reported that, in the safety
shot sites of NTS, the plutonium found in the soils surrounding the site was primarily
associated with coarse-silt (50 to 20 ;um) and fine-sand size (125 to 50 /jm) fractions. The
activity/size ratio of the coarse-silt fraction of two samples reported by Tamura (1976)
was approximately 7.7 and of the clay size was approximately 0.5. Since the size
segregation was based on the density of silicate particles (2.65 g/cm^ density), it could
not be ascertained whether the plutonium particles were of the designated sizes or
whether the plutonium particles were of a finer size and attached to the silicate surfaces.
These size associations are consistent with the findings of Mork (1970), who reported in
his studies of NTS soils that the major portion of the activity was associated with
particles larger than 44 iim.

In contrast to the NTS soils, the Pu in the bottom sediment originating from a
waste-transfer line leak at ML was primarily associated with particles less than 2 ^um in
diameter (Tamura, 1976). Interestingly, MuUer and Sprugel (1977) reported that small
amounts of plutonium released from stacks at ML and absorbed by soils were also
primarily associated with the <2-jum soil particles. They also found that fallout
plutonium in the environs of ML showed the same pattern of concentration in the various
size fractions.

A sample from the floodplain soil at ORNL showed that the plutonium distribution
followed the soil-particle size distribution (Tamura, 1976). This distribution would
indicate that the plutonium in the floodplain was part of settling sediment particles that
had reacted with the plutonium farther upstream. As noted earlier, upstream of the
floodplain is a waste pond that contains plutonium in the bottom sediment; the overflow
from this pond empties into the creek flowing to the floodplain. The activity/size ratio of
the floodplain sample was 1.40 in the clay size and 0.97 in the coarse-silt fractions; in

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS idi

comparison, the ratio for the two sizes in the ML samples were 2.34 and 0.29 in the
sediment sample and 2.42 and 0.60 in the soil sample. These results show that the ML
samples are enriched in the fine-clay size fraction.

Tamura (1976) reported on the size association of plutonium in a RF soil taken from
the 5- to 10-cm depth near the spill site. The activity/size ratio was 2.26 in the clay
fraction and 1.26 in the coarse-silt fraction. In another sample taken 1 km east of the spill
site, the ratios of the two sizes were 3.20 and 0.98, respectively (Tamura, 1977a).

The size associations of the plutonium show that in the safety shot sample of NTS the
clay-size fraction is not enriched in plutonium; at the other sites the clay size is relatively
enriched. However, the association with clay in the enriched samples does not necessarily
mean high acid solubility. The 8M nitric acid extraction at room temperature revealed
that the ML and ORNL samples are quite soluble (over 60%); the RF sample was less
soluble (15 to 20%). The difference in the acid solubility is likely due to the initial
soluble form in the ML and ORNL releases and the metallic nature of the RF sample.

Implication of Particle Size Association

Tamura (1977b) attempted to evaluate the significance of the size association of
plutonium on soil particles in terms of potential hazard due to resuspension and
inhalation of contaminated particles. He considered the soil particle size association of
the plutonium, the depositional character of the different particle sizes in the pulmonary
compartment of the lung, and the fraction of activity in the resuspendible fraction in the
soil. This initial attempt did not include considerations of soil erodibility, vegetation,
field size, and surface-rougliness factors, which are important in wind erosion of soils
(Skidmore, 1976).

Table 5 shows the soil plutonium indexes calculated from the three factors for the
four contaminated sites. The less than 125-/im size is considered to be the resuspendible
fracfion; others have suggested the less than lOO-jitm sizes (Chepil, 1945; Healy, 1974),
but the available data are given for the slightly larger size. The soil activity factor is
defined as the activity per unit weight of mass for each size fraction. This factor is derived
by dividing the activity portion of a given size by the mass contribution of that size; it
therefore weights the activity in the different potentially inhalable sizes.

Table 5 also gives the depositional fraction derived by the Task Group on Lung
Dynamics (1966) and the depositional factor derived as a product of the soil activity
factor and the depositional fraction. The depositional percentage of the larger
resuspendible sizes is relatively low; most of these particles are filtered by the upper
respiratory tract and have a short biological half-life (Task Group on Lung Dynamics,
1966). Also included in Table 5 is the fraction of the activity found in the < 125-/am sizes.
The high percentage results in a small effect on the final soil factor. The activity
distribution was detemiined by using water suspension and either ultrasonic treatment or
chemical dispersant (ML sample). Thus the actual association of the plutonium in the soil
may be different and should be evaluated.

The final soil index shows a range of 0.52 for the NTS sample to 1.26 for the RF
sample. This implies that the plutonium in the soil at the RF site is potenrially about 2.4
times as hazardous in terms of the inhalation pathway. It should be emphasized that the
number of samples is limited, and the factors may change with more information.
Furthermore, the erodibility and other factors of the soils were not evaluated; hence,
until these factors are evaluated and larger numbers of samples are investigated, the
indexes are only tentative.

i62 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

A soil-level standard of 2 (d/min) g~ ' of soil was tentatively set for the state of
Colorado (Johnson, Tidball, and Severson, 1976). The value of 2 (d/min) g' did not
specify the soil size fraction. If this value were used for tlie RF sample, then the
equivalent standard for the NTS would be 5 (d/min) g"' (1.26/0.52 x 2 = 5). Similar
calculations for ML and ORNL give 2 and 3 (d/min) g~ ' , respectively.

Healy (1974) suggested 500 (d/min) g~' for bare soil without reference to any
particular site; thus this suggested level miglit be interpreted as a general guide. The soil
factors in Table 5 show that NTS has the lowest value of 0.52. If 500 (d/min) g^' were
used as a general guide, the allowable concentration for NTS would be 960 (d/min) g^'
(1/0.52 X 500 = 960). Similarly, for RF, ML, and ORNL, the values would be 395, 425,
and 615 (d/min) g^' , respectively.

the importance of establishing a soil standard is related to long-term health risk for
exposed populations. For example, the U. S. Environmental Protection Agency (1977)
proposed a soil-screening level of 0.2 //Ci/m^ (surface 1-cm depth). This value was
established as a reasonable soil level whereby the resulting lung and bone doses to the
critical segment of the exposed population would be below proposed limits.

TABLE 5 Soil Factor Calculated from Soil Activity Factor,
Depositional Factor, and Resuspendible Fraction

Size, Soil Activity Soil activity Depositional Depositional Resuspendible Soil

jum fraction fraction factor fraction factor fraction Pu index

Nevada Test Site (Area 1 3)

0.04

0.03

0.75

0.40

0.30

2 to 5

0.03

0.04

1.33

0.12

0.16

5 to 125

0.43

0.92

2.19

0.03

0.07

0.50

0.99

4.27

Rocky Flats

0.5 3

0.12

0.28

2.33

0.40

0.93

2 to 5

0.04

0.14

3.50

0.12

0.42

5 to 125

0.34

0.49

1.44

0.03

0.04

0.99 0.52

0.50 0.91 7.27 1.39 0.91 1.26

Mound Laboratory*

<2 0.19 0.46 2.42 0.40 0.97

2 to 4 0.09 0.14 1.56 0.12 0.19

4tol25t 0.72 0.40 0.56 0.03 0.02

1.00 1.00 4.54 1.18 1.00 1.18

Oak Ridge National Laboratory

0.29

0.40

1.38

0.40

0.55

2 to 5

0.10

0.09

0.90

0.12

0.11

5 to 1 25

0.59

0.51

0.86

0.03

0.98 1.00 3.14 0.69 1.00 0.69

*Data from Muller and Sprugel, 1976.

t Assumes particles greater than 45 jum to be less than 1 25 jum.

ASSOCIATIONS OF Pu AND OTHER ACTINIDES IN SOILS 163

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164 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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and Certain Soil Components. CH3COOH-NH4 System, in Environmental Chemistry and Cycling

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of Selected Minerals, 7. Environ. Qual, 6(1): 89-95.
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Petroleum Geologists, Memoir 18.
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Radioecology of Plutonium and Other Transuranics in Desert Environments, M. G. White and P. B.

Dunaway (Eds.), USAEC Report NVO-153, pp. 27-41, Nevada Operations Office, NTIS.
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Nov. 17-21, 1975, pp. 213-229, STI/PUB/410, International Atomic Energy Agency, Vienna.
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Proceedings, GatUnburg, Tenn., Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), ERDA

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Sources of Variation in Soil Plutonium
Concentrations

JOHN E. PINDER III and DONALD PAINE

Variations in ^^^Pu and 239,240^^^ concentrations in surface soil near a nuclear-fuel
reprocessing facility were attributed to distance from the point of aerial release;
micro topographical heterogeneity in deposition; and sampling error, which included
aliquoting and analytical errors. Distance from the point of release accounted for
approximately 75% of the variation in concentrations of both nuclides, whereas sampling
error accounted for less than 5% of the variation. Microtopographical heterogeneity
accounted for approximately 20% of the variation in 239.240^^^ concentrations but only
5% of the variation in '^^^Pu concentrations. This difference may be due to different
histories of deposition of the nuclides at the site. Other sources of variation, errors in the
statistical models, and the implications for future sampling are discussed.

Concentrations of radionuclides in soils, plants, and animals are usually highly variable,
with coefficients of variation (standard deviation/mean) usually exceeding 1 .0
(Eberhardt, 1964; Remmenga and Whicker, 1967; Finder and Smith, 1975; Shanks and
De Selm, 1963). This is especially true of the isotopes of plutonium. Large coefficients of
variation in soil plutonium concentrations have been reported for plutonium contamina-
tion resulting from weapons testing (Nyhan. Miera, and Neher, 1976; Romney et al.,
1976), ''safety-shots," i.e., chemical explosions of nuclear weapons material (Gilbert
et al., 1976). aqueous discharges from industrial faciUties that handle plutonium
(Hakonson and Nyhan, this volume), and deposition of aerial releases from reprocessing
faciUties (Adriano, Corey, and Dahlman, this volume; Adriano and Finder, 1977;
McLendon, 1975; McLendon et al., 1976). A portion of this large variabiUty may be due
to the release of plutonium in particulate form and the analytical errors caused by
including various amounts of these particles in a sample (Doctor et ah, this volume;
Adriano, Wallace, and Romney, this volume); however, there must be a greater
understanding of the causes of this variation before the cycling processes of plutonium
can be fully understood or efficient sampling programs can be designed to estimate
plutonium concentrations or inventories. The purpose of this study was to evaluate the
relative importance of several potential sources of variation in soil concentrations of
^^^Pu and 2 3 9,2 4 op^ ^j^^^ ^^^d been released to the atmosphere from a reprocessing
facility at the U. S. Department of Energy's Savannah River Plant (SRP) near Aiken,
South Carolina.

We hypothesized three main components of variation in plutonium concentrations.
First, we expected the distance from the point of release to be important because soil
concentrations have been shown to decrease with increasing distance from the point of

165

166 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

release (McLendon, 1975; McLendon et al., 1976). Soil concentrations were negligibly
affected by reprocessing facilities at distances exceeding 10 km from the point of release
(McLendon etal., 1976). Second, we anticipated that plutonium concentrations in soils
collected in close proximity (i.e., within 2 ni) to one another might vary greatly owing to
microtopographical heterogeneity in the deposition of plutonium particles. Tlie occur-
rence of a single large particle at a site could greatly elevate the soil plutonium
concentration. Soil concentrations could also be elevated by an accumulation of
resuspended plutonium particles due to some feature of terrain, vegetation, or surface
roughness. The third source of variation was termed "sampling error." This is the
variation in plutonium concentrations observed among aliquots of a well-mixed soil and
includes both errors introduced in taking an aliquot for analysis and the analytical errors
associated with determining plutonium concentration in the aliquot.

Large plutonium particles may affect both the microtopographical heterogeneity and
the sampling error components of variation. If plutonium deposition from aerial releases
occurs as large particles that are resistant to weathering, the particle deposition will influ-
ence the sampling error term because the particle will occur in one aliquot and not in the
others. If, however, the deposition occurs as large particles that are easily weathered into
smaller independent particles which can be easily homogenized in the sample, then
aliquots of the soil will have similar concentrations. In that case, the effect of particulate
deposition would be expressed as large differences in concentrations among closely
spaced soils, i.e., as microtopographical heterogeneity. A mixture of particle sizes, or the
confounding effects of resuspension and redistribution of particles, may produce results
that are intermediate between the above two extremes.

Methods

To partition the total variation in soil plutonium concentration into components that are
due to distance from the source, microtopographical heterogeneity, and sampling error,
we collected three soils from each of five independent transects located from 183 to
436 m to the northwest of the point of release (a 62-m stack that exhausts filtered air
from the internal atmosphere of the reprocessing facility). Each transect was 2 m long,
and soil was collected at distances of 0, 1 and 2 m along the transect. Soil was obtained
from the upper 5 cm of the profile with a soil auger. For each point on each transect, we
homogenized the soil by vigorous, manual shaking in a cardboard ice-cream carton for
1 min and divided the soil into two samples of equal volume. The auger was cleaned
between sampling points, and a separate ice-cream carton was used at each point to
prevent cross-contamination. The samples were collected in the vicinity of H-area at the
SRP on a field that was being used for long-term studies of the transport and fate of
transuranic elements in agricultural ecosystems. The samples were collected before the
soil had been disturbed for agricultural purposes in November 1974.

The concentrations of ^^^Pu and 2 3 9,2 4 op^^ ^q^q determined by alpha spectrometry
under the direction of A. L. Boni at the Savannah River Laboratory (operated by E. I.
du Pont de Nemours & Co.). Ten-gram aliquots of soil were ashed and leached with HCl.
A triisooctylamine, 200- to 800-mesh solid ion-exchange resin was used to remove the
plutonium from the leachate by liquid ion exchange. Plutonium was leached from the
resin with H2 SO3 , electroplated on platinum disks, and counted. Plutonium-236 was used
as an internal standard in estimating recoveries of plutonium from the soil. Large soil
particles (>3 mm in diameter) were removed from the samples before aliquots were
drawn.

VARIATION IN SOIL PLUTONIUM CONCENTRATIONS 167

Mathematical Formulation

Comparisons of concentrations between samples at eacli transect point provide an
estimate of sampling error. Comparisons of samples among points on each transect
provide an estimate of microtopographical heterogeneity in soil concentrations, and
comparisons among transects provide an estimate of the importance of distance from the
point of release.

Let Y/y^- be the plutonium concentration in the A:th sample {k = 1 .2) drawn from the
/th transect point (j = 1 .2.3) in the /th transect (/ = 1 ,2, . . . ,5). Y^-yt may be partitioned
into components according to the statistical model,

Yifk=Ui + Di + Mij+eij/, (1)

where ju = grand mean of concentrations
D/ = effect of the /th transect location
M,y = effect of the/th position in the /th transect
ejji^ = sampling error

We assume that D/, M/y, and e/yy^ are normally distributed random variables with
/^D ~ i^M ~ A'e ~ 0 and variances Qq, a^^, and ol and that D/, My, and Cy^- are
independent or, in other words, that 0^), ajj^, and ol are constant for all combinations of
D and M. Later in this chapter we will test the validity of these assumptions and discuss
the inaccuracies introduced by the failure of the data to meet them. Random samples of
Y have expected mean jj. and expected variance o = o^y + o^ + o^ . The parameters a^,
a^, and ol are termed the variance components of a^ . Equation 1 represents a two-way
nested analysis of variance with random effects. Nested analyses and random-effect
models are discussed in greater detail by Scheffe (1959) and Searle (1971), who also give
procedures for estimating Qq, a^, and ol and calculating confidence intervals about the
estimates. The relative importances of distance, microtopographical heterogeneity, and
sampling error are given by the intraclass correlation coefficients, pq = Oy^/o^ ,
PM - ^m/^^ ' '^^^ Pe ~ (^l/o^ ' respectively (Scheffe, 1959). The estimated p^ for the ath
effect is given by Pa-o^/o^, where o^ is the sum of the estimates of the individual
variance components. Scheffe (1959) also gives procedures for testing the statistical null
hypothesis, Hq : o^ = 0, versus the alternative hypothesis, H/^ : o^ > 0. The formulas and
procedures outlined by Scheffe (1959) were used in the following analyses. The
Statistical Analysis System was used for the computations (Barr et al., 1976).

Results

Estimates of a^, a^j, and ol ; 95% confidence intervals about the estimates; and estimates
of intraclass correlation coefficients for the concentrations of ^^^Pu and ■^^^ '■^'*°Pu are
given in Table 1. All the variance components for 2 3 9.2 4 0pjj ^^^.^ statistically greater
(P<0.05) than the corresponding variance components for ^^^Pu. The o^ for
2 3 9,2 4 op^ was 1 1.853, whereas P for '^^Pu was only 0.129. Mean concentrations were
2.23 pCi/g for "^'^^°Pu and 0.481 pCi/g for ^^^Pu.

For both radionuclides, sampling error accounted for less than 5% of the total
variance, and o^) was the largest component of the total variance for both nuclides. The
major difference between the radionuchdes occurred in the relative importance of a^j.
Microtopographical heterogeneity was an important component of the variation in

168 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

239,240p^ concentrations and accounted for approximately 20% of the variation. The

95% confidence interval about a^ for Pu concentrations includes 0 and suggests that
Oy^ may not be greater than 0; however, the inclusion of 0 is the result of a slight
inaccuracy in confidence intervals calculated according to the formulas of Scheffe (1959,
pp. 231-235). An F-test of the null hypothesis, Hq : a^ = 0, indicated that a^ was
significantly greater than 0(F = 2.99; df= 10,15; P< 0.05).

Alternate Statistical Models

If each transect is considered separately, a one-way analysis of variance procedure can be
used to estimate a^ and o^ (Scheffe, 1959, pp. 221-224). These estimates, symbolized
Oy^ I and o\i, can be compared among transects to evaluate the assumptions of
independence of D/, My, and e,yyt- The a^,- and ol^t for ^^^Pu and 2 3 9,2 4 Op^
concentrations for each transect are compared in Table 2.

The aj^,/ for ^^^Pu concentrations range from -0.0003333 to 0.01674. Although
negative a^ ,• are impossible, negative a^ i can occur owing to the sampling variances of
mean-square' terms (Scheffe, 1959, pp. 228-229; Searle, 1971, pp. 406-408). The 95%
confidence intervals for all the a^ ,• overlapped and indicated that the estimates could
have been drawn from a common value of a^ ,• at all transects. The b\^i differed
significantly among transects (Fmax. = 85.6; df= 5,3; P < 0.05) (Kirk, 1968) and tended
to decrease as the mean concentration decreased.

The a^ ,■ for 239,240p^ concentrations ranged from -0.004167 to 12.16 and were
significantly greater than 0 for three of the five transects. Again, the 95% confidence
intervals for the a^ ,• overlapped. The broad confidence bands around the a^ ,■ in Table 2
for both ^-^^Pu and 2 3 9,2 4 0p^j concentrafions were due to the small number of degrees

TABLE 1 Estimated Variance Components, 95% Confidence Intervals About the

Estimates, and Intraclass Correlation Coefficients for the Model o^ = ap + a^ + Qg ,

Which Partitions the Total Variance of Soil Plutonium Concentrations (a^ ) into

Components due to Distance from the Source (a^), Microtopographical

Heterogeneity (a^), and Sampling Error (ag)*

Estimated

Degrees

95% Confidence interval

Estimated intracla,ss

variance

Lower

Upper

correlation

Component

component

Symbol

freedom

bound

coefficients

238pu

Distance

0.1168

;i2

0.03947

0.9830

0.904

Microtopographical

heterogeneity

0.006185

•^ 2

-0.0001446

0.02555

0.048

Sampling error

0.006220

0.003393

0.01490

0.048

2 3 9 ,2 4 Opj,

Distance

9.000

2.689

80.91

0.759

Microtopographical

heterogeneity

2.575

1.175

8.220

0.217

Sampling error

0.2771

«^e

0.1511

0.6639

0.023

*Estimates were computed from 30 determinations of ^^*Pu and 2 3 9.2 4 op^ Concentrations are
in picocuries per gram.

VARIATION IN SOIL PLUTONIUM CONCENTRATIONS 169

TABLE 2 Distances from the Source of Plutonium Release, Mean Plutonium
Concentrations in Soil, and Estimates of the Variance Components for the Model

,5)*

oj =0m,/

+ ol j for Each Transect (/ = 1 ,2, .

Distance
from

Mean

95% Confidence interval

source,

concentration.

Lower

Upper

Transect

pCi/g

bound

238py

183

1.02

0.009167

-0.196

0.664

0.1568

214

0.43

0.004742

-0.189

0.422

0.01218

275

0.55

0.01674

0.00388

0.675

0.0007167

406

0.19

0.0006083

-0.00054

0.0276

0.00018S3

436

0.20

-0.0003333

2 3 9 i2 4 Opy

-0.0881

0.0318

0.002333

183

7.81

12.16

0.986

505.7

1.324

214

0.40

-0.004167

-0.0882

1.34

0.01515

275

1.56

0.6259

0.155

25.1

0.01910

406

0.79

0.04060

-0.145

2.07

0.02415

436

0.61

0.05770

0.0132

2.33

0.002550

*Each CT^ j-has df = 2, and each a^ ■ has df = 3. Mean concentrations are computed from all six
samples at each transect.

of freedom associated with the a^ ^ and olj. The olj for 2 39 ,240 p^^ concentrations also
differed significantly among transects (Fmax. = 519,3; df= 5,3; P < 0,01) and tended to
decrease as the mean 2 3 9,2 4 op^^ concentrations decreased.

Although the a^ ^ did not differ significantly among transects, there appeared to be a
positive correlation between a^ ,• and mean concentration for both ^^*Pu and
2 39,240py concentrations. The apparent correlations of a^,- and blj with mean
concentration suggested that o^ ^ and Qg / were proportional to mean concentration.
Proportional relationships between means and variances can be expected because
concentrations varied over a broad range and analytical error was controlled to plus or
minus a percentage of the measured value. The components a^ ,- and olj may also vary
because the statistical model was inappropriate (e.g., a linear model for a nonlinear
process) or failed to contain important causes of variation, such as soil type or
disturbance. The variation in a^ y and olj recorded in Table 2 probably resulted from a
combination of factors, including the proportionality between analytical error and
concentration, the nonlinear relationship between concentration and distance from the
point of release, a change in soil type between transects 3 and 4, and soil disturbances.
The impacts of the first three of these factors on our interpretations of the relative
importance of microtopographical heterogeneity and sampling error were evaluated by
comparing the interpretations resulting from applying more-complex statistical models to
the original arithmetic data as well as logarithmic transformations of the data. None of
the results for these alternative models affected our conclusions that sampling error was a
small fraction of the total variation or that microtopographical heterogeneity was more
important for 2 3 9,2 4 Op^^ ^.j^j^ ^^j. 2 3 8p^ Because some readers may be interested in
specific alternate interpretations, we have added the raw data as Table 3.

/ 12 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

for both nuclides indicates that increased analytical precision would produce only a small
reduction in total variation.

Acknowledgments

This research was supported by contract EY-76-C-09-0819 between the U. S. Department
of Energy and the University of Georgia. We thank R. A. Geiger, R. M. Klein, E. H.
Lebetkin, and K. W. McLeod for assistance in either the field or the laboratory and J. G.
Wiener, J.J. Alberts, and K. W. McLeod for their review of an early draft of the
manuscript. J. Henry Horton, Jr., of the Savannah River Laboratory, kindly provided
data on the history of releases from the H-area facility.

References

Adriano, D. C, and J. E. Pinder III, 1977, Aerial Deposition of Plutonium in Mixed Forest Stands
from Nuclear Fuel Reprocessing,/ Environ. Qual., 6: 303-307.

Barr, A. J., J. H. Goodnight, J. P. Sail, and J. T. Helwig, 1976, A User's Guide to SAS 76, Sparks Press,
Raleigh, N. C.

Eberhardt, L. L., 1964, VariabiUty of the Strontium-90 and Cesium-137 Burden of Native Plants and
Animals, yVamre (London), 204: 238-240.

Gilbert, R. 0., L. L. Eberhardt, E. B. Fowler, E. H. Essington, and E. M. Romney, 1976, Statistical
Analyses and Design of Environmental Studies for Plutonium and Other Transuranics at NAEG
"Safety Shot" Sites, in Transuranium Nuclides in the Environment, Symposium Proceedings, San
Francisco, 1975, pp. 449-460, STI/PLIB/410, International Atomic Energy Agency, Vienna.

Kirk, R. E., 1968, Experimental Design: Procedures for the Behavioral Sciences, Brooks/Cole
Publishing Company, Belmont, Calif.

McLendon, H. R., 1975, Soil Monitoring for Plutonium at the Savannah River Plant, Health Phys., 28:
347-354.

, O. M. Stewart, A. L. Boni, J. C. Corey, K. W. McLeod, and J. E. Pinder, 1976, Relationships

Among Plutonium Contents of Soil, Vegetation and Animals Collected on and Adjacent to an
Integrated Nuclear Complex in the Humid Southeastern United States, in Transuranium Nuclides
in the Environment, Symposium Proceedings, San 1 rancisco, 1975, pp. 347-363, STI/PUB/4I0,
International Atomic Energy Agency, Vienna.

Miliiam, R. C, J. V. Schubert, J. R. Watts, A. L. Boni, and J. C. Corey, 1976, Measured Plutonium
Resuspension and Resulting Dose from Agricultural Operations on an Old Field at the Savannah
River Plant in the Southeastern United States of America, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Irancisco, 1975, pp. 409-421, STI/PUB/410, Inter-
national Atomic Energy Agency, Vienna.

Nyhan, J. W., I\ R. Miera, Jr., and R. E. Neher, 1976, Distribution of Plutonium in Trinity Soils After
28 Years,/ Environ. Qual., 4: 431-437.

Pinder, J. E., Ill, and M. H. Smith, 1975, I requency Distributions of Radiocesium Concentrations in
Soil and Biota, in Mineral Cycling in Southeastern Ecosystems, ERDA Symposium Series, Augusta,
Ga., May 1-3, 1974, E.G. Howell, J. B. Gentry, and M. H. Smith (Eds.), pp. 107-125,
CONI -740513, NTIS.

Remmenga, E. E., and !•. W. Whicker, 1967, Sampling Variability in Radionuclide Concentrations in
Plants Native to the Colorado front Range, Health Phys.. 13: 977-983.

Romney, E. M., A. Wallace, R. O. Gilbert, and J. E. Kinnear, 1976, 239,24op^j ^^^ 241^^.,
Contamination of Vegetation in Aged 1 ail-Out Areas, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Irancisco, 1975, pp. 479491, STI/PUB/410, Interna-
tional Atomic Energy Agency, Vienna.

Schefftf. H., 1959, The Analysis of Variance, John Wiley & Sons, Inc., New York.

Searle, S. R., 1971, Linear Models, John Wiley & Sons, Inc., New York.

Shanks, R. E., and H. R. De Selm, 1963, factors Related to Concentration of Radiocesium in Plants
Growing on a Radioactive Waste Disposal Area, in Radioecology, Proceedings of the first National
Symposium on Radioecology, I ort Collins, Colo., Sept. 10-15, I96I, V. Shultz and A. W.
Klement, Jr. ( Eds.), pp. 97-101 , Reinhold Publishing Corporation, New York.

Statistics and Sampling inTransuranic
Studies

L. L. EBERHARDT and R. 0. GILBERT

The existing data on transuranics in the environment exhibit a remarkably high variability
from sample to sample (coefficients of variation of 100% or greater). This chapter stresses
the necessity of adequate sample size and suggests various ways to increase sampling
efficiency. Objectives in sampling are regarded as being of great importance in making
decisions as to sampling methodology. Four different classes of sampling methods are
described: (1 ) descriptive sampling, (2) sampling for spatial pattern, (3) analytical
sampling, and (4) sampling for modeling. A number of research needs are identified in the
various sampling categories along with several probletns that appear to be common to two
or more such areas.

Most of the existing data on transuranic elements in the environment exhibits a
remarkably high variability from sample to sample. Since analytical procedures for these
elements are both complicated and expensive, many investigators use relatively few
replicates. In those few cases where moderately large samples have been taken, the
underlying frequency distributions generally have been badly skewed (nonsymmetrical).
The use of statistical methods in the design and analysis of studies and the use of efficient
sampling practices would help avoid the reporting of questionable conclusions.

This chapter identifies some sources of information on sampling and statistical
methods relevant to transuranic studies and suggests further research on particular
problems along these lines. We believe that too many studies of transuranic elements are
currently being conducted with unrealistically small samples. In many such cases,
statistical analyses are limited to reporting "counting errors"; thus the inadequacy of the
sampling goes unrecognized, at least by those preparing a report on its outcome. This is,
of course, not universally true but is all too often the case. As time goes on, it is to be
hoped that statistical measures of the adequacy of the sampling and chemical-analysis
procedures will become more widely used. The need for efficient statistical designs should
then become immediately apparent to investigators and sponsors. Efficient sampling
plans, however, require that rather definite objectives be specified for the study. We
consider objectives and the appropriate sampling plans in this chapter.

We will direct our discussion primarily to sampling designs rather than to
experimental designs since many of the "experiments" concerning transuranic elements
that we have encountered thus far are not replicated or have so few replicates that
statistical analysis of the results has little meaning. For purposes of this chapter, an
experiment occurs when the investigator controls, through randomization, the assignment

i73

/ 12 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

for both nuclides indicates that increased analytical precision would produce only a small
reduction in total variation.

Acknowledgments

References

Adriano, D. C, and J. E. Pinder III, 1977, Aerial Deposition of Plutonium in Mixed Forest Stands
from Nuclear Fuel Reprocessing,/. Environ. Quai, 6: 303-307.

Barr, A. J., J. H. Goodnight, J. P. Sail, and J. T. Helwig, 1976, A User's Guide to SAS 76, Sparks Press,
Raleigh, N. C.

Eberhardt, L. L., 1964, Variability of the Strontium-90 and Cesium-137 Burden of Native Plants and
AnimsLls, Nature (London), 204: 238-240.

Gilbert, R. O., L. L. Eberhardt, E. B. Fowler, E. H. Essington, and E. M. Roniney, 1976, Statistical
Analyses and Design of Environmental Studies for Plutonium and Other Transuranics at NAEG
"Safety Shot" Sites, in Transuranium Nuclides in the Environment, Symposium Proceedings, San
Francisco, 1975, pp. 449-460, STI/PUB/410, International Atomic Energy Agency, Vienna.

Kirk, R. E., 1968, Experimental Design: Procedures for the Behavioral Sciences, Brooks/Cole
Publishing Company, Belmont, Calif.

McLendon, H. R., 1975, Soil Monitoring for Plutonium at the Savannah River Plant, //ea/r/z Phys., 28:
347-354.

— -, O. M. Stewart, A. L. Boni, J. C. Corey, K. W. McLeod, and J. E. Pinder, 1976, Relationships
Among Plutonium Contents of Soil, Vegetation and Animals Collected on and Adjacent to an
Integrated Nuclear Complex in the Humid Southeastern United States, in Transuranium Nuclides
in the Environment, Symposium Proceedings, San Francisco, 1 975, pp. 347-363. STl/PUB/410,
International Atomic Energy Agency, Vienna.

Milham, R. C, J. I-. Schubert, J. R. Watts, A. L. Boni, and J. C. Corey, 1976, Measured Plutonium
Resuspension and Resulting Dose from Agricultural Operations on an Old I'ield at the Savannah
River Plant in the Southeastern United States of America, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San 1 rancisco, 1975, pp. 409-421, STI/PUB/410, Inter-
national Atomic Energy Agency, Vienna.

Nyhan, J. W., 1-. R. Miera, Jr., and R. E. Neher, 1976, Distribution of Plutonium in Trinity Soils After
28 Years, y. Environ. Quai. 4: 431-437.

Pinder, J. E., Ill, and M. H. Smith, 1975, frequency Distributions of Radiocesium Concentrations in
Soil and Biota, in Mineral Cycling in Southeastern Ecosystems. ERDA Symposium Series, Augusta,
Ga., May 1-3, 1974, E.G. Howell, J. B. Gentry, and M. H. Smith (Eds.), pp. 107-125,
CONF-740513,NTIS.

Remmenga, E. E., and I-. W. Whicker, 1967, Sampling Variability in Radionuclide Concentrations in
Plants Native to the Colorado front Range, Health Phys., 13: 977-983.

Romney, E. M., A. Wallace, R. O. Gilbert, and J. E. Kinnear, 1976, 239,240py ^^^ =41^^
Contamination of Vegetation in Aged I all-Out Areas, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San 1 rancisco, 1975, pp. 479491, STI/PUB/410, Interna-
tional Atomic Energy Agency, Vienna.

Schefft?. H., 1959, The Analysis of Variance, John Wiley & Sons, Inc., New York.

Searle, S. R., 1911 , Linear Models, John WUey & Sons, Inc., New York.

Shanks, R. E., and H. R. De Selm, 1963, factors Related to Concentration of Radiocesium in Plants
Growing on a Radioactive Waste Disposal Area, in Radioecology , Proceedings of the first National
Symposium on Radioecology, lort Collins. Colo., Sept. 10-15, 1961, V. Siiullz and AW.
Klement, Jr. (Eds.), pp. 97-101 , Reinhold Publishing Corporation, New York.

Statistics and Sampling inTransuranic
Studies

L. L. EBERHARDT and R. O. GILBERT

The existing data on tramiiranics in the environment exhibit a remarkably high variability
from sample to sample (coefficients of variation of 100% or greater). This chapter stresses
the necessity of adequate sample size and suggests various ways to increase sampling
efficiency. Objectives in sampling are regarded as being of great importance in making
decisions as to sampling methodology. Four different classes of sampling methods are
described: (1 j descriptive sampling, (2) sampling for spatial pattern, (3) analytical
sampling, and (4) sampling for modeling. A number of research needs are identified in the
various sampling categories along with several problems that appear to be common to two
or more such areas.

We will direct our discussion primarily to sampling designs rather than to
experimental designs since many of the ''experiments" concerning transuranic elements
that we have encountered thus far are not replicated or have so few replicates that
statistical analysis of the results has little meaning. For purposes of this chapter, an
experiment occurs when the investigator controls, through randomization, the assignment

173

1 74 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

of treatments to his experimental material. Studies, by random sampling, of uncontrolled
phenomena are not experiments in this sense.

Perhaps it is advisable to digress here to make a more detailed dis'tinction between a
sampling design and an experimental design. Consider a study of the effect of, say, four
levels of concentration of some chemical substance on a specific kind of plant. A suitable
experimental design might use, say, 25 individually potted plants kept in a growth
chamber. Each of 5 randomly selected pots would be treated with the same
concentration, and one lot of 5 plants would be a control. Another randomization would
be used in assigning the 25 pots to places in the growth chamber. There would be thus 5
replicates of each type of treatment (concentration of chemical) and 5 control pots.

In contrast, a sample survey design might be used to study the concentration of, say,
plutonium in a natural population of plants in the vicinity of a nuclear test site. Here we
can only observe, by sampling, the results of events over which we usually have no
control. There are no true replicates in the experimental sense. It is now fairly common
practice, however, to speak of "replicate samples," and we would only stress the need to
use the word "sample" to be precise. The investigator cannot control the physical
relationships involved; he has to take specimens from an existing population of plants in
the positions in which they occur. He can, of course, control the sampling process so that
he obtains several individuals of the same species growing rouglily the same distance from
ground zero and so on.

Since science is commonly thought of as being practically synonymous with the
experimental method, many people prefer to regard observations taken on some
uncontrolled process as experiments. Such a view is largely immaterial and irrelevant
insofar as the mathematical and computational aspects of statistical methods are
concerned. Only when we begin to draw inferences from analyses of the data does the
real distinction between "experiment" and "observation" become apparent; i.e., in a true
experiment we can use rather homogeneous material and, by randomization, ensure that
any effects due to position in the growth chamber, genetic factors, etc., are reflected by
the error term in the statistical analysis. Without this element of deliberate control of the
experiment, there is no assurance that unknown extraneous factors will not also influence
the factors under study. Hence the aiialysis of data obtained by sampling is a rather more
hazardous affair. However, exactly the same statistical analyses and an identical
mathematical model can be used for either an experiment or a sample survey.

For the present, many of the immediate needs for statistical guidance in transuranic
research programs can best be served from the sampling point of view. Many textbooks
and experienced statistical practitioners are available to aid in the design and analysis of
experiments. We are also interested in the design of experiments, but we have elected to
concentrate on sampling in this chapter. We will not try to be explicit as to the role of
source terms, but a long list of sources must, of course, be considered; e.g. ( 1) worldwide
fallout from nuclear events; (2) localized fallout from nuclear events; (3) localized
dispersion without a nuclear event, such as safety tests; (4) stack releases; (5) liquid
effluents; (6) accidents involving nuclear weapons; (7)burnup of SNAP devices;
(8) damage to other power sources; and (9) various kinds of eventualities concerning
stored wastes, transport, etc.

We will not attempt to deal with studies of the biological effects of exposures to
various substances. We have discussed the statistical problems in assessing the effects of
low-level, chronic pollutants elsewhere (Eberhardt, 1975a) and can only note here that
they are even more perplexing than those associated with sampling alone.

STATISTICS AND SAMPLING IN TRANS URANIC STUDIES 175

Objectives in Sampling

There are many ways in which the objectives of environmental studies can be arranged,
and we have suggested several such sets (Eberhardt, 1976; Eberhardt, 1977; Eberhardt
et al., 1976). In this chapter we consider four sampling methods to meet various
objectives: descriptive sampling, sampling for spatial patterii, analytical sampling, and
sampling for modeling. Although the four categories are neither mutually exclusive nor
all-inclusive, they do seem to serve as useful devices to cover most situations. Figure 1
illustrates the different objectives using the same physical example. Since the objectives

POINT SOURCE

ACCUMULATION IN ENVIRONMENT

INVENTORY
(VOLUME ESTIMATE)

PATTERN
(CONTOURS)

COMPARISONS

MODEL

Fig. 1 Sampling objectives.

/ 16 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

are quite diverse, illustrating them in the same example stretches the analogies somewhat
and does not allow adequate coverage of the wide range of possibilities. The figure is
meant to portray surface soil concentration of some contaminant emanating from a point
source, such as a power plant stack or cooling tower. Figure 2 is a different summary of
the four objectives.

Descriptive Sampling

Descriptive sampling is the classical approach to sampling. The objectives are to estimate a
total or a mean for some variable (or set of variables) over a definite population.
Textbooks on the subject, perhaps more generally known as survey sampling, have been
available since the early 1950s. The textbook by Cochran (1963) is best known to
biologists, but there may now be 10 or 12 books on descriptive sampling. We have
become accustomed to hearing this methodology called "sampling for inventory" in the
studies of plutonium in soil at the Nevada Test Site. Inventory, however, is not a good

1. INVENTORY

COMMONLY KNOWN AS "DESCRIPTIVE'
OR "SURVEY" SAMPLING

EXAMPLE

STRATIFIED RANDOM SAMPLING

2. PATTERN

PRESENT APPLICATIONS MAINLY IN
GEOLOGY (PETROLEUM AND MINING
EXPLORATION)

MAIN OUTCOME

CONTOUR MAPS

3. COMPARISON

LIMITED ATTENTION IN SURVEY
SAMPLING TEXTS ("ANALYTICAL
SAMPLING") . . . CLOSE RELATION
TO ANALYSIS AND DESIGN OF
EXPERIMENTS

EXAMPLES

• ANALYSIS OF VARIANCE

• ANALYSIS OF COVARIANCE

4. MODELING

MAINLY DEVELOPED IN INDUSTRIAL
EXPERIMENTATION (G.E.P. BOX
AND OTHERS)

EXAMPLE

ESTIMATE PARAMETERS IN TWO-
COMPARTMENT RETENTION MODEL

BODY
BURDEN

TIME

Fig. 2 Some relevant statistical methodology.

STATISTICS AND SAMPLING IN TRANSURANIC STUDIES 1 77

term to use here because it connotes results broader than those supplied in the usual
application of descriptive sampling. Most people think of an inventory as supplying
information on both quantity and location. When sampling is used, it is necessary to
decide which of these attributes should be emphasized. Descriptive sampling is concerned
with quantity.

In Fig. 1 descriptive sampling is illustrated by suggesting a volume-estimation
(integration) process. For graphic purposes, accumulation in the environment has been
shown as a surface. In fact, differential accumulation in, say, soil is, of course, reflected in
changes in concentration, and a total is usually estimated by averaging. Figure 3 shows a
common technique in descriptive sampling, i.e., stratification (which is described further
later in this chapter).

MAP OF STRATA

Proportion

of area
Stratum i" ^^^^tum

No.

1
2
3
4
5

(W,)

0.905
0.060
0.006
0.006
0.023

1.000

Standard

deviation

(S,)

4.23

42.68

221.70

719.37

Allocation (n^)
Inventory Pattern Comparison

31 (41)
21 (22)
11 (12)
35 (23)

90
6
1
1

25
25
25
25

Not relevant here ■
98 (98) 98

100

Fig. 3 Example showing how sampling intensity differs according to objectives.
Sampling for plutonium in surface soil at GMX site (area 5) at the Nevada Test Site.

Sampling for Spatial Pattern

When location is the major objective, the best sampling system may be quite different
from that prescribed in the survey-sampling textbooks for estimating a total. So far most
of the relevant results in this area have been produced in geology and geography and have
not really begun to show up in the statistical literature or the textbooks on sampling. We
(Eberhardt and Gilbert, 1976) have described some of the varied aspects of sampling for
spatial pattern relative to transuranic studies. The rather lengthy discussion following our
presentation in the work cited (see pp. 197-208) should be of interest in the present
context. A textbook on the subject is that of Agterberg (1974).

The basic method for describing pattern is that of drawing contour lines to show
regions of equal concentration (isopleths), as illustrated in Fig. 1.

A nalytical Sampling

Cochran (1963, p. 4) gives a good description of analytical sampling: "Comparisons are
made between different subgroups in the population, in order to discover whether

/ 78 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

differences exist among them that may enable us to form or to verify hypotheses about
the forces at work in the population." Not much space is devoted to this topic in the
textbooks, perhaps because it is very similar to the older body of methodology
encompassed under the heading of the analysis of variance and mostly (but not
necessarily) used in an experimental context. As mentioned in the introduction, the
mechanical details of analysis are very much the same whether an experiment or a survey
is involved. It may be, however, that survey sampling methodology has something
additional to offer in the way of ideas on allocating samples to "domains of study." We
have been impressed by the potential advantages of using auxiliary variables to increase
efficiency, i.e., to reduce costs by use of the analysis of covariance (Eberhardt, 1975b).

One of many comparisons that can be made is illustrated in Fig. 1 where the diagram
has been cut in two directions to suggest that an investigator may want to use statistical
methods to determine whether there are significant differences in concentrations along
transects in two (or more) directions.

Sampling for Modeling

Sampling for modeling covers a lot of ground. Perhaps one instance will serve to illustrate
the topic. Consider the uptake and retention of some radionuclide in an animal. Suppose
the analyses are expensive and involve sacrificing animals for each determination (ai, :
often the case for the transuranics). Since the process is a dynamic one, the outcomes
usually are represented by fitting a curve (modeling) and estimating rate constants (or
half-times). Some time, trouble, and money might be saved by studying ways to select
sampling times so as to obtain the "best" estimates of the parameters (rate constants). So
far as we can tell, this point has not been considered in the many thousands of laboratory
studies on radionucUdes and other trace substances over the last 30 yr in which
sampling times are either uniformly spaced or occasionally separated by geometrically
increasing intervals. One must, of course, also pay attention to determining the structure
of the models used and to various other points (Eberhardt et al., 1976; Eberhardt, 1978).
In Fig. 1 the basic idea is suggested by a heavy line denoting a specific model fitted to
concentrations in one direction from the source. In practice one might want to fit a
model to the entire surface. This might also involve comparisons to see whether a
directional component is needed in the model; so analytical aspects also may be involved.

Sampling Methods and Research Needs

In this section we cite additional references and identify problems that require additional
consideration. The discussions are arranged in the four categories of sampling methods
previously described and are followed by a section devoted to common problems.

Descriptive Sampling

The textbooks devoted to descriptive sampling discuss extensively a rather wide variety of
methods for estimating totals or means by sampling. We mention only those few methods
with which we have had some experience. The main method is stratified samphng,
wherein elements of the population to be sampled are assigned to one of a number of
strata. The basic idea is to assign elements to strata so that the elements in each are as
nearly alike as possible. If this is successful, then the variability within a given stratum is
kept small, and the costs of sampling are thereby reduced. Some advance knowledge on
which to base the stratification (the classification of population elements) is evidently

STATISTICS AND SAMPLING IN TRANS URANIC STUDIES 179

necessary. Often this prior information can be obtained by some relatively inexpensive
means of measurement, or it may be known or inferred from information about the
source of contamination etc.

Once the strata have been determined, the total sample must be allocated to the
several strata. Two general approaches have been used, proportional and optimum
allocation. In proportional allocation the sample is distributed simply in proportion to
the number of elements in each stratum (e.g., in soil sampling, to the area of the stratum).
This scheme is suitable if the variances are about the same in each stratum. Variances
associated with the transuranic elements, however, increase dramatically with mean
concentrations. We thus recommend optimum allocation, which is based on both size of
stratum and variability within the stratum.

Our initial efforts at stratification for sampling soil for plutonium at the Nevada Test
Site are described by Eberhardt and Gilbert (1972). Many of the details of our
subsequent experience appear in Gilbert et al. (1975). An alternative approach to
stratified sampling is to use an accurate but expensive method (such as chemical analyses
for plutonium) to "calibrate" a less accurate but cheaper method. Methods of this sort
fall under the heading of double sampling in textbooks. Some details of an application of
double sampling to sampling for plutonium appear in Gilbert and Eberhardt (1976a). The
method uses ratios or regressions of rather variable quantities and thus poses some
statistical problems (mentioned again later in this chapter). An important action in
designing a double-sampling scheme is to use a cost function to find the combination that
yields minimum cost (or that maximizes precision).

In soil sampling for inventory, sampling by depth needs further study. Much of
our work with stratified sampling has been concerned chiefly with a thin surface soil
layer. Since most of the plutonium is in that layer and resuspension questions focus there,
this is a logical approach. Some soil profiles, however, have been taken to investigate
vertical dispersion, and a detailed evaluation of allocation schemes for sampling in depth
is in order. The problem in profile sampling is, of course, the analytical costs. If 10
increments per profile are taken, costs of even a modest sampling scheme become
exorbitant.

In summary, descriptive (inventory) sampling has a well-known technology. Applica-
tions in any new area do, however, require statistical attention and a certain amount of
research. Unfortunately, methods designed for a specific application are often used in
other situations where they are not appropriate, e.g., the use of methods developed for
global fallout surveys for entirely unrelated purposes (Eberhardt, 1976, pp. 201-202).

Sampling for Spatial Pattern

In a variety of situations, the main objective in sampling is to determine a geographical
pattern rather than to simply estimate total quantities of a substance present in any
particular area. As noted previously the objectives of a study should determine the
sampling scheme. Different objectives, for example, may require very different allocations
of samples to strata (see Eberhardt and Gilbert, 1976). An example of the remarkable
contrast in the way samples might be assigned to several strata according to sampling
plans tailored to three different objectives is shown in Fig. 3. The box labeled
"allocation" gives the distribution of samples to strata appropriate under three of the
objectives of Figs. 1 and 2. Two sets of figures are given for "inventory." One is the
sampling pattern actually used, and the other is based on the results of the survey.

180 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Two applications of sampling for spatial pattern are the evaluation of resuspension,
where the pattern of surface contamination obviously is a controlling factor, and the
difficult matter of determining what portions of an area of high concentration should be
cleaned up (i.e., soU removed etc.). Wallace and Romney (1975) give a detailed review of
past experience, methods, and problems in cleaning up contaminated areas.

From a statistical point of view, many of the problems in designing a sampling plan
for measuring pattern remain unresolved or uncertain. The main technique for displaying
the results of sampling for pattern is a contour map. Such a map is prepared by computer
programs that interpolate between average concentrations assigned to points on a uniform
grid. The chief problems to be resolved are those associated with the weighting processes
used to transform the observed data into grid entries. At first glance, these problems may
seem to be minor — fine points of statistical technique. We have now compared enough
contours generated by various methods, however, to be able to demonstrate that the
differences are not minor ones; see, for example. Fig. 3 of the report by Gilbert et al.
(1976b, p. 457) and the series of graphs by Gilbert, Eberhardt, and Smith (1976). Some
technical aspects of the problems have been discussed by Gilbert (1976).

Mapping the concentrations is only the first stage in the cleanup problem. If
contaminated soils are removed, then it is usually necessary to make certain that the job
is done adequately; i.e.. Are there areas of unacceptably high concentration remaining?
Gilbert and Eberhardt (1977) have described one possible sampling scheme for this
purpose (acceptance sampling). Since cleanup operations are expensive and usually
damaging to the environment, this matter needs further study.

Analytical Sampling

The statistical technology of analytical sampling is closely related to traditional methods
of statisfical analysis. An alternate designation that we find useful is "sampling for
comparisons." The area is a broad one and includes changes in sp^ce and time, analysis of
spatial patterns, etc. Some examples of statistical analysis that involve sampling follow.

One of the more difficult features in analyzing data is the problem of dealing with
ratios of variable quantities. This well-known statistical problem is accentuated by the
very high variability associated with the transuranic elements. A common example is the
so-called "concentration factor," which is really a ratio. [The plant panel at the 1975
transuranic workshop in Seattle defined a "concentration ratio" (CR) and an "inventory
rafio" (IR), thus supplying accurate names to replace the former "concentration factor"
(Energy Research and Development Administration, 1976)]. The difficulties in dealing
with ratios have been addressed in a number of the reports cited previously, and a recent
evaluation is that of Doctor and Gilbert (1977). As these authors note, there are three
well-known ways ratios of variable quantities can be estimated: (1) by averaging ratios of
individual pairs of observafions, (2) by summing up the x and y observafions and
calculating a ratio of totals, and (3) by a calculation of the type used to obtain the slope
estimate in regression analysis (except that the "corrections for the means" are dropped
so that the slope is appropriate to a regression calculated through the origin, i.e., the
intercept is zero). They also describe two other possibilities and note that the several
methods can give quite different results. Thus there is not only the problem of which
method of estimation to use but also the issue of how to allocate sampling effort so as to
make comparisons (of ratios) that are as meaningful as possible.

Another problem in statistical analysis is associated with interlaboratory comparisons.
We have pointed out something of the uncertainties involved (Eberhardt and Gilbert,

STATISTICS AND SAMPLING IN TRANS URANIC STUDIES 181

1972) and have continued to study the issues (Gilbert and Eberhardt, 1976b). The main
difficuky again has to do with variability since rather large numbers of replicate
determinations are required to give reasonable assurance of detecting differences between
laboratories. Proper allocation of samples (replicates) to methods, elements, and
laboratories should help, but this has not been investigated in any detail.

In a statistical analysis of transuranic data, frequency distributions are often
dramatically skewed (asymmetrical). Thus consideration should be given to transforming
the data before analysis. We have looked into this option in some detail (Eberhardt and
Gilbert. 1973: Eberhardt et al., 1976). Our recommendation is to use a logarithmic
transformation of the data before doing any statistical analyses involving significance tests
based on the normal distribution. A feasible alternative is to consider nonparametric (or
distribution-free) methods. We have begun some limited investigations in this area. We
particulariy do not recommend estimating means (averages) by transforming back the
mean of log-transformed data (Link and Koch, 1975). If interest is directed chiefly to
estimating means on the original (untransfomied) scale, we recommend use of the
ordinary aritlimetic average of the untransformed data. If interest is chiefly in a statistical
analysis, then the resuUs should be discussed in terms of the transformed data. The
problem of how to allocate samples to accommodate both purposes, however, seems to us
to need more attention.

Sampling for Modeling

Many sampling and statistical problems must be dealt with if modeling is eventually to
achieve truly satisfactory status in environmental studies. Most of the present prospects
for models contain a substantial number of rate functions that are little more than
guesses. Rather than go into these problems, which transcend statistical and sampling
issues, we will only mention some simple models. A few details of methods for finding
optimum sampling times for a rather simple, but widely applicable, model are given by
Eberhardt (1978).

Three categories of simple models can alternatively be described as profiles of
concentration in time or space: (1) retention of some substance by an animal,
(2) measuring concentrations away from a point source, and (3) studying soil profiles.
Two facets of such studies may need to be considered in designing a study by sampling.
One is whether the investigator's main interest is in estimating rate constants or in
describing the profile itself since different sampling plans are then appropriate. A second
concerns the nature of replications. In retention studies individual animals can serve as
replicates, but. in the evaluation of soil profiles, the word "replicate" will not have the
same meaning; so sampling results may have rather different interpretations in the two
instances. Eberhardt (1978) gives some further details on sampling profiles. Essington
et al. ( 1976) give a number of details on actual soil profiles of several transuranics.

The notion oi sampling for modeling, which can also be described as sampling for
curve fitting (in a more restricfive sense), appears to be new in environmental studies. As
such it poses a number of problems that need further evaluation. The reader interested in
technical details might well start with the review by Cochran (1973). which provides
addifional references. Papers by Atkinson and Hunter ( 1968). Box and Lucas (1959), and
Box (1968: 1970: 1971) should also be consulted. As has already been noted, our
attention has been focused on finding the optimum times (or depths, or distances) for
sampling in the interest of obtaining a maximum amount of information for a given
samphng cost.

182 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Among the difficulties that need further review, we would like to mention the
following:

1. A model has to be assumed; so it is usually desirable, or necessary, to consider
several candidate models. This then brings in a need to try to decide from the data which
is "the" correct model and hence statistical analysis and selection of the best sampling
scheme for discriminating between models.

2. When several parameters are involved, the optimization process usually requires a
computer to perform the calculations.

3. In most cases advance estimates of the parameters are required. This may be seen as
a rather severe restriction, but, since most of the important models to be considered are
nonlinear, experimentation is not very practical without a fair amount of advance
knowledge of the system anyhow.

Some Common Problems

A number of problems common to the several kinds of sampling described here and to
studies of transuranics and other trace substances nee<l further consideration. Those
associated with "counting statistics" and "counting errors" have been the subject of
many investigations but continue to pose difficulties. Gilbert (1975) has prepared a
review relative to counting statistics in studies of the transuranics. One common problem
with low-level measurements has to do with the "below background," "not detectable,"
or "trace" measurements. An approach that we believe deserves further study is the use
of simple nonparametric methods. When a data set contains a number of below-
background measurements, simple averages are likely to be biased, the direction of the
bias depending on how the questionable measurements are handled. In such cases it seems
that the median may be a preferable measure of central tendency and that assessing
variability may be approached through order statistics. As an example, suppose one takes
30 samples and, after analysis for a transuranic, find that 10 results are reported as "not
detectable." If a mean is to be computed, one must then decide how to incorporate these
10 measurements. Should they all be assumed to contain none of the substance under
study and be assigned "zero" values? Usually this is not reasonable since a longer
counting time, larger sample mass, etc., would likely have turned up detectable levels in
some of the 10 samples. The procedure suggested here is simply not to use the mean but
to change over to the median, which does not require any decisions about the
troublesome 10 "not-detectable" samples (except that their levels were, in fact, less than
those for which levels were reported). Rather than calculating standard deviations, one
would need to calculate measures of variability based on order statistics (Conover, 1971;
Hollander and Wolfe, 1973). Random sampling is, however, also required for nonpara-
metric methods.

Another common problem is how to deal with the practice of taking aliquots
(subsamples). A related issue is the compositing of a group of samples. Both are related to
the analytical (chemical) process. In the former instance sample mass is reduced to
facilitate analysis; in the latter, a number of samples are combined in an effort to reduce
costs. The very high sample-to-sample variability characteristic of the transuranic
elements may cause trouble for the unwary investigator in either case. In both cases we
believe repHcate samples should be taken to provide a measure of the variability inherent
in the process. Of course, doing this increases costs. Compositing samples needs some
careful study to see whether it does, in fact, offer the gains that are usually assumed
without any checking. A major question about compositing soil samples is whether or not

STATISTICS AND SAMPLING IN TRANS URANIC STUDIES 183

they can be adequately mixed. The basic idea is that the composite sample will provide an
accurate average value for the individual samples used to make up the composite. If the
entire composite is used for analysis, there should be no problem with the concept. For
many transuranic analyses, however, only a relatively small mass is used; so a composite
itself may be subsampled (aliquoted) at the chemical-analysis stage. Whether compositing
is worthwhile, then, depends on how weU the sample is, or can be, mixed. The
hot-particle problem in plutonium analyses, along with the common practice of using a
small sample mass for analysis, suggests that compositing may not be very effective. Some
further details and suggestions have been reported by Eberhardt (1976) and Eberhardt
et al. (1976). A statistical evaluation of compositing (Rohde, 1976) suggests that the
correct measure of variability for the average of composited samples is difficult to obtain.

Transformation of skewed data has already been mentioned; it deserves some further
research. One of the difficulties has to do with transforming back; i.e., if a logarithmic
transformation is used for statistical analysis, many investigators prefer to express the
final results on an arithmetic scale. Simply using antilogarithms introduces a bias. In
many instances it may be quite feasible to simply stay on the logarithmic scale; the
consequences of doing so need to be further evaluated and explained (Agterberg, 1974,
pp. 289-300; Aitchison and Brown, 1966, pp. 44-48; Helen, 1968). An interesting
sidelight is that some investigators seem to believe that correlations calculated on
log- transformed data are not legitimate. Whether this is true or not depends on the
statistical model assumed; so that issue has to be resolved by specifying such a model.
There is, in fact, some reason to argue that correlations involving transuranics should be
done on log-transformed data. The usual model for correlation (bivariate normal
distribution) involves a linear model with normally distributed errors (deviations). As has
been pointed out several times, data on transuranics are generally not normally
distributed, and relationships between different variables may be nonlinear.

A matter of substantial importance is the choice between random and systematic
sampling. We have thus far largely advocated random sampling since it is the only
approach generally accepted as providing unbiased estimates of population parameters. As
Cochran (1963, Chap. 8) shows, systematic samples are vulnerable to unsuspected
periodicities in the variable being studied, and no widely trustworthy method for
estimating the variance is known. On the other hand, if adjacent samples have closely
correlated values, then under random sampling two sample points that fall close together
essentially duplicate the same information. Hence a systematic sample gives more for the
money spent, i.e., a smaller sampling error (even though we may not get a suitable
estimate of the sampling error). This point has been particularly emphasized in references
on sampling for pattern. Some recent work by Barnes, Gilbert, and Delfiner (1977), using
the field instrument for the determination of low-energy radiation (FIDLER) data from a
safety-shot site on the Nevada Test Site, suggests that readings taken on a grid resuh in
more precise estimates of americium contours than do comparable (FIDLER) readings
taken at random within activity strata. These analyses were made with "kriging"
techniques (see Delfiner and Delhomme, 1975), a method we are currently studying for
potential applications in transuranic studies.

In the first few years that we worked with field sampling for plutonium. very little
data on variabihty were available (Eberhardt and Gilbert, 1972). As more and more data
collected by random samphng have become available, we have come around to the
opinion that it will be worthwhile to look more carefully into systematic sampling and to
begin tests of its utility in a variety of field situations. Since most of the applications will

184 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

be two dimensional (area sampling), a logical approach is grid sampling. If we then use
stratification, we will, in most instances, want to vary the mesh of the grid between
strata. This can be done easily if mesh sizes change geometrically; i.e., for areas of the
highest concentration, use a grid size of one unit, for the next lowest concentration
stratum, use a grid twice as large, for the next four times as large, and so on down to the
lowest concentration stratum, which will have the coarsest grid mesh. With such a scheme
the problem of matching grid meshes at stratum boundaries is fairly simple. Some of the
flexibility of stratified sampling, however, is thus given up, and this probably reduces its
efficiency in situations where concentration gradients are known to change rapidly.
Hence a variety of field trials will be needed to work out details of this and other
problems.

Many other problems might be considered here, but we conclude by nofing one that
may go largely unnoticed, i.e., the practice of sieving soil samples and doing chemical
analyses on only certain sieve fractions. Such a practice can introduce important biases,
some of which are described by Gilbert et al. (1976a). Examples using plutonium data
from safety tests are given in the report by Gilbert and Eberhardt (1976b, pp. 131-153).

Conclusions

We have attempted to briefly mention some of the many facets of statistical and sampling
methodology which are relevant to studies of the transuranic elements. We believe that
too many investigators working with these elements are not sufficiently aware of the very
large component of "chance'' error inherent in their data. Much of our experience has
been with data on plutonium in soils, mostly that resulting from "safety shots" at the
Nevada Test Site, but we have also studied data from a number of other sites (Enewetak,
Los Alamos, Rocky Flats, and other locations). In other circumstances variability may be
much reduced. However, if it is not, then it is likely that many "significant" findings to
date are largely artifacts resulting from inadequate sampling.

References

Agterberg, F. P., 1974, Geomathematics, American Elsevier Publishing Co., New York.

Aitchison, J., and J. A. C. Brown, 1966, The Lognormal Distribution, Cambridge University Press,

London.
Atkinson, A. C, and W. G. Hunter, 1968, The Design of Experiments for Parameter Estimation,

Techno metrics. 10(2): 271-289.
Barnes, M., R. O. Gilbert, and P. Delfiner, 1977, A New Statistical Tool, Kriging, and Some

Applications to Area 13 Data, in Transuranic s in Desert Ecosystems, M. G. White, P. B. Dunaway,

and D. L. Wireman (Eds.J, USDOE Report NVO-181, pp. 431-445, Nevada Operations Office,

NTIS.
Box, M. J., 1968, The Occurrence of Replications in Optimal Designs of Experiments to Estimate

Parameters in Nonlinear Models, J. R. Stat. Sac. Ser B, 30: 290-302.

, 1970, Some Experiences with a Nonlinear Design Criterion, Technometrics, 12(3): 569-589.

, 1971, An Experimental Design Criterion for Precise Estimation of a Subset of the Parameters in a

Nonlinear Model, Biometrika, 5 8: 149-15 3.

, and'H. L. Lucas, 1959, Design of Experiments in Nonlinear Situations, Biometrika. 46: 77-90.

Cochran, W. G., 1963, Sampling Techniques, 2nd ed., John Wiley & Sons, Inc., New York.

, 1973, Experiments for Nonlinear Functions, / Am. Stat. Assoc. 68(344): 771-781.

Conover, W. J., \91 \, Practical Non-Parametric Statistics. John Wiley & Sons, Inc., New York.
Delfiner, P., and J. P. Delhomme, 1975, Optimum Interpolation by Kriging, in Display and Analysis of

Spatial Data. pp. 96-114, J. D. Davis and M. J. McCullagh (Eds.), John Wiley & Sons, Inc., New

York.

STATISTICS AND SAMPLING IN TRANSURANIC STUDIES 185

Doctor, P. G., and R. O. Gilbert, 1977, Ratio Estimation Techniques in the Analysis of Environmental
Transuranic Data, in Transuranics in Natural Environments, Symposium Proceedings, Gatlinburg,
Tenn., Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), USDOE Report NVO-178, Nevada
Operations Office, NTIS.

Eberhardt, L. L., 1975a, Some Problems in Measuring Ecological Effects of Chronic Low-Level
Pollutants, in Hearings Before the Subcommittee on Environment and the Atmosphere of the
Committee on Science and Technology, U. S. House of Representatives, 94th Congress,
pp. 565-586, No. 49, GPO.

, 1975b, Some Methodology for Appraising Contaminants in Aquatic Systems,/ Fish Res. Board

Can., 32(10): 1852-1859.

, 1976, Sampling for Radionuclides and Other Trace Substances, in Radiological Problems

Associated with the Development of Energy Sources, Fourth National Radioecology Symposium,
C. E. Gushing (Ed.), Dowden, Hutchinson and Ross, Inc., Stroudsburg, Pa.

, 1977, Applied Systems Ecology: Models, Data and Statistical Methods, inNew Directions in the

Analysis of Ecological Systems, pp. 43-55, G. S. Innis (Ed.), Simulation Councils Proceedings
Series, 5(1), Simulations Councils, Inc.

, 1978, Designing Ecological Studies of Trace Substances, in Environmental Chemistry and Cycling

Processes, DOE Symposium Series, Augusta, Ga., Apr. 28-May 1, 1976, D. C. Adriano and I. Lehr

Brisbin, Jr. (Eds.), pp. 8-33, CONF-760429, NTIS.
, and R. O. Gilbert, 1912, Statistical Analysis of Soil Plutonium Studies, Nevada Test Site, USAEC

Report BNWL-B-217, Battelle, Pacific Northwest Laboratories, NTIS.
, and R. O. Gilbert, 1973, Gamma and Lognormal Distributions As Models in Studying

Food-Chain Kinetics, USAEC Report BNWL-1747, Battelle, Pacific Northwest Laboratories, NTIS.
, and R. O. Gilbert, 1976, Samphng the Environs for Contamination, in Proceedings of the First

ERDA Statistical Symposium, Los Alamos, November 1975, ERDA Report BNWL-1986,

pp. 187-196, Battelle, Pacific Northwest Laboratories, NTIS.
, R. O. Gilbert, H. L. HoUister, and J. M. Thomas, 1976, Sampling for Contaminants in Ecological

Sysiemi, Environ. Set and Tech., 10: 917-925.
Energy Research and Development Administration, 1976, Workshop on Environmental Research for

Transuranic Elements, Proceedings of the Workshop, Battelle Seattle Research Center, Seattle,

Wash., Nov. 12-14, 1975, USAEC Report ERDA-76/134, NTIS.

Essington, E. H., E. B. Fowler, R. O. Gilbert, and L. L. Eberhardt, 1976, Plutonium, Americium, and
Uranium Concentrations in Nevada Test Site Soil Profiles, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Francisco, 1975, pp. 157-173, STI/PUB/410, Inter-
national Atomic Energy Agency, Vienna.

Gilbert, R. O., 1975, Recommendations Concerning the Computation and Reporting of Counting
Statistics for the Nevada Applied Ecology Group, USAEC Report BNWL-B-368, Battelle, Pacific
Northwest Laboratories, NTIS.

, 1976, Estimation of Spatial Pattern for Environmental Contaminants, draft of paper presented at

1976 Annual Meeting of the American Statistical Association, Boston, Aug. 23-26, 1976, USAEC
Report BNWL-SA-5 7 71, Battelle, Pacific Northwest Laboratories, NTIS.

, and L. L. Eberhardt, 1976a, An Evaluation of Double Sampling for Estimating Plutonium

Inventory in Surface Soil, in Radiological Problems Associated with the Development of Energy
Sources, Fourth National Radioecology Symposium, pp. 157-163, C. E. Gushing (Ed.), Dowden,
Hutchinson and Ross, Inc., Stroudsburg, Pa.

, and L. L. Eberhardt, 1976b, Statistical Analysis of "A Site" Data and Inierlaboratory

Comparisons for the Nevada Applied Ecology Group, \n Studies of Environmental Plutonium and
Other Transuranics in Desert Ecosystems, M. G. White and P. B. Dunaway (Eds.), ERDA Report
NVO-159, pp. 117-154, Nevada Operations Office, NTIS.

, and L. L. Eberhardt, 1977, Some Design Aspects of Transuranium Field Studies, in Transuranics

in Natural Environments, Symposium Proceedings, Gatlinburg, Tenn., Oct. 5-7, 1976, M. G.
White and P. B. Dunaway (Eds.), ERDA Report NVO-178, Nevada Operations Office, NTIS.

, L. L. Eberhardt, and D. O. Smith, 1976, An Initial Synthesis of Area 13 ^^'Pu Data and Other

Statistical Analyses, USAEC Report BNWL-SA-5 667, Battelle, Pacific Northwest Laboratories,
NTIS.

186 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

, L. L. Eberhardt, E. B. Fowler, and E. H. Essington, 1976a, Statistical Design Aspects of Sampling

Soil for Plutonium, in Atmosphere-Surface Exchange of Particulate and Gaseous Pollutants,
ERDA Symposium Series, Richland, Wash., Sept. 4-6, 1974, Rudolf J. Engelmann and George A.
Sehmel (Coordinators), pp. 689-708, CONF-7 40921, NTIS.

, L. L. Eberhardt, E. B. Fowler, E. H. Essington, and E. M. Romney, 1976b, Statistical Analysis

and Design of Environmental Studies for Plutonium and Other Transuranics at NAEG
"Safety-Shot" Sites, in Transuranium Nuclides in the Environment, Symposium Proceedings, San
Francisco, 1975, pp. 449^60, STI/PUB/410, International Atomic Energy Agency, Vienna.

, L. L. Eberhardt, E. B. Fowler, E. M. Romney, E. H. Essington, and J. E. Kinnear, 1975,

Statistical Analysis of 239-2 4 op^ ^nd ^ *' Am Contamination of Soil and Vegetation on NAEG
Study Sites, in Radioecology of Plutonium and Other Transuranics in Desert Environments, M. G.
White and P. B. Dunaway (Eds.), USAEC Report NVO-15 3, pp.339448, Nevada Operations
Office, NTIS.

Helen, D. M., 1968, A Note on Log-Linear Regression,/ Amer. Stat. Assoc, 63: 1034-1038.

Hollander, M., and D. A. Wolt^e, 1973, Nonparametric Statistical Methods, John Wiley & Sons, Inc.,
New York.

Link, R. F., and G. S. Koch, Jr., 1975, Some Consequences of Applying Log-Normal Theory to
Pseudobgnormal Distributions, Afflr/2. Geology, 7: 117-128.

Rohde, C. A., 1976, Composite Sampling, Biometrics, 32: 273-282.

Wallace, A., and E. M. Romney, 1975, Feasibility and Alternate Procedures for Decontamination and
Post-Treatment Management of Pu-Contaminated Areas in Nevada, in Radioecology of Plutonium
and Other Transuranics in Desert Environments, M. G. White and P. B. Dunaway (Eds.), USAEC
Report NVO-153, pp. 251-337, Nevada Operations Office, NTIS.

Appropriate Use of Ratios in Environmental
Transuranic Element Studies

p. G. DOCTOR, R. O. GILBERT, and J. E. PINDER III

Tliis chapter discusses some statistical aspects of two types of ratios used extensively in
environmental transuranic studies. Vie two ratios discussed, concentrations and pure
ratios, have different uses and different statistical problems. A concentration gives units
of numerator ( Y) per unit of denominator (X), e.g. , nanocuries of^^ ^Pu per gram of soil.
Concentrations are viewed as raw data and are used as input for further statistical
analysis. For this type of ratio, Y is assumed to be proportional to X. For environmental
radionuclide concentrations, variability between aliquots for small aliquot sizes tends to
become large. Tlie choice of aliquot size permitting a reliable estimate of concentration is
a major problem with this type of ratio. For a pure ratio the numerator and denominator
are measured in the same units, e.g., nanocuries of ^^^Pu over nanocuries of ^^'^ Pu. In
transuranic field studies both the numerator and denominator may vary considerably
among aliquots in the same sample. Pure ratios often appear as a ratio of concentrations,
e.g., concentration ratios and inventory ratios. However, pure ratios provide accurate
information on the relationship between Y and X only when Y is proportional to X. The
statistical problems of pure ratios center on an assessment of whether the multiplicative
assumption is valid. Multivariate statistical techniques offer alternatives to a pure ratio for
expressing the relationship between Y aiid X. The purpose of this chapter is not to
provide a catalog of statistical methods for ratio estimates but to stimulate critical
thinking about the use of ratios and to suggest approaches to the task of ratio estimation
compatible with the behavior of environmental radionuclide data.

Ratios are used extensively in scientific work, particularly in the environmental and life
sciences, to express the relationship between two independently measured attributes of,
for example, the same animal, soil sample, plant part, or geographic locaUty. Examples in
the field of environmental transuranic element research include ■^•'^Pu activity /weight for
a soil sample, the ^^^Pu/-^'*' Am ratio in a vegetation sample, and the ratio of 1 3^-1 37 (--^
activity in plant tissue to that in soil at a particular location.

A ratio is one of the simplest mathematical teclmiques for relating two numbers.
Another approach is to compute their difference. However, both techniques have precise
mathematical assumptions underlying their use. The use of a ratio implies that the
relationship is multipUcative; that is, if Y is the numerator and Xthe denominator of the
ratio, then

Y = 7X (1)

where 7 is the proportionality constant. The use of a difference implies that the
relationship is additive; i.e.,

Y = a + X

187

188 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

where a is the additive constant. Using ratios or differences without consideration for
these mathematical assumptions can produce misleading results.

Since the world does not behave exactly according to mathematical principles, the
decision of whether the multiplicative assumption holds is a substantive as well as a
statistical question. If the data do not support the notion that the relationship is
multiplicative, then other methods should be found to relate the two variables. More
sophisticated mathematical and statistical techniques for relating two variables include
linear and nonlinear regression, correlation and cluster analysis, and multivariate
regression and analysis of variance.

There are two main types of ratios. The first, called a dimensioned ratio by Simpson,
Roe, and Lewonton (1960, p. 13), expresses the amount of one variable per unit of a
second variable. The second variable is usually weight or volume. Examples are ^"^^Am
activity per gram of soil and ^^^Pu activity per Uter. Concentrations are so fundamental
to environmental radionucUde research that they are Uterally viewed as raw data.
However, the calculation of a concentration is a preprocessing step whose purpose is
scaling; i.e., the resulting value should be independent of the size of the sample on which
it was measured. The tacit assumption underlying this approach to scaling is the
multiplicative one that plutonium activity in a sample is proportional to the size of the
sample. Although this seems to be a reasonable assumption, it is not always true. Failure
to meet this assumption can produce severe problems when low-level concentrations
(picocurie and femtocurie range) and small amounts of sample material are present.

The second type of ratio is called dimensionless by Simpson, Roe, and Lewonton
(1960, p. 13). This ratio is unitless because the units of the numerator and denominator
are the same; so they algebraically cancel each other. Within tlie class of dimensionless
ratios, there are two subtypes. One is a percent or fraction, e.g., the amount of ^^^Pu
expressed as a fraction of plutonium in a soil sample. This ratio is special in that its values
are restricted to 0 to 100 for percents or 0.0 to 1.0 for fractions. Chayes (1971), in the
context of petrology, discusses the properties of and methods for dealing with
percentages. The second subtype of dimensionless ratio we call a pure ratio. The
denominator does not include the numerator; so the possible values of the ratio are
unconstrained. An example is the isotopic ratio, ^^^Pu/^^^Pu, in which the numerator
and denominator are measured on the same sample. Of the two types of dimensionless
ratios, pure ratios seem to be more prominently used in environmental radionuclide
research.

The extent to which ratios are considered essential to environmental radionuclide
research is evidenced by the compounding of ratios, i.e., a ratio of ratios. Two examples
are found on pp. 23-24 of the proceedings of the November 1975 Workshop on
Environmental Research for Transuranic Elements (U. S. Energy Research and Develop-
ment Administration, 1976), the concentration ratio (CR), defined as

Activity per weight of plant part
Activity per weight of substrate or reference material

and the inventory ratio (IR), defined as

Activity per unit area in product
Activity per unit area in source

RATIOS IN TRANSURANIC ELEMENT STUDIES 189

Tliese compound ratios are pure ratios whose numerator and denominator are
concentrations. The farther one is removed from the raw data by the compounding of
ratios, the harder it is to justify theoretically and statistically the multiplicative
assumption. Therefore, Simpson, Roe, and Lewonton (1960, p. 18) conclude that the
compounding of ratios should be done with great care.

The calculation of a ratio is simple; however, the ramifications as they affect the
statistical analyses are often complex (Sokal and Rolilf, 1969, pp. 17-19). First, ratios
magnify the inaccuracies of the component variables. For example, consider the average
ratio 1.0/2.0. Suppose the true measurements lie between 0.95 and 1.10 and between
1.90 and 2.10 for the numerator and denominator, respectively. There is a maximum
relative error of 10%= [(1.10 - 1 .00)/1.00] x 100 for the numerator. However, the
range for the ratio lies between 0.45 = 0.95/2.10 and 0.58 = 1.10/1 .90, giving a maximum
error of 16% = [(0.58 - 0.50)/0.50] x 100 for the ratio. Moreover, the midpoint of the
range of the ratio (in this case 0.52) is not the best estimate of the ratio.

Second, the frequency distribution of a ratio can be skewed or multimodal. This is
particularly true if either the numerator or denominator is a discrete random variable, i.e.,

it can take on only a small number of possible values (Simpson, Roe, and Lewonton,
1960, pp. 15-16). An example is a low-level concentration where the number of counts is
near zero. Multiplying by conversion factors and dividing by sample weight may produce
numbers that appear to represent a continuum, but the number of values the ratio can
take on is still small.

Third, taking ratios of two random variables does not preserve either of their
distributions. For example, the ratio of two normal random variables is not a normal
variable. This can present serious problems since most statistical methods require that the
data be at least approximately normally distributed. The underlying probability
distribution of the ratio (except in a few well-known situations) (Mielke and Flueck,
1976) cannot be inferred from the distributions of the two component variables.
However, one useful exception is that the ratio of two log-normal variables is
log-normally distributed. Finally, the ratio provides Uttle information on the relationship
between the component variables unless that relationship is multiplicative.

The problems of simple ratios (those composed of variables that are directly
measured) are magnified when the component variables are themselves ratios, for
example, CR's and IR's. Moreover, the generally unknown distributional properties of
ratios make their uncritical use as input for further statistical procedures problematic.
Chayes (1971) and Atchley, Gaskins, and Anderson (1976) discuss the behavior of ratios
of percents and correlated normal variables, respectively, when they are used as raw data
for some statistical procedures.

This chapter discusses some numerical and statistical problems encountered in using
concentrations and pure ratios in environmental radionuclide research. Its purpose is not
to provide a catalog of statistical methods for ratio estimates but to stimulate critical
thinking about the use of ratios and to suggest approaches to the task of ratio estimation
compatible with the behavior of environmental radionuclide data.

Concentrations

Recall that the purpose of the concentration is to eliminate the effect of the denominator
(aliquot size) on the numerator (radionucHde activity). This implies, in theory, that the
concentration can be represented as y

190 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

which is obtained from Eq. 1 by dividing both sides by X. The graphic representation is
given in Fig. 1, where the concentration is the same regardless of aliquot size.

The fundamental assumption that underlies the mathematical assumption of a
multiplicative relationship between activity and aliquot size is that the radionuclide
activity is homogeneously dispersed throughout the medium on which it is measured.
This is the justification for the aliquoting procedure in which, for example, a 1-g aliquot
is taken for analysis from a 100-g soil sample and the observed aliquot concentration is
ascribed to the entire sample.

o
(J

X (ALIQUOT SIZE)
Fig. 1 Theoretical relationship between concentration and aliquot size.

The classical statistical approach to estimating a ratio from a supposedly homoge-
neous set of data is linear regression. (See Snedecor and Cochran, 1967, pp. 167-171 , for
a complete discussion.) The model is

Y = 7X + e

(2)

where Y, X, and 7 are as in Eq. 1 and e is the deviation of the data from the fitted model.
The estimate of 7 is the ratio estimate.

An assumption underlying this approach is that Y is a random variable and X is
known without error. For a radionuclide concentration, tliis is a reasonable assumption.
The denominator is usually a weight or volume and is considered to vary through
measurement error only. This error is usually negligible compared with the sampling and
measurement errors associated with radionucHde activity. For that reason all the
variability in the ratio is attributed to the numerator.

Another statistical assumption underlying the usual unweighted least-squares fit of
Eq. 2 is that the variance of Y is the same at each value of X; i.e.,

- ^2

a'(Y|X)=a

Both Y and X are assumed to be positive, and, from Eq. 2, the fitted line is forced to pass
through the origin (Fig. 2). If both Xand Y are assumed to be positive, the consequence.

RA TIOS IN TRANSURANIC ELEMENT STUDIES 1 9 1

X (ALIQUOT SIZE)

Fig. 2 Theoretical relationship between activity and aliquot size. (Adapted from Doctor
and Gilbert, 1979.)

assuming a symmetric distribution for Y about the fitted line, is that the variance of Y
must go to zero as X goes to zero. This is usually stated: the variance of Y at a particular
X is some function of X; i.e.,

a2(Y|X) = f(X)

Snedecor and Cochran (1967, pp. 166-171) give methods for obtaining the weighted
least -squares estimates of 7 when f(X) = kX and kX' , where k is a constant. Doctor and
Gilbert (1^77) compared the behavior of these ratio estimates along with the sample
median of the ratios and the log-normal estimate of the median ratio for three sets of
transuranic data from Nevada Applied Ecology Group studies.

This classic approach to ratio estimation is often not applicable for estimating the
true mean concentration horn a set of environmental radionuclide data for two reasons.
First, as the amount of sample material decreases, the variability in observed radionuclide
activity tends to increase rather than decrease. Second, in the case in which the sample
size is under the researcher's control, e.g., soil samples, the usual laboratory practice is to
analyze only one size aliquot; so regression analysis is impossible. The variability problem
of radionuclide concentrations is discussed in the context of two examples: first, ^^^Pu
soil concentrations in a desert environment as the result of a nuclear test and, second,
1 34-1 3 7(^g vegetation concentrations taken from a stream bed receiving reactor effluents.

Soil Concentrations

The data consist of twenty ^'*' Am concentrations from each of five aliquot sizes (1, 10,
25, 50, and 100 g) taken from the same composite soil sample collected near nuclear site
201 at the Nevada Test Site. The aliquoting procedure (discussed by Doctor and Gilbert,
1979) was designed to ensure as homogeneous a dispersal of the americium as possible.
The concentrations are plotted in Fig. 3, where the solid lines delineate the range of the
data. The variability tends to increase as aliquot size decreases; the variability of the 1-g

192 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Aliquot size, g 1 10 25 50 100

N 20 20 20 20 20

Median 1.56 1.71 1.56 1.82 1.91

Mean 1.93 1.82 1.80 1.84 1.92
Standard deviation

of mean 0.34 0.11 0.12 0.05 0.04

Fig. 3 Americium concentrations in soil aliquots of different sizes from nuclear site
201, Nevada Test Site,

aliquots is much larger (standard error, 0.34) than that of the 10-g aUquots (standard
error, 0.11). The observed particulate nature of environmental transuranic element
concentrations suggests a hypothesis for tliis phenomenon. Assume that the americium
occurs as particles that are randomly dispersed throughout the soil. For a large (100-g)
aliquot size in which the probability of sampling a particle is high, the observed
concentrations will probably be reasonably consistent. However, as the aliquot size
decreases, the probability of getting a particle also decreases. This tends to violate the
homogeneous dispersal assumption. Tlie result is that the observed concentrations will
tend to show more variability, as evidenced in the extreme case by the 1-g aliquots.

The data show a positive skewness to high values which is most pronounced for the
smaller aliquot sizes. This behavior is consistent with the hypothesis of the previous
paragraph, in which the assumptions are similar to the axioms that generate the Poisson

RATIOS IN TRANSURANIC ELEMENT STUDIES 193

process (Parzen, 1967. p. 1 18). The Poisson distribution is a distribution of the number

of randomly dispersed particles in a given volume. The appropriateness of this approach

for describing the variability in observed radionuclide concentration as a function of

aliquot size depends on the relationship between radionuclide concentration and the

number of particles in a sample. We feel that the Poisson distribution as a model of

radionuclide concentration witliin a sample deserves study. This skewness to large values

is also characteristic of environmental radionuclide concentrations taken from samples at

different locations; so perhaps the Poisson distribution would find use in this situation as

well.

Although it is unusual to analyze 20 aliquots per sample, it is instructive to compare

the effects of this variability on the observed average concentration for each aliquot size.

The sample median (middle value of the observed concentrations) is denoted by an

asterisk in Fig. 3. Contrary to what might be expected, the medians show more variation

with aliquot size than the sample arithmetic means (connected by the broken line).

Skewness affects the median more than the mean, as evidenced by the 50- and 100-g data

for which skewness is least pronounced, and the mean and median are very close [see

Michels (1977) for theoretical justification] . Since aliquots of all sizes are all from the

same composite sample, theoretically the means should remain constant across aUquot

size. This appears to be the case here.

It should be emphasized that this example illustrates within-sample variability and not

variability due to different locations. However, in an environmental radionuclide study,

one is faced with between-location variability as well. One value of this study is that it

provides information on within-sample variabiUty which, in turn, permits an evaluation of

the amount of observed variability due to location differences. Similar studies might

precede a full-scale sampling effort that encompasses a new radionuclide, a new source, or

a new medium. Such studies provide a rationale for choosing an aliquot size that will

minimize within-sample variability under the constraints of laboratory capability and cost

(Doctor and Gilbert, 1979).

Vegetation Concentrations

The problems of obtaining reliable vegetation concentrations in the picocurie range are
illustrated by data on plant uptake of i34-i 3 7q collected at the Savannah River Plant
near Aiken, S. C. [See Sharitz et al. (1975) for a description of the site.] Here the sample
or ahquot size, unlike that for soil concentrations, cannot be controlled by the researcher.
The data consist of fifty-five ' ^^-i 3 v^^^ concentrations and sample weights of leaves from
Hypericum walteri growing on the floodplain of a South Carolina stream receiving reactor
effluents. Since some sampling designs may require that the sample be collected from a
species at a particular location without regard to the size of the individual, the same type
of data as that illustrated in Fig. 4 may result. Except for two high values at 0.4 g, the
variability is dramatically increased for samples weighing <0.2 g. It appears that for these
samples the assumption of uniform dispersal is seriously violated, which shows that
observed concentration is not independent of sample size (compare with Fig. 1). The
errors in measuring the cesium are large relative to the weight of the sample.
Furthermore, if negative readings occur and are either disregarded or reanalyzed until
positive values are obtained, a positive bias will be introduced, and this bias will be greater
for the smaller samples. Even if the accuracy of determining radionuclide content is
controlled, variation due to small sample weights may still be a problem if the range of
sample weights is large.

194 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

150 —

o
a.

z
o
o

100 —

Y/X = 50.1-6.2 X

• * *

Y/X = 74.7-52.8 X

0.2 0.4 0.6 0.8

SAMPLE WEIGHT, g

1.0

1.2

Fig. 4 Relationship between ' ^'' ' ^ ''Cs concentration and sample weight in Hypericum
walteri from a floodplain receiving reactor effluents.

In this case we do not have the luxury of taking a larger size sample to reduce
variability as we do for soil samples. Assuming that the concentration is constant, what is
the best way to estimate it? The problem can be viewed as twofold: to determine (I) the
minimum amount of sample required to yield a consistent estimate of the true
concentration and (2) the method for combining the data to give a best estimate of the
concentration.

A method of estimating the minimum sample weight required for an analysis
consistent with the homogeneous dispersal assumption is illustrated in Figs. 5 and 6. The
graph in Fig. 5 was obtained by arranging the samples in random order and computing the
variance for an increasing number of samples starting at 10. The initially higlily
fluctuating variance that then decreased to a final value of about 1950 is typical of the
plots obtained when this procedure is applied to positively skewed and leptokurtotic
(sharp peak) data. A similar procedure was used to generate the graph in Fig. 6 except
that the samples were arranged in order of decreasing weiglit. The form of the initial
phase of Fig. 6 (sharp increase followed by a steady decrease) is similar to that of Fig. 5,
and the variance appears to stabiUze at approximately 750 for samples weigliing >0.2 g;
however, when samples weigliing <0.2 g are added, the variance then increases
continually until the same tinal value as that in Fig. 5 is obtained. This suggests that

RATIOS IN TRANSURANIC ELEMENT STUDIES 195

3000

20 30 40

NUMBER OF SAMPLES

Fig. 5 Variance of '^""'^''Cs concentrations in Hypericum walteri as a function of
number of samples randomly ordered by weight.

2500

0.4 0,3

SAMPLE WEIGHT, g
0.2 0.1

0.05

20 30 40

NUMBER OF SAMPLES

Fig, 6 Variance of ' ^'* ' ^''Cs concentrations in Hypericum walteri as a function of the
number of samples ordered by decreasing weight.

196 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

concentrations from samples weighing <0.2 g provide rather inaccurate estimates of the
true concentration. Other sets of data may not present as distinct a choice for the
minimum weight, but this way of looking at the data can provide insights into how the
behavior of the data vis-a-vis the assumption of uniform dispersal affects the average
concentration.

We now discuss the problem of estimating the "true" concentration from this type of
data. Line 1 in Fig. 4 results when a straight Une (Y/X = a + j3X + e) is fit to all 55 data
points. Because of the large variability in concentration for the samples weighing <0.2 g,
the line has a statistically significant (at the 0.05 level) negative slope (j3=-52.8;
Fi ,5 3* = 4.02). This suggests that the mean of all 55 observations, 60.1 pCi/g with a
standard deviation of 44.2 pCi/g, is not a good estimate of the true concentration for this
set of data. However, when samples weighing <0.2 g are excluded, leaving 28 data points,
and the line is refit (hne 2), the slope, althouglr still negative, is not significantly different
from zero (j3 = —6.2; Fi ^2 6 - 0.06). This suggests that the mean of the 28 observations,
47.2 pCi/g, with a standard deviation of 27.1 pCi/g, is a more reasonable estimate of the
true concentration.

These two examples illustrate the fact that there is no one best method for estimating
a true concentration for a set of data. The importance of investigating the relationship
between concentration and aliquot size before assuming a constant concentration cannot
be overstated. Since the purpose of the concentration is to produce a value independent
of the size of the sample on which it is measured, disregarding that relationship can
produce misleading results in further statistical analyses.

Pure Ratios

Recall that for pure ratios the numerator and denominator are measured in the same
units, e.g., ^^^Pu/^^^Pu both in nanocuries, and ^^^Pu concentrations in vegetation and
soil measured in picocuries per gram. In this case both numerator and denominator are
random variables as compared with a concentration in which only the numerator is a
random variable. This distinction complicates the statistical treatment of this type of
data, but the basic assumption underlying the ratio in this situation, as for concentra-
tions, is still proportionality. If the assumption of proportionality cannot be supported
either theoretically or statistically, then other methods of relating the variables should be
found. In this section we discuss some statistical problems associated with pure ratios
encountered in environmental radionuclide research and illustrate the use of multivariate
techniques as a substitute for and a means of testing the validity of a ratio.

First, we discuss a situation particular to radionuclide research. In contradiction to a
statement in the introduction to this chapter that ratios tend to be more variable than the
component variables, there is a situation where the ratio is the stable variable. An
example is the ratio of ^^^Pu to ^"^^ Am observed at a safety -shot site on the Tonopah
Test Range cited in Doctor and GUbert (1977) (see Fig. 7). The "^Pu and ^^^Am
values were individually quite variable, but their ratio was essentially constant.

Sokal and Rohlf (1969, p, 17) suggest that a pure ratio should be used to explain the
relationship between two variables only if there is evidence that the process under study
is a function of (or operates on) the ratio of the two variables and not of the variables

*Observed value of the F statistic with 1 and 53 degrees of freedom (Snedecor and Cochran, 1967,
pp. 259-260).

RATIOS IN TRANSURANIC ELEMENT STUDIES 197

1 1

•

• -

•

—

b
^ 8

E
<

^ 6

•

•
•

0 0.08 0.16 0.24 0.32 0.40 0.48 0.56

239pu, nCi/g

Fig. 7 Relationship between ^ ^ ' Pu and ^ "• ' Am in soil at a safety shot site on the
Tonopah Test Range. (From Doctor and Gilbert, 1977.)

individually. An example is provided by plant uptake of technetium, which appears to be
related more to the ratio of pertechnetate to sulfate in the soil than to pertechnetate soil
concentration (Cataldo, Wildung, and Garland, 1978). It would seem that this criterion
would be met if the relationship between numerator and denominator is multiplicative.
Two more examples include the plutonium/americium ratio just mentioned and the
concentration ratio between two components of an ecosystem compartment model with
a constant transfer coefficient. In these situations the reason for using ratios far
outweighs their potential disadvantages. However, in some cases there may be more
informative and less misleading ways to relate the two variables than by the use of a ratio.
Whether or not one can assume that the relationship between two random variables is
multiplicative, the approach to these data should be a multivariate one. A first step is to
plot the numerator vs. the denominator as in Fig. 7. The Pearson product moment
correlation coefficient (Snedecor and Cochran, 1967, p. 172),

'■[A

.S (Xi - X)(Yi _ Y)
1=1

(Xi - X)^ .2 (Yi _ Y)^
1=1

measures the degree of linear association between two normally distributed random
variables. Although we cannot assume that radionuclide activity is normally distributed,
the correlation coefficient is still a useful piece of information. If the correlation is low,
the ratio will be highly variable. The correlation coefficient provides a measure of linear
association but not whether the relationship is multiplicative. That information can be
gained from a regression analysis.

198 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Since both numerator (Y) and denominator (X) are variables, the unconstrained
regression of X on Y

X = a' + i3'Y + e'

(Y assumed not variable) is as valid statistically as the regression of Y on X

Y = a + (5X + e (3)

where a', j3', and e' are different from a, |3, and e (Fig. 8). The two regression lines will
never coincide unless there is a perfect multiplicative functional relationship (Y = j3X)
between Y and X. When we constrain one of the regressions, say Y on X, to go through
the origin, Y = |3X + e, the other regression (X on Y) will not go through the origin unless

X = a' + ^'Y / ^

]

y = Q + 0X

/ /

Fig. 8 Relationship between linear regression of Y on X and that of X on Y.

X is a function (without error) of Y (Snedecor and Cochran, 1967, pp. 172-181). If the
multipUcative relationship seems valid, there are methods of taking a compromise slope
(Ricker, 1973), If both X and Y are normally distributed, then an estimator of the slope
(constant ratio), Y/X, has some nice statistical properties (Ricker, 1973; Creasy, 1956;
Cocliran, 1977, Chap. 6; Doctor and Gilbert, 1977). The sample median of the observed
ratios is another useful estimator of the constant ratio because it is not greatly affected
by extremely high or low values and because no assumption about the distribution of X
and Y or their ratio (Doctor and Gilbert, 1977) is required.

Suppose that the use of a ratio is justified, e.g., the isotopic ratio ^^^Pu/^'" Am for
the Tonopah Test Range data mentioned above (Fig. 7). Tliis is often the first step in a
series of statistical analyses; so several caveats should be mentioned. Making inferences
about ratios is risky because we are usually forced to make distributional assumptions
that ratios rarely fulfill. For example, a test of hypothesis regarding the difference
between two samples of ratios may assume that the ratio data are normally distributed.
Nonparametric techniques, such as the Wilcoxon rank sum test (Hollander and Wolfe,

RATIOS IN TRANSURANIC ELEMENT STUDIES 199

1973, pp. 67-74), may alleviate that problem to a certain extent. Ratios are often used as
the raw data for standard statistical methods, such as regression and analysis of variance.
The appropriateness of tliis practice is determined by the behavior of the ratios vis-a-vis
the assumptions underlying these methods.

A more general approach to a situation where a pure ratio can be used is a
multivariate one. This approach is appropriate whether or not the multiplicative
assumption is valid, and some multivariate techniques can be used as a check on that
assumption. The previously mentioned Pearson product moment correlation coefficient,
r, is a multivariate technique. The multivariate approach allows one to observe the
behavior of the two random variables simultaneously. A multivariate (bivariate if the
number of variables is two) variable can be most easily explained by an example. Let Y
and X be, respectively, the ^^^Pu and ^^ ' Am activity at various depths in a soil profile.
Both isotopes together can provide a more complete picture of the process of leaching of
radionuclides in soil than either can provide separately. The joint distribution of ^^^Pu
and '^^ ' Am activity at a particular profile depth may look like the two-dimensional curve
in Fig. 9. The points (x and y) under the highest part of the curve correspond to the

^^^^Am, nCi

Fig. 9 Hypothetical joint distribution of ^ ■" Am and ^ ^ * Pu.

combinations of ^^^Pu and ^''^Am activity most likely to occur. The relationship
between the two variables is completely specified by the joint distribution. Everything
that can be done statistically for single random variables — e.g., testing for differences
between groups, regression, and analysis of variance — can be done for multivariate
random variables (Morrison, 1967).

With this brief introduction to the rationale underlying multivariate statistical
techniques, we discuss two such techniques on two sets of environmental radionuclide
data. The first technique is multivariate regression and relates soil and vegetation
concentrations to distance from a point source of contamination. The second technique,

200 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

profile analysis, tests the hypothesis of different resuspension parameters for two
ecosystems. Goth techniques are based on the multivariate normal distribution. Despite
the caveats mentioned earlier regarding the use of the normal distribution for analyzing
environmental data, these techniques can still provide valuable insight into the behavior
of the data with respect to the multiplicative assumption. Also, transforming the data by
logarithms, as we have done here, tends to reduce variability and make the normality
assumption more tractable.

Multivariate Regression

The data consist of ^^^U vegetation and associated soil concentrations taken at various
distances from ground zero (GZ) at a safety -shot site (A site in Area 11) on the Nevada
Test Site (Gilbert and Eberhardt, 1976). The problem is to relate the soil and vegetation
concentrations jointly as a function of distance from GZ. There are at least two possible
approaches. We compare a univariate method that regresses the CR (vegetation/soil) on
distance with a multivariate one that regresses soil and vegetation concentrations jointly
on distance. A univariate regression with the CR as the dependent variable assumes that
the relationsliip between vegetation and soil concentrations is multiplicative. Multivariate
regression is not so constrained, and, because there are fewer assumptions on the
relationship of Y and X, it can be used to assess whether the relationship is multiphcative.
The data shown in Fig. 10 consist of 14 pairs of soil and vegetation concentrations of
^^^U taken from random locations within 300 ft of GZ. Figure 11 shows the ratio of
vegetation to soil as a function of distance from GZ. The straight line in Fig, 1 1 is the
least-squares fit of

In ( - 1 = a + i3D + e

where Y and X represent, respectively, vegetation and soil concentrations and D is the
distance from GZ. Direction with respect to GZ does not appear to be an important
factor; so it was ignored for this example. Although the fit in Fig. 1 1 looks reasonable by
the R^ criteria* (R^ = 0.56), note that the range of the ratio for distances greater than
250 ft spans three orders of magnitude.

Compare this with the linear multivariate fit in Fig. 10 (Anderson, 1958,
pp. 179-187), which is the simultaneous least -squares fit of

hi(Y) = ai +/3iD + ei (4)

In (X) = a2 +1320 + ^2 (5)

*R2 = 1

I (Zi-Zi)'

.2 (Zi-Z)^

1=1

where zj is the estimate of z, R^ is a measure of the amount of variability accounted for by the model,
R^ = 1 implies a perfect functional relationsliip, and R^ = 0 implies no linear relationship (Snedecor
and Cochran, 1967, p. 402).

RATIOS IN TRANSURANIC ELEMENT STUDIES 201

o
z
o
o

0.01

0.001

0.023D

ln(Y) = -2.66 - 0.00898D

0.1

z
o
o

(/I

3
ui
n

0.01

0.001

100 200 300

DISTANCE FROM GZ, ft

400

Fig. 10 Uranium-235 concentrations in vegetation and soil as a function of distance
from GZ at A site, Area 11, Nevada Test Site.

where ai , |3i , ei and tta > i^a . ^2 represent, respectively, the parameters and error terms of
the regression of In vegetation concentration on distance and In soil concentration on
distance. Equation 4 corresponds to line 1 in Fig. 10 and Eq. 5 to hne 2. This approach
assumes that ei and €2 are related; they represent, respectively, the deviations from the
models given by Eiqs. 4 and 5 on the same unit, in this case location. Again the
straight -line fits look reasonable; for vegetation the univariate R^ =0.45 and for soil
R^ =0.77.

The multivariate approach has two advantages: (1) it can be used to check the
assumption of proportionaHty that underlies the use of the ratio and (2) it can help

202 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

h-
<
H
LU
CJ
LU
>

0.01

200 300

DISTANCE FROM GZ, ft

400

Fig. 11 Ratio of ^^^U concentrations in vegetation and soil as a function of distance
from GZ at A site, area 11, Nevada Test Site.

explain the observed variability in the ratios. Most important is the question of
proportionality. If vegetation concentration is proportional to soil concentration, lines 1
and 2 in Fig. 10 should be parallel. A comparison of their slopes would be a test of that
assumption. The slopes for the vegetation and soil concentrations are —0.00898 and
-0.023, respectively. In Fig. 10 the two regression lines do not appear to be parallel.
Simultaneous 95% confidence intervals for j3i and (^2 , which are rather large because there
are so few data points, are given by

-0.017 <|3i < -0.00098

-0.032 <|32 < -0.013

(Morrison, 1967, pp. 107-109). Since the two intervals overlap, one cannot claim that j3i
and ^2 differ significantly, although making a judgment on so few observations is risky.

RATIOS IN TRANS URANIC ELEMENT STUDIES 203

The second advantage to the multivariate approach is that it can explain the
variability in the ratios. A large ratio can be caused by either a large numerator or a small
denominator and a small ratio by the converse. The uncritical use of ratios can obscure
information. For example, the apparently strong trend of increasing ratio with increasing
distance in Fig. 11 is, in part, due to the two samples at 273 and 306 ft, for which the
vegetation concentration is larger than the soil concentration. Admittedly these anomalies
may be traceable to the near background levels of ^ ^^ U at greater distances from GZ, but
this example illustrates the need for caution when using ratios as input to further
statistical analyses.

Profile Analysis

In this example a technique called profile analysis by Morrison (1967, pp. 186-197) is
used to compare inventory ratios (IR) between ecosystem components at several sites.
The elements of the profile are ^^^Pu standing crops in the various ecosystem
components at one location. Comparative studies of radionuclide inventories in
ecosystems are often based on models such as the simple three-compartment one
illustrated in Fig, 12. The box-and-arrow model represents the aboveground components

RESUSPENDIBLE
SURFACE MATERIALS

LIVE
VEGETATION

LITTER

—

Fig. 12 Simple three-compartment model of aboveground components of herbaceous
plant community.

of a herbaceous plant community on an abandoned field. The boxes denote ecosystem
compartments (resuspendible surface materials, live vegetation, and Utter), and the arrows
denote fluxes of ^^^Pu. Inputs to the model (arrows entering the larger dashed box)
represent aerial deposition of ^^^Pu from a reprocessing facility. Outputs (arrows exiting
the dashed box) represent either wind dispersal of resuspended ^^^Pu or ^■^^Pu
incorporated into the soil. Other potentially important compartments and fluxes have
been omitted for simplicity. Resuspendibles are those materials which can be resuspended
into the atmosphere by a 6 m/sec wind velocity at ground level (McLendon et al., 1976).
The question often asked is whether the radionuclide distribution among compart-
ments is the same at each site aUhough the amount of radionuclide per unit area may
differ between sites. The question can be stated another way: Are the IR's of the amount

204 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

of radionuclide in one compartment to that in another the same at the various sites?
Assume for ease of exposition that there are two sites and three compartments. Let v\^
(i= 1,2; j = 1,2,3) represent the true radionucUde content of the ']th compartment at
site i. Then the IR of compartment 1 to compartment 2 at site 1 can be expressed as
V\ ilv\2- The rephrased question can be expressed as the statistical null hypothesis

Ho:

>i r

'V2x'

V\2

J^2 2

j^i 1

V2V

^^13

i^2 3

Vx2

J^2 2

y\3_

/2 3_

(6)

Recall that In (X/Y) = In (X) - In (Y). By defining /iy = In v,^ , then

i^y-"

1 -Mi2

If the data are collected so that a measurement of radionuclide content for each
compartment is made at each sampling location, the hypothesis (Eq.6) can then be
expressed as

H'o:

A^i 1 -A'i2

1^12 - f^l 3

f^2l -A'22
M22 -^23

(7)

The terms (/ij j — ^j j+ , ) are estimated by

^ --^ 1 V

k=l

where Xj j ^ is the radionuclide content observed at the ktfi sampling location in the jr/z
compartment at site i. Note that

Vx..j+i.k/

Inxij^k -lnx,,j + i,k

so consequently (/ij j -^(ij+, ) is the mean of the logarithms of the nj observed IR's of
compartment] to compartment] + 1 for site i.

There are several reasons for doing this. First, instead of a nonlinear hypothesis on the
ratios of random variables, we now have a linear hypothesis tor which multivariate
techniques currently exist. Second, taking logarithms tends to equalize variability and
make the normality assumption more tractable; both of these conditions are assumed by
the test of H'q, which is discussed below.

RATIOS IN TRANS URANIC ELEMENT STUDIES 205

Note that for three compartments there are three possible IR's, which are listed in
Eq. 6. For H'o the IR's are expressed as Afjj — jUjj+i; so three compartments can be
represented by only two IR's. A nice property of the following test of H'q is that the
ordering of the compartments is independent of the test, and compartment j + 1 need not
be the source for compartment j. The hypothesis H'o is tested by the statistic

nin; ^

Ml, 2), (Ml, 2 -Ml, 3)] - [(M2,1 -M2,2),(M2,2 -M2,3)]}| '

(8)

where S is the pooled sites covariance matrix of the (jUi j — /ij j+j ) terms. T^ is distributed
as Hotelling's T^ (Morrison, 1967, p. 1 17), and

F =

(n, +n, -3)T^
2(ni + n2 - 2)

(9)

where F has the F distribution with 2 and ni + n2 — 3 degrees of freedom if H'o is true.
Detailed procedures for testing Hq (test of paralleUsm) are given by Morrison (1967,
pp. 143 and 188).

The data for this example are from unpublished data supplied by A. L. Boni, J. C.
Corey, H. H. Horton, and M. H. Smith of the Savannah River Ecology Laboratory, Aiken,
S. C. They consist of ^^^Pu inventories (measured in picocuries per square meter) for the
three compartments for two sites located at 0.23 km (community 1) and 0.43 km
(community 2) from the point of aerial release of ^^^Pu from a reprocessing facility at
the U. S. Department of Energy Savannah River Plant. The data are given in Fig. 13,
where the horizontal and vertical bars denote, respectively, the aritlimetic means and 95%
confidence intervals computed from 17 samples in community 1 and 12 samples in
community 2. For tliis set of data,

-2 _

17x 12

17+ 12

^ {[4.63, 1.04] - [3.44,2.04]

0.578 0.263
0.263 0.773

]%MA

= 6.8

(10)

With F=3.27 (computed using Eq. 9), the null hypothesis is rejected at the O!=0.10
level. These results indicate that a greater proportion of the ^^^Pu occurs in the
resuspendible compartment at the more highly contaminated site (community 1),
whereas a greater proportion of the ■^^^Pu occurs in the litter at the less contaminated
site (community 2).

The method is easily extended to more than three compartments or more than two
locations (Morrison, 1967, p. 188). It can also be applied to concentration ratios.

Assume that there are p compartments. It might appear easier to choose p — 1 IR's of
interest (denoted by Kjj, j = 1, . . ., p - 1) from the p compartments and test the null
hypothesis

206 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Ho:

K^i ,2

K2,l

^2,2

•

K2,p-1_

L^i,p-U L

using the means of the observed IR's for each site. This test is, however, dependent on
which p - 1 of the possible p(p — 1) IR's is selected. This is not true of Eq. 10, where the
T^ value obtained is independent of the ordering of the compartments and the
acceptance of H'o implies that all the possible p(p — 1) IR's in the two locations have not
been demonstrated to be unequal.

From these two examples we have seen that multivariate statistical techniques can be
used to address such questions as the assumption of a muUiphcative relationship between
variables and the simultaneous test of equality of IR's, which are not easily dealt with by
the use of ratios or traditional univariate techniques.

10'

,4 _

O 103 -

I-
<

I-

o
o

2 _

■\

1 1

- 1

> 95% Confidence interval

> Mean

—

- 1

- 2

~ 1

- 2

—

RESUSPENDIBLES

LIVE
VEGETATION

COMPARTMENTS

LITTER

Fig. 13 Plutonium-238 standing crop in three ecosystem compartments for two sites.

RATIOS IN TRANS URANIC ELEMENT STUDIES 207

Summary

The appropriate use of ratios in environmental radionuclide research demands some
knowledge of the relationship between the two variables composing the ratio.
Concentrations and pure ratios are two types of ratios used extensively in environmental
transuranic element research. A concentration is the amount of activity per unit of weight
or volume, e.g., nanocuries per gram. A pure ratio, say ^^*Pu/^^^Pu or ^^*Pu in
vegetation/^ ^^Pu in soil, is used to express a relationship between the two random
variables. In either case the ratio is a valid description of that relationship only if the
relationship is approximately multiplicative. For concentrations a major problem
(addressed in tliis cliapter using two sets of environmental radionuclide data) is
determining the aliquot size for which activity is homogeneously dispersed throughout
the medium. For pure ratios the problem is assessing whether the relationship is
multiplicative. If not, more meaningful methods, such as multivariate statistical
techniques, must be found for relating the two variables. This chapter discusses two
multivariate techniques, regression and profile analysis, which might be used as a test of
the multipUcative assumption.

Acknowledgment

This paper was written under U. S. Department of Energy contract EY-76-C -06-1 830.

References

Anderson, T. W., 1958, An Introduction to Multivariate Statistical Analysis, John Wiley & Sons, Inc.,

New York.
Atchley, W. R., C. T. Gaskins, and D. Anderson, 1976, Statistical Properties of Ratios. I. Empirical

Results, 5vsr. Zool., 25: 137-148.
Cataldo, D. A., R. E. Wildung, and T. R. Garland, 1978, Technetium Accumulation, Fate, and

Behavior in Plants, in Environmental Otemistry and Cycling Processes, DOE Symposium Series,

No. 45, Augusta, Ga., Apr. 28-May 1, 1976, D. C. Adriano and I. L. Brisbin, Jr. (Eds.),

pp. 538-549, CONF-760429, NTIS.
Chayes, P., 1911, Ratio Correlation, University of Chicago Press, Chicago.
Cochran, W. G., 1977, Sampling Techniques, 3rd ed., John Wiley & Sons, Inc., New York.
Creasy, M. A., 1956, Confidence Limits for the Gradient in the Linear Functional Relationship, 7. R.

Stat. Soc, Ser. 5, 18: 65-69.
Doctor, P. G., and R. O. Gilbert, 1977, Ratio Estimation Techniques in the Analysis of Environmental

Transuranic Data, in Transuranics in Natural Environments, Symposium Proceedings, GatUnburg,

Tenn., Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178,

pp. 601-619, Nevada Operations Office, NTIS.
, and R. O. Gilbert, 1979, Two Studies in Variability for Soil Concentrations: With Aliquot Size

and with Distance, in Selected Environmental Plutonium Research Reports of the NAEG, M. G.

White and P. B. Dunaway (Eds.), USDOE Report NVO-192, Nevada Operations Office, NTIS.
Gilbert, R. O., and L. L. Eberhardt, 1976, Statistical Analysis of "A Site" Data and Interlaboratory

Comparisons for the Nevada Applied Ecology Group, xn Studies of Environmental Plutonium and

Other Transuranics in Desert Ecosystems, M. G. White and P. B. Dunaway (Eds.), ERDA Report

NVO-159, pp. 117-154, Nevada Operations Office, NTIS.
Hollander, M., and D. A. Wolfe, 1973, Non-Parametric Statistical Methods, John Wiley & Sons, Inc.,

New York.
McLendon, H. R., et al., 1976, Relationships Among Plutonium Contents of Soil, Vegetation and

Animals Collected on and Adjacent to an Integrated Nuclear Complex in the Humid Southeastern

United States of America, in Transuranium Nuclides in the Environment, Symposium Proceedings,

San Francisco, Nov. 17-21, 1975, pp. 347-363, STI/PUB/410, International Atomic Energy

Agency, Vienna.

208 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Michels, D. E., 1977, Sample Size Effect on Geometric Average Concentrations tor Log-Normally

Distributed Contaminants, £'nv/ra«. 5c/. TechnoL, 11: 300-302.
Mielke, P. W., Jr., and J. A. Flueck, 1976. Distribution of Ratios for Some Selected Bivariate

Probability Functions, in 1976 Proceedings of Social Statistics Section, American Statistical

Association, Washington, D. C.
Morrison, D. F,, 1967, Multivariate Statistical Methods, McGraw-Hill Book Company, New York.
Parzen, E., 1967, Stochastic Processes, Holden-Day, Inc., San Francisco, Calif.
Ricker, W. E., 1973, Linear Regression Fisheries Research, J. Fish. Res. Board Can., 30: 409-434.
Sharitz, R. R., S. L. Scott, J. E. Pinder III, and S. K. Woods, 1975, Uptake of Radiocesium from

Contaminated Floodplain Sediments by Herbaceous Plants, Health Phys., 28: 23-28.
Simpson, G. G., A. Roe, and R. C. Lewonton, 1960, Quantitative Zoology, Harcourt Brace

Jovanovich, Inc., New York.
Snedecor, G. W., and W. G. Cochran, 1967, Statistical Methods, 6th ed., Iowa State University Press,

Ames, Iowa.
Sokal, R. R., and F. J. Rohlf, 1969, Biometry, W. H. Freeman and Company, San Francisco.
U. S. Energy Research and Development Administration, 1976, Workshop on Environmental Research

for Transuranic Elements, Battelle Seattle Research Center, Seattle, Wash., Nov. 12-14, 1975,

ERDA Report ERDA-76/134, NTIS.

Review of Resuspension Models

J. W. HEALY

Resuspension has been recognized as a potential mode of human exposure to
contaminants in the soil for a number of years. However, methods for expressing the
resulting concentrations quantitatively have been crude; in most cases, the resuspension
factor was used. In this chapter we distinguish three main types of resuspension:
wind-borne, mechanical disturbance, and local. The three main models for estimating
concentrations (resuspension factor, resuspension rate, and mass loading) are described,
and the data applicable to each are reviewed.

The studies of wind erosion of desert sands and of agricultural soils liave provided
much of the information available on the mechanisms involved in wind resuspension.
However, such a body of evidence is not available for mechanical disturbance of the soil,
although recent experiments have provided some data. Information on concentrations in
the vicinity of the individual causing the disturbance is still poor.

Considerable progress has been made in the past few years on wind resuspension. For
both wind and mechanical resuspension, scientists are on the verge of providing improved
estimates using the resuspension rate at specific locations. However, for generic studies
the mass-loading approach is recommended.

Resuspension from soils and subsequent inhalation of the resuspended material has long
been considered the chief source of exposure to transuranium elements deposited in soils.
In spite of the obvious importance of this pathway, research has been limited; thus it is
difficult to obtain a reasonable prediction of the resulting concentrations and inhalation.
In fact, Slinn (1978) indicates that he does not trust resuspension factors, rates, or
velocities to within many orders of magnitude. In the following text we will review the
data available and attempt to arrive at as reasonable an answer as possible.

Types of Resuspension Considered

The overall resuspension problem can be divided into three types for conceptual
understanding and calculations (Healy, 1977a; 1977b): (1) wind-driven resuspension, (2)
mechanical resuspension, and (3) local resuspension. For wind resuspension the energy
required to dislodge the particles arises from the wind, and the particles then disperse
downwind, depositing on surfaces at a rate depending on the aerodynamic properties of
the particles and the nature of the terrain. Both mechanical resuspension and local
resuspension result from mechanical disturbance of the soil. However, in the mechanical
resuspension case the concern is with concentrations downwind following dispersion and
deposition. Local resuspension, on the other hand, describes the exposure in the
immediate vicinity of the individual before dispersion occurs.

209

210 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Another type of resuspension that will be considered only briefly is transfer
resuspension. This involves the transfer of the contaminant from its place of deposit to
another place where inhalation may be more probable. Unfortunately data to describe
this type of resuspension are extremely limited.

Each of these types of resuspension wiU be considered in deriving the best estimate of
exposure to an individual in an area contaminated by the spread of transuranium
elements in the soil.

Resuspension Modeling

The three basic techniques in use for resuspension modeling are (1) the resuspension
factor, (2) the resuspension rate, and (3) the mass-loading approach. Each method has its
strengths and its weaknesses, particularly in view of the state of the technology at this
time. Each of the techniques is described, and its advantages and disadvantages are
discussed.

Resuspension Factor

The resuspension factor is defined as the ratio of the concentration in the air at a
reference height (usually 1 m) to the quantity of the contaminant per unit area on the
surface of the ground. The usual units are meters"^ . Its chief advantages are its simplicity'
and the fact that most measurements in the past have been expressed in this form; so
values are available for calculation.. The chief disadvantages are that it is a completely
empirical formulation, and thus it is difficult to extrapolate from one terrain to another,
and it ignores the distribution of the contamination over the are;^ and the size of the
contaminated area involved. Thus the denominator contains the local quantity on the
ground, and the numerator is an undefined function of the air concentration resulting
from upwind contaminated areas and activities.

A problem common to this model, as well as to all others, is the uncertainty resulting
from the depth to be used in assessing the quantity per unit area to be used in the
denominator. For a uniform profile in the soils, the quantity per unit area will increase in
direct proportion to the depth of the sample used. For nonuniform profiles the estimate
of the quantity per unit area will also change, depending on the depth used. For wind
resuspension it can be assumed that the appropriate depth is small, perhaps a miUimeter
or less, although it is likely that this depth may be variable, depending on wind speed and
the degree of saltation allowed by the size of the area and the nature of the soils. For
mechanical disturbance the depth will be some function of the depth to which the
disturbance occurs, the function depending on the relative ease with which the particles
can escape from the soil to the atmosphere.

Mishima (1964) has tabulated resuspension factors measured over a variety of
conditions. It is frequently noted that these values range over eleven orders of magnitude.
However, the values quoted represent both outdoor and indoor conditions, with and
without mechanical disturbance, and at various times after the contaminant has been
deposited. In a brief review of Mishima 's table, it is noted that the values for mechanical
disturbance range from about 2 x 10~^ to 7 x 10~^ m"^ (with one value of 10~^ m"^
based on uranium contaminant with dust stirred up and sampling at a height of 1 ft
ignored). For periods of no activity, with relatively freshly deposited material, the values
generally range from 10"^ to 2 X 10~^ m~^ , whereas for aged material they range from
6 X 10~'° to 10~'^ m~^ . It is difficult to generalize these numbers because the exact
value will depend on the degree of disturbance, the placement of the sampler, the

REVIEW OF RESUSPENSION MODELS 2\ 1

meteorological conditions at the time, and the nature of the soils. However, many of the
measurements were made in d6sert areas with low soil moisture where resuspension would
be expected to be highest.

An apparent reduction in the resuspension factor with time was indicated by Wilson,
Thomas, and Stannard (1960) from measurements at the Nevada Test Site (NTS). Here
samples were taken at three different distances from the center of a safety shot starting
about 1 month after the contamination occurred. The investigators noted that the data
were too erratic to establish half-times for the decay of the air concentration beyond a
very crude estimate. This estimate was made by plotting the medians of the data at each
sampling distance, and it provided a value of 5 weeks for the concentration half-time.
This value was used by Langham (1969) in assessing future hazards from plutonium
contamination. A somewhat larger half-life of 45 days was used by Kathren (1968) in a
study of acceptable contamination levels for plutonium in soils. Anspaugh et al. (1973)
measured the decrease in air concentration with time immediately following a cratering
event in Nevada and following the venting of an underground shot. For the cratering
event sampling was carried out for 6 weeks following the event with a measured half-time
of 38 days. Tlie venting experiment started 3 months after the event and continued for 9
to 10 months, with the most predominant radionucUdes being the isotopes of ruthenium.
Here a half-life of 66 days was found.

However, the consequences of continuing such half-times over a long period were
pointed out by Healy (1974) and Anspaugh (1974). Healy noted that samples taken at
the same location as those taken by Wilson, Thomas, and Stannard (1960) but about 1 yr
after the conclusion of the Wilson, Thomas, and Stannard study (Olafson and Larson,
1961) gave values up to several orders of magnitude greater than would be predicted by
the 35-day half-Ufe. Anspaugh (1974) indicated that the functional nature of the decrease
in resuspension rate with time cannot be confidently extrapolated and that previously
pubHshed models should not be applied to calculations many years after the
contamination event. He also cited two sets of measurements at NTS where the area had
been contaminated with plutonium by high-explosive detonations some 20 yr earlier.
These studies gave values for the resuspension factor of 3 x 10"^^ m~^ and 2 x 10~^
m~' . These data indicate unequivocally that resuspension does occur after this period of
time, although predictions using the short half-life following deposition would result in
unmeasurable values of air concentration.

Anspaugh, Shinn, and Wilson (1974) used the available information to derive an
empirical expression for the resuspension factor which allows the resuspension factor to
decrease with time. In their derivation they used four constraints: (1) the apparent
half-time of decrease during the first 10 weeks should approximate a value of 5 weeks, (2)
this half-Ufe should about double over the next 30 weeks, (3) the initial resuspension
factor should be 10""* m~\ and (4) the resuspension factor 17yr after the
contaminating event should approximate 10~^ m~^. The value for the resuspension
factor in the aged source resulted from 23 individual air concentration measurements at a
location contaminated with plutonium 17 yr earlier where the average resuspension factor
was found to be 10~^ m"^ . An expression that approximates these constraints is given as

R(t) = 10-^ exp [-0.15 (t)'^] + 10-^ (1)

where R(t) is the resuspension factor (meters"^) and t is the time since the
contaminating event (days).

212 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Oksza-Chocimowski (1977) provided a "generalized" model for the change in
half-time of the resuspension factor which allows the ratio of the resuspension factor at
time zero to the resuspension factor at a long time [R(0)/R(o°)] to vary:

T^(t)=Aln(l+B + CtD)+-^|^ (2)

In — -^^

R(oo)

where A = constant coefficient, days
B = constant

C = constant coefficient, days"
D = constant exponent
R(0') = initial resuspension factor, m~ '
R(oo) = final resuspension factor, m~^

t = time since contaminating event, days

This expression was evaluated for a range of values of the constants with limiting
values corresponding to ratios of R(0)/R(°°) of 10 and 10''. However, data are not
available to make adequate choices for a given area.

In both the Anspaugh, Shinn, and Wilson and the Oksza-Qiocimowski models, the
constants are evaluated on the basis of data from a hmited number of areas and
conditions. In Anspaugh 's model, for example, it is apparent that the final resuspension
factor after long aging is based on wind resuspension only in a desert area. Whether
disturbance in the area would cause an increase is unknown, although it appears likely to
do so. Sehmel and Orgill (1974) made measurements downwind from the original
oil-storage pad at Rocky Flats and related the concentrations found to the 2,1 power of
the wind speed. Subsequent measurements were higher, however, owing to the digging of
a ditch between the oil storage area and the sampler. This work also involved increased
vehicular activity. This disturbance caused about an order of magnitude increase in
concentrations at the samplers which persisted after the work had been completed
presumably because of a change in the character of the surface. The half-time for decrease
over the next 7 montlis appeared to be about 9 months. Thus it is possible that
disturbance could not only increase the concentration at the time that it occurs but also
could result in increased wind-driven resuspension factors for some time thereafter.

Sehmel and Lloyd (1976a) and Sehmel (1977a) have also provided data on the
wind-borne resuspension of a submicrometer tracer, calcium molybdate, that was applied
to a test area by spraying. In this experiment, using cascade impactors arranged to operate
in different wind conditions, they noted that the resuspension factor at 1.8 m height
increased as the 6.5 power of the wind speed. However, this slope was determined by
drawing the line through the bottom end of uneven wind-speed ranges, a procedure that
could well result in an overestimate. They also noted, for this system, that there seemed
to be little, if any, decrease in the resuspension over a period of 3 yr. Whether this is due
to a "preaging" by the apphcation in water is unknown.

In concluding the discussion of the resuspension factor, it is apparent that this
empirical approach does not inherently incorporate many of the variables, and present
estimates are relatively crude. In particular, the present estimates appear to be based on
short-term experiments with Uttle attempt to provide a factor applicable to long-term
(say, 1 yr) averages.

REVIEW OF RESUSPENSION MODELS 213

Resuspension Rate

The resuspension rate is defined as the fraction of the contaminant present on the ground
that is resuspended per unit time by either winds or mechanical disturbance. Once
obtained, it can be used to describe concentrations at any point around a nonuniform
contaminated area by the use of point-source dispersion and deposition equations and
integration over the area. This potential use was illustrated by applying it to an area
contaminated with plutonium by a safety shot (Healy, 1974) and the inverse use at the
same area to obtain resuspension rates from measured air concentrations (Anspaugh et al.,
1975). It was introduced for use in resuspension calculations by Healy and Fuquay
(1958), although in a crude form.

Slinn (1978) has pointed out that the resuspension rate can be converted to a
resuspension velocity by multiplying by the ratio of the quantity of contaminant per unit
area and dividing by the volumetric concentration of the contaminant in the soil. Such^
velocity is analogous to the deposition velocity with, however, a negative sign when
conditions are such that net resuspension occurs.

Three techniques have been used to measure resuspension rates for a given area: (1)
measurement of air concentrations resulting from a known pattern of a tracer material on
the ground and inferring the resuspension rate from height profiles, which gives the total
transport, or from use of dispersion equations in known meteorology; (2) measurement
of air concentrations from an existing contaminated area and obtaining the resuspension
rate as given in 1 ; and (3) measurement of natural dust fluxes and relating these fluxes to
some association between the concentration of the contaminant in the soil and the dust
flux. The last method has been used only for wind resuspension.

Wind Resuspension. A detailed body of knowledge exists on the mechanisms of the
movement of soils by wind through the classic studies of Bagnold (1943) on desert sands
and the detailed studies of Chepil (1941; 1945a; 1945b; 1945c; 1951a; 1951b; 1956;
1957; 1960), Bisal and Hsieh (1966), Woodruff and Siddoway (1965), U. S. Department
of Agriculture (1968), and MaUna (1941) on agricultural soils. These studies wall not be
reviewed in detail since much of the information is appUcable to the limited condition of
erosion of erodible soils. There are, however, data and concepts applicable to the
resuspension process, at least for the limited conditions of agricultural soil, and a brief
review of these is in order.

The relationsliip between erosion and winds is complex; a large number of variables
affect the outcome, Chepil (1945a) listed the most important of the factors as related to
the three categories given in Table 1 . In the following discussion I will briefly describe
some of the more important findings applicable to the general problem of resuspension
from the extensive work on soil erosion.

Soil movement across an eroding field is primarily from movement of the smaller
particles, usually less than about 1 mm in size. There are three mechanisms for
movement, and the particular size for each is somewhat dependent on the wind speed.
The heaviest particles move by surface creep or movement along the surface. Chepil
(1945a) noted that these grains were too heavy to be moved by the direct pressure of the
wind but were propelled by the impacts of smaller grains moving in the second method of
movement, saltation. In saltation the grains, after being rolled by the winds, suddenly
leap almost vertically. Some grains rise only a short distance, whereas others, depending
on the wind speed, can rise to several feet. They are then carried forward by the winds

214 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Most Important Factors in Wind Erosion
of Agricultural Soils

I. Air

II. Ground

Velocity

Turbulence

Density
Temperature
Pressure
Humidity

Viscosity

Roughness

Cover

Obstructions

Temperature

Topographic features

*Based on data from Chepil (1945a).

III. Soil

Structure
Organic matter
Lime content
Texture
Specific gravity
Moisture content

and settle by gravity until they strike the ground. Chepil (1945a) attributes this sudden
rise to the spinning of the grain as it rolls along the surface, with the Bernoulli effect
causing a difference in pressure at the top and bottom of the grain. The third method of
movement is, then, suspension of the small soil particles by the wind. In the last case the
particles must be small enough to be kept airborne by the turbulent forces in the
atmosphere overcoming the force of gravity. Since the turbulence varies with atmospheric
stability and, to some extent, with wind speed, one would expect that larger particles
would be suspended in strong turbulent winds and that these particles would settle out as
the winds decrease. However, the fme particles can remain suspended for long times and
can cover large distances. It is the suspension fraction, and particularly the smaller
particles, that is of interest in resuspension since resuspension is defined as the suspension
of a previously deposited contaminant.

The airflow characteristics over the surface of the soil are important in transmitting
force to the soil grains and in determining the velocity at wliich they start moving. In a
neutral atmosphere (i.e., temperature decrease with height is adiabatic), the velocity
profile is logarithmic with height; so one can write (Prestley, 1959),

K \zo/

(3)

where Uz = wind speed at a height z

zo = height at wliich the wind speed is zero
U:^ = friction velocity (or drag velocity)
K = von Karman constant, which has been found by integration to equal 0.4

Zo is a characteristic of the surface, increasing as the roughness of the surface increases.
The friction velocity is of importance in determining the force exerted on any object
protruding above the laminar layer of the atmosphere and has been shown to be the
characteristic wind speed tliat affects soil movement (Bagnold, 1943; Chepil, 1945b;
1945c).

Bagnold (1943) has shown that in severe erosion conditions the profile of wind speed
with height is changed by the momentum transfer to the particles in saltation. It has been
observed that the wind velocity must exceed some threshold value to induce movement
of the soil. Chepil (1945b) has studied the movement of particular sizes of grains and has
noted that the curve of grain diameter times the specific gravity vs. threshold friction

REVIEW OF RESUSPENSION MODELS 215

velocity for initial movement has a minimum at about 0.15 mm (150 idm) with a friction
velocity of about 0.15 m/sec; i.e., grains smaller than this size and grains larger than this
size require higher friction velocities to initiate any movement. The stabiUty of the finer
grains is illustrated by the simple experiment of Bagnold (1943), in which he placed a pile
of talcum powder on a smooth surface and exposed it to winds. Tlie layer was stable at
relatively liigh wind speeds. However, a few grains of larger particles sprinkled on the pile
resulted in rapid dispersion at a wind speed much lower than would serve to disperse the
particles without tliis added factor. It is believed that the relative stability of the small
particles is due to the fact that they do not protrude above the laminar layer; thus no
drag force is exerted on them. The minimum in the curve of velocity required to institute
movement and particle size, then, is due to the balance between the increasing drag force
as the particle increases in size and the increasing downwind force from gravity as the
particle becomes larger. Above about 0.15 to 0.2 mm, the threshold velocity required to
initiate movement increases as the square root of the product of the particle diameter and
its specific gravity (Bagnold, 1943; Chepil, 1945a).

As a result of this threshold friction velocity, it is apparent that direct aero-
dynamic pickup of small particles from the soil is unlikely. Instead, the process of
saltation is the key to producing the suspended fraction because the impact of these
particles as they strike the ground provides the energy to propel the smaller particles
above the laminar layer into the wind stream where they are transported by eddies in the
wind. Tlius, in the talcum powder example, the sand particles sprinkled on the talcum
powder served the function of the saltating particles. In a field, knolls, ridges, sand
pockets, or other areas most exposed to the wind and/or containing the easily erodible
grains start to erode at a lower velocity than the rest of the field. Once the erosion starts,
it spreads downwind, and the bombarding action of the particles in saltation causes
movement in other parts of the field that normally would not be eroded (Chepil, 1945a).
The threshold velocity of the field is therefore the threshold of the most exposed or most
erosive spots in the field. Since the avalanching effect of saltation increases down the field
in the direction of the wind, the length of the field is also a factor in the degree of
erosion.

An important consequence of the role of sahation in the production of resuspension
is that there will be no dust, or particles flowing in suspension, unless the wind speeds are
great enough to produce sahation under the conditions of the tleld. This places a
threshold condition on the wind speeds required to resuspend particles from the ground.

Bagnold (1943) measured the rate of soil flow for desert sands and found that these
rates could be described by an equation of the form of

q = Cu|^ (4)

where q = rate of soil flow (grams per centimeter width per second)
p - density of the air
g = acceleration of gravity
C = constant that differs for differing soils and forms of erosion

Bagnold concluded that on the desert sands the tlow in suspension was small, about y2o^/'
of the total flow, as compared with saltation and surface creep. Chepil (1945a; 1945b)
made similar measurements on agricultural soils in a wind tunnel. His results indicated

216 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

that all three methods of flow appear to follow Bagnold's law of the cube of the wind
velocity, at least for the soils tested, a fine sandy loam and a heavy clay soil. The constant
C for total soil flow varied widely for different soils; the range in these experiments was
1.0 to 3.1. Chepil also measured the proportion of each type of flow on four widely
different soils. These results are given in Table 2.

TABLE 2 Relative Portion of Three Types of Flow on

Different Soils

Percent of flow in

Soil type

Surface creep

Saltation

Suspension

Sceptre heavy clay
Haverhill loam
Hatton fine sandy loam
Fine dune sand

24.9

7.4

12.7

15.7

71.9

54.5
54.7
67.7

3.2
38.1
32.6
16.6

*Based on data from Chepil (1 945b). ■

It is apparent that the fraction of the total flow carried in suspension is considerably
higher on agricultural soils than on desert sands (presumably because of the availability of
the smaller particles). These studies showed that the logarithm of the flow plotted against
height was essentially a straight line and that relative concentrations of soil particles at
different heights remained the same with wind velocities ranging from 13 to 30 miles/hr
(6 to 13.5 m/sec). Presumably, then, the relative flows also remained constant.

Chepil (1945c) also provided the sizes of the particles in the soils studied, between
<0.1 and 0.83 mm. The relative suspension flow vs. the fraction of particles <0.1 mm is
plotted in Fig. 1. The use of any other particle size range or cumulative percentages gave
erratic results. This may indicate the importance of the fraction of the smaller particles in
the soil in producing the suspension fraction.

An important factor in the suspension fraction is the aggregate state of the smaller
particles in the soil. Particles in the submicron size range rarely exist as such in the soils
because they tend to either clump together or to adhere to larger particles and thus
become small aggregates. In fact, Chepil (1945b) states that particles smaller than 0.005
mm (5 )Um) do not exist as such in ordinary soils because they are aggregated into larger
individual grains. He adds also that single grains or aggregates 0.05 to 0.5 mm in diameter
have Uttle or no cohesive properties and are easily carried by the winds. This means that
contaminants in the soils, either as fine particulates or absorbed on the surface of soil
particles, will largely exist as soil aggregates and will behave in the same manner as the
soil. Chepil (1957) demonstrated the aggregation of material carried in suspension at
heiglits of 4 to 8 ft in a dust storm by sizing particles by sedimentation in CCI4 , a
nonpolar solvent that tends to preserve aggregates, and then repeating in water following
dispersion with sodium hexametaphosphate. The curves show the percentage of particles
smaller than a given value reaching zero at about 5 jum in diameter in a 1954 storm and
about 10 iJtm in a 1955 storm. By contrast, the dispersed samples showed 15 to 25% of
the particles smaller than 5 iim.

An important factor in aggregation is the moisture content of the soil. This has been
investigated by Chepil (1956) and by Bisal and Hsieh (1966). Chepil (1956) has provided
a formulation for the soil flow taking into account the increased resistance to movement

REVIEW OF RESUSPENSION MODELS 21 7

10 20

PERCENT < 0,1 mm

Fig. 1 Suspension flow vs. percent of grains in soil <0.1 mm.

due to the cohesion of the absorbed water films. However, it is noted that the water
content of a field can decrease rapidly following a rain owing to the drying actions of the
winds.

Many other factors influence the erosion of the field, such as the presence or absence
of ridges, the quantity of vegetative cover, and the presence or absence of a surface crust.
These have been combined to give a soil-erosion equation (Chepil, 1960; Woodruff and
Siddoway, 1965)

E = f(l',C',K',L',V)

(5)

where E = potential average annual soil loss (tons per year)

r = soil and knoll erodibility
C' = local wind-erosion cUmatic factor
K' = soil ridge roughness factor
L' = field-length factor

V = equivalent quantity of vegetation

Mathematical relations have been established between the individual variables. The
relationships, however, are so complex that the individual factors are evaluated separately
in a form in which combinations of the factors can be evaluated graphically.

218 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

We will not attempt here a complete discussion of this equation along with numerical
factors but will discuss each of the factors in a quaUtative fashion because it appears that
many of these could be of importance in any case of wind resuspension.

The soil and knoll erodibility consists of two terms. The soil-erodibility index is
related to the percentage of dry aggregates greater than 0.84 mm in diameter. The higher
this percentage, the lower the soil erodibility. Conversely, the higher the percentage of
aggregates less than 0.84 mm in diameter, the greater the erodibility. The knoll erodibility
is expressed as the percentage of the level-ground erosion that occurs at various slopes of
the knolls in the field. This factor ranges from 0 for a level field to about 650% at the top
of a knoll having a slope of 10% and about 360% from that portion of the windward
slope where the drag velocity is the same as the top of the knoll (about the upper third of
the slope). A surface crust stability factor is usually ignored because the crust
disintegrates rapidly as a result of abrasion once the wind erosion starts.

The soil ridge roughness is a measure of the surface roughness other than that caused
by clods or vegetation. Ridges of 2 to 4 in. in height have been found to be the most
effective in controlling erosion; the erosion for ridges of this height is about 50% of that
over a smooth surface.

The wind-erosion chmatic factor includes the influence of wind speed and moisture.
The rate of soil movement varies directly as the cube of the wind velocity. In this factor
the mean annual wind velocity, corrected to a standard height of 30 ft, is used. Since
atmospheric wind velocities are normally distributed, the probability of obtaining high
winds is higher with liigher mean velocities. The rate of soil movement varies
approximately as the square of effective soil moisture. The wind-erosion climatic factor
has been given for a number of locations by the U. S. Department of Agriculture (1968)
for each month of the year. To illustrate the differences in this factor from one place to
another, I have given rough ranges for four locations. For tlie State of Washington, the
chmatic index varies from about 1 to 50; the higliest values are for the months of March
through May. For eastern New Mexico, along the northeastern border, the climatic factor
remains the highest in the nation throughout the year; values range from about 70 to 300.
For the State of Ohio, the factor ranges from 1 to 10, and for the State of Georgia, about
1 to 5. Thus there are widely differing wind and moisture factors throughout the country;
the east, in particular, has low factors as compared with the west. It could be predicted
that the resuspension of contaminants from the soil will also be lower in the areas of low
climatic factor for erosion.

The field-length factor again has two parts: the distance across the field and the
sheltered distance. The distance across the field is measured along the prevailing
wind-erosion direction. On an unprotected field, the rate of soil flow is zero on the
windward edge and increases with distance downwind until, for a large field, the flow
reaches a maximum that the wind of a given velocity can maintain. The sheltered distance
is that distance along the prevailing wind-erosion direction that is sheltered by any
barrier.

Vegetative cover is an important factor in controUing wind erosion. Three different
factors are included in the equivalent quantity of vegetative cover. The first is the
quantity of the vegetative cover expressed as clean, air-dried residue. The second denotes
the total cross-sectional area of the vegetative material. The finer the material and the
greater the surface area, the more it reduces the wind velocity and the more it reduces
wind erosion. The tliird factor is the orientation of the cover. The more erect the

REVIEW OF RESUSPENSION MODELS 219

vegetation, the higher it stands above the ground, the more it reduces wind velocity near
the ground, and the lower the erosion.

Gillette and his collaborators (1972; 1973; 1974; 1976) have been studying the
vertical flux of dust from agricultural soils under the influence of the winds. In a study of
particle size distribution at 1.5 and 6.0 m along with the horizontal velocity, it was
concluded that the upper limit of the ratio of the settling velocity to the friction velocity
for aerosols having the potential for long-range transport is approximately 0,2 or sUghtly
greater (Gillette and BUfford, 1974). Also, by studying the sizes of the aerosols produced
at different values of the friction velocity, it was concluded that the dominant injection
mechanism of soil aggregates less than 10 iim in diameter is not direct aerodynamic
entrainment (Gillette and Blifford, 1974). It was noted that the size distribution of the
vertical flux [expressed as dN/d (log r)] was proportional to r^ for particle sizes greater
than 2 pim. A similar particle size distribution was noted for the parent soil when the size
distribution was measured with liquid Freon dispersal. Since the dielectric constant of
liquid Freon is close to that of air, it tends to preserve the aggregate state of the soil.
However, if the aggregates are broken up, the number of particles less than 10 /im greatly
increases (Gillette and Blifford, 1974), showing once again the importance of aggregation
of the smaller particles in the soil.

In a study of the vertical flux above several soils, Gillette (1974) noted that the
production of a flux of particles less than 10 Mm in size increased more rapidly with
friction velocity on a soU that had a higher percentage of silt and clay than it did on a soil
with a relatively low percentage of these materials. On both soils extrapolation of the
curves to the point of intersection indicates a threshold friction velocity of about 0.18
m/sec, about the same value as that for the impact threshold discussed earUer. For the
soils containing 3,5% clay (< 1 /jm) and either 0,5 or 1 .0% silt (<25 jum), the vertical flux
of particles < 1 jum increased as the ratio of the friction velocity to the threshold friction
velocity raised to the 5.14 power. For the soil containing 3,5% clay and 1% silt, the
vertical flux increased as the 9.62 power of the ratio of friction velocity to the threshold
friction velocity. This provides a model for the vertical flux of ,r

Fa = Const. (u*/u*threshold) - (6)

For the two soils evaluated, the constant at the point of intersection was about
6x 10^* cm^ see"' cm~^, or, assuming an aggregate density of 2.5 g/cm^ , about
1.5 X 10~^ g see"' m~^. An important conclusion from these studies was that the
increase in the vertical flux at powers of the friction velocity much greater than the
observed cube for the horizontal flux was due to the breakup of the aggregates by
sandblasting and release of the smaller particles to suspension.

In a further study of the vertical flux resulting from eight different soils (Gillette,
1976), a regular pattern of increase with friction velocity appeared to exist with
considerable spread. A curve that I fitted by eye to the bulk of the data indicated an
average increase of vertical flux as the 5th power of the friction velocity with a constant
at the assumed threshold friction velocity as given by the earlier study of two soils. A
comparison of the size distribution of a parent sandy soil with the size distribution of the
horizontal flux to a height of 1 .3 cm indicated that the size distribution of the horizontal
flux is very similar to the size distribution of the parent soU aggregates. The aerosol at 1
m in height showed a mode for the particles greater than 20 /im centered around 50 jum.
As height increased, the particles less than 20 [im became an increasingly larger

220 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

proportion of the total aerosol. It was also noted that the concentration of larger particles
increased with the wind speed.

Travis (1975; 1976) has developed a model for the redistribution of wind-eroding
soil-contaniinant mixtures using the Gillette and Blifford (1972) and Gillette, Blifford,
and Fryrear (1973) relationships for the vertical and horizontal flux. Tliis model assumes
that the contaminant is closely associated with the soil aggregates and moves in the same
manner as the soil. Since it incorporates only studies from eroding agricultural soils, it
should be Umited to these soil conditions.

Shinn et al. (1976) measured the vertical profile of dust in the atmosphere at two
sites and related these profiles to the vertical dust flux by the eddy-correlation method.
The two sites were an area at NTS where plutonium contamination had occurred during a
series of nonnuclear tests over 20 yr earlier (GMX Area) and in an agricultural field in
west Texas. Measurements were made simultaneously of the dust flux by an optical
particle detector and the mean wind and temperature profiles.

It was noted that the mass-concentration particle size distributions at both sites had a
maximum at about 4 or 5 iim mass median aerodynamic diameter, and tliis decreased by
an order of magnitude at 1 and 10 jLim. This distribution does not agree with the data of
Chepil and Woodruff (1957) or Sehmel (1976a), which show significant quantities of
particles much greater than 10 lum at elevations up to 30 m above the ground. Shinn et al.
(1976) indicate that this could be due to reaggregation after sampling because the
sedimentation velocity of such particles would be greater than u^. However, this remains
an uncertainty tliat requires investigation.

The wind-profile measurements showed the roughness length (zq) at the Texas site to
be 0.044 cm and at the GMX site to be 2.0 cm. This resulted in a drag coefficient
referenced to the wind speed at 2 m (U:^/u2oo) of 0-05 at the Texas site and of 0.10 at
the GMX site. It was noted that the dust profiles in this study, as well as in previous ones,
fit a power law with exponents of either —0.2 or —0.35. Tliis is due to the fact that all are
nearly bare surfaces and the measurements were made in dynamic neutral atmospheric
stability conditions. The dust flux calculation was parameterized by several simple
relations.

F = k| (7)

where F - flux
Z = height

X = dust concentration
K - eddy diffusivity

Since

K = u^ kz (8)

where k is the van Karman constant (k = 0.4). Since the dust concentration follows a
power-law distribution with height,

dz z

REVIEW OF RESUSPENSION MODELS 221

Since the power, P, is about 3 and the concentration over the heights from 0.7 to 2 m
deviates only about 20% from the 1-m reference velocity, one obtains, by combining the
above,

F^-0.12u=,Xi (10)

The data at GMX and Texas, respectively, give values for Xi of

Xi=6.1u^«^ (11)

and

Xi=522u^3^ (12)

For GMX and Texas, respectively, these values then result in fluxes of

F = 0.73u^°^ (13)

and

F = 62.6u^3« (14)

A tentative model of the upward dust flux was derived from profile and soil erosion
data from a number of locations. This was titled the Gillette-Shinn model and is
expressed by

where F is the flux, the reference wind speed is 1 m/sec, Fq is a reference dust flux at
U:^ = 0, and 7 is the power in the dust profile. From a series of experiments by Gillette
and Goodwin (1974), a tentative relationship between the parameters in the Gillette-
Shinn model and the soil-erosion index was derived (Fig. 2).

An initial attempt to assess resuspension from the ground was made by Healy and
Fuquay (1958) and Healy (1974) using data from Hilst (1955) and Hilst and Nickola
(1959) who were using zinc sulfide (ZnS) particles to estimate the effects of various
surfaces on wind erosion. At this time the high rate of increase with wind speed due to
breakup of aggregates was not known, and it was assumed that the rate increased as the
square of the wind speed. Later, the rates were converted to a cube relation with the wind
speed (Healy, 1977a), although it was noted that for these results the square appeared to
give less variance in the data. This indicated a pickup rate of 5 x 10~^ u^ per second.
Since the wind speeds were measured at a 2-m height (Hilst, 1955) and the Hanford area,
the site of the experiments, has a Zq of about 2 cm, this value would be about 4 x 10~^ '
u* per second. This can be only a preliminary estimate, however, because of the
alteration of the natural roughness feature by the change in courses.

Sehmel (1977b; 1977c) and Sehmel and Lloyd (1976a) have studied the resuspension
of a tracer, submicrometer calcium molybdate, sprayed as a suspension over a liglitly
vegetated area with a roughness height of 3.4 cm. The area sprayed was within a circle of
29-m radius with a sampling tower in the middle. The average surface concentration on

222 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

400 600 800

SOIL-ERODIBILITY INDEX

Fig. 2 Tentative relation for parameterization of dust flux for the Gillette-Shinn
model.

the ground was 0.62 g of molybdenum per square meter. Measurements were made with a
cowled impactor, which always faced into the wind, at heights to 6.1 m; some
measurements were made at specific wind speeds as measured at a height of 2.1 m above
the ground. Resuspension rates were calculated from a mass balance calculated from the
profile. Those particles depositing in the cowl were considered as "nonrespirable," and
those entering the impactor were considered as "respirable." The impactor separated the
respirable particles into nominal diameters for unit-density spheres of 7, 3.3, 2.0, and 1 .1
/nm and smaller particles on the backup filters.

In these experiments the resuspension rates ranged from about 10~^ ^ to 10~^ sec~^
(Sehmel, 1977a). Hots of all the data showed that the resuspension rates increased with
the 1.0 to 4.8 power of the wind speed. However, these results included wind intervals
with wide speed increments; thus an uncertainty as to the actual wind speed to be used
existed. When these intervals were eUminated, the resuspension rates for all sizes in the
impactor and the backup filter increased with wind speed to the 4.8 power. It may be
noted, however, that these experiments were run over a finite period of time; so
differences in stability would occur. This would cause differences in the wind profile and
a change in u^:. Thus, if the majority of the lower speed winds were in the stable
condition, where u* would be lower than for the neutral condition, and if the majority of
the higher speed winds were in the unstable condition, where u* was greater than for the
neutral condition, an exaggeration of the slope of the curve would occur. Since lower
wind speeds frequently occur in the stable condition and higher wind speeds in the
unstable condition, it can be postulated that such an effect has influenced these
relationships, albeit to an unknown degree. It is noted, however, that dust loadings by the

REVIEW OF RESUSPENSION MODELS 223

same technique appear to increase as the 2.9 power of the wind speed for the higher wind
speeds (Sehmel, 1977c).

The data from the backup filters, wliich showed an increase as the 4.8 power of the
wind speed and for which both the small interval and wide interval of sampling fitted the
curve, gave a fit to

RR= 1.96X 10"'^ u*-^^ (16)

where RR is the resuspension rate. If one assumes that the logarithmic wind profile
existed throughout the period, this becomes

RR = 2 X 10-*^ ut-^2 (17)

when the threshold velocity is ignored.

Sehmel (1972) also measured the resuspension rate of zinc sulfide tracer particles
from an asphalt surface. Average resuspension rates were determined to range from
5 X 10~^ to 6 X 10~* sec~^ for average wind speeds from 2 to 9 mph (0.9 to 4.2
m/sec). The dependency of resuspension rate on wind speed was not determined, but
there was some indication that wind gusts greater than about 15 mph (7 m/sec) rapidly
suspended particles.

In a similar experiment but in an area of cheatgrass, a resuspension rate of

3.4 X 10~^ sec~^ occurred in an area of surface roughness of 4 cm and a friction
velocity of 0.52 m/sec (Sehmel, 1976b). After truck traffic that removed 0.35% of the
ZnS, the surface roughness was reduced to 3 cm, and a resuspension rate of 1 .25 x 10^^
sec~^ was measured in a friction velocity of 0.5 1 m/sec.

Sehmel (1975) measured the resuspension rate of 10-/im-diameter uranine particles
deposited on the inner surface of an aluminum tube with a 2.93-cm inside diameter. The
resuspension rates ranged from 10~^ to 10~^ sec~^ and were dependent on airflow rates
and resuspension time. Orgill, Petersen, and Sehmel (1976) measured the resuspension of
DDT from forests in the Pacific northwest. The DDT was sprayed in the early morning
with low wind speeds, and sampling was conducted by an aircraft -mounted sampler for 5
days. Calculated resuspension rates on three of the sampUng days were 1 .0 x 10~^,

2.5 X 10"^ and 7.7 x 10"^ Sehmel (1976b; 1977c) proposed a correlation between
resuspension rate and the roughness height Zq using the data from the experiments on the
aluminum tube, the ZnS from the asphalt surface, the molybdenum tracer from desert
soil, and the DDT from the forest. This curve shows a decrease in resuspension rate from
the aluminum-tube data to the molybdenum tracer with the asphalt surface falling on the
Unes between these two points. However, the DDT from the forest, with a large roughness
height, had an increase by 2.5 orders of magnitude from the soil data. It is not clear how
these resuspension-rate data were corrected for the differing wind speeds or values of u^^
that existed in each of the experiments. It is further noted that the surfaces and, possibly,
mechanisms of wind pickup from the surface were different. For example, the
resuspension of DDT from the forest could have been primarily a result of the mechanical
movement of leaves, needles, and branches rather than the types of force found on the
soil surface.

There have been many measurements of the air concentration in or near an area
contaminated with radioactive materials, but most of these are not suitable for estimating
resuspension rates or dependency on wind speed because of the lack of detailed
meteorological data at the time of the measurement or the lack of a detailed knowledge

224 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

of the configuration and contamination level of the source. Anspaugh et al. (1975; 1976)
studied the resuspension of plutonium at the GMX Area in Nevada using an ultrahigh
flow-rate sampler so that samples could be obtained in a relatively short time while
meteorological conditions were relatively consistent. The resuspension rates measured
ranged from 2.7 x 10"^ ^ to 4.8 x 10~* ^ see"' . There was a strong correlation between
the resuspension rate and u|. Dividing the resuspension rate by ul greatly reduced the
variance found in the resuspension rate itself. This result is consistent with Shinn's value
for the dependence of dust flux on u* at this location. The value of the normalized ratio
to u^ ranged from 1 .5 X 10~^ ' to 10~^. It is noted that the pattern of plutonium in the
soil used for these derivations was determined by field instrument for the determination
of low-energy radiation (FIDLER) measurements of the ^'*' Am gamma (Healy, 1974)
converted to plutonium concentrations by statistical comparison of the plutonium and
americium content of soil samples (Eberhardt and Gilbert, 1972). This should provide an
estimate of the plutonium in the surface layers of the soil; so correction for plutonium
that has migrated to some depth is not required.

The dust samples used by Shinn et al. (1976) in parameterization of the dust flux in
this area were also analyzed for plutonium (Anspaugh et al., 1976). The profile of
plutonium concentration at a distance of 100 m from ground zero (GZ) agreed with the
shape of the dust flux to a height of about 1 m and then showed lower concentratic^^
than expected. At the 730-m point, the deviation is smaller. The deviations were as would
be expected for a limited source with rather abrupt discontinuities. However, the close fit
of the plutonium concentrations to the dust profile at the lower two heights indicates
that the plutonium concentration at a height less than 1 m is closely coupled to the
ground concentration, even though the soil contamination is less by two orders of
magnitude at the greater distance downwind.

Saltation tluxes were also measured at the GMX Area (Anspaugh et al., 1976). Values
ranged from 3 X 10"'' to 8 x 10"^ g cm~^ sec~\ which are 10~^ to 10""* of those
measured by Chepil (1945a) for wind-eroded fields. In this connection Shinn (1977)
points out that NTS, the location of the GMX Area, is unique in that the natural
resuspension rate owing to convective winds is very low compared with more erodible
sites in the western United States. He has concluded that the natural desert shrub land,
covered by a "desert pavement," or the dry lakes, covered by a crust after a rain, are not
subject to wind erosion unless they are physically disturbed.

Sehmel (1977a; 1977b) has measured the resuspension of plutonium at Rocky Flats
and of plutonium and cesium at the Hanford plant. However, the source areas are poorly
characterized; so resuspension. rates cannot be estimated. Various values of the power of
the increase of concentration with wind speed ranging from unity to 9.3 have been
obtained in these experiments; so it is difficult to draw conclusions from these data.

Mechanical Resuspension. Mechanical resuspension is that caused by forces other
than the wind. Such forces could range from the movement of small animals on the
surface, through humans walking, to the movement of heavy equipment or plows across
the ground. There are several differences between mechanical resuspension and wind
resuspension, chief of which is the fact that the resuspending force is independent of the
wind speed (although dilution downwind will increase at higher wind speeds). Instead, the
resuspension rate will depend on the magnitude of the force applied as well, perhaps, as
on the nature of the force. A second difference is the depth from which resuspension can
possibly occur. In the case of wind resuspension, the layer from which particles can

REVIEW OF RESUSPENSION MODELS 225

contribute to the airstream is limited by the depth to which the saltating particles can
cause ejection. For mechanical disturbance the depths over which the forces can be
applied varies with the means of disturbance but could reach depths of 1 or 2 ft in
plowing. Of course, the probability of resuspension is not the same at all depths, but no
data are available to indicate possible variations. The extreme example would be the
excavation of a hole, such as a basement, where material could be ejected into the air
from considerable depths in the ground.

Another difference arises from the fact that most mechanical disturbances are a point
source; i.e., the disturbance occurs over a fairly Umited area at any one time. There could
be multiple disturbances that could result in an approximate area source or the
disturbance could move with time, which would result in an average that resembled a line
source or an area source. An example of the line source would be traffic moving along a
road. The average resuspension rate from the road would be the product of the
resuspension per vehicle times the number of vehicles divided by the time over which this
number of vehicles passed. An area source would be the average result of a farmer
plowing a field and producing a resuspension rate at each point. Here the average
resuspension rate would be the instantaneous resuspension rate that occurs at each point
divided by the time required to plow the field. Such relationships allow the derivation of
calculational methods for finding the average concentration downwind if the resuspension
rate from the disturbance can be defined.

There are also similarities with wind resuspension. One would expect the dust flux to
be greater in fields that contain a larger quantity of small aggregates. Many of the factors,
such as moisture or vegetative cover that inhibit erosion, would be expected to minimize
mechanical disturbance. Thus, when digging in contaminated soil, it is common practice
to keep the soil damp to minimize resuspension. However, local areas of low saltation
(vegetated strips in the field) will not affect the mechanical-resuspension rate.

A few measurements of mechanical resuspension can be used to give an order-of-
magnitude estimate of the rate of resuspension under different conditions. Sehmel (1972)
measured the resuspension caused by an individual walking along a 50-ft length of asphalt
10 ft wide that had been previously seeded with zinc sulfide tracer. He reports that
1 X 10"^ to 7 X 10"'* (at wind speeds of 3 to 18 mph)of the tracer was resuspended per
walk-through. Assuming a walking speed of 3 mph, this would result in resuspension rates
of 9 X 10~^ to 9 X 10~^ sec"'. Such values have much uncertainty because of the
width of the seeded area, but it is noted that they are about two orders of magnitude
greater than the wind-resuspension rates with wind speeds of 2 to 9 mph.

In a continuation of these same experiments, Sehmel (1973) reports the results of
driving vehicles in the adjacent lane and through the tracer material. Both a car and a
three-quarter-ton truck were used. His values, reported as fractional resuspension per pass,
were converted to a resuspension rate tlirough use of the length of the seeded area and
the speed of the vehicle. The resuspension rate varied from 10~^ to 8 x 10~^ sec~\
depending primarily on the speed of the vehicle. Sehmel (1973) had reported that the
resuspension was proportional to the square of the speed. The resuspension rate caused
by the truck was greater than that caused by the car, presumably because of the greater
turbulence from the truck. A rapid weathering of the particles was also noted. In this
calculation it was assumed that there was no removal by the winds during the 30 days of
the experiment; so the latter rates are the lower Umit of the resuspension rates. However,
the resuspension rate 30 days after application was two to three orders of magnitude

226 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

lower than the initial rate. These data indicate that there will be rapid depletion of the
source for materials deposited on such pavements if any significant traffic occurs.

Sehmel (1976b) also performed a similar experiment with the zinc sulfide tracer
placed on a strip of cheatgrass. The course was the same size as the asphalt course. A
^4 -ton truck was driven through the area at different speeds. The results indicated a
relatively high resuspension rate of 3.7 x 10~^ sec"' on the first pass at 2.2 m/sec. For
the second pass at 6.7 m/sec, the resuspension rate had decreased to 4 x 10~^ sec"^ ,
whereas for the tliird pass at 13.4 m/sec, the rate increased to 8 x 10~^ sec"' . With the
final pass at 17.9 m/sec, the resuspension rate increased to its highest value of 9.4 x 10"^
sec"'. The high resuspension rate at the low speed was caused by removing the most
readily suspendible tracer from the cheatgrass. By the time the higlier speeds were
attained, it is likely that tliis more readily suspendible material was removed and the
resuspension was from the soil surface.

Milham et al. (1976) described the results of air sampling during the agricultural
preparation of two fields having small concentrations of plutonium accumulated 25 to 30
yr earlier as the result of a release from a nearby stack. Samples were taken at several
locations during operations in the field. Healy (1977a) converted these to approximate
resuspension rates by use of the field sizes and meteorological parameters given in the
paper. Tliese results are given in Table 3.

TABLE 3 Resuspension Rates from Agricultural Operations*

Estimated resusper

ision rates,

sec

North field

South field

7.6 m

30.5 m

7.6 m

30.5 m

Milham et al. (1976)

Bush hogging

9x 10-*

2x 10"'

1 x 10-«

8x 10-*

Disking

4x 10-*

6x 10"*

Subsoiling

7x 10"'

3x 10-*

3x 10"*

3x 10-'

Fertilizing

2x 10-*

3x 10-*

1 X 10"*

1 X 10-*

Planting

1 X 10-«

4x 10-*

6x 10"'

2x 10-*

Myers etal. (1976)

Rototilling

9x 10-*

*From Healy (1977a).

A somewhat similar experiment was performed by Myers et al. (1976). Here the
plutonium was applied to a small field in the form of digested sewage-plant sludge
containing a small amount of plutonium. The sludge was allowed to dry for 4 weeks
without rain, and the area was rototilled. Tlie rototiller was 2 m wide with the dust cover
removed and was pulled beliind a tractor. The sampling results were, again, converted to
an approximate resuspension rate by Healy (1977a), and the result is given in Table 3.

Mass Loading

The mass-loading concept is an attempt to bypass the details of the soil characteristics
and the resuspension process and to relate directly measured soil concentrations of the
contaminant to the air concentration by use of the mass of soil particulates in the air.

REVIEW OF RESUSPENSION MODELS 227

Thus the air concentration of the contaminant is given by the product of the
concentration of the contaminant in the soil and the concentration of the soil particulates
in the air. If the quantity of particulates in the air is known from other data, one need,
theoretically, only measure the soils in the region to provide an estimate of the air
concentration of the contaminant.

Two parameters, the dust loading in the atmosphere and the appropriate concentra-
tion of the contaiTiinant in the soil, are needed to provide estimates by this method.
Healy (1974) used an average value of 120 jUg/m^ of dust in a generic analysis of limits
for plutonium in the soil. Tliis was derived from the Federal Secondary Standard for
particulates in the air expressed as a geometric mean of 60 A^g/m^ assuming a geometric
standard deviation of 2. Anspaugh (1974) explored a reasonable mass loading in several
ways. Tlie lower bound is quoted as about 10 i^g/m^ . Examination of the data on the
levels in mine atmospheres which have led to a considerable prevalence of pneumoco-
niosis in the workers indicates that standards on the range of 1 to 10 mg/m^ have a very
small, if any, margin of safety. Anspaugh (1974) also quotes some British data which
indicate that dust levels in excess of 1 mg/m^ could lead to considerable public health
problems. He also used the data on ambient mass loading for 1966 from the National Air
Surveillance Network to show that the average for urban stations ranged from 33 to 254
jUg/m^ with a mean for all nonurban locations of 38 idg/m^ . For the nonurban stations
the average ranged from 9 to 79 idg/m^ . From these studies he chose an average of 100
jUg/m^ as reasonable for predictive purposes (Anspaugh, Sliinn, and Wilson, 1974;
Anspaugh et al., 1975). The U.S. Environmental Protection Agency (1977) also
examined the data from the nonurban stations from the National Air Surveillance
Network for the years of 1964 and 1965. Their map of these data indicates values ranging
from 9 iUg/m^ in southern Montana to 56 /Ug/m^ in western Pennsylvania, 57 Mg/m^ on
the southern Oregon coast, and 59 lug/m^ on the North Carolina coast. However, the
prevalence of high values in the east would indicate the possible inclusion of industrial
particulates in these samples. The U. S. Environmental Protection Agency used a value of
100 jUg/m^ in the calculation of their screening level.

Several uncertainties appear in the use of the data from the National Air Surveillance
Network. The first was pointed out above in that the particulates that are collected can
include a portion of those generated by industrial operations; so the values could be high.
The second problem arises from the fact that the samplers are frequently in positions,
such as on top of buildings; so they do not measure the air actually breathed by people.
Associated with this question is the potential for people engaged in various activities to
generate their own dust. Tliis would result in local concentrations in excess of the
ambient value measured by the network. However, the 100 /ig/m^ value still appears
reasonable from the standpoint that it is an average over a full year, and people do not
work or play in dusty operations all the time. For example, if we assume that an
individual spends 8 hr per day, 5 days per week, 50 weeks per year in a concentration five
times the maximum value noted in the ambient air measurements, or 300 jug/m^ , the
average concentration through the year would be only 115 )Ug/m^ . Although some
individuals, such as farmers, work longer hours during the week, their exposure to dust is
limited to fewer weeks per year, and a portion of their time in the field is during periods
of high moisture or vegetation in the soil when dusty conditions are limited.

Associated with the question of the concentration of soil particles in the air is the
question of the origin of the particles. Once airborne, the smaller particles can travel very

228 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

long distances. For example, Carlson and Prospero (1972) have reported the movement of
dust from the Sahara desert over the northern equatorial Atlantic Ocean, and Clayton
et al. (1972) have reported evidence of transport across the Pacific Ocean to Hawaii.
Since most contaminated areas are relatively small in area, one would expect that only a
fraction of the dust in the air would originate from the area. Because of preferential
deposition of the larger particles from sources some distance away, this "background
dust" would contain a higher percentage of the smaller particles that are more readily
deposited in the lung than the dust originating from the local, contaminated area. Thus
Anspaugh and Phelps (1974) report that measurements at the GMX area with Anderson
high-volume cascade impactors for about 1 month indicate that the mass distribution of
sizes is about 1.6 iJ.m MM AD with a geometric standard deviation (a^) of about 15, and
the plutonium and ^"^^ Am had an activity median aerodynamic diameter of about 3 idm
with a Og of about 7. It was also noted that the average activity of the soil was about
one-third that found in the soil in close proximity to the sampler. It is noted that even
higher activity was upwind.

From this we conclude that a direct comparison of the size distribution of
contaminated particles in the air with those in the soil is probably valid only for very
large areas. For the more usual size of contaminated area, the dilution of the total mass in
the air, particularly in the smaller particle sizes, could be significant. It is noted, however,
that, for resuspension by mechanical disturbance, this dilution may be of lower
importance because of the frequently higher concentrations resulting from such
disturbances.

The second question, that of the appropriate concentration of the contaminant in the
soil, is more subtle. As was discussed earlier, soil particles that are carried in suspension
are the smaller ones because the larger ones will settle rapidly. Tlius, if the concentration
of the contaminant in the soil fraction containing the small particles is greatly different
from that in the other particle sizes, it would appear that the concentration predicted by
the mass-loading approach using the total soil concentration would theoretically be low.

Tamura (1977) has analyzed the particle sizes and their associated plutonium content
in samples from several existing plutonium-contaminated areas and has shown that
fractionation of the plutonium content by particle size does exist. Analyses were done
using water as the suspending agent, and the effect of this, as compared with the carbon
tetrachloride used by Chepil or the liquid Freon used by Gillette, on the aggregate size is
unknown. However, in two samples from the NTS, the aggregates less than 20 ;um had
plutonium concentration three and five times greater than the total soil mass. (At the
NTS the bulk of the activity appears to be in the 20- to 53-iJim size range.) In a bottom
sediment from the canal at Mound Laboratory, the concentration in the fraction lower
than 20 /im was 1.8 times as high as the total; at the floodplain at ORNL, the soil
concentration in the fraction lower than 20 ^m was 1.1 times as higli as the total; and, in a
sample from Rocky Flats, the soil concentration in the fraction less than 20 jim was
about 3 times as high as the total. It is of interest that these distributions reflect both the
method of contamination and the soil type. At the NTS the plutonium was mechanically
dispersed by explosive material, and the pa'ticle size distribution reflects the largest
amount of the plutonium in the 53- to 125-/jm size range, although the higliest
concentration was in the smaller particle sizes. The Mound Laboratory and ORNL
samples reflect the distribution expected by adsorption on the smaller particles in the
sample, whereas Rocky Flats is intermediate, wliich retlects, perhaps, some adsorption as

REVIEW OF RESUSPENSION MODELS 229

well as direct contamination of the larger soil particles by the plutonium-bearing oil that
was the source of the contamination.

Tlie U. S. Environmental Protection Agency (1977) proposed the use of an
"enrichment factor" to include these data in resuspension calculations. This is defined as
the summation of the products of gi , the ratio of the fraction of the total activity
contained witliin the size increment i to the fraction of the total mass in the size, and f,
the fraction of the airborne mass within each increment of particle size in the air. For the
distribution in soil sizes at Rocky Flats, they calculate an enrichment factor of 1.5.
Tamura (1977) has defined a "soil plutonium index" which accounts for the size
distribution as well as the lung deposition. This is given as

SI = SA X LD X RA (18)

where SI = soil plutonium index
SA = soil activity factor
LD = lung deposition factor
RA = resuspendible activity factor

Tlie soil activity factor is the fraction of the activity in a given mass fraction divided by
the mass fraction for particles less than 100 /jm. (Tamura used 125 iJ.m in evaluating this
factor because this was the sieve size used in his analysis.) Values of this factor range from
3.14 for the ORNL sample to 7.27 for the Rocky Flats sample. The lung deposition
factor is the deposition in the pulmonary region as defined by the International
Commission on Radiological Protection (1966). The final factor, the resuspendible
activity factor, is the fraction of the total soil plutonium index activity in the
resuspendible fraction. Indexes derived from available data give 0.52 for Area 13 at NTS,
1.26 for Rocky Flats, 1.18 for Mound Laboratory, and 0.69 for ORNL.

Another approach to the use of the smaller particles is that of Johnson, Tidball.and
Severson (1970). Their sampling technique was to brush the surface dust into a container.
The 5-/jm or smaller particle sizes were then separated from the sample after aggregates
had been broken up, and plutonium analyses were performed on this fraction. They
found that the concentration in these small particles was 4 to about 300 times as large as
that in similar samples taken to a depth of Vg in. Their conclusion was that these results
provided a better indication of the hazard than the conventional sample, although they
did not explore mechanisms of breaking down the aggregates found in the soil, which
severely limit the quantities of particles of this size found in natural soils, nor did they
examine pathways o\' this material to man.

These relations have never been actually tested to show their validity. The work on
soil erosion indicates the many additional factors that will influence wind erosion and
resuspension. These include the soil texture, the moisture content o\' the soil, the
presence or absence of vegetation in vegetative residues, and the characteristic surface
roughness. In the case of NTS. the desert pavement undoubtedly has more influence on
either wind or mechanical resuspension than the other factors. We would believe that
wind erosion, in particular, is more complex than these relations would indicate.
However, it is possible that such concepts may be more applicable to mechanical
disturbance.

A direct test of the mass-loading technique has been made by Anspaugh, Shinn,and
Wilson (1974) and Anspaugh et al. ( l'-)75). The measured concentrations of a number oi'

230 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 Comparison of the Predicted Concentration in Air Using a Mass Loading
of 100 idg/m^ with Measured Concentrations*

Nuclide

Air concentration

Location

Predicted

Measured

Nevada Test Site

NE, Anspaugh and Phelps (1974)

"'Pu

7.2 X 10-^ pCi/m'

6.6 X 10-3 pCi/m^

GZ, Anspaugh and Phelps (1974)

2 3 9pu

0.12pCi/m'

0.023 pCi/m^

Lawrence Livermore Laboratory

GudiksenetaL (1973a)

238U

150pg/m^

52pg/m3

Gudiksenetal. (1973b)

238U

150pg/m^

lOOpg/m^

Silver etal. (1974)

238U

150pg/m^

86 pg/m^

SilveretaL(1974)

""K

10-3 pCi/m^

9.8 X 10-* pCi/m'

Argonne National Laboratory

Sedlet, Golchert, and Duffy

232-pi^

320 pglm'

240 pg/m'

(1973)

Sedlet, Golchert, and Duffy

Natural

215 pg/m^

170pg/m^

(1973)

uranium

Sutton, England

Hamilton (1970)

Natural
uranium

llOpg/m^

62 pg/m^

*Based on data from Anspaugh, Shinn, and Wilson (1974) and Anspaugh et al. (1975).

nuclides in the air were compared with a concentration calculated from the quantity of
the nucUde in the soil assuming a mass loading of 100 /ig/m^. These results are given in
Table 4.

The agreement between calculated and predicted values is good. Of course, the
sources for the natural isotopes are large in area. However, the values at NTS show
reasonable agreement between calculation and prediction. It is believed that the soil
concentration at the point of sampUng was used for the predicted values. If this were the
case, the discrepancy between the two results at GZ is explainable on the basis that the
concentration in the soil is highest at this point and the measured dust arose from
surrounding areas of lower concentration.

Discussion

This review covered concepts and numerical values related to tlie resuspension problem
and did not include the important conceptual and modeling studies that have been carried
out by several individuals, including Amato (1971), Trevino (1972), Horst (1976), and
Slinn (1978). This was done dehberately in order to focus on the nonmathematical
aspects of the problem and to attempt to bring the factors of importance into focus.

It is apparent that a gratifying amount of progress has been made on the
determination of resuspension in the past few years. The studies by Anspaugh, Shinn, and
Wilson (1974) and Anspaugh etal. (1975) at NTS have shown the feasibiUty of
measurement of the resuspension rate in a contaminated area, and their application of the
mass-loading concept has added greatly to the understanding of this model. The work of
Gillette (1974; 1976) and Shinn et al. (1976) on the dust flux has given new insights into
methodology and the phenomena concerned. The studies by Selimel (1977b; 1977c),

REVIEW OF RESUSPENSION MODELS 231

with the tracer particles, have given values that are extremely useful for appUcation. Slinn
(1978) has provided parameterization concepts that aid in understanding.

This is not to say that additional work is not needed. Further studies of both the dust
flux and resuspension rate at contaminated areas in various regions and types of soil are
definitely needed along with models, such as those of Gillette (1974) and Shinn et al.
(1976), which provide relationships between the resuspension and readily measured
parameters that can be used to estimate resuspension rates. Concurrent studies of the dust
flux and resuspension of a contaminant are badly needed, particularly in undisturbed
areas apart from agricultural soils. The dust-flux model requires assumptions as to the
connection between resuspension of a contaminant and the dust flux. The only
checkpoint now available is the measurements at the GMX Area of the dust flux by Shinn
etal. (1976) and the plutonium resuspension rate by Anspaugh et al. (1975). In
particular, additional data are needed on resuspension by mechanical disturbance. Few
appropriate experiments are available, and it is frequently difficult to interpret them in a
manner that provides useful results. A particular area of concern for which very few data
are available is the possibility of contamination while playing and working in an area with
subsequent transfer to a place where inhalation is more probable. Extreme examples of
this would be pulling a contaminated garment over one's head or contaminating pillows
or other bed clotliing. Although one could feel that this could not be a major source of
exposure, we cannot tell until the experiments are done.

A primary purpose of this chapter is to choose resuspension parameters to be used in
the calculation of the dose to individuals in a contaminated area. The study has
reinforced our previous prejudice that the resuspension factor is not the method for use
because of the failure of this method to account for many of the variables and because
the conditions of the measurement are seldom described in sufficient detail to allow
intelligent extrapolation to areas different from those in which the measurements were
made. There is some usefulness to this technique, however, in describing the exposure of
the individual causing the disturbance.

The resuspension rate has been our favorite method because of the capability of
integrating over a contaminated area using accepted dispersion and deposition parameters
to provide concentration isopleths around the area. For a specific situation in which the
soil and meteorological parameters can be defined, this is still the preferred method, and
the state of the art is rapidly approaching sufficient detail that this can be done.

However, for a generic study the mass-loading approach seems to be best. The work
of Tamura (1977), the U. S. Environmental Protection Agency (1977), and Johnson,
Tidball, and Severson (1970) all indicate that even this approach requires revision for the
distribution of contamination in the soil. However, as has been pointed out, there are
factors that tend to compensate for this, such as the size of the area, and the magnitude
of the correction factor proposed by the U. S. Environmental Protection Agency (1977)
and Tamura (1977) is only on the order of 1.5 to 2, a value that tends to get lost in the
noise of the other uncertainties. In addition, the success shown by Anspaugh (1974) in
predicting the concentrations of several nucHdes in widely different climates and soil
types is encouraging.

Anspaugh (1974) used a mass loading of 100 jUg/m^ in his comparison. The measured
values were primarily ambient air and included no component for mechanical
disturbance. In view of the agreement found in his study, we would propose the use of
200 Afg/m^ for generic studies to make allowance for these other types of exposure. This

232 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

seems to be unrealistically high when compared with air-samphng results. This may well
be due to the factors proposed by Tamura, with the actual mass loading of little
importance as compared with the correlation found by Anspaugh using an arbitrary value
of 100iUg/m^

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REVIEW OF RESUSPENSION MODELS 233

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Models for Internal Dosimetry of the Human Respiratory Tract, Health Phys., 12: 173-207.
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Surface of Soil, Science, 193: 488.
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234 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Milliam, R. C, J. F Schubert, J. R. Watts, A. L. Boni, and J. C. Corey, 1976, Measured Plutonium
Resuspension and Resulting Dose from Agricultural Operations on an Old Field at tlie Savannah
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Mishima, J., 1964, A Review of Research on Phi ionium Releases During Overheating and Fires,
USAEC Report HW-83668, Hanford Laboratories, NTIS.

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Olafson, J. H., and K. R. Larson, 1961, Plutonium, Its Biology and Environmental Persistence, USAEC
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Prestley, C. H. B., 1959, Turbulent Transfer in the Lower Atmosphere , Tlie University of Chicago
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Sedlet, J., N. W. Golchert, and T. L. Duffy, 1973, Environmental Monitoring at Argonne National
Laboratory, Annual Report for 1972, USAEC Report ANL-8007, Argonne National Laboratory,
NTIS.

Sehmel, G. A., 1972, Tracer Particle Resuspension Caused by Wind Forces upon an Asphalt Surface, in
Pacific Northwest Laboratory Annual Report for 1976 to the USAEC Division of Biology and
Medicine, Part 1, Atmospheric Sciences, USAEC Report BNWL-1651(Pt.l), pp. 136-138, Battelle,
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, 1973, Particle Resuspension from an Asphalt Road Caused by Car and Truck Traffic, Atmos.

Environ., 1: 291-309.

— ^, 1975, Initial Correlation of Particle Resuspension Rates As a Function of Surface Roughness
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, 1976a, The Influence of Soil Insertion on Atmospheric Particle Size Distributions, in Pacific

Northwest laboratory Animal Report for 1975 to the USERDA Division of Biomedical and
Environmental Research, Part 3, Atmospheric Sciences, ERDA Report BNWL-2000(Pt. 3),
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- — , 1976b, Particle Resuspension from Truck Traffic in a Cheat Grass Area, in Pacific Northwest
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Pacific Northwest Laboratories, NTIS.

, 1976c, Particle Resuspension from an Asphalt Road Caused by Car and Truck Traffic, m

Atmosphere-Surface Exchange of Particulate and Gaseous Pollutants (1974), ERDA Symposium
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, 1977a, Plutonium and Tracer Particle Resuspension: An Overview of Selected Battelle-

Northwest Experiments, in Transuranics in Naniral Environments, Symposium Proceedings,
Gatlinburg, Tenn., Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report
NVai78, pp. 181-210, Nevada Operations Office, NTIS.

— -, 1977b, Transuranic and Tracer Simulant Resuspension, ERDA Report BNWL-SA-6236, Battelle,
Pacific Northwest Laboratories, NTIS.

, 1977c, Plutonium and Tracer Particle Resuspension: An Overview of Selected Battelle-

Northwest Experiments, in Transuranics in Natural Environments, Symposium Proceedings,
Gatlinburg, Tenn., Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report
NVO-128, Nevada Operations Oftlce, NTIS.

REVIEW OF RESUSPENSION MODELS 235

, 1977d, Radioactive Particle Resiispension Research Experiments on the Hanford Reservation,

ERDA Report BNWL-2081, Battelle, Pacific Northwest Laboratories, NTIS.
- — , 1978, Plutonium Concentrations in Airborne Soil at Rocky Flats and Hanford Determined

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1974, pp. 846-858, CONF-740921, NTIS.
— -, and F. D. Lloyd, 1976b, Resuspension of Plutonium at Rocky Flats, in Atmosphere-Surface

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757-779, CONF-740921, NTIS.
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Pacific Nortliwest Laboratories, NTIS.
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Environmental Levels of Radioactivity in the Vicinity of the Lawrence Livermore Laboratory ,

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(Coordinators), pp. 906-944, CONF-740921, NTIS.
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602-608.

Transuranic and Tracer Simulant
Resuspension

G. A. SEHMEL

Plutonium resuspension results are summarized for experiments conducted at Rocky
Flats, on site on the Han ford reservation, and for winds blowing from off site onto the
Hanford reservation near the Prosser barricade boundary. In each case plutonium
resuspension was shown by increased airborne plutonium concentrations as a function of
either wind speed or as compared with fallout levels. All measured airborne concentra-
tions were below maximum permissible concentrations.

Both plutonium and cesium concentrations on airborne soil were normalized by the
quantity of airborne soil sampled. Airborne radionuclide concentrations (in microcuries
per gram) were related to published values for radionuclide concentrations on surface
soils. For this ratio of radionuclide concentration per gram on airborne soil divided by
that for ground-surface soil, there are seven orders of magnitude uncertainty from 10"^
to 10^ . This uncertainty in the equality between plutonium concentrations per gram on
airborne and surface soils is caused by only a fraction of the collected airborne soil being
transported from off site rather than all being resuspended from each study site and also
by the great variabilities in surface contamination.

Horizontal plutonium fluxes on airborne nonrespirable soils at all three sites were
bracketed within the same three to four orders of magnitude from 10~'^ to 10~^ iiCim" ^
day~^ for ^^^Pu and 10'^ to 10~^ idCi m~^ day~^ for ^^^Pu. These represent the
entire experimental base for nonrespirable airborne plutonium transport.

Airborne respirable ^^^Pu concentrations increased with wind speed for a southwest
wind direction coming from off site near the Hanford reservation Prosser barricade.
Airborne plutonium fluxes on nonrespirable particles had isotopic ratios,
^^ ^Pu/^ ^ ^ ^'^^Pu, similar to weapons-grade plutonium rather than to fallout plutonium.

Resuspension rates were summarized for controlled inert-particle-tracer simulant
experiments. Wind resuspension rates for tracers increased with wind speed to about the
fifth power. This wind-speed dependency is comparable to that measured for off-site
plutonium resuspension near the Prosser barricade However, plutonium resuspension
data near the U-Pond Area showed an air concentration dependency on wind speed to the
1.5 power. There is still uncertainty in the wind-speed dependency of airborne
concentrations at different sites.

The weathering half-life is the average time required for airborne concentrations from
resuspension sites to decrease by one-half when airborne concentrations are averaged over
all meteorological conditions. Airborne plutonium and cesium concentrations measured
at Hanford as well as tracer resuspension experiments show that the weathering half-life is
much greater than that usually reported in the literature: 5 months or much longer rather
than only 35 to 45 days

Resuspension rates for local mechanical resuspension of inert tracer particles caused
by vehicular and pedestrian traffic are summarized.

236

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 237

Resuspension occurs when particles on a surface are disturbed and carried up into the air
by air currents. Wind-caused resuspension is the process by which wind blows particles
from a surface into the air and transports them downwind. For radionuclide-
contaminated surfaces, wind might cause radionuclide particles to be resuspended and
transported to other sites. Resuspension occurs at radionuclide-contaminated sites on the
Hanford reservation in Washington (Sehmel, 1977c; Pacific Northwest Laboratory,
September 1973-October 1974), at Rocky Flats in Colorado (Johnson, Tiball, and
Severson, 1976; Krey et al., 1976b; 1976c; Sehmel, 1976a; Sehmel and Lloyd, 1976b;
Volchok, Knuth, and Klemman, 1972), at the Nevada Test Site (Anspaugh et al., 1969;
Wilson, Thomas, and Stannard, 1961), at the Savannah River Laboratory reservation in
South Carolina (Milham et al., 1976), and at other sites (Mishima, 1964; Stewart, 1967).
However, with our present knowledge (Horst, 1976; Oksza-Chocimowski, 1976), amounts
of wind-caused resuspension and its effects cannot be adequately predicted.

Radioactive particles deposited on natural or man-made surfaces are resuspended by
both wind and mechanical activity. Wind resuspension can occur over a wide area as well
as over a local area. In contrast, mechanical-activity resuspension is usually more localized
and can present an immediate inhalation problem to the worker in a contaminated zone.
Although mechanical activity is frequently at a point, integration of mechanical activity
over time could result in an area source. For example, an area source could be generated
during the plowing of a field. In both wide-area and local resuspension of radionuclide
particles, particles transported downwind could become a potential radiological concern
to man. Sources for resuspended particles include radioactive fallout as well as releases
from nuclear faciHties. At present the significance of fallout resuspension is unknown.
Data are needed to define the relative inhalation hazard of fallout-particle resuspension
vs. the direct delivery of stratospheric debris.

Radioactive-particle resuspension is probably more important at nuclear faciHties
where the surrounding environment has been contaminated with radioactive particles.
These particles can be resuspended by both wind stresses and mechanical disturbances.
However, resuspension mechanisms are poorly understood, and consequently resus-
pension rates and . potential airborne inhalation hazards cannot now be adequately
predicted.

The need for such predictions is not new: for many years resuspension has been
known to be occurring at nuclear sites. Some of the earliest data were obtained
(Anspaugh et al., 1969; Wilson, Thomas, and Stannard, 1961) at the Nevada Test Site.
Ground radioactivity contours were determined as a function of time after a test
detonation. Initially, ground-surface concentrations were caused by plume deposition.
Subsequent ground radioactivity contours showed (Anspaugh et al, 1969) a migration of
radionucUdes from the Test Site which indicated that resuspension had occurred.
Similarly, aerial surveys at Hanford (Bruns, 1976) have shown transport of ^"^^ Am by
wind resuspension.

Resuspension is of considerable interest at the Rocky Flats nuclear plant in Colorado
where ground surfaces were contaminated with plutonium from leaking storage barrels
containing plutonium-contaminated cutting oil (Johnson, Tiball, and Severson, 1976;
Krey et al., 1976b; 1976c; Sehmel, 1976a; Sehmel and Lloyd, 1976b; Volchok, Knuth,
and Klemman, 1972). After the leakage was discovered, the barrels were removed and
corrective actions were taken, but plutonium resuspension from residually contaminated
soil surfaces is still occurring.

238 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

More recently, resuspension has been reported at study sites on the Hanford
reservation (Sehmel, 1977c; Pacific Northwest Laboratory, September 1973— October
1974). These sites were low-level liquid-waste disposal sites.

Although environmental plutonium resuspension is receiving attention, resuspension
physics is poorly understood. Resuspension was early characterized by a "resuspension
factor." The resuspension factor is defined as the ratio of airborne pollutant
concentration (amount per cubic meter) at breathing height divided by the ground-
surface contamination level (amount per square meter). Thus the resuspension factor has
units of meters" ^ . Reported resuspension factors vary many orders of magnitude with
values from 10"* ^ up to 600 m"* (Mishima, 1964; Stewart, 1967; Sehmel and Lloyd,
1976a). Resuspension-factor variations have not been adequately explained as a function
of experimental conditions.

Resuspension factors from about 10"^ to 10"^ m" * are often used in hazard
evaluations. The resuspension factor is useful since a worker's inhalation hazard is most
Ukely related to the local resuspension caused by his work activities within a
contaminated zone; however, resuspension factors are only a very rough estimate of the
potential airborne contaminant concentration since resuspension factors cannot be
accurately predicted. In addition to local resuspension, airborne contaminated particles
can reach workers from upwind contaminated areas. Hence both local and upwind
resuspension should be considered, but resuspension factors in either case cannot be used
in downwind transport models.

The resuspension factor is an index of only the potential inhalation concentration and
not the total resuspension release rate from a surface-contaminated area. Resuspension
release rates are needed for source terms in calculating total downwind diffusion and
transport of resuspended particles. Only recently have particle resuspension rates been
measured (Sehmel, 1973b; 1975; 1977b; Sehmel and Lloyd, 1976a; 1976c).

The objective of this chapter is to summarize reported resuspension rates (Sehmel,
1976a; Sehmel and Lloyd, 1976b) and parameters (Sehmel, 1977b; 1977c) determined at
the Pacific Northwest Laboratory between 1971 and early 1977. These include
plutonium resuspension measurements at Rocky Flats and at Hanford as well as results
from controlled tracer simulant source resuspension experiments.

In these experiments airborne concentrations were measured as functions of wind
speed, airborne particle size, and wind direction, and the collected radionuchdes or tracer
simulants were determined per gram of airborne soil or solids. Particulate air samples were
collected as a function of wind speed to determine whether airborne radionuclide
concentrations increased at higher wind speeds, and concentrations as a function of
particle size were measured to determine the distribution of radionuclide particles
resuspended as individual particles or attached to host soil and sohd particles. In addition,
airborne radionuclides were normalized by the total amount of airborne solids to relate
concentration per gram of airborne soHd to concentration per gram of radionuclide on
the ground.

Experiments

The experiments for measuring particle resuspension reported here have been reported in
fuller detail in the following references:

• Plutonium and americium from resuspension study sites at Hanford (Sehmel,

1977c) (Fig. 1).

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 239

HANFORD
METEOROLOGICAL
STATION TOWER

U-POND AREA^X
S-16

GABLE

MOUNTAIN

POND

Fig. 1 Location of Hanford study sites.

ORIGINAL Security

°'L FENCE -V

STORAGE \

AREA- ■^

' 61

0.3 D— |/A2
0.3 D-|' 30
^A3

SITE A

V- CATTLE
\ FENCE
14 \|

SITE B

0.0, SAMPLING ELEVATION, m
0.0, DISTANCE BETWEEN SITES, m

SAMPLER OPERATION

O CONTINUOUS, ALL WINDS
A 4.1- to 6.3-m/sec WINDS
< 6.3- to 9.8-m/sec WINDS

CASCADE IMPACTOR
WITH COWLS, 0.57 m^/min

n CONTINUOUS HIGH-VOLUME SAMPLER, 1.4 m-^/min

Fig. 2 Rocky Flats tower locations.

• Plutonium from contaminated environmental surfaces at Rocky Flats (Sehmel,
1976a: Sehmel and Lloyd, 1976b) (Fig. 2).

• Controlled inert simulant tracer particles from selected surfaces on the Hanford area
(Sehmel, 1973b; 1975; 1976c; 1977b; Sehmel and Lloyd, 1976a; 1976c).

Plutonium resuspension results from off-site Hanford near the Prosser barricade are
reported here for the first time. Most wind-caused resuspension research concerns
resuspension from vegetated areas. Experiments concerning local resuspension caused by
mechanical activity include tracer studies of resuspension rates for a man walking across
an asphalt strip and for cars and trucks driven on asphalt or cheat grass.

240 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Two different resuspension rates are used. For wind-caused resuspension, resus-
pension rates are reported as the fraction of particles resuspended per second. Thus the
total wind-caused resuspension is a product of the surface contamination level, the
duration of resuspension, and the resuspension rate. For local mechanical disturbances by
vehicular or pedestrian traffic, resuspension was measured each time a car, ^4-ton truck,
or person passed across the length of a 3-m-wide tracer-contaminated area. Thus traffic
resuspension rates are reported as the fraction of particles resuspended per pass.

Particles

Resuspension was measured for several types of particles. The plutonium particle size
distributions on soils at Rocky Flats and Hanford were uncontrolled. A forest-spray
operation provided an opportunity to measure resuspension of DDT as tracer particles
not specifically controlled for size. The controlled, inert tracer particles used were
submicrometer CaMo04 particles and ZnS particles with an 8-//m mass aerodynamic
equivalent diameter.

Air Samplers

Airborne resuspended particles were either sampled with total air samplers* or sized while
airborne with particle cascade impactors.f Particle cascade impactors were used for
plutonium and CaMo04 particles.

The particle cascade impactor for sampling respirable particles was attached to a
rotating cowl, which allowed simultaneous sampling of larger nonrespirable particles. The
cowl-impactor system (Sehmel, 1973a) shown in Fig. 3 was evaluated by Wedding,
McFarland, and Cermak, 1977. Particles entering the 15-cm-diameter cylindrical sampler
inlet of the cowl either settled on the cowl floor or were drawn up into the impactor.
Particles settling on the cowl floor are called "nonrespirable" in this chapter. Respirable
particles entering the particle cascade impactor were separated into nominal aerodynamic
diameter ranges of 7, 3.3, 2.0, and 1.1 m, which are impactor stage 50% cutoff diameters
for unit-density spheres. Smaller particles were collected on an impactor backup filter.

Results and Discussion

Airborne radionuclide concentrations were determined at transuranic resuspension study
sites at Rocky Flats, Colo. (Sehmel, 1976a; Sehmel and Lloyd, 1976b) and the Hanford
area in Washington (Sehmel, 1977c; Pacific Northwest Laboratory, September 1973-
October 1974). In addition, some cesium resuspension data are reported for Hanford
(Sehmel, 1977c). In contrast to transuranic resuspension, tracer simulants (Sehmel,
1977b) were used to determine particle resuspension rates. Results for each set of
experiments are discussed separately.

Radionuclide-Particle Resuspension On Site

Airborne plutonium concentrations at Rocky Flats and Hanford were measured as a
function of particle diameter, wind speed, and sampUng site. Radionuclide concentrations
per gram of airborne solid were determined.

^General Metal Works, Inc., model GMWL-2000-high-voltage air sampler with filter holder,
t Andersen 2000, Inc., model 65-100 high-volume sampler head.

TRANSURANIC AND TRACER SIMULANT RES USPENSION 24 1

HIGH-VOLUME SAMPLER

SYSTEM
SUPPORT

ARM

WIND-

CYLINDRICAL
SAMPLE INLET

'}>~~^£-—^^aU CASCADE IMPACTOR

CYLINDRICAL
COWL BODY

•SPINDLE EXTENSION

WIND-ORIENTATION
TAIL FIN

SPINDLE BEARING
ASSEMBLY BOLT

WIND-
DIRECTION
SENSITIVE
ROTATING
COWL

Fig. 3 Rotating cowl and impactor.

Plutonium Resuspension Research at Rocky Flats. Plutonium resuspension at Rocky
Flats was investigated experimentally (Sehmel, 1976a; Sehmel and Lloyd, 1976b). In
early work an empirical resuspension model was developed (Sehmel and Orgill, 1973)
which was based on published weekly plutonium concentrations at Health and Safety
Laboratory sampling station S-8 along the site's eastern security fence. The plutonium
data were analyzed in terms of the meteorology during sampling times. Collected airborne
plutonium was related to hourly average wind speeds and wind directions. Model results
showed that airborne plutonium concentrations increased as the 2.1 power of wind speed.
Subsequently airborne concentrations were predicted for the succeeding time period.
These results showed a wide difference between predictions and experimental results. The
interpretation of these differences was that the plutonium resuspension source charac-
teristics had changed (Sehmel and Orgill, 1974).

Battelle— Northwest experimental measurements of plutonium resuspension at Rocky
Flats were made in July 1973 (Sehmel, 1976a; Sehmel and Lloyd, 1976b). As shown in
Fig. 2, airborne plutonium concentrations were measured at three sampling sites east of
the plant. The first sampling site was along the eastern security fence. This site was called
sampling site A. Sampling site B was along the eastern cattle fence, and sampling site AB
was between sites A and B. The distance from site A to site AB was 227 m. Airborne
plutonium at these sites was sampled and analyzed as a function of sampling height,
particle size, and wind speed. For comparison, a particle cascade impactor sample was
simultaneously collected at a background site 13 km west in the mountains.

242 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Fallout levels of ^^^Pu entering the area were estimated from the cascade impactor
operated in the mountains. There was no detectable ^■'^Pu activity on the 7-, 33-. and
1.1 -)nm impactor stages, and there was no radiochemical result for tlie l-jim stage. The
only detectable background plutonium activity was on the backup filter that nominally
collects submicrometer particles. Airborne ^^^Pu concentration at the background
station was 4 ± 3.5 X 10" ' ^ AtCi/cm^ , which corresponds to 0.7 ± 0.62 X 10~^ /jCi/g of
airborne soil on the backup filter. Error limits are the 2a radiochemical counting limits.

In on-site research, airborne particles were separated in the sampling process into two
main fractions. One sample contained particles collected by gravity settling in the inlet
cowl section of the sampler as shown in Fig. 3. The second fraction contained those
particles passing through the inlet section and collected within the high-volume cascade
impactor. The smallest particles collected in the inlet cowl section were about lO/imin
diameter. This fraction was assayed for ^^^Pu and ^^^Pu. In some cases nonrespirable
particles were sieved into smaller size fractions, and these fractions were also assayed for
^^^Pu and ^'^^Pu. Data for respirable and nonrespirable particles are discussed separately.

Respirable Plutonium Concentrations at Rocky Flats. Airborne •^ ^ ^ Pu concentrations at
the three Rocky Flats sampling stations were reported (Sehmel and Lloyd, 1976b) in
microcuries per cubic centimeter of air and microcuries per gram of airborne soil. The
maximum airborne ^^^Pu concentration was 3.7 X 10^' ^ jnCi/cm^ . The maxima. u
^^^Pu concentration on the airborne soil was 5 x 10~^ AiCi/g total airborne soil and
7 X 10~^ AiCi/g for the respirable fraction of airborne soil collected on the 2-/nm-particle
impactor stage. All airborne ^^^Pu concentrations were significantly less than MPC's of
soluble ^^^Pu in air for occupational exposure in a 40-hr work week (2 x 10~^^
AtCi/cm^) or nonoccupational exposure in a 168-hr week period (6 X 10~^^ juCi/cm^)
(International Commission on Radiological Protection, 1959).

Airborne plutonium concentrations were a function of both sampling height and
particle diameter. Airborne concentrations are shown for site AB in Fig. 4 for each
particle cascade impactor stage. In contrast to simple modehng concepts, airborne
concentrations did not always decrease with an increase in height. There were
unexpectedly high ^^^Pu concentrations at this site for several particle diameters and
heights.

Plutonium was associated with particles collected on each particle cascade impactor
stage. Since there was no plutonium in the upper stages of the impactor at the
background mountain site, the ^^^Pu found in upper stages of impactors at Rocky Flats
sampling sites indicates that some plutonium was resuspended while attached to larger
particles. Resuspension of submicrometer particles also occurred at Rocky Flats.

The general trend of the complete airborne ^^^Pu concentration data is a decrease in
concentration with increasing distance eastward from site A (Sehmel and Lloyd, 1976b).
As might be expected, this decrease in concentration corresponded to increasing distance
from the original oil storage area, which was the principal source of ground
contamination. However, significant deviations did occur in concentration profiles of
airborne '^^^Pu with both distance and height. These deviations might be attributed to
sampling some more-active-than-normal particles or clusters of particles. These increases
in average airborne ^^^Pu concentrations were present at both sites AB and B.

As indicated in Fig. 5 for site AB, which was over 392 m from the oil storage area and
which was on the flat terrain, some more-active-than-normal particles or clusters of
particles (hot) may have been present in the 2.0-/im size range. In tliis case the

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 243

-\ 1 — I — I — r—i—rr

I ' ' "1

Impactor 50% cutoff

diameter, ;jm

3.3

2.0

•

1.1

Backup

filter

▲

Total

impactor

^ O 1-* — Mountain site

0.1

10'

_l I I I I ' I '

-I I I I I I I 1 1

J I I I ' ' I '

-17

1-16

10"

■15

AIRBORNE 239pu CONCENTRATION, ^Ci/cm^ (ALL WINDS, SITE AB)

Fig. 4 Airborne ^^'Pu concentration at site AB at Rocky Flats as a function of
impactor collection site.

I-

< 1

tx.
<

0.1
10-18

1 1 M 1 1 Ml 1 1 1 1 M II 1 1 1 1 II II 1 1

bH>H

■

b^ ,

=|- ^-i^ —

1\ I

\ V '^^ 1

^ 1— I 1 ^^"■^•^..^

•^-□^ 1 ^-^..^^

^s,^ ^^^

^^ ^V-\_i

— ^:::hcn —

^^ t _

~— ^^^^ / _

^^ /

^ /

~ ^ /

^ rr^ 1 1 rr 1 O Site A at security fence

^ LJ 1 1 ^*j^n ' —

D Site AB between A and B

~ A Site B at cattle fence

1 1 1 II M 1 1 1 1 1 II 1 1 1 1 1 1 1 1 II 1 1 1

10'

AIRBORNE 239pu CONCENTRATION, /uCi/cm^ (ALL WINDS)

Fig. 5 Airborne ^ ^' Pu concentrations from impactor 2.0-Aim stage collections at Rocky
Flats.

244 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

DC
<

Impactor operation
O Continuous, all winds
A 4.1- to 6.3-m/sec winds
D 6.3- to 9.8-m/sec winds

J I I I I ml \ I I I I I III \ I II 1 1 III \ L

0.1

10-18 10^17 10-16 10-15 10-14

AIRBORNE 233pu CONCENTRATION, AiCi/cm^ (FROM TOTAL IMPACTOR SAMPLE)

Fig. 6 Total airborne ^^'Pu concentration at AB site at Rocky Rats as a function of
wind speed.

concentration at the 1-m height of site AB is one to two orders of magnitude greater than
at otlier heights for this site. More important to the hot particle concept is the
concentration at the 10-m height of site B. This concentration of 2.3 X 10^'^ juCi/cm^
was the largest ^^^Pu concentration for 2-)um particles measured at any Rocky Flats
location. This relatively high concentration was unexpected since this sampling location
was the most remote from both the ground and the original oil storage area. This suggests
that other relatively hot particles could also be escaping from the plant boundaries;
however, due caution is indicated in interpreting this hot particle concept. The total of
6 d/min collected on the 2-jum stage, or 2.3 x 10^'^ /uCi/cm^ , is much less than the MPC
in air of 2 X 10"'^ juCi/cm^ (occupational). It is conceivable that the majority of this
hot plutonium was attached to one soil particle.

The functional relationship between airborne plutonium resuspension concentrations
and wind speed could not be developed as unequivocally as initially anticipated (Sehmel
and Uoyd, 1976b). This was due in part to the inadvertent loss of about a fifth of the
collected filter samples during radiochemical analysis. Unfortunately, most samples from
the higher wind speeds were lost. Even with the Umited plutonium data collected in this
experiment, it was evident that airborne ^'^^Pu concentrations increased with an increase
in wind speed. In Fig. 6 total airborne concentrations are shown for air sampled at all
wind speeds (average wind speed of 0.9 m/sec), at wind speeds from 4.1 to 6.3 m/sec, and
at wind speeds from 6.3 to 9.8 m/sec. Airborne ^^^Pu concentrations at wind speeds
from 4.1 to 6.3 m/sec are definitely larger than average airborne concentrations for
continuous air sampling. However, the 2a radiochemical counting statistics error limits
are too large to determine the wind-speed dependency. Nevertheless, an attempt to
approximate airborne ^^^Pu concentrations and consequently the resuspension-rate

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 245

dependency on wind speed was made for the 7-/im-diameter particles. This approximation
was for the 0.3-m height at sampling site AB. For the three data points taken at the 0.3-m
height, ■^^^Pu concentrations increased with the 5.9 power of wind speed. The
uncertainty in this exponent is too large to make a valid comparison between airborne
plutonium and soil concentrations.

The July 1973 plutonium resuspension experiment at Rocky Flats showed resus-
pension of both ^^*^Pu and ^■^^Pu. However, all airborne plutonium concentrations were
significantly below MPC's in air. Since '^^^Pu was collected on each particle cascade
impactor stage, the suggestion is that most plutonium was attached to soil particles when
the plutonium was resuspended.

Respirable Plutonium Concentrations at Hanford. Extensive data were obtained on
airborne radionuclide concentrations around resuspension sites studied (Sehmel, 1977c).
These concentrations were expressed both in microcuries per cubic centimeter of filtered
air and microcuries per gram of airborne solids. Tliis report summarized ranges of data
collected but did not detail data for each experiment.

Airborne plutonium concentrations for both ^^*Pu and ^^^Pu measured (Sehmel,
1977c) in resuspension experiments are shown in Fig. 7 and are compared with Hanford
300 Area fallout levels (Thomas, 1976) approximately 30 km distant. The data
represented experiments conducted over various time periods. For each data symbol the
vertical line is plotted at the mid-time of the resuspension experiment, and the
experiment duration is shown by horizontal lines drawn at both maximum and minimum
measured airborne concentrations. Airborne peak plutonium concentrations at resus-
pension study sites were significantly greater than 300 Area fallout levels (Thomas,
1976), and airborne ^^^Pu concentrations, in general, were greater than airborne ^^^Pu
concentrations. However, althougli resuspension was and is still probably occurring at
these sites, measured airborne concentrations were significantly less than MPC's
(International Commission on Radiological Protection, 1959).

Airborne plutonium concentrations at the U-Pond Area tended to remain constant as
a function of time. This constancy indicates that the weathering (or fixation) half-life for
surface contamination available for resuspension at this site is on the order of years. Tliis
is much greater than the 35 to 40 days often quoted (Wilson, Thomas, and Stannard,
1961) in Uterature on resuspension. However, the year weathering half-life at the U-Pond
could be a manifestation of some resuspension surface renewal process since this is an
active waste-disposal site. The explanations are unclear for differences in weathering
half-life.

The maximum airborne ^^^Pu concentration measured was 8 x 10~'^ juCi/cm^ near
the Hanford meteorological station (HMS) tower on Jan. 11, 1972. All other plutonium
concentrations except one were at least one order of magnitude lower. This one
exception was measured 6.1 m above ground at the U-Pond during October 1973. In
comparison with other October data, the concentration for this sample was about one
and one-half orders of magnitude greater than any other sample. We hypothesized that
some more-active-than-normal particles or clusters of particles (hot) were resuspended
and collected on this filter.

Nonrespirable Airborne Plutonium Fluxes at Rocky Flats and Hanford

Nonrespirable airborne plutonium fluxes were calculated for both ^^*^Pu and ^^^Pu. The
Rocky Flats data (Sehmel, 1976a) are shown in Figs. 8 and 9.

246 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

■14

8 10-^^

_I
Q.

DC
O

tr in- 17

10'

■18

I I I I I I I I M I I I I I M I I I I I I I I I I I I I I 1 I I I I I I I

TTTTTT

622 R

near HMS tower

Tu: MPC^Q = 2 X 10"'^ /jCi/cm-^
MPC,^„ = 6 X lO"''-' iiC\lcn?

loo

238pu: MPC^Q = 2 X 10"^^ /jCi/cm^
7 X 10'''^ ^Ci/cm^

MPC

168

■*- U Area

^■^^Pu at 6.1-m height

Gable

Mountain
pond (239pu)

' I I I I I I I I I I I I I I I

300 Area
— fallout levels

I I I I I I I II I I I

I I I I I I I I

JAJO|JAJOJAJOJAJO
1972 1973 1974 1975

DATE

Fig. 7 Range of airborne plutonium concentrations at on-site Hanford resuspension
sites compared with fallout levels.

In Fig. 8, the nonrespirable airborne ■^^^Pu horizontal flux is shown as a function of
sampling distance and sampling height. As might be expected, the maximum airborne
^^^Pu flux on nonrespirable particles was at site A near the original oil storage area. The
maximum airborne ^^^Pu flux was 6 x 10"'* y,Q'\ m~^ day"' . The airborne '^^^Pu flux
decreased with both distance and sampling height. At site A the ^^^Pu flux decreased
over one order of magnitude as the sampling height was increased from 0.3 to 2 m above
ground level. Similarly, at site AB the nonrespirable airborne ^^^Pu flux again decreased
about one order of magnitude as the sampling height was increased from 0.3 to 1 to 2 m
above ground level. Airborne ^^^Pu fluxes on nonrespirable particles decreased almost
two orders of magnitude between sampling sites A and AB. However, between sampling
sites AB and B, airborne ^^^Pu fluxes on nonrespirable particles did not show a
significant variation with distance.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 247

10"

3 10-5

lO"'' r

Q.
O)
LU
OC

z
o

10-' r

10"

Site A

Sampling height, m

♦ 0.3

A 10

T t

Site AB Site B

0 200 400 600 800 1000

DISTANCE FROM SITE A, m

Fig. 8 Decrease with distance of total ^ ^ ' Pu flux on nonrespirable particles at Rocky
Flats.

From a comparison of the data for the three different Rocky Flats sites, the
conclusion is that the airborne nonrespirable ^■^^Pu flux does not decay as a simple
exponential function of distance from site A. In addition, data for sampling heights above
1 m at sites AB and B show that the airborne ^^^Pu flux did not significantly decrease
for heights greater than 1 m up to 10 m. The nonrespirable particle plume height above
10 m is unknown.

Similar results are shown in Fig. 9 for total ^^^Pu flux on nonrespirable particles at
Rocky Flats as a function of a sampling site and sampling height. The maximum
nonrespirable airborne ^^^Pu flux was 1.2 x 10~^ /nCim"^ day~^ and was at the 0.3-m
sampUng height at site A. Again, at site A as well as site AB, airborne ^^^Pu fluxes
decreased rapidly as the sampUng height increased from 0.3 up to 10 m. However, from

248 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10"

-5

3
O-
co

10"

< 10-^

Q-

10"

Site A

Sampling height, m

♦ 0.3

■ 1

• 2
A 10

t t

Site AB Site B

I I

200 400 600 800 1000

DISTANCE FROM SITE A, m

Fig. 9 Decrease with distance of total ^ ^ ' Pu flux on nonrespirable particles at Rocky
Flats.

site AB to B, an unexplained observation was made. Airborne ^^^Pu fluxes at 2 and 10 m
heights at site B were greater than those at site AB. An explanation for this increase is not
apparent, but the increase is supported by comparing plutonium analyses uncertainties.
Error bars for site B show a ^^^Pu tlux range significantly above error bars around
site AB.

Nonrespirable airborne fluxes at Rocky Flats were greatest near the original oil
storage area (source of contaminated leakage) and near ground level. Fluxes of ^^^Pu
ranged from 10^^ up to 10"^ jjC'x m"^ day" ' . hi contrast, fluxes of ^^^Pu ranged from

10~^ to 10"^ MCi

day'

Nonrespirable airborne plutonium fluxes around the U-Pond within the Hantord area
are shown in Fig. 10. The ^^^Pu flux was less than the ^-^^Pu flux. This decrease is

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 249

100

10'

-| — I I I

I I I I I I 1 1 1 — I — I — i—T-r

O East side of U-Pond 1 ,„^

L 480-m site

A West side of U-Pond [ separation

I ti — I

238

Pu <r

-^ 239p^

1 0 limits

J I I I

J 1 I I I I I I

10"

-5

AVERAGE AIRBORNE PLUTONIUM FLUX, /LiCi m'^ day"''

Fig. 1 0 Airborne ^ ^ ' Pu and ^ ^ ^ Pu fluxes on nonrespirable particles at Hanford U-Pond
during Feb. 27 to Nov. 1 0, 1975 (particles collected in cowls).

similar to the Rocky Flats data. However, the U-Pond data show that the nonrespirable
plutonium flux extends at least up to 30 m above ground level. Also, there was a greater
airborne plutonium flux east of the U-Pond than west of the U-Pond. This is to be
expected since prevailing winds are from the west.

At the U-Pond the airborne "^Pu flux ranged from 10"^ to 10"^ /iCim"^ day"^
which is within the midrange of 10"^ to about 10~^ ^.C\ m~^ day~^ measured at
Rocky Flats. Similarly, the ^^^Pu flux at the U-Pond ranged from about 10^^ to
10"^ /LzCi m"^ day"^ , which is within the 10"'' to 10~^ /jCi m"^ day"^ measured at
Rocky Flats. The bracketing of the nonrespirable airborne particle fluxes near the U-Pond
and at Rocky Flats even within three to four orders of magnitude may be coincidental
since surface sources and other factors are pecuHar to each site.

The on-site data reported are the first results to quantify the range of nonrespirable
airborne plutonium fluxes. Ground contamination on nonrespirable particles at both sites

250 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

is poorly defin-^d or nonuniform (Corley, Robertson, and Brauer, 1976; Krey et al.,
1976b; Maxfield, 1974; Mishima, 1973; Mishima and Schwendiman, 1973; 1974; Nees
and Corley, 1975); hence the data cannot at present be analyzed to reflect resuspension
rates or resUspension factors for nonrespirable particles.

Plutonium Concentration per Gram of Airborne Soil

Airborne plutonium concentrations were normalized to the soil collected with the
airborne plutonium. Plutonium concentrations (in microcuries per gram) were determined
as a function of particle diameter as determined with both particle cascade impactors for
respirable particle diameters and sieve sizes for nonrespirable particles. Resuspended
plutonium is attached to nonrespirable as well as to respirable particles. Hence
nonrespirable soil particles may contribute significantly to downwind airborne plutonium
concentrations and represent one mechanism for transporting plutonium to surrounding
land.

For Rocky Flats nonrespirable soil collected at 0.3 m above ground level was sieve
sized (Sehmel, 1976a) into twelve different size increments. Each size increment was
analyzed for '^^^Pu and ^^^Pu. Plutonium concentrations were normalized (micrograms
per gram) to the grams of soil collected within each size increment. Results are shown in
Fig. 11 for ^^^Pu as a function of particle size at sites A and AB. Plutonium-239 was
associated with all particle sizes. The maximum concentration was about 10""* /uCi/g for
particle sizes between 10 and 20 jum. For larger particle diameters up to 230 jum,
concentrations tended to decrease with an increase in particle diameter. Concentrations at
site A were greater than those at site AB. This is expected since site A was closer to the
original oil storage area at which plutonium leakage occurred.

At each site plutonium concentrations (in microcuries per gram) indicate general
continuous relationships as a function of particle diameter, which might be used to infer
how plutonium is attached to airborne-soil particles. For nonrespirable particle diameter
ranges determined from sieve sizes, the data could be approximated by a straight Une
inversely proportional to particle diameter. The relationship is complicated by the
collection of both contaminated on-site and uncontaminated off-site nonrespirable
particles within the cowls.

For respirable particles, the ^^^Pu microcuries per gram was nearly independent of
particle diameter. This independence might suggest that plutonium attachments are
volume phenomena for these respirable particles. In contrast, plutonium particle
attachment to soil particles is expected to be controlled by available soil particle surface
area for nonrespirable particles. Additional data are required to conclude how plutonium
particles are attached to airborne particles in both respirable and nonrespirable size
ranges.

Plutonium-238 concentrations on airborne soil are shown in Fig. 12. In this case only
nonrespirable particle diameter ranges are shown. There was insufficient ^^^Pu collected
in the particle cascade impactor samples to yield positive results in respirable particle
diameter ranges. SimUar to ^^^Pu, '^■^^Pu nonrespirable concentrations were greater at
site A than at site B and also showed an inverse relationship with particle diameter.
However, there is fine structure showing deviation around any apparent inverse
relationship. This fine structure indicates that there is yet much to be learned about
plutonium resuspension and plutonium attachment to nonrespirable as well as to
respirable host particles.

TRANSURANIC AND TRACER SIMULANT RES USPENSION 251

z
o

i^j

10'

,-3

10"

O 10"

10"

t-
z

O 10"

Q.
01

10"

-I 1 r-

-I 1 1 ] — MM]

-| 1 1 I I I

— Site A

— Site AB

1 «

j:.

Backup
filter

Aerodynamic

size I Sieve size

_i I ' I I I 1 1

J III

10 100

PARTICLE DIAMETER, Mm

1000

Fig. 11 Plutonium-239 concentration on airborne soil as a function of particle diameter
at Rocky Flats.

Both ^^^Pu and ^^^Pu concentrations on airborne soil decreased from site A to
site AB. Site AB/site A ratios are shown in Fig. 13. Concentrations per gram decreased by
a factor of up to about 10"^ in the intervening 392-m distance separating sampling
sites A and AB. Concentration ratios for ^''^Pu on respirable particles were about 10~^
and were independent of particle size. In contrast, for nonrespirable particle sizes "^^^Pu
ratios between sites decreased nearly linearly as the particle diameter increased.
Plutonium-239 concentrations on nonrespirable particles decreased at rates greater than
those for the ^^^Pu concentrations. For ^^^Pu the concentration ratios for
site AB/site A were nearly one order of magnitude greater than for ^^^Pu. These larger
ratios suggest that ^^^Pu resuspended more readily relative to ^^^Pu.

252 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10"

1 1 1 I M I I I 1 1 1 I I I

-\ 1 I M I

4 1-

-6

-j' 10'

—

..„_

(£

■ 7

O 10

111

o
o 10

■8

—

" Site AB

-i

T ^
4- -r-f- J-

•i- T ;

: -c-

■»■ i T

' 1 1 i 1 1 1

I I I I ' 11)1*

1 1 1 1 1

10 100

PARTICLE DIAMETER, nm

1000

Fig. 12 Plutonium-238 concentration on airborne soil as a function of particle diameter
at Rocky Flats.

Concentrations for on-site Hanford experiments are shown in Fig. 14 for collection
on filters. Concentrations of ^^^Pu were somewhat greater than those of ^^^Pu.
Plutonium concentrations on airborne solids ranged from 10~^ to 10~^ AiCi/g. The only
exception was the October 1973 single sample described above, for which the
concentration was 6x 10^^ /jCi/g. Otherwise plutonium concentrations on airborne
solids around the U-Pond appeared to be nearly independent of time.

Plutonium concentrations on airborne nonrespirable particles were also determined
(Sehmel, 1977a) near the U-Pond on the Hanford reservation for sampling heights from
0.3 up to 30 m above ground. Airborne solids were sampled continuously for all wind
directions at sites both east and west of the U-Pond. The distance between sampling sites
was 480 m. Samples were analyzed for both ^^^Pu and ^^^Pu. Calculated results shown
in Fig. 15 are for nonrespirable airborne solids collected within cowls. Results show that
plutonium concentrations on nonrespirable airborne solids were approximately one order
of magnitude higher east as compared with those west of the U-Pond. This increase is
caused by prevailing west winds, which caused resuspension from this low-level
hquid-waste disposal area. East of the U-Pond plutonium concentrations on nonrespirable
airborne solids tended to be uniform with height up to 30 m. The plume height above
30 m is unknown. Plutonium concentrations on nonrespirable airborne solids are within
the range shown in Fig. 14 for smaller particles collected on filters. In both cases ^^^Pu

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 253

I-
<

10"

■D
_i
a.

cc.
O
S 10-3

-4

Backup filter

}

239pu
238p^

Aerodynamic size

T 1 1 — I I I M

-| — r

^l-

_ i i

J. ±

•f T

if'*

' f

Sieve size

I I I I I I

I M ij

J L

J L.

10 100

PARTICLE DIAMETER, /urn

1000

Fig. 13 Decrease in ^^*Pu and ^^'Pu concentrations on nonrespirable particles from
site A to site AB at Rocky Flats.

concentrations ranged from 10""^ to 10~^ /jCi/g. However, ^^*Pu concentrations on
respirable solids collected on filters tended to be greater than those on the nonrespirable
particles.

Both ^^^Pu and ^^^Pu concentrations on nonrespirable airborne solids near the
U-Pond were less than concentrations determined at Rocky Flats site A. This comparison
can be seen by comparing data in Figs. 11 and 12 with data in Fig. 15. However,
plutonium concentrations on nonrespirable airborne solids at Hanford's U-Pond and
Rocky Flats site AB tended to be comparable.

Airborne and Ground Plutonium Ratios

There might be some relationship between plutonium concentrations on airborne soil and
those on ground-surface soil if all airborne soil came from local resuspension. At Rocky
Flats ground-surface soils were characterized (Krey et al., 1976b) for the same time
period these nonrespirable airborne samples were collected. Ground-surface sample results

254 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10"

o
z
o
o

10"

^ 10"^

IXI

QC
O
CD

10"«:r

M I I I I I I I M I I M M I I M ! M M I M I I I I M M I I M

^^^Pu at 6.1-m height

239

238

■♦■ U-Pond area

622 R

near HMS tower

10"

Q I *l I I I I I I I I 1 I I I I I I I M I I I I I I I I I I I M I I I I I I I

I I I I I

A J 0 |J A J O |J A J 0 |J A J 0
1972 1973 1974 1975

DATE

Fig. 14 Range of plutonium concentration on airborne solids at on-site Hanford
resuspension sites.

for a 5-cm sampling depth are summarized in Table 1 for sampling sites near sites A, AB,
and B. Ground-surface ^^^Pu/^'^^Pu ratios are also shown.

Ground and airborne plutonium soil-sample results are compared in Table 2 for both
^^^Pu and ^^^Pu. Comparisons are for plutonium in ground-surface samples 5 cm deep
vs. airborne nonrespirable particle concentrations in particle diameter ranges. Airborne
concentrations were taken from Figs. 11 and 12 data limits. From these data activity
ratios of airborne to surface concentrations (microcuries per gram) were calculated. Only
the maximum ranges are reported. Maximum ranges of the ratio of airborne/ground-
surface soil concentrations are shown in the last two columns. These ratios range from
1 X 10"^'* up to 2. Thus in hazard evaluations (Johnson, Tiball, and Severson, 1976) one
might consider maximum plutonium concentrations on airborne soil to be comparable to
plutonium concentrations on ground-surface soils. This is indicated by the ratio 2.
However, in most cases plutonium concentrations on airborne soil were significantly less
than those on ground-surface soils.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 255

100

E 10 —

10"

side of U-Pond 1 _

t 480-m separation
side of U-Pond J —

O East side of U-Pond
A West

T
/
/
/
/
/

;

238

Pu < 1 ^ 239p^

J L

1 a limits
I I I II I I

10"

lO''^ 10 ^ 10'*^

AVERAGE PLUTONIUM CONCENTRATION ON AIRBORNE SOIL, AiCi.g

-5

Fig. 15 Airborne nonrespirable ^^'^Pu and ^^'Pu on nonrespirable airborne solids at
Hanford U-Pond during Feb. 27 to Nov. 10. 1975 (particles collected in cowls).

TABLE 1 Results of Selected Surface Soil Samples
at Rocky Flats*

"'Pu,

2 3 8 pu^

Sample ratio,

Location

MCi/g

2 38py/2 3 9py

Near site A

3.10 X 10-^

5.77 X 10- =

0.019

6.89 X 10-"

1.24 X 10- =

0.018

Near site AB

7.70x 10- =

1.56 X 10-*

0.020

3.86 x 10-'

7.66 X 10-^

0.020

Near site B

2.66 X 10-*

5.09 X 10-*

0.019

3.85 X 10- =

7.12 X 10-"

0.018

*Microcuries per gram of dry soil ±% standard deviation;
samples from 2000-cm^ area at 5-cm sampling depth. Two
samples per site.

Identification of relationships between sites of radionuclide concentrations on
airborne solids and contaminated ground solids would be useful in establishing criteria for
releasing contaminated areas for other uses. Concentrations on Hanford ground surfaces
obtained from the literature (Corley, Robertson, and Brauer, 1976; Maxfield. 1974;
Mishima and Schwendiman, 1973; Nees and Corley, 1975) are shown in Table 3 for

256

TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Ratio of Plutonium Concentration per Gram of Airborne Soil to That
per Gram of Ground-Surface Soil at Rocky Flats

Range, MCi/g

Ratio

UCilg a
\ MCi/g !

range
irborneN

239py

"«Pu

surface /

Site

Airborne* Surfacef

Airborne 1 Surfacef

"'Pu

238py

A
AB

1.3 X 10-" to 3.10 X 10-' to
4.6x10-= 6.89x10-"

3.2 X 10-* to 7.70 X 10-= to
8.5x10-' 3.85x10- =

7.2 X 10-* to 5.77 X 10"= to
3.4x10-" 1.24x10- =

1.8 X 10-* to 1.57 X 10-* to
1.4x10-' 7.66x10-"

0.04 to

0.2
0.0001 to

0.08

0.006 to

0.6
0.0009 to

*From Fig. 11 as a function of particle diameter.

|From Table 1 for total ground-surface sample at 5-cm depth.

:j:From Fig. 12 as a function of particle diameter.

TABLE 3 Concentrations of Plutonium and Cesium in
Hanford Ground-Surface Solid Samples Reported in the

Literature

Concentration, MCi/g solids

Location

Year

Range

Average

Range

On site (Nees and

Corley, 1975)
Inside 200 Areas

(Corley, Robertson,

and Brauer, 1976)
BC Area(Maxfield,

1974; Mishima and

Schwendiman, 1973;

Nees and Corley,

1975)
Within Hanford

site boundary

(3 to 26 km)

(Corley, Robertson,

and Brauer, 1976)
Site perimeter

(Nees and Codey,

1975)
Off site

(19 to 34 km)

(Codey, Robertson,

and Brauer, 1976)

1973 LAL* to 9.7x10-" LAL* to

4 X 10"* 2.4 X 10-*

1971 2.6 X 10-« to Not reported
6.9 X 10-"

1974

1971 7.2 X 10-' to
1.2 X 10-"

1973 LAL* to
4x 10"'

1971 LAL* to

7.6 X 10-

7 X 10-"

2.1 X 10-= to
2.5 X 10-^

Not reported

LAL* to
1.5 X 10-*

Not reported

*LAL, less than radiochemical analytical limit.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 25 7

^^^Pu and ^^^Cs. The range for ^^^Pu was from less than radiochemical analytical limits
to 6.9 X 10^^ MCi/g of surface solids. The maximum reported ■^^^Pu concentrations for
surface solids were within the 200 Areas. The minimum ^'^^Pu concentrations on surface
solids reported (Corley, Robertson, and Brauer, 1976) in the literature occurred at 19 to
34 km from Hanford. Surface contamination levels were reported only out to 34 km. The
^^^Cs concentrations ranged from 3 x 10~^ to 2.5 X 10^ ■^ i^Ci/g. These contamination
levels are used in Table 4.

Although there are only limited data for comparing the ratio of airborne soil to that
of surface soils, these ratios were calculated from the available Hanford data summarized
in Table 4. Table 4 is a summary of Hanford airborne solids concentrations (Sehmel,
1977c) for ^^^Pu, ■^'*' Am, and '^^Cs. Plutonium and americium concentrations were
obtained from Figs. 14 and 19. Table 4 also shows ground-surface ratios of airborne
solids. From these, ratios of airborne solids were determined. The last column shows
maximum ratio ranges. Ratios vary from 1 X 10"^ to 1.5 x lO"', which indicates that
contamination levels on airborne solids can be either much less than or much greater than
contamination levels on contaminated surface solids.

Caution should be used in interpreting these data. The ground-surface contamination
data are limited in quantity and were not necessarily obtained in the same areas where
resuspension experiments were performed. Airborne-particle and ground contamination
levels are shown for the BC-Crib Area in the central columns of Table 4. In this case the
BC-Crib Area was sampled (Mishima, 1973) in ten 1-m^ areas. Data from these 10 squares
indicated that surface contamination levels varied by about a factor of 100. Ratio ranges

TABLE 4 Ratio of Airborne to Ground -Surface Radioactivity Concentrations

per Gram of Solids at Hanford

Concentration, ^Ci/g solids

Maximum

Ground

l-surface

ratio range*

Airborne solids

contamination

/ MCi/g air \

Material

Minimum

Maximum

Minimum

Maximum

I /uCi/g surface y

239p^j

2x 10-**

6 X 10" =

LALt

4x 10~*
(Nees and
Corley, 1975)

5 X 10"' to
1.5 X 10^

^"'Am

1 x 10-"

7 X 10- =

NRt

238p^j

1 X 10"'

1 X 10-'

NRt

'^^Cs

2 X 10-=

1 X 10-'

NRt

'^^Csat

BC-Crib Area (10

1-m^ areas)

3x 10- =

7 X 10-"

2.1 X 10- =

1.2 x 10-^

1 X 10-^

(Mishima and

Schwendiinan,

1973)

Maximum reported

value (Maxtield,

2.5 x 10-'

3 X 10'

1974)

*Ratios from positive reported ground contamination values.

tLAL, less than radiochemical analytical limit; NR, surface contamination levels not reported for
all areas. If available, data for each area are reported separately.

258 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

CL in-l

10"

GO
EC

Site B

Site AB

10"

200 400 600 800

DISTANCE FROM SITE A, m

1000

Fig. 16 Ratio of ^^'Pu/^^'Pu on nonrespirable airborne particles as a function of
distance at Rocky Flats.

calculated from these data may be more representative than ranges calculated for ^^^Pu.
The ^^''Cs data for the BC-Crib Area ranged from 10"^ to 30. The magnitude of this
range is important since one might as a first approximation assume air contamination
levels per gram of solids to be equal to surface contamination levels per gram of solids.

Airborne ^ ^ ^Pu/^ ^ ^Pu Ratios

Airborne ^^^Pu/^^^I*u ratios at Rocky Flats consistently changed (Sehmel, 1976a) from
sampling site A to site B. Ratios shown in Fig. 16 are for total plutonium collected within
each cowl at each sampUng site and sampling height. At site A the ^•^^Pu/^'^^Pu ratio on
airborne nonrespirable soil is comparable with ratios determined from 5-cm-deep
surface-soil samples. The surface-soil range is shown as the crosshatched soil sample range.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 259

Q-

D.
00

10"

I I I I I I

1 1 \ — [Mil

T — r

I MIL

Site A

Site AB

Soil sample
range-

J I I I I 1 1 1

10 100

PARTICLE DIAMETER, pm

J I'll

1000

Fig. 17 Ratio of ^'^Pu/^^'Pu on nonrespirable airborne soil as a function of particle
diameter at Rocky Flats.

This range (Kiey et al., 1976b) represents surface-soil samples taken between the eastern
security fence and beyond the eastern cattle fence to Indiana Avenue.

At sites AB and B, airborne ^■^^Pu/^^^Pu ratios on nonrespirable particles are over
one order of magnitude greater than similar ratios from 5-cm-deep surface-soil samples.
These ratios do not appear to be significantly affected by sampling height between 0.3
and 10 m above ground level. However, considerations of the data-counting-statistics
error bars at sites A and AB tend to indicate liigher '^^^Pu/^^^Pu ratios closer to the
ground.

The ^^^Pu/^^^Pu ratio was determined for all nonrespirable particle diameter ranges
at sites A and AB at the 0.3-m sampling height. These results are shown in Fig. 17 along
with the total ground-surface-soil sample range. At site A the ^^^Pu/^^^Pu ratio is nearly
independent of particle diameter for all particle diameters above 20 iim. It is only for
particles less than 20 /nm that the ^^^Pu/^^^Pu ratio is significantly elevated at site A.
The ^^^Pu/^^^Pu ratios on nonrespirable particles greater than 20 /Lim at site A are all
comparable with ratios in the 5-cm-deep surface-soil samples. This similarity would be
expected if there were no preferential '^^^Pu to ^^^Pu separation on soil surface. In
contrast to site A, ^^^Pu/^^^Pu ratios for site AB are much different and are
significantly elevated for all particle diameters above the surface-soil sample range.

260 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

One can only hypothesize as to why there should be a difference in ^^^Pu/^^^Pu
ratios at sites AB and A as compared with soil-surface samples. One difference is that the
surface activity level at site A is greater than that at site AB. As was shown in Table 1 , the
ground-soil-surface activity level at site A was 8 to 80 times as great as that at site AB. If
plutonium particles were attacked by microorganisms (Wildung and Garland, 1977) in the
soil, microorganism activity might be decreased by the increased activity level at site A. If
microorganisms preferentially attacked ^^^Pu at site AB, which had a lower plutonium
contamination, ■^■^^Pu on surface soils might become more readily available for
resuspension. Other possibilities for increased resuspendibility of surface ^^^Pu at site AB
might be differences in soil chemistry between sites A and AB or preferential ^^^Pu
ejection (Oksza-Chocimowski, 1976) from particles during decay. Many possibilities exist,
but the reasons for the elevated ^^^Pu/^^^Pu ratios at site AB are uncertain. Additional
research is needed to determine causes of elevated ^^^Pu/^^^Pu ratios at site AB.

Estimation of Relative Plutonium Fluxes for Respirable
and Nonrespirable Particles

Direct comparisons between airborne fluxes on respirable and nonrespirable particles
were not made since respirable and nonrespirable samples were sampled differently.
Respirable particles at Rocky Flats were sampled at a constant flow rate of 0.57 m'^/min,
and nonrespirable particles were collected by inertial collection within cowls. To calculate
the relative flux on respirable and nonrespirable particles, one needs to know (l)the
average wind speed to determine the average flux on respirable particles and (2) how
particles were collected within the cowl by inertial impaction. In this case inertial
impaction means that particles would enter the cowl as if cowl sampling were isokinetic.
Actually there is flow divergence around the cowl inlet. Consequently correction factors
are needed for calculating true airborne particle fluxes for nonisokinetic sampling.
However, nonisokinetic correction factors are not available. Even with these qualifica-
tions, one might still be interested in approximating the relative plutonium fluxes for
respirable and nonrespirable particles. Consequently a simple calculation was made using
the Rocky Flats data to illustrate the relative orders of magnitude for respirable and
nonrespirable plutonium fluxes.

The horizontal plutonium flux can be estimated from airborne soil fluxes and
plutonium concentrations on airborne soil. Since soil fluxes are reported (Sehmel,
1976a), plutonium concentrations on airborne soil will be discussed before plutonium
fluxes. Plutonium concentrations of airborne respirable and nonrespirable soil are shown
in Table 5. This table summarizes both total plutonium concentrations per gram of soil
collected within cowls and respirable concentrations per gram for cases in which air was
sampled continuously with particle cascade impactors. Concentrations of ^^^Pu ranged
from 2 X 10~^ to 6.2 x 10"^ /nCi/gon respirable airborne soil. Concentrations of ^^^Pu on
respirable soil were less than radiochemical analytical limits. On nonrespirable soil ^^^Pu
concentrations ranged from 1 X 10~^ up to 3 x lO"'* /jCi/g, and ^^^Pu concentrations
ranged from 2 x 10~^ up to 5 x 10^^ A^Ci/g. Respirable and nonrespirable concentra-
tions were combined in the last two columns to estimate the average plutonium
concentration on airborne soil. For this calculation isokinetic sampling was assumed. This
assumption is discussed further in a later section. Total plutonium concentrations on
airborne soil ranged from 1 X 10~^ up to 1.9 x lO"'* ^Ci/g for ^^^Pu and from
2 X 10"' up to 3 x 10"^ juCi/g for "^Pu.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 261

TABLE 5 Plutonium Concentration on Airborne Soil Collected
During Continuous Air Sampling at Rocky Flats

Sampling
height.

Plutonium concentration, pCi/g (10

^' iuCi/g)

Sampling
site

Respiiable

* Nonrespirable

Total, assuming
isokinetic sampling

2 39pu

239py

2 38py

239pu

238pu

0.3

62.3

203

4.2

188

3.8

4.6

313

5.1

98.3

0.3

5.0

1
11.5
12.4

0.2
1.6
1.6

1.6

0.2

2.1

10.2

2.4

3.1

0.2

38.2

10.0

4.2

36.4

0.2

*Respirable samples were collected for wind-speed increments rather than continuous sampling,
2 3 8 py ^2s less than radiochemical analytical detection limits, 1 (d/min) g" ' - 0.45 pCi/g.

TABLE 6 Average Plutonium Flux Entering Cowl-Impactor
System During Continuous Air Sampling at Rocky Flats

Sampling

height,

Plutonium flux, ^Ci m ^ day"

-1

Sampling
site

Respirabk

'.*

Nonrespirable

239pu

233pu

239pu

238pu

AB
B

0.3
2

0.3
2
10

4.06 X 10-'
2.43 X 10-'
3.22xlO-«
1.16 X 10-«
1.93 xlO-'

LALt
LALt
LALt
LALt
LALt

1.02 X 10- =

4.03 X 10-'
5.66 xlO- =
8.23 X 10-'
3.18 X 10-'

2.11 X 10-'
6.68 xlO-'
1.13 xlO-«
1.85 xlO-'
1.32 xlO-'

*Respirable is all material collected within a particle cascade impactor.
tLAL, less than radiochemical analytical limit.

Average airborne plutonium fluxes entering cowl inlets at Rocky Flats were
calculated from airborne soil fluxes and plutonium concentrations on airborne soil. As
shown in Table 6, plutonium fluxes on both respirable and nonrespirable particles were
calculated. The calculation was based on collected plutonium, the sampling time, and the
cross-sectional area of the cowl inlet. The cowl-inlet diameter was 15.2 cm. The
maximum ^^^Pu flux was 1 x lO^^juCi m"^ day"\ and the minimum flux was
3 X 10^^ /iCi m~^ day"^^. On the basis of these limited data, these calculated fluxes
might be used to estimate the total plutonium fluxes over larger integrated areas.
However, such estimates should be made with caution since the flux variability at other
sites and evaluations is unknown.

Average plutonium fluxes entering the cowl impactor system are used to estimate the
respirable fraction of airborne plutonium. Estimates given (Sehmel, 1976a) in Table 7 are
based on an isokinetic sampling assumption. Respirable fractions ranged from 3 to 98% of
total airborne plutonium. However, one should use these numbers with caution. Plu-
tonium collected v^thin particle cascade impactors contained particles of 7-, 3.3-, 2.0-,

262 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Estimated Respirable Percent of Total
Airborne ^^^Pu Flux at Rocky Flats

Sampling

hei^t,

Assumed

site

isokinetic*

correctionf

0.3

3.8

37.7

0.3

36.3

58.5

98.4

*Isokinetic =

= respirable/despirable + nonrespi

rable).

respirable) (Ug/Us)

(respirable) {\}.J\J) + nonrespirable

where respirable = (d/min) min~' collected in impactor
nonrespirable - (d/min) min"' collected in cowl

Ug = average flow rate through cowl inlet,

0.36 m/sec
Ua - average wind speed, 0.9 m/sec

and 1.1 -/nm diameter (which are impactor-stage 50% cutoff diameters for unit-density
spheres) as well as smaller particles collected on the impactor backup filter. From the
inhalation standpoint, particles collected on the l-jim stage should not be included within
the respirable particle size range. Only the smaller particles are usually considered
respirable. However, in the present comparison between cowl-collected nonrespirable
particles and impactor-collected respirable particles, there is much uncertainty in
calculating the relative 1-^xm particle concentration as compared with the cowl-collected
particle concentration.

The better estimate of the respirable plutonium fraction at Rocky Flats is shown in
the last column of Table 7. On the basis of the assumed correction factor, the fraction of
respirable airborne plutonium changed from 4 to 98% for an isokinetic sampling
assumption to 9 to 99% of the total airborne plutonium. True fractions of respirable
plutonium should be between these limiting values.

Fractions of respirable plutonium have been reported (Volchok, Knuth, and
Klemman, 1972) as 25% for plutonium collected wdthin particle cascade impactors at
Rocky Flats. Since the present results show that plutonium is also attached to particles in
much larger size ranges for cowl-collected samples than for particles collected on the
intial stage of an impactor, published fractions of respirable plutonium are probably
indicative of maximum fractions at each sampling location rather than a true fraction.

Even if all plutonium collected in the cowl-impactor sampling systems were
respirable, airborne plutonium concentrations were still below the MFC (International
Commission on Radiological Protection, 1959). The MPC168 hr-air for ^^^Pu is
6 X 10~'^ juCi/cm^. The smallest fraction of respirable plutonium in Table 7 is 3.8%.
This sample at 0.3 m for site A also represents die largest airborne plutonium
concentration measured in our experiments. For this sample the real respirable ■^ ^ ^ Pu
concentration (<3.3 pm diameter) was 2 X 10"'^ /iCi/cm^, whereas the concentration
was 5 X 10"^^ ]uCi/cm^ if the l-^im particles of the particle cascade impactor were

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 263

E 10-
o

10'

O
O

E
<

ULI

I 10

-17

10"

I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I M I I I I I I I.

-Pond area

HMS tower

Z-

-19

K 10-^2 ^ci/cm^

MPC^o

MPCigg = 2 X 10"^2 nC\lcrr?

300 Area
fallout levels
(•••• less than)

ial I I I I I I I I I II I I I I I II I I I II I I I I I I I I I I I I I I I I I M I I I II

JAJOJAJOJAJOJAJO
1972 1973 1974 1975

DATE

Fig. 18 Range of airborne ^'"Am concentrations (above radiochemical detection
limits) at on-site Hanford resuspension sites compared with fallout levels.

included. If all airborne plutonium is assumed respirable and soluble, a maximum
calculated ^^^Pu concentration is (5 X 10"'V0.038) 1.3 X lO"*"^ iuCi/cm^ which is
still significantly less than the MPC of 6 X 10 * ^ ^iCi/cm^ for occupational exposure.

Americium Resuspension at Hanford

At Hanford airborne ^"^ Am concentrations measured (Sehmel, 1977c) at U-Pond, Z-19,
S-16, and HMS tower Areas ranged from about 10"'^ to 10~^^ jLtCi/cm^ . However,
other filter samples indicated that the total ■^'*'Am collected was below radiochemical
detection limits. In Fig. 18 airborne ^"^^ Am concentrations are compared with the 300
Area fallout levels (Thomas, 1976). However, the comparisons are incomplete since data
for •^^^ Am fallout level were not reported for all time periods. In addition, many data
for fallout level were below radiochemical detection limits; these are indicated by short
horizontal dotted lines. Obviously, actual fallout levels for these time periods could have
been significantly less.

Airborne ^"^^ Am concentrations at the U-Pond were of the same order of magnitude
as reported fallout levels. In contrast, airborne concentrations at the Z-19 and S-16 Areas
were significantly above fallout levels during February to May 1975. Comparisons of
upwind and downwind tower air sample results at the S-16 Area showed that increased
^'^ ^ Am airborne concentrations were from ^'* ' Am resuspension from the dry S-16 Area.
(This area has since been covered.) Maximum ^^^Am concentrations measured in

264 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

-4

O 10

H
<
CC

z
o
o

E
<

,-5

10'

10"

I I I I I I M I I I I I I I M I I

U-Pond area

HMS tower

Z-19

I I I I I I I I -

S-16

I I I I I I I

A J 0

1973

I I I i

I I

A J

1974
DATE

A J

1975

Fig. 19 Range of ^"'Am concentrations on airborne solids at on-site Hanford
resuspension sites.

November 1975 at the HMS tower are also comparable to concentrations measured at the
S-16 and Z-19 Areas. However, ^^^ Am was above radiochemical detection Umits on only
two HMS tower air filters for this time period.

Concentrations of ^'^^Am on airborne solids at each resuspension site (shown in
Fig. 19) ranged from about 10"'' to lO"'* AtCi/g of airborne solids. Airborne concen-
trations were least for the U-Pond Area and greatest for the Z-19 and S-16 Areas. For the
HMS tower data, all except two air filters collected less than radiochemical detection
limits. Nevertheless, maximum ^'*^ Am concentrations per gram of airborne solids at the
HMS tower for these two samples were comparable to concentration ranges measured at
the Z-19 and S-16 Areas.

Plutonium Resuspension from Off Site Near Hanford

Resuspension of plutonium off site near the Prosser barricade on the Hanford site was
studied. The Prosser barricade is located about 19 to 20 km southeast (130 to 160°) of
the fuel processing areas. Airborne solids were collected by sampUng with particle cascade
impactors and rotating cowl systems. Air sampHng was only when wind was blowing from
190 to 260°. This range of southwest (225°) winds came from off site toward the

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 265

-15

n
I 10

z
O

I-

•16

-17

^10

z
cc
o

E 10 ^^
<

10'

•19

Particles

on backup filter

- r,y^ S/f

__ u

Vi '

0.4 to 1 pCi/g

I ...I

1.1 plus 2.0 Mm
stages

1
0.4 to 1.7 pCi/g

3.3 plus 7 Mfn
stages

~i — I — I — I — III'

Total impactor collection

239

Pu ON AIRBORNE SOLIDS

I i :. ^1 I I , III!

5 7

3 5 7 3 5 7

WIND SPEED, m/sec

0.04 to 0.8 pCi/g

j Sampling height, m

^ 2.0

D 5.8

Wind-speed

increments, m/sec

3 to 5
5 to 7
7 to 11

0.4 pCi/g av.

3 5 7

Fig. 20 Airborne ^ " Pu concentrations near Prosser barricade at Hanford from Apr. 12
to June 29, 1976, when sampling only 190 to 270° winds.

Hanford area. All southwest winds were continuously sampled with rotating cowl systems
for nonrespirable particles, whereas respirable particles were sampled with particle
cascade impactors for wind-speed increments 3 to 5, 5 to 7, and 7 to 1 1 m/sec at a height
of 1.5 m.

Airborne plutonium concentrations blowing in from off site are shown in Fig. 20 for
the particle cascade impactor data. Airborne concentrations in both air and collected
soUd are given. Airborne plutonium concentrations determined with particle cascade
impactors are shown as a function of wind-speed increments for plutonium collected on
the impactor backup filter, the 1.1- plus l.O-jjLm stages, and the 3.3- plus 7-/jm stages.

Airborne plutonium concentrations increased with increasing wind speed. Concentra-
tions increased up to about two orders of magnitude as wind speed increased from 3 to 5
up to 7 to 11 m/sec. Straight lines are drawn through data to direct attention to the
wind-speed tendency of the data. For the plutonium collected on the cascade impactor
backup filter, lines proportional to wind speed to the 1.1 and 4.2 power are shown. For
wind speeds below about 5 m/sec, airborne plutonium concentrations tended to increase
nearly Unearly with wind speed. However, above 5 m/sec, plutonium concentrations
increased with wind speed to the 4.2 power. At the present time it is unknown whether

266 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

these indicated relationsliips would suggest a threshold wind speed of 5 m/sec for
resuspension or whether a smooth curve should be drawn tlirough all data points.

In other portions of this figure, selected straight lines suggesting wind-speed
dependencies are shown only for wind speeds above 5 m/sec. For the 1.1- plus 2.0-fim
impactor stages, airborne concentrations increased with wind speed to the 9.3 power. For
the 3.3- plus 7-/im impactor stages, airborne concentrations increased with wind speed to
the 5.2 power. For total plutonium collection within particle cascade impactors, a range
of wind-speed dependency is shown on the right side of the figure. For a sampling height
of 0.3 m, air concentrations increased witli wind speed to the 4.4 power. However, at a
sampUng height of 1.8 m, airborne concentrations increased with wind speed to the 3.0
power.

These data in Fig. 20 are the first to show that plutonium is resuspended from off-site
locations. In addition, airborne plutonium concentrations show a very high wind-speed
dependency for this off-site plutonium resuspension. As will be discussed later, tracer
wind resuspension rates suggest a wind-speed dependency to the 4.8 power. Tliis is similar
to the wind-speed dependency shown by this off-site plutonium resuspension data.
However, other data on plutonium resuspension show a different wind-speed dependency.
West of the U-Pond on the Hanford reservation, airborne plutonium concentrations
increased (Sehmel, 1977c) with wind speed to only the 1.5 power. Reasons for these
differences in the wind-speed dependency of on-site vs. off-site plutonium resuspension
are unknown. Possibly a threshold wind velocity above which resuspension increases
rapidly with wind speed was not exceeded at the U-Pond.

Nonrespirable airborne plutonium blowing from off site onto the Hanford reservation
was also measured. In this case sampling direction was controlled by placing stops that
allowed the rotating cowl (Fig. 3) inlet to rotate only with the range of 190 to 260°.
Plutonium analysis was for the total nonrespirable solids collection in cowls at each
height rather than as a function of particle size, as was done for Rocky Flats (see Figs. 1 1
and 13).

Plutonium-239 concentrations and fluxes for nonrespirable particles blowing from off
site near the Hanford Prosser barricade are shown in Table 8. Plutonium concentrations
on nonrespirable airborne solids ranged from 1.3 X 10""^ up to 2.1 X 10~^ AtCi/g. These
concentrations are similar to those shown in Fig. 15 for airborne nonrespirable particles
collected west of tlie U-Pond.

TABLE 8 Plutonium Transport on Nonrespirable Particles from
Off Site near the Prosser Barricade on the Hanford Reservation

Airborne ^ ^ ' Pu
nonrespirable flux,

/iCi m~' day"'

2 3 9

Pu on airborne

Only for 190

Sampling to 260° winds,

height, m (d/min)g~' juCi/g 3 to 11 m/sec For total time in field

0.3

0.29

1.3 X 10-^

3.9 X 10-*

8.0x10-^

0.46

2.1 xlO-"

4.0 X 10-*

8.3 x 10-'

5.8

0.32

1.5 xlO-'

1.4 X 10-*

2.8 X 10-'

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 267

Airborne nonrespirable ^^^Pu fluxes also were calculated for these Prosser barricade
samples. Horizontal flux calculations were made for both the total time wind was
between 3 and 1 1 m/sec and 190 to 260° and for the total time cowl air samplers were in
the field. Fluxes are shown in the last two columns of Table 8. When ihe shorter time
period (3 to 11 m/sec winds) is used for calculating the horizontal plutonium flux, fluxes
range from 3.9 x 10"^ to 1.4 X 10~^ /aCi m~^ day"^ . This Prosser barricade flux range
is within the range measured near site A at Rocky Flats (shown in Fig. 8). However, if
the total time cowl air samplers were in the field is used for calculation, Prosser barricade
airborne nonrespirable off-site plutonium fluxes were lower and comparable to those
measured at Rocky Flats sites AB and B (see Fig. 8).

These cross-comparison data show that there is a comparable plutonium flux on
nonrespirable particles off site at Hanford, on site at Hanford U-Pond, and on site at
Rocky Flats for the fime periods invesfigated. Comparable fluxes may be caused by more
soil being transported from off site at the Prosser barricade site. As shown by the range of
^^^Pu concentrations on airborne soil from 1.3 x 10~^ up to 2.1 x 10~^ iJ-Ci/g, this
range is greater than fallout levels in soil-surface samples. As shown in Table 3, reported
■^■^^Pu concentrations in surface samples 19 to 34 km from Hanford had a range from
3.6 X 10"^ to 7.6 X 10~^ AfCi/g. These last values are similar to a fallout level of
3.8 X 10"^ juCi/g measured (Hardy, 1974) at North Eastham, Mass.

Most plutonium collected appears not to have originated from fallout. Rather, most
plutonium collected on these airborne nonrespirable particles near the Prosser barricade
resembles weapons-grade plutonium (Krey, 1976; Krey et al., 1976a). Plutonium isotopic
ratios (^^°Pu/'^^^''^^°Pu) (in atom percent) for these nonrespirable samples were
6.10 ± 0.02 at 0.3-m heiglit, 6.31 ± 0.02 at 2-m height, and 6.28 ± 0.03 at 5.8-m height.
In comparison, the isotopic ratio determined from a sample of forest-fire smoke plume
near Mt. St. Helens, Wash., was 13.82 ±0.05. Isotopic ratios for respirable particles
sampled near the Prosser barricade were not determined. Although plutonium was
blowing from off site near the Prosser barricade, airbome respirable plutonium
concentrations were below MPC's, as is shown in Fig. 20.

Relative Amounts ef Radionuclide "Clusters " on Particles Resuspended

Data from these studies indicate that occasionally some more-radioactive-than-normal
particles or clusters of radioactive particles were resuspended and collected on sampling
filters. In the October 1973 plutonium data for Hanford shown in Figs. 7 and 14, one
filter at 6.1-m heiglit coUected 6.5 x 10"^ ^iCi/g of airborne sohd (1.2 x 10"^^ )i;Ci/cm^
of filtered air). The plutonium measurement was 36 times as great as the maximum value
of 1.8 X 10"^ fiCilg of airborne soHd (4.0 x 10"^^iuCi/cm^ of filtered air) collected on
other filters simultaneously sampling at heiglits of 6.1 and 0.3 m. We hypothesize that
this relatively high plutonium collection on this filter was due to collecfion of one or
more larger (more radioactive than normal usually resuspended and sampled) particles or
clusters of particles.

The size of the larger particle(s) cannot be measured since the filter samples were
dissolved for plutonium analysis. Nevertheless, tlie relative size can be estimated from the
ratio of radioactivities collected for "normal" and "larger" particle sizes. Assuming that
the radioactivity of a particle is proportional to its volume, then flie filter with a
plutonium activity 36 times as great as the next highest measured activity may have
collected larger particles of plutonium activity 36 times as great as the activity of normal

268 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

particles. This would correspond to a diameter for the larger, less-frequently resuspended
particle(s), which may be three times the diameter of normally resuspended plutonium
particles. Only since extensive air samples have been collected has the resuspension of
these unusually active particles been suggested. However, the frequency of their
resuspension appears to be very low.

The presence of a more-radioactive-than-normal resuspended particle is also indicated
by tlie ^'"Am data shown in Figs. 18 and 19. The two positive ^'''Am air samples for
the HMS tower may indicate that larger-than-normal ^'^'Am particles or clusters of
particles were collected on these filters.

As discussed earlier for resuspended particles at Rocky Flats, a more-radioactive-
than-average particle (or cluster of particles) was collected on the 2-/jm stage of a particle
cascade impactor.

Resuspension Factors at Hanford

Resuspension factors have been used to describe resuspension air concentrations. The
resuspension factor (expressed in units of meters"') is defined as the airborne
concentration of contaminant per cubic meter divided by the surface contamination level
per square meter immediately below the point where the airborne concentration was
measured. Often air concentrations for determining resuspension factors have been
measured from about 1 to \% m above ground. However, airborne concentrations are a
function of the upwind contamination level, not a contamination level immediately
below the air-concentration measurement site. It is the transport from upwind
contamination sites to the concentration measurement site that determines the airborne
concentration.

Although the validity of resuspension factors is questionable, they were for a long
time the only method for estimating air concentrations. Consequently resuspension
factors were estimated from data obtained at Hanford (Sehmel, 1977c). Resuspension
factors and the basis for their calculation are shown in Table 9 for both ^^^Pu and

1 ABLE 9 Resuspension Factors at Hanford

Material

Area

Air concentration

Minimum

Maximum

Surface contamination
(G), MCi/m'

Minimum

Maximum

Resuspension
factor* range,

'Pu Inside diemical separation
areas (Corley, Robertson,
and Brauer, 1976)

'Cs BC (Bruns, 1976;Mishima,
1973)

7x10'* 8x10-

4,9 X 10"

2x10-" ZxlQ-'" 0.29

1.2 X 10

55.4

6 X 10-
2 X 10-'

4 X 10-'
7 X 10-

■ Resuspension factor = 1 0' x /<j-

'^^Cs; literature values are used as an indication of ground-surface contamination levels.
Airborne concentrations (Sehmel, 1977c) are shown in Fig. 9. Resuspension factors
calculated from air concentrations and ground -surface contamination levels are shown in
the last column. Resuspension factors for ^^^Pu range from 6 x 10^'° to 2 x 10~^
m~ ' . Resuspension factors for '^''Cs range from 4 x 10 '' to 7 X 10"^ m"'. These
ranges, from 10"^' to 10~^ m~ ' , are within ranges reported in the literature (Mishima,
1964; Sehmel and Lloyd, 1976a).

TRANS URANfC AND TRACER SIMULANT RESUSPENSION 269

1-11

E
o

I10-12

F 10'^^

8 10-1^

(- 10-15

g 10-16

-17

IQ-

■^ r — 1 ' 1 1 M 1 1 — 1^ 'I'M 1 ^ ' >! 1 III \ \ — 1^ 1 M M r-

— I

MPC

^^^ ^ '' Assume soil = 2 g/cm-^ ^ --

40 hr^X' * ^ ^ ^X' ^ -'

■

-^10-5 ^- ^^

^ :_

^ MPC,68hr " ^" ^^ .-' ^

^-^10 6 ^ ^^ ^^

^ ^^^ ^ ^^^^

y„ ^^^ y ^^^

;

^^Q-' .-" ^^

"-_ ^

^^^ ^ ^^^

—

^^^ ^ ^^^ y

- ^

^^^ ^ .^^^ *^

^^ ^- 10-8 ^y^ ^^

^^^ ^ ^^^ y

.^^ — =

^^^ ^ ^^^ ^

^^ ^,^'^^10-^ '' ^

\ ^^^

X' ^^^ ^ ^^^

.-- ^y^ ^^-',o-»^,^^

- .' ^ ' ^10-"

—

PARAMETER; RESUSPENSION FACTOR, m^X^ x^

^y^ ^^^ ^^"^"^ ^^""l0-12

^^^ ^ -^^^ ^

•_

'' .^^ ^ ^

-y^

-13

'' ^xn^ ^1 ^^^

1 L-i'Ti 1 1 i 1 L^^i 1 ^ 1 un'Tii 1 1 — \ — i L/ii il j_

i 1 i U i

10"

10-^ 10"^ 1

PLUTONIUM ON GROUND SOIL, nCi'g

1Q^

-J 1 I 1 1 1 1 1

-5 10-'* 10-2

PLUTONIUM ON GROUND SOIL, ^Ci/g

10-

,0-j

I i i ! I I I i

J I 11)1

102 10^ 10"

PLUTONIUM ON GROUND SOIL (d/mm) g'^

10^

Fig. 21 Equivalency of airborne piutonium concentrations and ground -surface concen-
trations based on the resuspension-factor concept.

Prediction of Airborne Radionuclide Concentrations from Resuspension Factors

Airborne concentrations can be predicted from resuspension factors and surface
contamination levels if both values are known. Equivalencies of airborne concentrations
and ground-surface concentrations are shown in Fig. 21. The reader can estimate two of
the parameters and use this tlgure to predict the third parameter. However, as indicated in
the last section, the range of experimental resuspension factors is very large. Conse-
quently realistically predicting the relationship between surface and airborne concentra-
tion is fraught with uncertainties.

Prediction of Airborne Radionuclide Concentrations from Airborne Solids at Hanford

Airborne concentrations can either be determined experimentally or calculated on the
basis of simplifying assumptions. For example, one assumption is that radionuclide
concentrations on airborne solids are equal to radionuclide concentrations per gram of
ground-surface contaminated solids. As is shown in Table 4 for very limited data, this
assumption is usually not valid since the ratio of radionucHde concentration per gram of
airborne solids to the radionuclide concentration per gram of surface sohds ranged from

21 Q TRANSURANIC ELEMENTS IN THE ENVIRONMENT

AIRBORNE RESPIRABLE SOIL CONCENTRATION, pg/m^

Fig. 22 Equivalencies between plutonium on airborne soil and airborne plutonium
concentrations.

10~^ to 10^. Nevertheless, two equality assumptions are used in this section to predict
airborne radionuclide concentrations.

Airborne plutonium concentrations can be predicted by assuming both parameters of
airborne soil concentration and plutonium concentration per gram on that airborne soil.
In this case the equivalency between these two parameters and airborne plutonium
concentrations is shown in Fig. 22. As a point of reference, airborne concentrations at
Hanford are (Sehmel, 1977b) about 80 )Ug/m^ for wind speeds of 5 m/sec. Hence, in using
Fig. 22, a plutonium concentration on airborne soil of approximately 10~^ /aCi/g would
be required (assuming a surface contamination depth of 1 cm and a soil density of 2
g/cm^) at an airborne soil concentration of 80 jug/m^ to exceed or approach maximum
permissible airborne ^^^Pu concentrations.

As is shown in Tables 2 and 3, there is no experimental basis for adequately
predicting plutonium concentration on total airborne soil since airborne soil usually
consists of both uncontaminated soil blowing in from off site and the resuspended
contaminated soil. At our present state of knowledge, there are only limited data for
ratios of plutonium on total airborne soil compared with total surface soil. Moreover,
there is no experimental resuspension data to relate plutonium concentrations on
respirable surface soils vs. plutonium concentrations on respirable airborne soils.

Airborne solids concentration levels were estimated from data (Sehmel, 1976b;
1977c) shown in Fig. 23. Airborne particle volume distributions were determined at the
Hanford area using both an optical particle counter and a cowl-impactor system (Sehmel,

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 271

lO^r

10^

102

—

yp?

Fig. 23 Airborne particle volume
distributions at Hanford.

1 10 10^

PARTICLE DIAMETER (D), /im

1973a). The range obtained with the optical particle counter is shown as a crosshatched
area. The cowl-impactor data have an upper limit, increasing with wind speed and
decreasing height (the April 1972 data), and a lower limit for other test periods, indicated
by the lines described by 3.46 D°' ' and 1 160 D" ^ ■''^, where D is the particle diameter.
The upper, or maximum, limits of the curves for any particle diameter were integrated as
a function of particle diameter to determine maximum airborne mass loadings. The lower
limits were also integrated as a function of particle diameter. These integrations predicted
the solids mass loading per unit volume of air (shown in Table 10) as a function of four
different particle diameter ranges: 0.16 to 1, 1 to 10, 10 to 100. and 100 to 230 (jtm.

Particles less than 10 /im in diameter are frequently considered respirable (i.e.. small
enougli to be inhaled into the lungs), even though 3.5 nm appears to be more exact. In
the following discussion, particles with diameters less than 10 /jm are considered
respirable and those larger dian lOjum are nonrespirable.

212 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 10 Calculated Airborne Concentrations from Airborne Solids Concentrations

and Surface Contamination

Particle diameter range, /jm

Respirable

Nonrespirable

Airborne partides

0.16 to 1

1 to 10

0.16 to 10

10 to 100

100 to 230

10 to 230

Total volume, Mm'/cm'

Upper limit

4.58

3.5 X 10'

354.6

2.22 X 10"

9.37 X 10"

1.16 X 10-

Lower limit

0.036

3.83

3.87

9.39

0.609

10.0

Solids mass loading

Lipper limit, mg/m'

0.00916

0.704

0.713

44.4

187.4

231.8

Lower limit, Mg/m^

0.072

7.66

7.73

18.78

1.218

20.0

Calculated* extreme

maximum airborne

concentration from

limits of airborne

solids concentrations.

MCi/cm'

"'Pu, Upper Imiit

4.92 X 10-"

1.60 X 10-"

Lower limit

5.33x 10-'*

1.38 X 10-"

' "Cs, Upper limit

1.78 x 10~' '

5.79 X 10-'

Lower limit

1.93 xlO-'^

S.OOx 10-"

*Concentration on aii

borne solids

assumed equ

al to maximum reported in Table 3 for concentration on groiu"'

solids:

>»»Pu = 6.9x 10-''

MCi/g

'''Cs=2.5 X 10"'

MCi/g

where

'^'Pu: MPC.ohr = 2 x 10"' ' MCi/cm'
MPC, ,,hr = 6 X 10-" MCi/cm^

" ■' Cs: MPC, 0 hr = 6 X 1 0-" MCi/cm'
MPC,,,hr = 2 X 10-' nGlcm^

Mass loadings in these respirable and nonrespirable ranges are shown as upper limits of
0.7 mg/m^ for respirable particles and 231.8 mg/m^ for nonrespirable particles. The
lower limits are 7.7 jug/m^ for respirable particles and 20 mg/m^ for nonrespirable
particles. These mass loadings were multiplied by maximum surface contamination levels
for ^^^Pu and *^^Cs as shown in Table 3. This approach yielded a predicted maximum
airborne ^^^Pu concentration for respirable particles of 4.92 x 10^'^ jiQ\lcm^ and a
predicted lower limit of 5 .33 X 10"'^ AtCi/cm"' .

Upper and lower limits of '^^Cs concentrations on respirable and nonrespirable
particles were calculated similarly. However, a comparison of airborne ^^^Cs and ^^^Pu
concentrations predicted by this method and measured airborne concentrations shows the
shortcomings of this calculational approach. The predicted respirable range of ^^^Pu
concentrations from 5.33 x 10"^^ to 4.92 X 10~'^ )uCi/cm^ is within the lower two
orders of magnitude of die 10~^^ to 10"''* )uCi/cm^ experimental range shown in
Fig. 7. This simUarity in range is considered fortuitous when one also compares the
predicted '^"^Cs respirable range from 1.93 x 10"'^ to 1.78 x 10"' ' juCi/cm^ with the
experimental range from 2 x 10"'^ to 3 x 10"'^ /iCi/cm^ (Sehmel, 1977c). For '^^Cs
the minimum predicted airborne concentration of 1.93 x 10 '^ /jCi/cm^ is 0.6 of the
maximum experimental concentration of 3 x 10 '^ ^tCi/cm^. However, the maximum
predicted airborne concentration for '^''Cs of 1.78 x 10'' /.(Ci/cm^ is 60 times the
maximum experimental concentration of 3 x 10"'^ /iCi/cm^. These ratio comparisons

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 273

for ^^^Pu and ^^^Cs of predicted maximum to the experimental maximum airborne
concentration show a ratio range from 0.05 for ^^^Pu to 60 for ^^^Cs. Thus there are
at least three orders of magnitude uncertainty in using the mass-loading approach in
calculating the maximum expected airborne radionuclide concentration.

This airborne-particle mass-loading approach to calculated airborne radionuclide
concentrations does not distinguish the sources of airborne material. For simple wind
resuspension, airborne solids included contaminated solids (lower or higher radionuclide
concentrations per gram of soUd) blown in from the surrounding area as well as solids
resuspended from the prime resuspension study site. In contrast, airborne solids above a
mechanically disturbed area are resuspended from tlie study site. In this case possibly
such an equality assumption of radionuclide concentration per gram of solids would be
appropriate for mechanical disturbances (Milliam et al., 1976) of contaminated soil area.

Tracer-Particle Resuspension

Tracer particles placed on selected surfaces were used to measure (Sehmel, 1977b)
resuspension rates caused by both mechanical and wind resuspension to determine
particle resuspension rates. Mechanical resuspension was measured for vehicular traffic on
asphalt and cheat grass areas and for pedestrian traffic on an asphalt area. Wind
resuspension was measured as a function of v/ind speed and also as a function of
respirable and nonrespirable particle diameters.

Mechanical Resuspension Rates. Mechanical resuspension includes both vehicular
resuspension and pedestrian resuspension.

Vehicular Resuspension. A /4-ton truck and a car were driven over ZnS tracer
particles (8-/jm mass median diameter) placed on one lane of asphalt road. Resuspended
tracer was measured to determine resuspension rates (Sehmel, 1973b). Results are shown
in Fig. 24 for particle resuspension rates at vehicle speeds of 5, 15, 30. and 50 mph. The
resuspension rate is tlie fraction of particles resuspended from the tracer lane each time
the vehicle was driven down the road (fraction resuspended per pass). Wlien a car was
driven through the tracer lane at speeds up to 30 mph, resuspension rates increased with
the square of car speed from about 10""* to lO"'^ fraction resuspended per pass. This
means that these resuspension rates were proportional to car-generated turbulence. When
tlie car was driven on the lane adjacent to the tracer lane, resuspension rates were lower
for each vehicle speed but increased with vehicle speed from about 10~^ to 10~^
fraction resuspended per pass.

Resuspension was also measured when a ^^-ton truck was driven on the tracer lane.
Resuspension rates for truck passage increased from about 10"^ to 10^^ fraction
resuspended per pass. Since resuspension rates were higher, truck-generated surface-stress
turbulence appears to have been much greater than that for car-generated turbulence. For
vehicle speeds above 20 mph, resuspension rates for car and truck passage are comparable.
This similarity might be caused by tire surface-stress turbulence rather than by air
turbulence.

Resuspension rates were also a function o{ the time tracer particles were on the
asphalt road. As shown in Fig. 25, particle resuspension rates decreased as a function of
time. For tliese data the tracer had been on the road for 4 days. Vehicle-generated
resuspension rates increased from about 10^^ to about 10~^ fraction resuspended per
pass as vehicle speed increased from 5 to 50 mph. For both vehicles resuspension was

274 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

LU
Q.

£lO-3

O
an
u.

a.
W

o 10
<

I-
o
<
cc

,-4

10"

1 I I I I I I

"I 1 1 — I — I I I I

Car driven
through tracer.

Truck driven
through tracer

Car driven
past tracer

J L

10
VEHICLE SPEED, mph

100

Fig. 24 Rates of tracer particle resuspension caused by vehicle passage over an asphalt
road.

greater when the vehicle was driven through the tracer lane than when driven on the larie
adjacent to the tracer lane.

Resuspension caused by truck passage througli a cheat grass area was also measured
(Sehmel, 1976c; 1977b). Results are shown in Fig. 26 along with resuspension rates from
the asphalt road. Truck-caused resuspension from the cheat grass area was always less
than that from the asphalt road. This decrease is attributed to the protective action of
cheat grass in hindering truck-generated turbulence from reaching the ground and
resuspending the tracer.

Resuspension from the cheat grass area decreased for truck speeds from 5 to 30 mph.
This decrease is attributed to the sequence of experimental truck speeds. The initial truck
speed was 5 mph. Apparently the relatively larger resuspension rate at 5 mph was caused
by the most readily resuspended particles being removed from the cheat grass. In
succeeding experiments at increasing truck speeds up to 15 mph, and possibly 30 mph, all
readily resuspended tracer was removed from the cheat grass foliage. When the truck

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 275

10"

Q
Z t^

Q- <
to Q-

LU <

_l Q-

^? ,0-

QCO
<(r

I- a:

CJLL

<
tr

1-5

I I I I I I I I

Truck driven
through tracer

> I I I I I 1 1

Car driven
through tracer

10 100

TRUCK SPEED, mph

1000

Fig. 25 Rates of tracer particle resuspension caused by vehicle passage over an asphalt
road 4 days after particle deposition.

10-^

O
GC

o
cc
u.

D 10-3

2 to
m <-

^^
Z) u-l
W -I
LU U

QC -

wuj 10""

ocn
<

° 10-5

H
O
<

10"

"^T^

I I I

Truck, asphalt road— \

A^_,

Truck, cheat grass road

J ^ 1 I I I I I

J i I I L_L_L

10
VEHICLE SPEED, mph

Fig. 26 Rate of tracer particle resuspension caused by vehicle passage over asphalt and
cheat grass roads.

276 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

speed was subsequently increased from 30 to 40 mph, resuspension per pass also
increased. Apparently increased air turbulence at the base of the cheat grass increased
resuspension rates.

Pedestrian Resuspension. Resuspension caused by a man walking along the ZnS
tracer lane of tlie asphalt road was also measured (Sehmel, 1977b; Sehmel and Lloyd,
1972). The man walked across the tracer area in a leisurely fashion; the stride and paces
per second were not measured. For tracer on a 3-m-wide road lane, the reported
resuspension rate was the fraction of particles resuspended each time the person walked
down the length of tlie tracer lane. With wind speeds of 3 to 4 m/sec, pedestrian-caused
resuspension rates were from 1 x 10~^ to 7 x lO""* fraction resuspended per pass along
the tracer lane. This pedestrian-generated resuspension was greater than wind resuspen-
sion during the experiment.

Wind-Caused Resuspension. Experimental values of wind-caused resuspension rates
of tracer particles from environmental surfaces have not been experimentally determined
from mass balance techniques other than for the present data (Orgill, Petersen, and
Sehmel, 1976; Orgill, Sehmel, and Petersen, 1976; Sehmel, 1975; 1977b; Sehmel and
Lloyd, 1972; 1976a; 1976c). Some data were initially obtained using S-^tm mass-median-
diameter (MMD) ZnS particles and average wind speeds from 1 to 5 m/sec. More
extensive data as a function of wind speed were obtained using submicrometer CaMo04
particles. Average resuspension rates for ZnS particles were measured for resuspension
from an asphalt surface (Sehmel and Lloyd, 1972) and a cheat grass surface (Sehmel.
1976c). For average wind speeds of 1 to 4 m/sec, wind resuspension rates from an asphalt
surface ranged from 5 x 10 ^ to 6 x 10^^ fraction resuspended per second. For average
wind speeds of 1 to 5 m/sec, wind resuspension rates from a cheat grass surface ranged
from 5 X 10"^ to 6 X 10"^ fraction resuspended per second.

Wind-caused resuspension was measured for submicrometer CaMo04 particles
deposited in a liglitly vegetated area on the Hanford area. Tracer particles were deposited
in a circular area of 23-m radius around a centrally located air-sampling tower.
Resuspended particles were measured at the tower as a function of wind-speed increments
for respirable particle diameters and at all wind speeds for nonrespirable particles.
Respirable particles were collected within particle cascade impactors (Fig. 3), and
nonrespirable particles were collected by impaction and gravity settHng within cowls.
Resuspension rates for each size range airborne were calculated by assuming that the
entire tracer source was also in tliat same size range.

Wind-caused resuspension rates for the tracer— host soil particles as resuspended are
sliown in Fig. 27 as a function of wind speed. Resuspension rates ranged from about
10 '^ to 10"^ fraction resuspended per second. Different functional dependencies of
resuspension rates on wind speed can be obtained from these data, depending on which
set of wind-speed increments is used. During the January to February period, air sampling
was for large wind-speed increments; in subsequent experiments wind-speed increments
were smaller. The straight lines shown in Fig. 27 were drawn through all data points. In
these cases resuspension rates increased with the 1.0 to 4.8 power of wind speed.
However, when only data points for smaller wind-speed increments were used,
wind-caused resuspension rates increased with wind speed to the 4.8 power for 7-, 3.3-,
2.0-, and l.l-/.tm-diameter particles as well as for the smaller particles collected on the
cascade impactor backup filter.

277

—J —

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t' ;

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^~ — -~^^

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-- — .

(£

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18, 1

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Jan.
Mar.
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21^ TRANSURANIC ELEMENTS IN THE ENVIRONMENT

For comparable wind-speed increments, tracer resuspension rates were nearly
independent of time the tracer was on tlie ground surface. However, it is often assumed in
theoretical modeUng that particles become less available for resuspension with time. The
assumption in these models is that pollutant particles become fixed or attached to soil
particles and subsequently "migrate" into the ground surface. This process is called
weathering but is poorly understood.

The independence of wind-caused resuspension rates with time is a significantly
different observation than some others have made. Literature on radioactive resuspension
indicates that airborne radioactivity concentrations decrease with a weathering half-life of
35 days (Anspaugh et al., 1969; Wilson, Thomas, and Stannard, 1961). In contrast, if
there is a weathering half-life for the controlled tracer experiments described above, the
weathering half-life must be on tlie order of years. Some differences in reported
weathering might be explained by how air samples were collected. In work reported by
others, air concentrations were measured continuously. In contrast, in our tracer
experiments air concentrations and hence resuspension rates were measured as a function
of wind speed. Even these differences in determining weathering half-lives illustrate how
poorly weathering is understood.

Average wind-caused tracer resuspension rates are reported for both respirable and
nonrespirable particles in Fig. 28. In these cases respirable refers to all particles collected
within cascade impactors and nonrespirable refers to particles collected within cowls.
Nonrespirable particle resuspension rates were nearly independent of time and were of
the order of 10^ ^ ^ fraction resuspended per second.

Resuspension rates for respirable particles ranged from about 10~^ ^ to 10~^ fraction
resuspended per second. These resuspension rates did not decrease with time. For the first
two sampling periods, resuspension was measured for all wind speeds. In succeeding
experiments resuspension rates were measured only for wind speeds above 1 and above
4 m/sec. The upper, or solid Une, portion of the respirable particle curve corresponds to
resuspension rates calculated for the wind samphng periods. These periods correspond to
wind speeds above 1 and above 4 m/sec. The lower limit of the respirable particle curve
corresponds to resuspension rates calculated by assuming that resuspension time
corresponds to the total time that cascade impactors were in the field (i.e., time included
for winds less than 1 and less than 4 m/sec).

Initial Generalized Wind-Resuspension-Rate Correlation

Guidelines based on exisfing experimental resuspension-rate data are needed for
estimating resuspension rates. An initial correlation (Sehmel, 1975; 1977b) was developed
from data reported for uranine particles resuspended from a smooth surface, ZnS from an
asphalt surface (Sehmel and Lloyd, 1972), submicrometer molybdenum tracer from a
vegetated desert soil (Sehmel and Lloyd, 1976a), and DDT from a forest (Orgill, Sehmel,
and Petersen, 1976; Orgill, Petersen, and Sehmel, 1976). Each of these surfaces has a
much different estimated aerodynamic surface-roughness height, Zq, ranging from
4 X 10~^ cm for the smooth surface to 1 m for a forest. Roughness height is calculated
from the log-Unear velocity profile and is the height at which the extrapolated velocity
profile reaches zero velocity.

Ranges of measured average resuspension rates were correlated (Sehmel, 1975;
1977b) as a function of measured or estimated surface-roughness heiglits (zo)in Fig. 29.
Resuspension rates range seven orders of magnitude from 10~'° to lO"'' fraction

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 279

-7

I-
<

z
g

UJ
CC

10"

I I I I I

Respirable

I I I I I I I I I I

Wind-speed increment,
m/sec

X O All

A 1.3 to 20.1

D 4.5 to 9.4

Lower limit from
total time in field

1— X 1

Nonrespirable

I X '

17I I I I L_L

J I I L

I I

_L_L

0 J

1972

J
1974

0 J

DATE

J
1975

J
1976

Fig. 28 "Average" wind-caused tracer resuspension rates (lightly vegetated desert on
Hanford reservation).

resuspended per second. The practical significance of these numbers can be made
apparent by noting that a year is 3.2 x 10^ sec.

This initial resuspension-rate correlation shows that resuspension rates decrease as
surface roughness increases, at least for the three smaller roughness heights. However,
measured resuspension rates for DDT sprayed on a forest are two orders of magnitude
greater than rates for the desert soil.

This is an unexpected and unexplained increase in resuspension rates. A possible
explanation of the increase might be increased resuspension caused by tree movement in
the wind. Also, various other gross differences in controlling variables and experimental
factors may have influenced results. Since the data are so extremely limited, this apparent
correlation should be used with extreme caution until correlations based upon several
physical parameters instead of only Zq are developed. Nevertheless, this initial correlation
does give some justification for estimating resuspension rates until better correlations are
developed.

280 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

-2

-3

10"

10'

1-4

6 _

^ 10"^

LU
CO

a
^ 10-

10"

10-10-

,-11

,-3

Note: 1 yr = 3.2 x lO'' sec

•10-/im uranine from
2.9-cm-i.d. aluminum tube

ZnS from
aspfialt surface •

Molybdenum tracer
from desert soil

10"

,-1

102

ROUGHNESS HEIGHT (z^), cm
Fig. 29 Initial correlation of wind-caused tracer resuspension rates.

Conclusions

This review of 1971 to 1977 resuspension data determined at the Pacific Northwest
Laboratory indicates the following problem areas:

• There are more theoretical resuspension models available for prediction than data to
vaUdate or to use in those models. Tlieoretical model development is limited by
availability of experimental data.

• The data base discussed for relating plutonium contamination of surface soils to
that of airborne soil is based on gross surface-soil and airborne -soil samples. Data have not
been collected to determine any relationship between plutonium size distributions and
concentrations on airborne soil and those on surface soils.

• Resuspended plutonium is transported on both respirable and nonrespirable soil
particles. Data reported are the entire data base for plutonium transport on airborne
nonrespirable soil. Additional data are needed to describe plutonium transport on

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 281

nonrespirable particles and the subsequent degradation to a respirable resuspension
source.

• One possible assumption for describing resuspension concerns the prediction of
plutonium concentration per gram of airborne soil vs. plutonium concentration per gram
of surface soil. However, ranges of airborne concentration per gram vs. surface-soil
concentration per gram indicate a wide discrepancy between airborne vs. surface soils.
This wide discrepancy shows that there seems to be no justification for assuming any
equalities between plutonium concentrations on surface soils vs. airborne soils.

• The rate of change of average airborne radionuclide concentrations with time has
been described by a weathering half-life. However, the weathering half-life is not well
known. Data shown for radionuclides as well as for inert tracer particles indicate that the
half-life is from 5 months or longer rather than the often quoted 35 to 40 days.

• At Rocky Flats the ^^^Pu/^-'^Pu ratio can be much greater on airborne
nonrespirable soils than on surface soils. Thus there may be a preferential resuspensioh
transport of '^^Pu vs. '^^Pu. For migration witliin surface soils, there are data showing
preferential migration with depth as well as of ^^^Pu compared with ^^^Pu.

Botli ^^^Pu and ^^^Pu resuspension occurred on site at Rocky Flats and the Hanford
area, but all airborne plutonium concentrations were significantly below maximum
permissible concentrations (MPC's) in air. In addition, plutonium was resuspended from
off site near the Hanford reservation. In ail cases plutonium was deposited on each stage
of particle cascade impactors, which indicates that most plutonium was resuspended
while attached to larger host soil particles.

Plutonium was transported on both respirable and nonrespirable airborne-soil
particles. In most resuspension research reported by other researchers as well as in
air-monitoring activifies, airbome concentrations of particles have been measured without
sampling both respirable and all nonrespirable particle sizes present. Only respirable or
near-respirable size particles are frequently measured since the usual air-sampling
techniques tend to keep larger, nonrespirable particles from being collected. Conse-
quently total airborne plutonium concentrations could be greater than normally reported
using most existing s^mpUng equipment systems. Results from those systems are a
conservative estimate (high concentration) of airborne respirable plutonium concentra-
tions. Nevertheless, plutonium transport on nonrespirable particles may be a significant
factor in total plutonium transport. Larger than respirable particles are resuspended and
may not travel too far downwind before redepositing again. In contrast, respirable
particles remain airborne for a much longer distance. Additional research is needed to
clarify the relative significance of plutonium transport on respirable as compared with
nonrespirable particles.

Plutonium concentrations per gram of both respirable and nonrespirable airborne soils
discussed in this chapter are summarized in Table 1 1. Ranges are within several orders of
magnitude, from 2 x 10"^ to 6 x 10"^ /jCi/g for respirable ^^^Pu and from 1 x 10' ''
to 3 X 10"'^ //Ci/g for nonrespirable ^^^Pu.

In all cases plutonium concentrations per gram of soil were calculated on the basis of
total soil samples. In addition, there is no proven method to predict the ratios of
concentration per gram of airborne soil to concentration per gram of surface soil. As
shown in Tables 2 and 4, this ratio ranges seven orders of magnitude from 10"'* to 10^.
This uncertainty range is almost as large as the uncertainty range of 10~^ ^ to 600 m ^
for resuspension factors (Mishima, 1964; Stewart, 1967; Sehmel and Lloyd, 1976a).

282 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 1 Summary of Plutonium Concentrations on Total Airborne Solids

Total airborne concen

itration range, MCi/g

2 38p„

239pu

Site

Respirable

Nonrespirable

Respirable*

Nonrespirable

Rocky Flats

NRt

2x 10-' to

2.1 X 10-* to

1 X 10-* to

(Sehmel, 1976a;

5.1 x 10-*

6.2 X 10- =

3.1 X 10-^

Sehmel and

Lloyd, 1976b)

Hanford reservation

On site (Sehmel,

1 X 10-' to

4 X 10-' to

2 X 10-« to

1 X 10-' to

1977c)

1 X 10-*

1 X 10-'

6 X 10- =

4 X 10-*

From off site

NRt

5 X 10-« to

1.3 X 10-' to

near Prosser

1 X 10-*

2.1 X 10-'

barricade

* Respirable as used in this chapter refers to those particles found on all filter and impactor stages
as contrasted to nonrespirable particles collected by gravity in rotating cowls.
tNR, no radiochemical results.

TABLE 12 Summary of Plutonium Total Transport
Fluxes on Nonrespirable* Particles

Range of total plutonium fluxes,

^lC\ m~

-^ day-'

Site

238pu

2 39pu

Rocky Flats

1 X 10-« to

1 X 10-^ to

(Sehmel, 1976a)

1 X 10-=

6x 10-^

Hanford reservation

On site

2x 10-» to

4 X 10-' to

2x 10-'

4x 10-*

From off site

NRt

1.4 x 10-* to

near Prosser

3.9 X 10-*

barricade for 1 90

to 260° winds

*Nonrespirable as used in this chapter refers to those
particles collected within the rotating cowl shown in Fig. 3.
tNR, no radiochemical results.

Airborne plutonium transport fluxes on nonrespirable particles are summarized in
Table 12. These data and the decrease of flux with distance shown in Figs. 8 and 9
constitute the present knowledge on this subject. There is surprisingly reasonable
agreement within several orders of magnitude for nonrespirable airborne plutonium
fluxes. However, the agreement may be caused, in part, by relatively more soil transport,
with a lower plutonium-on-soil concentration being comparable to a lower soil transport
and concurrent higher plutonium-on-soil concentration. Nevertheless, these ranges of
horizontal plutonium fluxes on nonrespirable particles could be used in modeling efforts
and hazards analysis until a greater data base is experimentally obtained.

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 283

Airborne concentrations of ^^^Pu, ^"^^Am, ^^^Pu, ^-^"^Cs, and ^°Sr increased
(Sehmel, 1977c) as a function of wind speed to the one to sixth power for on-site
resuspension study sites. Airborne plutonium concentrations for off-site resuspension
increased as a power function of wind speed. Above a wind speed of about 5 m/sec,
airborne plutonium concentrations increased with wind speed to the third to fifth power.
Explanations for differences in exponents are needed. Differences might be attributed to
source characteristics and extent.

In contrast, for controlled source experiments, tracer-particle resuspension rates
increase with about the fifth power of wind speed. This fifth power is similar to off-site
plutonium resuspension.

Plutonium concentrafions on collected airborne soil at Rocky Flats ranged from a
maximum of twice the concentration on ground-surface soil to as low as 10""* of the
concentration on surface soils. The maximum reported (Krey et al., 1976b) surface-soil
concentrafion was 3 x 10"^ /iCi/g. Even at tliis relatively high plutonium surface-soil
concentrafion, airborne respirable plutonium concentrations were significantly below
airborne MPC's on a yearly basis. Similarly, at the Hanford resuspension study sites
(Sehmel, 1977c), maxim.um concentrations per gram of airborne solid for transuranics
were: for^^^Pu, 6 X IQ-"^ iuCi/g; for "^Pu, 1 x 10"^ /uCi/g; and for ^^ ^ Am, 7 x lO^^
AtCi/g.

There is much uncertainty in the relafionships between radionuclide concentrations
per gram of airborne solid and per gram of surface-soil sofids. For tlie transuranic-element
data reported at Rooky Flats, the ratio ranges from 10~^ to 2, and at Hanford tlie ratio
ranges from 0.5 to 1.5 X 10^. Uncertainties are probably complicated by spatial
distributions of surface contamination.

Radionuclide particles can be resuspended either as individual particles or. more
probably, attached to host soil or solid particles. An average-activity, or normal-activity,
radionuclide particle distribution is usually collected on sampling filters. However, at
both Rocky Flats and Hanford, one filter sample collected in each case showed
significantly greater plutonium concentration than tlie maximum for all other samples
collected during the same time period. These anomalous higher concentrafions are
attributed to one or more plutonium particles of unusually higher activity than those
normally, or most frequently, resuspended.

There is increasing, but still conflicting, data that ^^^Pu might be more mobile tlian
^^^Pu. The isotopic-ratio data reported for airborne plutonium transport on nonrespira-
ble particles at Rocky Flats support the greater mobility concept. Tliis conclusion was
obtained by comparing airborne ^^^Pu/^^^Pu rafios on nonrespirable soil near the
eastem security fence and the eastern cattle fence. At the eastern security fence, airborne

Pu/ Pu rafios were similar to ground surface ratios. However, at the eastern catfie
fence, the ^^^Pu/^^^Pu rafios were significantly greater than those measured on local
surface soils (see Fig. 11). Consequenfiy an explanafion is needed for the increased
^^*Pu/^^^Pu rafio at distances from the original oil storage area. Possible explanafions
include preferential biodegradation of ^^^Pu compared with ^^^Pu and preferential
ejecfion of ^^^Pu at the eastern security fence and at the eastern catfie fence. These
possibilifies exist but are not definifive. Further research is needed to explain the higher
relative airborne ^^^Pu/^^^Pu concentrafions near the eastern catfie fence.

As used in this chapter, the weathering half-life is the time required for airborne
concentrafions at a resuspension site to decrease by one-half. Weathering is probably a

284 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

function of the source characteristics. However, very little is knovv'n about predicting
weathering. For fallout previous hterature has indicated that this half-life is between 30
and 45 days. However, results of '^^Cs and transuranic air concentration measurements
at resuspension study sites at Hanford reported here indicate a weathering half-life of
tVom 5 months (Sehmel, 1977c) to 1 yr or greater. Knowledge of this ill -de fined half-life
is important in resuspension modeling efforts to describe airborne effects of surface
contamination. Changes in surface contamination availability with time must be known if
models are to predict airborne concentrations. This range of from 5 months to 1 yr or
greater for a weathering half-Ufe must be considered when evaluating resuspension
changes with time.

Transuranic-element resuspension rates have not been directly measured at surface
contaminated sites other than inferred from nonvalidated models since published
characteristics of the contaminated surface sources are not adequate for direct
measurement of particle resuspension rates. Consequently resuspension rates were
measured with controlled tracer-particle simulants using a uniform surface contamination
source. On the basis of those measurements, the following conclusions were reached.

Particle resuspension rates are a function of at least wind speed and mechanical
disturbances. Mechanical disturbances, such as vehicular traffic or a man walking, can
cause high local resuspension rates. In comparison, average wind resuspension rates from a
local area could be less important per unit area than local mechanical-disturbance
resuspension. However, wind-caused resuspension rates apply to the entire contaminated
area. In the comparison of relative resuspension from wind-caused and mechanical
disturbances, one would need to know the total surface contamination area for wind
resuspension vs. small localized surface contamination levels for mechanical-disturbance
resuspension rates. Botli mechanisms, however, do resuspend and transport potentially
hazardous respirable particles.

Resuspension rates for respirable and nonrespirable particles are needed for inclusion
as source terms in atmospheric diffusion and transport equations; however, model
predictions are no better than the uncertainty in the source data. In the case of
resuspension rates, uncertainties are very large. Much research is yet needed to develop
resuspension models to predict particle resuspension rates for any situation.

Wind-caused resuspension rates from a sparsely vegetated area have only been directly
measured with submicrometer tracer particles and estimated for tracer particles larger
than 1 idm (Healy and Fuquay, 1958; 1959). The potential etTects of different particle
diameters and chemical properties on resuspension rates are unknown. It miglit be
hypothesized that similar results would be expected for other submicrometer particles of
interest since submicrometer particles are probably attached to host soil particles when
particles are resuspended. If the particles of interest were much larger, it is unknown
whether the particles would be resuspended attached to host soil particles or resuspended
as discrete parficles.

The change in airborne concentration of a pollutant as a function of time is often
attributed to a weathering half-life, the fixation of the pollutant particle into the
ground-surface soil. In contrast, weathering half-lives for respirable tracer particles are
now estimated here as being on the order of years. Predictions using weathering half-lives
of months vs. years could have a significant implication in environmental hazards
evaluations. At tlie present time, credit for decreased airborne radioactivity from
resuspension could be attributed to a weathering halt~-life of months. If a weathering
half-life of years were applicable for transuranic elements, the potential downwind

TRANSURANIC AND TRACER SIMULANT RESUSPENSION 285

inhalation hazard from resuspended particles would be significantly increased. Additional
experimental data are needed to determine what weathering half-life or variation of
half-life with time should be used in hazards evaluations.

Vegetation on a resuspension surface will decrease resuspension rates. This is rather
obvious but can be detlnitely concluded from decreased tracer-particle resuspension rates
for vehicles driven on a cheat grass area compared with those for vehicles driven on an
asphalt area. Since vegetation does decrease resuspension, vegetation should be retained
on all areas of potential surface contamination and existing surface contaminated areas
until constructive cleanup operations can be initiated.

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Behavior of Transuranic Elements in Soils, Plants, and Anitnals, ERDA Report PNL-2416, Pacific

Northwest Laboratory, NTIS.
Wilson, R. H., R. G. Thomas, and J. N. Stannard, 1961, Biomedical and Aerosol Studies Associated

with a Field Release of Plutonium, USAEC Report WT-1511, University of Rochester, Atomic

Energy Project, NTIS.

Interaction of Airborne Plutonium
with Plant Foliage

D. A. CATALDO and B. E. VAUGHAN

The interaction of airborne pollutants with the foliage of terrestrial plants has been
investigated from many aspects, including interception, retention, and absorption.
Although interception parameters for both gaseous and particulate pollutants have been
effectively modeled, the behavior and fate of pollutants, especially particulates, after
foliar interception are not known. Particles with diameters of 10 to 200 pm exhibit
retention half-times of 10 to 24 days. Direct and indirect data, however, suggest that
submicronic particles are more effectively retained on plant foliage than are larger
particles. Studies are presented to describe the retention behavior of sub micron-size
particles deposited on foliage of bush bean and sugar beet plants. Simulated rainfall \yus
used to evaluate retention efficiency. These studies showed submicronic particles to be
increasingly less available for leaching with increasing residence time on the leaf; e.g.,
more than 90% of the foliar plutonium deposits were firmly held to the leaf surface.
Retention mechanisms are discussed in terms of leaf morphology and the leaching regimes
used. The absorption of foliar plutonium and its subsequent translocation to seed and
root tissues were dependent on a number of parameters, including chemical form and the
presence or absence of a solution vector.

The behavior of the transuranic elements in the environment and their potential for
transfer in the food chain have been the subject of extensive study over the past 25 years.
Although there is a general understanding of many problems concerning atmospheric
transport (Slinn, 1975; 1976) and of the behavior of plutonium in specific ecosystems
(Nevada Test Site), little is known of the controlling mechanisms that influence the
bioavailability of plutonium and the other transuranic elements and their subsequent
transfers along the food web to man. With the current stratospheric depletion of fallout
plutonium (Bennett, 1976), the importance of the inhalation route to man is greatly
reduced. This then suggests that the major sources of transuranic elements in the future
will result from resuspension of fallout-contaminated soils on a global basis, resuspension
from highly contaminated local sources, accident situations, and low-level releases from
nuclear facilities.

Present radiological safety estimations frequently discount foliar sorption and
emphasize the soil-to-root pathway for the entry of transuranic and other radioelements
to the food chain (Vaughan, Wildung, and Fuquay, 1976). Typical dose-assessment codes
assume a rapidly declining exponential loss of material from leaves (Soldat, 1971). This is
certainly not a general situation. It does not apply to the behavior of plutonium aerosols
described here and probably applies only to very large particles and to certain gaseous

288

AIRBORNE PLUTONIUM 289

radioelements. such as iodine (Markee, 1971). Our studies indicate that the fate of
transuranic elements following foliar interception is influenced by particle size (mass).

The importance of the foliar-entry pathway compared to root absorption for
worldwide fallout was recognized long ago (Chamberlain, 1970; Russell, 1965). In later
studies of particles of probably wind-resuspended origin, 87% of the ^''Sr, 81% of the
'^^Cs, and 73% of the ''^'^Ce in forage plants were derived from foliar contamination
(Romney et al., 1973). This was shown by comparing plants grown inside plastic
enclosures with those grown with no cover at the Nevada Test Site. Plutonium may
behave similarly, but unfortunately there are no well-controlled field observations.
Recently, the importance of foliar-to-root pathways in the plant was defined in the
Liquid Metal Fast Breeder Reactor final environmental statement (U. S. Atomic Energy
Commission, 1974). Despite inconsistencies with respect to other dose-assessment codes,
risk tends to be minimized by specifying extremely conservative limits at points where
radioelements actually enter the human body, i.e., air and food. As a matter of systema'tic
practice, an improved quantitative understanding of the basic environmental processes is
required. This becomes important in situations where new technology may lead to
different physical (size) and chemical characteristics of the source term for release,
especially in nuclear-fuel reprocessing plants, and where comparatively large increases in
the handling of radioelements are projected for the future (Energy Research and
Development Administration, 1975).

The following discussion reviews current knowledge regarding the retention and
absorption potential of foliar surfaces and describes the fate of transuranic particles
following plant-foliage interception as deduced from the extrapolation of information on
the behavior of other particles and the limited information on plutonium.

The Problem of Retention of Particles on Foliar Surfaces

The retention of particles on foliar surfaces depends on many parameters associated with
the foliar surface and the physical aspects of the particle. Leaf factors affecting the
efficiency of particle entrapment include components of the leaf that affect roughness
(Holloway, 1971), namely, venation, surface features of epidermal cells, nature of the
cuticle surface, nature and frequency of trichomes, and the microstructure of surface
wax. Each of the microtopographical features of the leaf may contribute to the
entrainment and retention of particles. Other factors affecting retention include surface
stickiness from organic and inorganic secretion, leaf wetness and charge attraction
between particles, and surface waxes or components. In addition, retention is dependent
on particle size, particle density, wind speed within the boundary layer, and, when a
particular element comprising or contained in a particle is being considered, solubility.

Available information on foliar retention is sometimes disconcerting and contra-
dictory when one tries to reconcile the retention and behavior of relatively insoluble
particles with early fallout data on soluble or volatile fission products. Early fallout work
with respect to fission products has been reviev/ed by Chamberlain (1970) and Russell
(1965). In general, these reviews indicate a retention half-time of 10 to 14 days for
soluble fission products, with losses resulting from reentrainment of carrier particles,
sloughing of surface wax (Moorby and Squire, 1963), and rainfall (Middleton, 1959).
Except for radioiodine (Markee, 1971), such a short retention half-time is probably
characteristic only of large aerosol particles, as described later.

290 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Several studies approached the problem of particle retention by using smiulated quartz
fallout material containing adsorbed fission products. Witherspoon and Taylor (1'569)
found that over a 33-day period 88- to 177-^(m-diameter (MMD)* particles were more
effectively entrained by pine foliage than by oak. Although after only 1 lir wind
resuspension accounted for a 90% reduction in the number of particles in oak leaves as
compared with a 10% reduction in pine, the first rainfall (to + 1 day) accounted for a
15% reduction of particle activity remaining at 1 hr. Retention half-times reported for
periods of 0 to 1, 1 to 7, and 7 to 37 days were 0.12, 1.4, and 25 days for oaks, and 0.26,
4.5, and 21 days for pines, respectively.

A similar study by Witherspoon and Taylor (1970) presented data for the retention of
44- to 88- and 88- to 175-iL(m-lVIMD particles by various agricultural plants. These studies
indicate that average wind speeds of 0.5 mph over the initial 12-hr period following
contamination are more effective in removing the smaller particles (21.1% vs. 15.8%)
and that average wind speeds of 1.1 mph over a 12- to 36-hr period resulted in a higher
loss of the larger particles (21 .6% vs. 15.4%). Subsequent to to + 6 days, varying amounts
of rainfall resulted in a marked reduction in retention of particles of both size ranges. The
resuspension behavior of these relatively large particles is in keeping with theoretical and
empirical measurements on the inertial forces within the boundary layer required to
resuspend spores from leaves (Aylor, 1975; 1976; Aylor and Parlange. 1975).

Subsequent studies by Witherspoon and Taylor (1971) with 1- to 44-|Um-MMD
simulated fallout particles showed longer weathering half-lives for 1- to 44-/jm particles
(17.9 days) than those reported earlier for 44- to 88- (15.7 days) and 88- to MS-iim (15.1
days) particles. Loss rates were also less affected by time or rainfall after a particle
residence time of 7 days. This suggests that particle size does, in fact, play an important
role in the extent of toliar retention.

Although these studies aid in our understanding o^ the interception and retention of
larger particles (>10jL(m) analogous to close-in fallout, questions arise as to the behavior
of fallout particles of submicron size. Both Iranzo (1968) and Romney et al. (1975)
reported that plutonium-containing material resuspended in field situations is difficult to
remove from contaminated foliage; as much as 50%- is tenaciously held on foliar surfaces.
This suggests that retention is affected by factors other than the passive association of
particles with relatively flat foliar surfaces where only inertial forces influence their
removal or resuspension.

In studies o\~ 6.77 ± 0.02-ium-AMAD+ uranine particles, where the primary particle
had an MMD in the submicron range. Wedding et al. ( 1975) have shown that deposition is
related to the roughness o^ the leaf surface. By analogy, the leaf-roughness factors
affecting deposition should also atTect retention. The etTect of wind and rainfall on
foliady deposited PbCI; particles (1- to 3-iUm MMD) was evaluated by Carlson et al.
(1976). These studies showed that lead particles remained fixed to leaves under
controlled conditions for up to 4 weeks after fumigation; reentrainment wind speeds of
up to 6.7 m/sec were inefTective in removing surface deposits. Losses due to simulated
rainfall were proportional to the amount of rainfall; mists were more efTective than
droplets in removing lead deposited on leaf surfaces; only 15 and 5% of the foliar
deposits, respectively, were leachable.

*Mass median diameter (MMD) assumed; particles pliysically measured.
tActivity median aerodynamic diameter.

AIRBORNE PLUTONIUM 291

Total

t ■ • ■] Insoluble
I I Soluble

TIME OF LEACHING AFTER CONTAMINATION, days

Fig. 1 Leachability of plutonium from bush bean foliage. Sets of four plants each were
leached at 1, 7, 14, or 21 days after contamination; v ± SEM (« = 4). (a) Fresh
plutonium dioxide, (b) Water-aged plutonium dioxide.

Data on the retention of plutonium by foliar surfaces are limited. The data that do
exist are based on laboratory studies in which a low-windspeed exposure chamber was
used to contaminate plant canopies (Cataldo, Klepper. and Craig, 1976). Figure 1
illustrates the leachability of two forms of '^^^Pu dioxide as a function of residence time
on the foliage of the bush bean following a simulated rainfall of 0.4 cm in 7 min. The
particles deposited on the foliage had aerodynamic sizes (activity median aerodynamic
diameter ± geometric standard deviation) of 1.274 ^(m± 1.63 and 0.734 /im ± 2.16 for
freshly prepared and water-aged oxides, respectively. The count modes for the log-normal
distributions were 0.142 and 0.019 /jm, respectively. These latter values represent the
particle diameter (absolute size) with the highest frequency within the family of particles.
The plutonium retained on foliage after mild leaching ranged from 92 to 99%. These data
are qualitatively similar to those obtained for 1- to 3-ium lead particles (Carlson et al.,
1976) but are contrary to data obtained with larger simulated fallout particles
(Witherspoon and Taylor, 1969; 1970; 1971). Both the fresh and the hydrated oxide
exhibit a reduced leachability with increased residence time on the leaf. The retention
mechanism may be related to physical entrapment of the submicron-size particles in small
fissures on the leaf surface or to charge adsorption between particles and the leaf surface.
The inability to readily remove plutonium from foliar surfaces has been noted (Hanson,
1975; Romney et al., 1975; Iranzo, 1968); the mechanisms controlling retention,
however, are not clear.

Little (1973) used weakly acidic solutions to study the physical processes of ion
exchange involved in the retention of heavy metal particles, such as lead, on foliage.
Table 1 compares the leachability of foliar plutonium by synthetic rainwater with and
without 0.1% HNO3. Leaching with acidic solution results in a moderate increase in
insoluble plutonium leached from leaves contaminated with fresh PuO^ but a substantial
increase from leaves contaminated with the hydrated oxide. The large increase in the

292 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Effect of Acid Solution on Leachability of Foliar Plutonium 7 Days After
Exposure of Bush Bean Plants to Fresh and Hydrated Plutonium Dioxide*

Foliar plutonium leached,t %

Synthetic

rainwater +

Leached

Synthetic

0.17r HNO,

Compound

component

rainwater

(pH 2.0)

^'^'Pu dioxide

Soluble

0.5 ± 0.1

1.0 ± 0.1

Insoluble

1.6 ± 1.0

2.5 ± 0.6

Total

2.1 ± 1.1

3.5 ± 0.7

^ ^ * Pu dioxide

Soluble

0.6 ± 0.1

4.5 ± 0.2

(hydrated)

Insoluble

2.0 ± 1.3

5.3 ± 0.5

Total

2.6 ± 1.4

9.8 ± 0.7

*Plant fohage was exposed to polydispersed aerosols. The freshly prepared oxide had an
AM AD of 1.27 Mm and a GSD of 1.63; the hydrated oxide had an AMAD of 0.73 /Lzm and a GSD
of 2.16. Count modes for the aerosols were 0.140 and 0.018 ^m for the fresh and hydrated
oxides, respectively. Plants were leached with 200 ml of solution (equivalent to a 7-min rainfall
of 0.4 cm).

t Leachability expressed as microcuries of leachate/(microcuries leached + microcuries
remaining on leaves) x 100; four replicate samples, .v ± SE.

soluble components may result from a solubilization of noncrystalline plutonium on the
surface of the particles. The increased leachability of the hydrated oxide (0.019 iim), as
compared with the fresh oxide (0.142 jum). may be related to the larger surface area
available for reaction.

Even though much of the foliar-deposited plutonium is unavailable for leaching with
weakly ionic pH 5.8 solution, the increased removal of both soluble and insoluble
components with acidic solutions may indicate that a portion of the submicron particles
intercepted by foliage may be held on the leaf surface by charge phenomena and by
physical entrapment and not necessarily buried in waxy plates. Table 2 compares the
leaching behavior of plutonium from two plant species with different surface roughnesses.
Plants were leached with 800 ml of solution, and the leachate was collected in 50-ml
fractions. Since total plutonium in the leachate decreased logarithmically (plutonium
activity in the last few leachate fractions approached background levels), reported
retention values represent plutonium not readily leachable.

Scanning-electron-microscope micrographs of the leaf blades show that bush bean
leaves have moderate to low surface relief and sugar beet leaves are relatively flat. The
difference in surface microtopography between these two species is related primarily to
patterns of wax deposition and the presence of trichomes in the bush bean leaves. The
surface wax of bush bean leaves is laid down in such a way as to form high longitudinally
oriented ridges with deep crevasses, and the surface of the waxy plates is relatively
smooth. By comparison, the surface wax of sugar beet leaves forms relatively shallow
irregular convolutions, and the surface of the wax deposits is rougher than that of the
bush bean. The trichomes of the bush bean leaves, which are approximately 150-/im high
and spaced approximately 190 /./m apart, provide additional surface relief. This
microtopography and its effect on particle entrapment and leaf-surface wettability may
provide a basis for understanding the processes involved in particle retention.

AIRBORNE PLUTONIUM 293

In general, the leaching data for sugar beets and bush beans suggest that both surface
roughness and particle size affect the retention of particles on foliar surface. With the
larger fresh-oxide particles (count mode approximately 0.142 /im), substantially more of
the plutonium is leachable from smooth leaf surfaces under both leaching conditions.
This may be the result of physical entrapment of particles in comparatively deep fissures
or crevices contributing to surface roughness in the bush bean leaf, especially if it is
assumed that a particle must be suspended in a water droplet to be removed from the leaf
surface. Similarly, the effect of acid leachate may be in alleviating the attractive forces
holding particles to leaf surfaces, particularly in the case of the sugar beet. The retention
behavior of the smaller hydrated oxide particles (count mode, approximately 0.019 jum)
is slightly different from that of the fresh oxide. The synthetic rainwater was about
equally effective in removing particles from both the bush bean and sugar beet; the acid
leach was slightly more effective with the bush bean.

Although the gross surface structure of bush bean and sugar beet leaves is obviously
different, the microtopography of the surface itself may not be as different with respect
to the retention of very small particles (0.02 jum). This may explain similarities in the
retention behavior of plutonium deposited onto the foliage of sugar beets and bush beans.
It is impossible with limited data to generalize as to mechanisms controlling the fate of
particles on foliar surfaces. For the small hydrated-oxide particles, however, it appears
that leachability and retention are not only dependent on particle size with respect to leaf
topography and physical attraction, such as charge, but also on the ability of a water
droplet to contact the particle; thus wettability and contact angle become important
(Gregory, 1971). along with other environmental factors (Hull, Morton, and Wharrie,
1975) that influence the physical and chemical nature of the leaf surface.

Aside from our lack of understanding of mechanism, it is important to note that the
behavior of small particles, such as that of plutonium on leaf surfaces with respect to

TABLE 2 Effect of Continuous Leaching Regimes on the Removal of
Plutonium Particles from Leaves of Bush Bean and Sugar Beet*t

Plutonium retained on leaves

after leaching4 %

Synthetic

rainwater +

Leaf

Synthetic

O.l^HNOj

Plant species

roughness

Plutonium form

rainwater

(pH 2.0)

Phaseohis vulgaris

Moderate

Fresh ^^^PuOj

97.6 ± 0.9

97.0 ± 0.4

(Bush bean)

Hydrated' 3 «PuOj

95.5 + 1.2

71.6 ±6.7

Beta vulgaris

Smooth

Fresh'^'PuOj

82.0 ±4.9

64.7 ±9.8

(Sugar beet)

Hydrated ^'^PuOj

95.7 ± 1.1

83.0 ± 3.5

*D. A. Cataldo, unpu Wished data.

t Plant foliage was exposed to polydispersed aerosols. Particle-size data for bush bean
are given in Table 1. For sugar beet, fresh oxide had an AMAD of 1.59 ^m and a GSD of
1.76; the hydrated oxide had an AMAD of 0.75 /um and a GSD of 1.84. Count modes for
the aerosols were 0.130 and 0.048 mhi for the fresh and hydrated oxides, respectively.
Plants were leached with 800 ml of solution (equivalent to a 28-min rainfall of 1.7 cm) 7
days after exposure.

|Four replicate samples, ;c ± SE.

294 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

retention half-time, differs markedly from that commonly reported for fission products
and larger particles {>\Q ^m). The tenacity of plutonium retention observed by Romney
et al. (1975) and Iranzo (1968) tends to reinforce the laboratory studies with plutonium
described here.

Availability of Foliar Deposits for Uptake and Transport to Other Plant Tissues

Foliar structures are a source of many organic and inorganic substances that either
migrate to the surface by diffusion and mass flow or are actively exuded by secretory
structures. This may provide a chemical environment on the leaf surface which enables
readily hydrolyzable species to be complexed or chemically stabilized and therefore made
more available for foliar absorption. Since foliar surfaces represent an efficient absorptive
structure (Wittwer, Bukovac, and Tukey, 1963; Franke, 1967; 1971), fohar application of
micronutrients to correct nutrient deficiencies is effective, especially in situations where a
specific nutrient tends to be immobile and not as available for plant uptake from soil
(Krantz et al., 1962; Franke, 1967). The actual mechanisms involved in foliar absorption
are not totally understood. Available information indicates that, although the cuticle of
the leaf is hydrophobic in nature, penetration is facilitated via intermolecular spaces
(Fisher and Boyer, 1972), modification in cutin composition at anticlinal epidermal walls,
and the presence of ectodesmata (Franke, 1967; 1971) and trichomes (Benzing and Burt,
1970). The role of stomates as a route of foliar penetration under normal conditions is in
question and is currently considered of negligible importance (Greene and Bukovac,
1974).

The relative importance of foliar absorption as compared to root absorption as a
route of entry into plant tissues depends on several factors. For soluble species that
remain relatively available in soil solution, root-absorption processes are as effective as, or
more effective than, foliar-absorption processes. This does not imply that foliar surfaces
are not effective sites of absorption. Elements reported to be absorbed and transported
from fohar surfaces include inorganic N, Rb, K, Na, Cs, P, CI, S, Zn, Cu, B, Mn, Fe, and
Mo (Wittwer, Bukovac, and Tukey, 1963). For specific nutrient deficiencies, foliar
application is sometimes the method of choice (Bradford, 1966; Labanauskas, 1966).
Tliis is especially true for nutrilites that tend to hydrolyze readily in soil solution or are
rapidly adsorbed to soil particles and therefore are not so available for root absorption.

The fate, with respect to foliar absorption, of relatively insoluble elements, such as
plutonium, which make up or are carried on discrete particles, is in some way analogous
to the behavior of micronutrients, such as iron, which tend to form relatively insoluble
products in aqueous environments. If we can assume that small particles containing
plutonium (<1.0Mm) can be retained in foliar surfaces over an extended period of time,
the question arises as to the absorptive capacity of foliar surfaces for available plutonium.
Since absorption of a particular element is a function of tlie concentration of the
available or soluble component, an extended residence time on plant foliage may provide
the time necessary for soluble components to be chemically modified and/or absorbed by
internal tissues. This may represent a more efficient route of entry than root absorption
because in root absorption the same finite amount of plutonium deposited in soil may be
insolubilized and adsorbed to soil particles, which, of course, reduces the concentration
available for root absorption.

Absorption data from laboratory studies with bush bean plants contaminated with
aerosolized plutonium are given in Tables 3 and 4; the protocol for this study was

AIRBORNE PLUTONIUM 295

TABLE 3 Extent of Translocation of ^ ^ ^Pu from Contaminated Foliage of
Bush Bean to Seed Tissues in the Absence and Presence of a Solution Vector*!

Time of leaching
after

Transport ratio:}:

Stage

Pu oxide

contamination

of development

(fresh)

(hydra ted) §

Pu citrate

Pu nitrate

No leach

<4.5 X 10-*

1.1 X 10- =

5.4 X 10-*

6.8 X 10-*

Day 1

Pre flowering

2.6 X 10-=

4.1 X 10- =

8.4 X 10-=

2.6 X lO-'*

Day 7

Flowering; seed
development

1.7 X 10-5

1.8 X 10- =

1.8 X 10-"

1.4 X 10-'

Dav 14

Seed filling

1.8 X 10-=

4.4 X 10-=

3.5 X 10-=

1.4 X lO-''

Day 21

Seed filling
completed

<5.0x 10-'

4.3 X 10- =

4.7 X 10-=

4.0 X 10-*

Average for

1.5 X 10-=

3.7 X 10- =

8.7 X 10-=

4.2 X 10-"

leached plants

*D. A. Cataldo, unpublished data.

t All compounds were supplied from solutions at pH 5.8 to 7.0; aerosol characteristics for the fresh
and hydrated oxides are given in Table 1 ; aerosols of plutonium citrate had an AM AD of 1.61 Mm and
a GSD of 1.86, and the count mode was 0.200 jum; plutonium nitrate had an AM AD of 2.29 /nm and a
GSD of 1.91, and the count mode was 0.152 ^m. Plants were leached with 200 ml of solution
(equivalent to a 7-min rainfall of 0.4 cm).

^Transport ratio = picocuries per gram of seed tissue/picocuries per gram of contaminated leaf
tissue; average of four replicate samples.

§Aged in H^O at pH 7.0 for 10 months.

TABLE 4 Extent of Translocation of ^ ^ ^ Pu from Contaminated Foliage of
Bush Bean to Root Tissues in the Absence and Presence of a Solution Vector*t

Time of leaching

Transport

ratio t

after

Stage

Pu oxide

contamination

of development

(fresh)

(hydrated) g

Pu citrate

Pu nitrate

No leach

<6.6x 10-*

<4.8 X 10- =

7.3 X 10-*

3.3 X 10-=

Day 1

Preflowering

2.2 X 10-=

1.6 X 10-=

4.7 X 10-=

1.1 X 10-"

Day 7

Flowering; seed
development

1.4 X 10-=

2.2 X 10-=

5.7 X 10-=

4.6 X 10-"

Day 14

Seed filling

1.1 X 10-=

4.3 X 10-=

7.1 X 10-=

1.4 X 10-=

Day 21

Seed filling
completed

1.6 X 10-=

1.9 X 10-=

1.7 X 10-"

3.3 X 10-=

Average for

1.6 X 10-=

2.5 X 10-=

8.6 X 10- =

1.5 X 10-"

leached plants

*D. A. Cataldo, unpublished data.

t All compounds were supplied from solutions at pH 5.8 to 7.0; aerosol characteristics are given in
Table 3. Plants were leached with 200 ml of solution (equivalent to a 7-min rainfall of 0.4 cm).

:}: Transport ratio = picocuries per gram of root tissue/picocuries per gram of contaminated leaf
tissue; average of four replicate samples.

§ Aged in HjO at pH 7.0 for 10 months.

296 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

reported by Cataldo, Klepper, and Craig (1976). The objective was to evaluate, on the
basis of chemical form supplied and the presence or absence of a solution vector
(simulated rainfall), the extent of absorption and translocation of foliarly applied
plutonium. All plants were exposed to plutonium at 20 days from planting (preflowering)
and were held for an additional 28 days to allow time for both absorption of plutonium
and seed filling. During this 28-day period, contaminated plants were either maintained in
the absence of a simulated rainfall or subjected to a simulated rainfall at 1,7, 14, or 21
days after contamination. This simulated rainfall provided a solution vector on the
surfaces of contaminated leaves to enable diffusion and absorption of mobile plutonium
forms. The pots containing soil and root were double bagged with polyethylene and
sealed at the lower stem; the seed tissue was contained in pods that were formed after
exposure. The effect of both the chemical form of the plutonium and the timing of
simulated rainfall on the absorption of plutonium from foliar surfaces and its
translocation was determined by analysis of uncontaminated seed and root tissues.
Quantitation was by means of transport ratios: TR = picocuries of ^'^^Pu per gram (dry
weight) of seed or root ^ picocuries of ^ ^ ^ Pu per gram (dry weight) of contaminated leaf
tissue. Although there is a tendency to compare TR values with classical concentration
ratios derived from plants grown on contaminated soils, this comparison is inappropriate.
At harvest the dry weights of contaminated leaves and uncontaminated roots and seeds of
individual plants were approximately 0.6, 0.5, and 0.3 g, respectively. Plants were
contaminated with 0.25 to 1.0 x 10^ d/min ■^^^Pu. Transport ratios (TR values) for root
and seed tissues from plants not subjected to leaching (solution vector) ranged from
< 4.5 X 10~^ to 3.3 X 10~'^. Application of a simulated rainfall to provide a solution
vector for diffusion and absorption of soluble components on the leaf surface increased
uptake and transport of plutonium to seed and root tissues for all compounds of
plutonium studied (1.5 X 10~^ to 4.2 X 10"''). Apparent differences in translocation
between the various plutonium forms may result from the relative size of the soluble
fraction. The fresh plutonium dioxide was truly particulate at the time of contamination,
the aged oxide consisted of particles with a fractured crystal lattice (Park et al., 1974),
the citrate represented a relatively stable soluble complex, and the nitrate represented an
unstable complex that rapidly hydrolyzed on dilution to form colloidal hydroxides. The
order of bioavailability for transport to seed and root was plutonium nitrate
(hydroxide) > plutonium citrate > aged oxide > fresh oxide. (Tables 3 and 4 show
average values for leaching treatment.)

An interesting aspect of these data is that maximum TR values are obtained when the
simulated rainfall occurs at day 7 or 14, the time of rapid seed development. This
phenomenon is of interest from the standpoint of the mobility of plutonium within the
plant and the chemical form of the plutonium. It is generally accepted that materials
must be transported out of mature leaves in the phloem. The entry of molecules into this
transport conduit is metabolically regulated, and the loading process is highly specific for
individual organic metabolites and inorganic elements (Crafts and Crisp, 1971). There is
growing evidence that many inorganic nutrilites, especially multivalent cations, are
transported as organic complexes in both the xylem (Tiffin, 1967; 1971 ; Bradfield, 1976)
and phloem (van Goor and Wiersma, 1976). By analogy to the behavior of nutrilites,
plutonium must be transported out of the contaminated leaves via the phloem. Similarly,
it is unUkely that inorganic plutonium could remain soluble at the pH of phloem cell sap
(pH 7.2 to 8.5) (Ziegler, 1975). Therefore it is possible that the mobile plutonium which
was deposited in seed and root tissues may have been complexed with phloem-mobile

AIRBORNE PLUTONIUM 297

organic species. This would explain the apparent increase in TR values seen during the
time of seed development. During this period there is a significant change in both the
composition and quantity of specific metabolites being produced by leaves and being
exported to metabolic sinks, such as seeds and roots. This change in metabolism may
increase the potential for soluble species of plutonium to become complexed with organic
metabolites and subsequently to be exported to metabolic sinks. Although this is a
tentative judgment and subject to substantive studies, this interpretation serves to explain
the observed results on the basis of known metabolic aspects of plant function.

Conclusions

The ability of terrestrial plants to accumulate potentially hazardous elements from soils
via root absorption and the relative importance of these elements in the food web to man
has prompted numerous studies over the past 25 years. The majority of these
investigations have been concerned with soil— plant transfer rates since the soil represents
a major repository for pollutants released to the environment and since the plant root is
an efficient solute-absorbing structure. Until recently the foliar portions of plants were
considered to play a minor, transient role at best with respect to dose-assessment
problems.

Our current understanding of the aerodynamic behavior of particles and anticipated
reductions in particle-size distributions of materials such as plutonium through an
expanded nuclear energy program suggests that a reevaluation of the role of plant foliage
in particle interception and absorption of materials contained on airborne particles is in
order.

This need is supported by both early investigations and studies currently under way.
Early studies of worldwide fallout and current work on contaminated soils resuspended
by wind indicate that foliar retention and foliar absorption may be as important as, and
in some cases exceed, the role of root absorption with respect to food-chain transport. A
critical evaluation of past literature on the leaching of foliar deposits suggests that aerosol
polydispersity and large particle size (e.g., 45 ^m MMD) may explain the comparatively
large degree of leaching or weathering reported for comparatively large particles
(Witherspoon and Taylor, 1969; 1970; 1971). This view is reinforced by data reported for
well characterized particles of lead and plutonium in laboratory studies and field
observations for fallout plutonium. These latter investigations indicate that a sizeable
fraction (>80%) of submicron-size particles deposited onto foliage are tenaciously held
on leaf surfaces under varied conditions (e.g., simulated rainfall and wind). Aside from
the potential health implications associated with increased foliar retention, the problem
of foliar absorption must be considered. In the reported studies, a substantial fraction of
the foliar plutonium deposits was transported to seed and roots. Transport ratios were
affected by both the presence of a solution vector (simulated rainfall) and the timing of
its application with respect to the stage of plant development.

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298 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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The Relationship of Microbial Processes
to the Fate and Behavior
of Transuranic Elements in Soils, Plants,
and Animals

R. E. WILDUNG and T. R. GARLAND

Soil physico chemical and microbial processes will influence the long-term solubility,
form, and bioavailability of plutonium and other transuranic elements important in the
nuclear fuel cycle. Consideration is given to the chemistry /microbiology of the
transuranic elements in soil, emphasizing possible organic complexation reactions in soils
and plants and the relationship of these phenomena to gastrointestinal absorption.

Initial solubility of the transuranic elements in soil is governed largely by hydrolysis
with soil sorption in the order Pu > Am — Cm > Np. Soluble (<0.01 jim), diffusible
plutonium in soils (usually less than 0.1% of total) appears to be largely present as
particulates of hydrated oxide, but several lines of evidence indicate that microorganisms
influence the solubility and plant availability of plutonium and that the nonparticulate
plant-available fraction is stabilized in solution by inorganic or organic ligands of limited
concentration in soil. Vie possible role of soil microorganisms in influencing the
solubility, form, and plant availability of the transuranic elements is discussed on the
basis of the (1 J known chemistry of organic ligands in soils: (2) effects on the soil
microflora; and (3) principal microbial transformation mechanisms, including direct
alteration (valence state and alkylation), indirect alteration (metabolite interactions and
influence on the physico chemical environment), and cycling processes (biological uptake
and release on decomposition of tissues).

The toxicity of plutonium to microorganisms depends on plutonium solubility in soil.
However, soil microorganisms are generally resistant to plutonium; toxicity is due
principally to radiation rather than to chemical effects. Highly resistant bacteria, fungi,
and actinomycetes have been isolated from soil, and these organisms liave been shown to
be capable of transporting plutonium into the cell and altering its form in the cell and in
solution. Vie resulting soluble plutonium complexes exhibit a range of mobilities in soil
and tend to be of higher molecular weight than simple complexes (plutonium-
diethylenetriaminepentaacetic acid) and negatively charged. Vie forms of plutonium
complexes, although not well defined, are dependent on organism type, carbon source,
and time of plutonium exposure during growth. Viese factors, in turn, are a function of
plutonium source, soil properties, and soil environmental conditions. Knowledge of the
relative influence of these factors serves as a valuable basis for predicting the long-term
behavior of plutonium and other transuranic elements in soils. There is growing evidence
that these phenomena also markedly influence the availability of plutonium to plants and
animals.

Plutonium present in solution as an organic complex is readily assimilated by the
plant in the Pu(IV) state. Evidence to date indicates that soil sorption rather than plant
discrimination limits plant uptake of plutonium and that organometal complexes serve
mainly to deliver plutonium to the root membrane; i.e., the ligands are not taken up by

300

RELATIONSHIP OF MICROBIAL PROCESSES 301

the plant stoichio metrically with the metal. After passing the root membrane, Pn(IV) is
translocated to the shoots in the xylem through formation of a number of organic
complexes with plant ligands. The form of plutonium differs in leaves and stems, but
greater than 90% of the soluble plutonium associated with these tissues was present as
complexed Pu(IV) after growth on soil to which ionic Pu(IV) had been added.

A reevaluation of plant -to -animal transfer coefficients used presently in dose
assessments may be required since plutonium incorporated in plant tissues is markedly
more available to animals than Pu(IV) gavaged in the inorganic plutonium solutions that
were used for previous measurements. Differences in gastrointestinal transfer of
plutonium in stems and leaves of alfalfa are related to differences in plutonium solubility
in these tissues. Tims the form of plutonium in soils and plants may be closely related to
plutonium availability to 'animals.

Although information leading to an understanding of the complex biochemical
interrelationships that exist between soils, plants, and animals is rapidly developing, these
phenomena are not sufficiently understood at present to be described by simple models.

The major factor governing availability of the transuranic elements to plants will be their
solubility in soil since, for root uptake to occur, a soluble species must exist adjacent to
the root membrane for some finite period. The form of this soluble species will have a
strong influence on its stability in soil solution, on its mobility in soils, and on the rate
and extent of uptake and, perhaps, on its mobility and toxicity in the plant.
Furthermore, the results of preliminary studies discussed in this chapter suggest that the
concentration and chemical form of the element in the plant play a major role in
influencing its availabiUty to animals on ingestion. Thus any assessment of the long-term
behavior of the transuranic elements in the terrestrial environment must be based on the
determination of the factors influencing solubility and on the form of soluble species in
soil. These factors, illustrated in Fig. 1, include the concentration and chemical form of
the element entering soil; the influence of soil properties on the elemental distribution
between the solid and Hquid phase; and the effect of soil processes, such as microbial
activity, on the kinetics of sorption reactions, transuranic concentration, and the form of
soluble and insoluble chemical species.

Portions of the soil chemical and microbiological sections of this chapter have been
published elsewhere by the U. S. Department of Energy (Wildung, Drucker, and Au,
1977) and the American Society of Agronomy (Keeney and Wildung, 1977). However, at
the request of the editor these have been reiterated to provide a complete treatment of
the subject.

ANIMALS
PLANTS

SOIL ^ / *'-'"- \ M SOIL

PROCESSES^ /SOLUTION \ ^PROPERTIES

SOURCE
Fig. 1 Factors influencing transuranic behavior in the terrestrial environment.

302 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Transuranic Qiemistry in Soil

Sources of the Transuranic Elements

The transuranic elements of principal importance in the nuclear fuel cycle (plutonium,
americium, curium, and neptunium) can enter the soil through several avenues (Vaughan,
Wildung, and Fuquay, 1976), including (1) fallout from atmospheric testing, (2) possible
escape of airborne particulates and liquid effluents during reprocessing of spent fuels and
fuel fabrication, and (3) leaching from waste-storage facilities. The major sources of the
transuranic elements can be classified according to expected initial solubility in soil:
Insoluble source terms

MOx + L ^ ML

Soluble source terms
Hydrolyzable

(1) M(N03)x + L + H2 0

MOx • nH2 0

or
. M(OH)x

l + ML

Nonhydrolyzable

(2) M(N03)x + L + H2 O ^ MO; + ML

Organic complexes

(3) MLi +L2 +H2 0^MLi +ML2 + MLi,2

MLi + L2 + H2 O ^ as m (1 ) or (2)

where M represents transuranic elements and L represents inorganic and organic ligands
capable of reacting with transuranic elements and forming soluble or insoluble products.
Particulate oxides of the transuranic elements initially can be expected to be largely
insoluble in the soil solution. Ultimately, solubility is expected to be a function of the
composition, configuration, and equivalent diameter of the particle as well as soil
properties and processes. Oxide particles of the highest specific activity and containing
the highest concentrations of impurities in the crystal lattice may exhibit the greatest
solubility. The combination of configuration and equivalent diameter as reflected in
surface area exposed to solution will be the other main factor governing oxide solubility.
Once solubilized, the transuranic elements will be subject to the chemical reactions
governing soluble salts. Hydrolyzable transuranic elements entering the soil in acid
solutions sufficiently concentrated to maintain soluble ions can be expected to be rapidly
insolubilized as a result of hydrolysis on dilution and subsequent precipitation on particle
surfaces. These include Pu(in, IV, and VI), Am(III), Cm(III), and Np(III, IV, and VI).
Conversely, transuranic elements not subject to marked hydrolysis can be initially more
soluble. These include Pu(V) and Np(V). Immobilization of these chemical species (PuO:

RELATIONSHIP OF MICROBIAL PROCESSES 303

or Np02) can occur through cation-exchange reactions with particulate surfaces.
CompHcating this situation, disproportionation and complexation reactions may occur

concurrently.

Transuranic elements entering the soil as stable organocomplexes. such as might occur
in the vicinity of a spent -fuel separation facility, may be initially highly soluble (Wildung
and Garland, 1975). The duration of solubility and mobility in the soil will be a function
of the stability of the complex to substitution by major competing ions, such as calcium
and hydrogen (Lahav and Hochberg, 1976; Lindsay, 1972; Norvell, 1972), and the
stability of the organic ligand to microbial decomposition (Wildung and Garland, 1975).
The disruption of the complex may lead to a marked reduction in transuranic-element
solubility through hydrolysis and precipitation reactions, as described for acid solutions
on dilution. A portion of the ion released may react with other, perhaps more stable,
ligands in soil. The mobility of the intact complexes, in turn, will be principally a
function of their chemical and microbiological stability and the charge on the complex,
which will govern the degree of sorption on soil particulates.

Further generalizations of transuranic-element behavior on the basis of source terms
are complicated by the overwhelming importance of soil properties and processes in
influencing transuranic-element behavior on a regional and local basis. This chapter
considers, in detail, the influence of soil properties and abiotic and biotic processes on the
long-term solubility of the transuranic elements entering soils. Consideration is also given
to the implications of these processes in terms of transuranic-element plant and animal
availability. Principal emphasis is directed toward the role of soil microorganisms in this
phenomenon. Microorganisms, in intimate association with soil particles, are known to
play an important role in effecting solubilization of elements considered insoluble in soils
strictly on the basis of their inorganic chemistry. To date studies of the microbiology of
the transuranic elements have been limited principally to plutonium. This chapter will
emphasize plutonium. but, where possible, the available information is used as a
framework for broader discussions encompassing the long-term behavior of other
transuranic elements.

Chemical Reactions Influencing Plutonium Behavior

The principal chemical reactions likely influencing plutonium behavior in soil are:

• Four oxidation states

Pu3\ Pu'\ PuOr. PuOr (Pu'\ Pu^')

• Disproportionation

Pu''' + PuOr^Pu'' + PuOr

• Hydrolysis

Pu^' + 4H2 0^Pu(OH)4 +4H^(Ksp ^10^'^

• Complex formation

2Pu^' + 3 DTPA^Pu.DTPAj (logK~ 18)

304 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Plutonium ions may commonly exist in aqueous solution in valence states III, IV,

V(Pu02), and VKPuOf*). Other valence states are known (II and VII) and predicted

(VIII), but these occur under unique conditions (Cleveland, 1970). Disproportionation

reactions are common, and, due to kinetic factors, plutonium is unique among the

chemical elements in that it may simultaneously exist in all the common valence states.

The tendency of plutonium to hydrolyze in aqueous solutions of low acidity follows the

order Pu'*'' > PuOi^ > Pu^"" > ?\xO\ (Cleveland, 1970). Hydrolysis, which occurs in a

stepwise fashion, is likely the major mechanism whereby plutonium is insolubilized in the

environment. At high (grams per liter) plutonium concentrations, hydrolysis of Pu"*^ may

lead to the formation of a colloidal plutonium polymer. At these concentrations the

polymer is characterized by a distinct absorption spectrum. Although the polymer has

not been fully characterized, it is generally thought to be an intermediate hydrolysis

product of Pu^"^ containing oxide or hydroxide bridges with an absorption spectrum

different from that of Pu(0H)4. However, studies by Lloyd and Haire (1973) indicated

that the polymer may be aggregates of small, discrete, amorphous or crystalline, primary

particles of 5 to 20 A in diameter. It is of interest that X-ray-diffraction patterns of the

polymeric plutonium and that of Pu(0H)4 (Ockenden and Welch, 1956) showed a

pattern characteristic of the cubic Pu02 lattice, which suggests that the polymer and the

hydroxide of Pu'*'^ may be hydrated Pu02 with differences occurring in primary particle

size and crystallinity (Lloyd and Haire, 1973). The formation of the hydrated PuOo is

likely directly related to Pu'*" concentration and inversely related to the acid

concentration.

Plutonium also tends to form many complexes with a range of stabilities. The

strongest complexes are generally formed by reaction of organic ligands with Pu^*.

However, many inorganic complexes and organic complexes of all valences may be stable

under appropriate conditions. The presence of organic ligands in soils likely influences the

equilibrium and concentration form of plutonium in solution through complexation and

subsequent inhibition of hydrolysis, polymerization, or disproportionation. These

reactions, in various highly complex combinations resulting from differences in source

term, soil properties, and processes, govern plutonium solubility in soil and availability to

plants.

Soil chernical reactions are important in governing the behavior of the various torms

of plutonium entering soil. Initially, soluble forms entering soil have the potential for

undergoing a range of chemical transformations. Insoluble plutonium, such as high-fired

oxide, entering soil likely will be solubilized with time, provided that soluble, stable

complexes are formed. However, regardless of the form of plutonium entering soil, its

ultimate solubility will be controlled by its aqueous chemistry and by soil factors. Soil

physicochemical properties can be expected to have complex, interdependent effects on

plutonium solubility. The long-term behavior of plutonium in soil will be a function of

the kinetics of these reactions.

On the basis of research with other trace metals, recently summarized by Keeney and

Wildung( 1977), and limited information on the transuranic elements, it can be concluded

that the soil physicochemical parameters most important in influencing the solubility of

the transuranic elements include (1) solution composition. Eh and pH; (2) type and

density of charge on soil colloids; and (3) reactive surface area. These phenomena will, in

turn, be dependent on soil properties, including particle size distribution, organic-matter

content, particle mineralogy, degree of aeration, and microbial activity. The delineation

ot the influence of these factors on plutonium solubility is difficult owing to the complex

chemistry of plutonium.

RELATIONSHIP OF MICROBIAL PROCESSES 305

A reasonable approach to the study of the chemistry of plutonium in soil is to direct
initial attention to the factors influencing its solubility in soil. However, plutonium
solubility in soil is difficult to define because solubility will depend on the method of
measurement and because solubility must be arbitrarily evaluated because of the sorption
of plutonium on submicron clay particles and the formation of submicron particles of
hydrated plutonium oxide, which are difficult to centrifuge and may pass membrane
filters. These effects can be illustrated by comparison of the differences in the solubility
of plutonium in soils [100 days after amendment as Pu(N03)4] as determined by water
extraction and subsequent membrane filtration with the use of membranes of different
average pore sizes (Table 1). The major fraction of the plutonium added was sorbed on

TABLE 1 Solubilities of Plutonium in Water Extracts

of a Ritzville Silt Loam as Determined by Filtration

with Membranes of Different Pore Sizes*

Membrane pore size,

PiutOR

ium solubility ,t

Pg/g

60,000

0.45

20,000

0.01

4,000

0.0015

1,000

0.0012

300

0.0010

*I'rom Garland and Wildung (1977).

f Plutonium added at a level of 620,000 pg/g of soil.

the soil since a maximum of 10% of the extracted plutonium passed through the 5-/jm
membrane. Successive filtration through membranes with decreasing pore size resulted in
decreases in plutonium concentration in the filtrate. Thus plutonium in the aqueous
extract appeared to-be in a wide range of particle sizes. Although membranes with pore
sizes of 0.45 ;um are commonly used to separate soluble matter from particulate matter,
plutonium in these filtrates may be in colloidal forms. The plutonium in the 0.001 0-/im
filtrate appeared soluble, was stable in solution, and approximated the quantity of
plutonium taken up by plants (Wildung and Garland, 1974). Of the soluble plutonium
forms likely to enter soils (previous section), Pu(N03)4 and plutonium-diethylenetri-
aminepentaacetic acid (DTPA) represent, in their respective chemistries, the range in soil
behavior likely to occur. The water solubility (<0.01 /jm) of ^^^Pu and ^^^Pu amended
to a Ritzville silt loam (organic C content. 0.7%: pH 6.2) in the Pu(N03 )4 and Pu-DTPA
forms differs markedly (Wildung and Garland, 1975). The DTPA complexes of both
isotopes were water soluble in soil and appeared to be stable over the tlrst 40 days of
incubation (Fig. 2). After 7 days of incubation, the ^-^ ^Pu-DTPA appeared to be slightly
less soluble than the ^^"^Pu-DTPA. After 95 days of incubation, both isotopes, initially
added as the complex, appeared to decrease in solubility, perhaps as a result of microbial
degradation of the organic moiety and the development of new chemical equilibria.

Equilibrium concentrations of soluble plutonium added as the nitrate were not
obtained until 7 to 10 days. The solubility of ^''^Pu and ^^^Pu added to the soil as
nitrates was much lower than the DTPA complexes, which likely reflects hydrolysis to

306 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

100

LLI

_l
03
D
_I
O

I-

0.1

0.03

0 239pjj_DJPA (144 ^g/g)

A 239p^_[3TPA (1 ^g/g)

O 238pu_DTPA (0.6 Mg/g)

H 238pu_pHrate (0.6 mq/q)

□ 239p^J_p|trate (144 Mg/g)

20 40 60 80

INCUBATION TIME, days

100

120

Fig. 2 Solubility of plutonium with time after addition to soil as the nitrate and the
DTPA complex. [From Wildung and Garland ( 1975).]

the largely insoluble hydrated oxide. It is clear that organic ligands have a pronounced
effect on plutonium solubility in soil. The rate of decrease in the solubility of each
isotope added as the nitrate was similar. However, in contrast to the slightly lower
solubility of the ^^^Pu-DTPA compared with the ^^'^Pu-DTPA. ^^^Pu added as the
nitrate was a consistent factor of 2 to 3 times as soluble as ^■'^Pu initially added as the
nitrate. This difference probably resulted from the formation of larger hydrated oxide
particles at the higher plutonium concentration (^^^Pu), but it may also have reflected
the presence of soil components, such as organic ligands, which stabilized plutonium in
solution but were present in limited concentrations and became important only at lower
plutonium concentrations.

The water solubility of ^■'^Pu, when incorporated in relatively large plutonium oxide
particles (>1 /jm), would be expected to be greater than the solubility of '^^^Pu oxide
particles of similar size as a result of crystal damage and radiolysis arising from the greater
specific activity of the ^^^Pu (an approximate factor of 270). However, the behavior of
the two isotopes in soil on solubilization of the oxide might be expected to follow a
course similar to that exhibited by the nitrates (Fig. 2).

Equilibrium solubility after 6 days of incubation (Garland. Wildung, and Routson,
1976) of plutonium, added as Pu(N03)4, in soils of different properties occurred after
approximately 20 hr (Fig. 3). The quantities of plutonium soluble at equilibrium in water

RELA TI ON SHIP OF MICR OBI A L PR OCESSES 307

6200 -

g 4960

O 3720

I-
D

D
_l
O

2480

1240

Muscatine

- /

>*^^

/ ^

/^ Ritzville

_ — a

• — ^

/ ^

n //

1 J< ./^

rfy^

Muscatine

^» —

— ■ — •

111,1

20 40

TIME OF CONTACT, hr

Fig. 3 Quantity of soluble plutonium removed from three soils by 0.0171/ CaCK . Soil :

solution = 1 : 100. o o, distilled water. • •, O.OlAf, CaCl.. [From Garland,

Wildung, and Routson (1976).]

and 0.0 IM CaCl2 differed with soil type. In the CaCli solution, solubility was lowest in
the Muscatine soil, which exhibited higher silt and clay content than the other soils.
Importantly, at equilibrium there was more plutonium extracted by water than by O.OIM
CaClo in the Muscatine soil. The Hesson and Ritzville soils did not exhibit this property.
This may be related to a difference in the dispersibility of fine colloids in this soil and/or
the presence of higher concentrations of stabilizing ligands. However, the lack of a
proportional dilution effect (not shown in Fig. 3) in the water extractability of
plutonium at lower solution-to-soil ratios in this soil, as compared with that in the
Ritzville and Hesson soils, provided presumptive evidence for the presence of a dispersible
ligand in higher concentration in the Muscatine soil.

Applying diffusion principles to characterization of mobile plutonium species in soils.
Garland and Wildung (1977) estimated the concentrations and molecular weights of
mobile plutonium in five surface soils representing a range in particle size distributions.
pH (4.4 to 6.2), organic C (0.7 to 12.5%), and cation-exchange capacities (14 to
45 meq/100 g). Diffusion coefficients were calculated from measurements of the rate and
extent of plutonium migration from soil through an agar matrix. The diffusion
coefficients calculated for the most mobile species in the five soils varied from 1.5 to
3.0 X 10"^ cm^/sec (Table 2). Estimated concentrations and molecular weights of the
most mobile plutonium components in the five soils ranged from 9 to 55 pg/g and from
5000 to 21,000 g/mole, respectively. Thus estimated concentrations of the most mobile
plutonium species were of the same order of magnitude as those observed by water
extraction and subsequent ultrafiltration through the 0.001 0-)Um membrane (Table I).
This membrane retained Pu-DTPA (molecular weight. 1700). Hypothetical globular
peptides of molecular weights of less than 500 would pass through this membrane.
However, if the molecule were a hydrated PuO^ sphere of similar dimensions, it would

308 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Estimated Concentrations and Molecular Weights of ^Mobile Plutonium
in Soils from Measured Diffusion Coefficients*

Most mobile

species

Least mobile

species

Diffusion

coefficient,

, Molecular

Soil

coefficient.

Molecular

Soil

cm ^ /sec

weight.

concentration,+

cm ^ /sec

weight,

concentration,t

Treatment

(x IQ-*)

g/mole

Pg/g

(x lO"')

g/mole

Pg/g

Control

Pu-DTPA

5.8

1,700

Soils

■

Ritzville

3.0

5,000

2.3

0.9 X 10«

150

Quillayute

2.5

7,200

2.7

0.7 X 10«

1,200

Hesson

2.4

8,100

2.7

0.6 X 10*

330

Salkum

1.5

21,000

2.3

0.8 X 10*

340

Muscatine

1.9

13,000

3.1

0.5 X 10*

170

*From Garland and Wildung (1977).

fPlutonium added at a level of 620,000 pg/g of soil.

have a molecular weight of between 10,000 and 25,000, which approximates the
molecular weights of the most mobile plutonium species as determined from diffusion
coefficients. This fraction, therefore, likely consisted of small particles of Pu(0H)4 or
hydrated oxide.

The estimated diffusion coefficients for the least mobile plutonium components
ranged from 2.3 to 3.1 x 10"^ cm' /sec with corresponding soil concentrations of 150 to
1200 pg/g (Table 2). This concentration of plutonium in soil approximated the quantity
of water-soluble plutonium passing the 0.001 5-jum ultrafiltration membrane (Table 1).
Hypothetical globular proteins in this size range would have average molecular weights of
< 10,000. Particles of Pu(0H)4 or hydrated oxides would have molecular weights of
200,000 to 500,000. Estimated molecular weights for these least mobile species
calculated from diffusion coefficients were between 600,000 and 900.000. Thus it would
appear, as in the case of the most mobile species, that the least mobile species of
plutonium were particulate Pu(0H)4 or hydrated oxides.

The comparison of filtration and diffusion data indicates that the mobile plutonium
in incubated soils was in the form of hydrated oxide or hydroxide in a continuum of
sizes. If it can be assumed that plutonium in particulate form was not available to plants,
it is possible that the small fraction of plutonium taken up by plants was present in soil as
reaction or dissolution products with insufficient stability and/or concentration to be
detected by the methods used. Insight into this possibility was not provided by
comparison of plutonium behavior in different soils, as might be expected, because the
estimated concentrations and molecular weights of the mobile species were not related to
the soil properties measured.

Several conclusions can be drawn from studies of the soil chemistry of plutonium
which have important implications in terms of the potential role of the soil microbiota in
influencing plutonium behavior in soil. The definition of plutonium solubility by
filtration or diffusion alone is compHcated by plutonium chemistry, but, in conjunction,
the measurements suggest that mobile plutonium is largely particulate. However, a
fraction of the mobile plutonium is available to plants.

RELATIONSHIP OF MICROBIAL PROCESSES 309

This material is obviously not particulate but is present in insufficient concentration
for characterization with current methods. The question remains, "What is the form of
the small quantity of plutonium available to plants?" This information is essential to
understanding the mechanisms whereby plutonium can be resupplied to solution from the
solid phase in a range of soils and to predictions of the long-term availability of
plutonium to plants. From investigations of plutonium valence state in a neutral,
0.0004M NH4HCO3 solution equilibrated with PUO2 microspheres and in burial-ground
leachates, Bondietti and Reynolds (1976) concluded that Pu(VI) may be stable in
significant quantities in solution and suggested that monomeric Pu(VI) and its complexes
may be important in plutonium mobilization. In the present studies, evidence was
presented which suggested that plutonium ions are stabilized in soil solution by inorganic
or organic ligands for subsequent uptake by the plant. Furthermore, equilibration of
weathered plutonium-contaminated soil with chelating resins has been shown (Bondietti,
Reynolds, and Shanks, 1976) to result in significant desorption of plutonium from the
solid phase. It is known that organic ligands result in the most stable plutonium
complexes. Soluble organic ligands in soil are generally derived from microbial processes.

Chemical Reactions Influencing the Behavior of Other Transuranic Elements

Other transuranic isotopes of concern in the nuclear fuel cycle include ^^^ Am, ^'*^Am,
^'^^Cm, ^"^^Cm, ^'*'*Cm, and ^^''Np. Althougli detailed studies of their interaction with
soils are lacking, some information has become available in recent years. Furthermore, the
aqueous chemistries of these elements have been fairly well established (Katz and
Seaborg, 1957). The most stable ions of americium and curium in aqueous solutions are
the cations (III); neptunium is most stable as the oxyion (Np02). Disproportionation is
not common with these elements. Thus major differences in their environmental behavior
as compared with that of plutonium would be expected. Hydrolysis reactions may still be
a primary factor governing the environmental behavior of americium and curium, but
greater mobility and plant availability in soils might be predicted on the basis of greater
solubility of the hydroxides in comparison with Pu(0H)4. The neptunium oxycation
should not be subject to significant hydrolysis at environmental pH values (Burney and
Harbour, 1974). Of the transuranic elements, neptunium has been the least studied, but,
because of its chemical characteristics, it may be the most available to the biota. A
comparison of plutonium, americium, and neptunium sorption in several soils (Routson,
Jansen, and Robinson, 1975) indicated sorption in the order Pu > Am > Np. The
chemistry of curium should be very similar to that of americium if present at equal mass
concentrations.

Organic Complexation Reactions

Research to date on the chemistry of the transuranic elements in soil has pointed to the
importance of understanding transuranic-element organic complexation reactions in soil,
particularly in surface soils and aquatic sediments where organic-matter content is
generally highest or in subsoils where the transuranic elements may be dispersed in
conjunction with synthetic complexing agents. Very little information is available
concerning the interaction of the transuranic elements with the soil organic fraction.
However, despite the difficulties in characterization of soil organic complexes, much is
known both theoretically and experimentally regarding the interactions of metals with
functional groups of soil organic matter (Keeney and Wildung, 1977). Much of this

310 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

information concerns micronutrients of greatest agronomic importance (B, Co, Cu, Fe,
Mn, Mo, Se, and Zn) and has been the subject of a number of excellent reviews over the
last two decades (Mitchell, 1964; 1972; Mortensen, 1963; Hodgson, 1963; Stevenson and
Ardakani, 1972). Earlier studies generally emphasized metal interactions with intact soil
or with the higher molecular-weight humic components of soil, whereas recent studies
emphasize the more soluble components of soil.

It is practical to categorize metal complexes in soil in terms of their solubility since,
in general, it is this factor, as previously noted, that most influences their mobility and
plant availability. Three principal categories have been proposed (Hodgson. 1963),
althougli the complexity of the soil system results in considerable overlap between
categories. These categories include the (1) relatively high-molecular-weight humic
substances containing condensed aromatic nuclei in complex polymers derived from
secondary syntheses which have a high affinity for metals but are largely insoluble in soil,
(2) low-molecular-weight organic acids and bases classified as nonhumic substances and
derived largely from microbial cells and metabolism which demonstrate relatively high
solubility in association with metals, and (3) soluble ligands that are precipitated on
reaction with metals.

Humic Substances. Humic substances are generally divided into three categories based
on their solubilities (Felbeck, 1965). The humin (alkaU and acid insoluble) fraction is
soluble only under drastic conditions and is apparently of the highest molecular weight.
The humate (alkali soluble and acid insoluble) and fulvate (alkali and acid soluble)
fractions of soil may constitute up to 90% of the soil organic fraction (Kononova, 1966).
The humates and fulvates are characterized, in part, by a high charge density due to acidic
functional groups (Stevenson and Ardakani, 1972; Felbeck, 1965). This property leads to
a high degree of reactivity, and these materials exhibit a strong pH-dependent affinity for
cations in solution and are likely strongly bound to soil minerals and other organic
constituents in soil (Greenland, 1965). The acidic functional groups consist principally (in
general order of acidity) of carboxyl, hydroxyl (phenolic and alcoholic), enolic, and
carbonyl groups (Broadbent and Bradford, 1952; Felbeck, 1965; Schnitzer, Shearer, and
Wright, 1959). Total acidity has been estimated to range between 500 to 900 and 900 to
1400meq/100g for humic acids and fulvic acids, respectively (Stevenson and Butler,
1969). The acidic hydrogen of humic acids was differentiated by Thompson (1965) into
diree groups at 100 to 200, 500 to 700, and 1000 to 1200 meq/100 g using nonaqueous
titration methods. Basic functional groups, likely amides and heterocyclic nitrogen
compounds (Bremner, 1965), probably also contribute to retention of metals but are of
much less importance than acidic groups at most soil pH values.

In batch equilibration studies (Bondietti, 1974), calcium-saturated humates removed
greater than 94% of the Pu(IV) from pH 6.5 aqueous solutions (compositions not given).
It is unclear whether the humates represented a surface for precipitation of hydrolyzed
species or were involved in complexation of plutonium. However, in studies of plutonium
desorption from humates and reference clays, citrate removed 10 to 30% of sorbed
plutonium from the clays but less than 1% of that from the humic acids. Ligands forming
stronger complexes with plutonium [DTPA and ethylenediaminetetraacetic acid
(ETDA)] were required to remove significant quantities (up to 30%) of the plutonium
from the humate complex.

Although humic and fulvic acids likely account for most of the metal immobilization
attributed to the soil organic matter (e.g., Hodgson, 1963; Stevenson and Ardakani,

RELATIONSHIP OF MICROBIAL PROCESSES 311

1972), they have the potential for formation of soluble complexes with metals,
particularly in dilute solutions. Small quantities of metal fulvates, thought to be of lower
molecular weight than the humates. may be present in soil solution. A nondialyzable
material with infrared absorption spectra and elemental analyses similar to fulvic acids
was isolated from a dilute salt (O.OIM KBr) extract of a mineral soil by Geering and
Hodgson (1969). The material exhibited a concentration equivalent to 2.5% of a
dialyzable fraction but was more effective in complexing copper and zinc.

Nonhumic Substances with Potential for Metal Complexation. Lower molecular-weight
biochemicals of recent origin have been implicated in metal complexation and
solubilization in soil. These materials represent (1) components of living cells of
microorganisms and plant roots and their exudates and (2) the entire spectrum of
degradation products which uUimately serve as the building units of the soil humic
fraction. The quantity and composition of these materials will vary with soil, vegetation,
and environmental conditions (Alexander, 1961; 1971). Readily decomposable wastes
disposed to soil under conditions appropriate for microbial growth may, for example,
result in immediate and marked increases in organic materials identified in (1) and longer
term increases of materials in (2). Conversely, toxic materials may have the opposite
effects. The specific compounds produced will be dependent on the properties of the
waste and soil environmental conditions after disposal (Routson and Wildung, 1969).

Althougli the concentration of the transuranic elements and other metals soluble in
the soil solution or in mild extractants is low, often near-minimum detectable levels, the
major portions of copper and zinc were shown to be associated with low-molecular-
weight components. Most of the titratable acidity of this fraction was attributed (Geering
and Hodgson, 1969) to aliphatic acids (<pH 7.0) and amino acids (>pH 7.0).

The production, distribution, and action of organic acids in soil were reviewed by
Stevenson (1967). A wide range of organic acids is produced by microorganisms known
to be present in soil. These include (1) simple acids, such as acetic, propionic, and
butyric, which are produced in largest quantities by bacteria under anaerobic conditions;
(2) carboxyHc acids derived from monosaccharides, such as gluconic, glucuronic, and
a-ketogluconic acids,, which are produced by both bacteria and fungi; (3) products of the
citric acid cycle, such as succinic, fumaric, maUc, and citric acid, which are common
metabolic excretory products of fungi; and (4) aromatic acids, such as p-hydroxybenzoic,
vanillic, and syringic acids, wliich are thought to be fungal decomposition products of
plant lignins. A variety of organic acids have also been reported in root exudates.

Amino acids are the other important group of compounds identified in significant
quantities in the soil solution by Geering and Hodgson (1969) which may be expected to
exhibit strong affinity for metals. The qualitative and quantitative aspects of amino acids
and other nitrogenous components in soils have been reviewed by Bremner (1967). It was
concluded that soil acid hydrolysates do not differ greatly in amino acid composition, but
quantitative differences may occur with differences in soil, cHmatic, and cuhural
practices. A number of acidic and basic amino acids have been reported in soil. However,
it appears that the major portion of amino acid-N that is present in hydrolysates is in
(1) the neutral amino acids, glycine, alanine, serine, threonine, valine, leucine, isoleucine,
and proline; (2) the acidic amino acids, aspartic acid and glutamic acid; and (3) the basic
amino acids, lysine and arginine. Most of the amino acids detected in soil hydrolysates
have also been shown to exist free in small quantities in soils with levels seldom exceeding
2/ig/g. In the soil solution (Geering and Hodgson, 1969) neutral amino acids also

312 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

appeared to predominate. Basic amino acids were not detected, although two acidic
amino acids (aspartic and glutamic acids) were present.

Stevenson and Ardakani (1972) concluded that organic acids and amino acids,
although present only in small quantities in soil, were present in sufficient quantities in
water-soluble forms to play a significant role in the solubilization of mineral matter in
soil. Small quantities of a number of other complexing agents, such as nucleotide
phosphates, polyphenols, phytic acid, porphyrins, and auxins, also exist in soil (pertinent
references summarized by Mortensen, 1963). Complexation with biochemicals of recent
origin would likely be the principal mechanism for microbial mobilization of the
transuranic elements in soil.

Microbial Transformation of the Transuranic Elements in Soil

Potential Mechanisms of Transformation

From the resuhs of limited studies of soil chemistry, microbiology, and plant availability
of transuranic elements in soils and by inference from studies of complexation of other
trace metals in soils, it can be concluded that the soil microflora will play a significant
role in transformations governing the form, and, ultimately, the long-term solubility and
behavior of transuranic elements in soil. Four general mechanisms whereby micro-
organisms may alter the form of trace metals in soil (Alexander, 1961; Wood, 1974) are
(1) indirect mechanisms resulting from metal interactions with microbial metabolites or
changes in pH and Eh; (2) direct transformations, such as alkylation and aUeration of the
valence state through microbial oxidation (use of the metal as an energy source) or
microbial reduction (use of the metal as an electron acceptor in the absence of oxygen);
(3) immobihzation by incorporation into microbial tissues; and (4) release of metals on
decomposition of organic residues.

All these mechanisms may be operational in transformations of transuranic elements
in soils. However, on the basis of present knowledge, it is not possible to draw
conclusions as to their relative importance in affecting the long-term behavior of the
transuranic elements. Since there is a paucity of information available, these mechanisms
will be addressed around a framework of current information that is limited principally to
plutonium.

Microbial Alteration of Solubility in Soil

To provide a preliminary assessment of the potential for microbial alteration of
plutonium solubility in soil under aerobic conditions, Wildung, Garland, and Drucker
(1973; 1974) measured soil respiration rate (an index of soil microbial activity), microbial
types and numbers, and plutonium water solubility in sterile (gamma irradiation) and
nonsterile soils that contained 10 ^Ci (Pu)/g (soil) [added as Pu(N03)4] . Carbon dioxide
evolution was used as a measure of soil respiration rate. For a measure of plutonium
solubility, the soil was subsampled at intervals during incubation over a 65-day period,
and the subsamples (1 g) were suspended in 1 liter of distilled water. After a 4-hr
equilibration period, an aUquot of the soil suspension was filtered through 5-, 0.45-, and
0.01-jum filters. The plutonium in the 0.45- and 0.01 -^tm filtrates was designated water
soluble, although it was recognized that plutonium likely was present as fine colloids
(previous section).

RELATIONSHIP OF MICROBIAL PROCESSES 313

Changes in the soil respiration rate and plutonium solubility during the 65-day
incubation period are shown in Fig. 4(a). The lack of CO2 evolution from the
gamma-irradiated soil verified its sterility. The increased CO2 evolution rates in the
nonsterile soil over the 4- to 12-day period reflected logarithmic growth for all classes of
organisms. The concentration of plutonium in the 0.01 -/um tlltrate during the incubation
period ranged from approximately 0.04 to 0.14% of the plutonium initially applied.
Solubility of plutonium was essentially identical in the sterile and nonsterile soils,
decreasing with time.

In a subsequent experiment, the plutonium-containing sterile soil was inoculated with
the plutonium-treated nonsterile soil (1 g) which had been previously incubated for 65
days [Fig. 4(a)] , and the respiration rate and solubility of plutonium in the inoculated soil
were measured for a period of 30 days [Fig. 4(b)] . When the sterile soils were inoculated
with nonsterile soil, CO2 evolution increased at a much more rapid rate without a lag
phase, and this was followed by a factor of 2 increase in water solubility (<0.01 jum) 0/
plutonium beginning after 5 days of incubation, which suggests the development of a
microbial population in the plutonium-containing nonsterile soil that was particularly
capable of alteration of plutonium solubility.

An analogous set of experiments was conducted with amended (carbon and nitrogen)
sterile and nonsterile soils. The carbon and nitrogen were added to increase microbial
activity and assess the effect of increased activity on plutonium solubility. The
amendments markedly increased microbial activity (respiration rate, microbial types/
numbers) in the nonsterile soil but did not increase solubility in the <0.01 fim fraction
compared to unamended soils. However, there was a significant increase in plutonium
solubility in the <0.45-/im fraction of the nonsterile soil on initial incubation. As in the
case of the soil that was not amended with carbon and nitrogen, reinoculation of the
sterile soil with the nonsterile soil markedly increased solubility in the <0.01-/um fraction.

At least under the conditions of this study, the evidence strongly suggested that the
solubility of plutonium in soil was influenced by the activity of the soil microflora. The
potential mechanisms affecting the change in solubility include mechanisms (1) and (2)
described on page 312, i.e., indirectly through the production of organic acids that may
complex plutonium or the alteration of the solution pH and/or Eh near the soil colloid
without measurable effects on the overall soil pH or directly through the reduction of
plutonium to Pu(III). Oxidation to Pu(VI) would likely increase solubility, but recent
evidence suggests that this does not occur in these systems. Of course, a combination of
the mechanisms is possible, i.e., alteration of affinity for organic ligands through a change
in valence state. If the mechanism of solubiHzation was indirect, the results might be
applicable to other transuranic elements; e.g., from consideration of the aqueous
chemistry, a reduction in pH would be expected to increase the solubility of the other
transuranic elements as well as that of plutonium.

Increased water solubility of plutonium on incubation under optimum conditions for
microbial activity may increase plutonium uptake by plants provided that the limiting
factor is not discrimination at the root membrane. To determine if the increased
solubility on incubation resuhed in increased plutonium uptake by plants, the incubated
soils were planted with barley and cultured by a spht-root technique that allowed
measurement of the uptake, sites of deposition, and chemical forms of plutonium in plant
shoots and roots (Wildung and Garland, 1974). The resuhs were compared with the
results of similar plant studies in which the soils had not been incubated.

314 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

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RELATIONSHIP OF MICROBIAL PROCESSES 315

Prior incubation, which in microbial studies was shown to increase the solubihty of
plutonium in soil, increased plutonium and americium uptake by shoots compared with
the unincubated controls (Table 3). The effect was greatly accentuated in the case of the
soil-free roots, and incubation increased the soil-to-plant concentration ratios by up to
37 times relative to the unincubated control, depending on plutonium soil concentration
level.

TABLE 3 Uptake of Plutonium and Americium in Barley as a Function
of Prior Incubation in a Ritzville Silt Loam Soil

Concentration ratio*

Unincu

ibated

incubated-i-

Plant component

Plutonium

Americium

Plutonium

Americium

Shoots

Roots (soil free)

0.019
0.060

0.042
0.13

0.071

2.2

0.12
3.6

*(Microcuries of plutonium per gram of oven-dry plant tissue per microcuries
of plutonium per gram of oven-dry soil) x 10~^. Initial concentrations of ^"Pu
and ^'''Am were 0.5 ^Ci/g and 0.03 juCi/g, respectively, on a dry-weight basis.
Mean standard errors (n = 3) were ±39 and ±10% for plutonium and americium,
respectively.

tSoil was incubated 30 days after amendment with carbon and nitrogen to
provide optimal microbial activity.

Effect on the Soil Microflora

Soil microorganisms may be exposed to relatively high transuranic-element concentra-
tions even when total transuranic-element soil concentration is low. Soil organisms may
be expected to be present at highest levels in the immediate vicinity of soil colloids
(Alexander, 1961) where, from the aqueous chemistry of the transuranic elements and on
the basis of recent information on transuranic-element chemistry in soil (previous
section), the transuranic elements are likely to be concentrated. It is therefore necessary
to determine the toxicity of the transuranic elements to soil microorganisms since
microorganisms exhibiting resistance to the chemical effects of the transuranic elements
may have the highest potential for participating in alteration of transuranic-element form.
However, the transuranic-element series does not contain stable isotopes, and organisms
chemically resistant to these elements must exhibit a degree of radiation resistance, which
is dependent, in large part, on the radiochemistry of the isotope. Resistance to the
chemical effects of transuranic elements can occur by three general mechanisms,
including (1) inability of the transuranic elements to produce a toxic effect on cell
metabolism at the cytoplasmic or exocytoplasmic levels; (2) inability of organisms to
transport the transuranic elements; or (3) ability of the organisms to convert transuranic
elements, by the direct and indirect mechanisms discussed in a previous section, to a form
that is either less capable of entering the cell or is not toxic to the cell. The last
mechanism is likely the most important in the alteration of transuranic-element form in
soil.

Effect on Microbial Types, Numbers, and Activity. The etTect of soil plutonium
concentration on the soil microflora has been measured as a function of changes in

316 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

microbial types and numbers and soil respiration rate (Wildung, Garland, and Drucker,
1973; Wildung et al., 1974). A noncalcareous Ritzville silt loam (pH 6.7) was amended
with ^^^Pu (N03)4 at levels of 0.05, 0.5, and lO^uCi/g and with starch, nitrogen, and
water to provide optimal microbial activity. Subsamples of soil were periodically removed
to determine the changes in types and numbers of soil microflora with time. During this
period soil respiration rate was monitored by continuous collection of soil-evolved COj.
The growth curve of fungi (Fig. 5) was generally typical of the growth response for
other classes of microorganisms. Total microbial numbers were compared at the end of

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-"o)

1,000

- i f // — -

____^_^

— :

~Z ra

1/1/

a.c

/ // /

2 9:

<///

3 >

n//

/ ///

J Vt/

-i C

(

i /y

Plutonium

CO 0)

u. o

100

0.0 AiCi/g

—

O &

0,05 MCi/g

a:
m

'^ ^^^y>

0.50 AiCi/g

- ^~~~-~Z3^^^^^--^

10.0 MCi/g

D
Z

1111

1 1 1 1 1

1 1 ]

INCUBATION TIME,

days

Fig. 5 Influence of plutonium concentration on the growth of fungi in soil (Wildung
et al., 1974). Arrows denote time intervals at which growth rates and total numbers were
compared with other microbial types (Table 4).

logarithmic growth. The organisms generally reached this stage after 8 to 14 days of
incubation. Growth rates were compared over the intervals of maximum microbial growth
for each organism at each plutonium concentration. The results are summarized in
Table 4.

The plutonium did not generally affect the rate of growth, but it decreased the total
numbers of all classes of microorganisms at levels as low as 0.05 juCi/g or 7 /Jg/g. The
fungi were the exception, differing from the controls only at a plutonium concentration
of 1 0 juCi/g or 144 /ig/g. Thus the plutonium did not affect the maximum generation rate
but rather it affected the kg period or onset of the stationary phase, which limited
microbial numbers.

The accumulative COi curve generally corresponded to the growth curve of the fungi.
For the other classes of organisms, maximum logarithmic growth occurred before the rate
of CO2 evolution reached minimum levels. The rate of CO2 evolution and cumulative
CO2 over the incubation period were significantly reduced only at the 144-jUg/g level of
plutonium amendment, although numbers of all classes of organisms except the fungi
were depressed below this level (Table 4). This is in marked contrast to the results of
studies with a number of other heavy metals (Drucker et al., 1973), such as silver and
mercury, in which respiration rate was a sensitive measure of metal effect at levels as low
as 1 jug/g in soil. Differences in the effects of the metals may be related to differences in

RELATIONSHIP OF MICROBIAL PROCESSES 317

TABLE 4 Effects of Plutonium at Several Soil Concentration Levels
on the Distribution of Microorsanisms in Soil Relative to Controls*

Effect (p < 0.05)t of plutonium or

Growth rate at plutonium

Total nu

mbers at

plutonium

concentrations

(MCi/g) of

concentrations

(^Ci/g) of

Microbial type

0.05

0.5

10.0

0.05

0.5

10.0

Bacteria

Aerobic and microaerophillic

Nonspore formers

Spore formers

Anaerobic and facultative

anaerobic

Nonspore formers

Spore formers

Fungi

Actinomycetes

•*From Wildung, Garland, and Drucker (1973; 1974).

tPositive sign denotes significant effect. Zero indicates that there was no significant

effect.

soil solubility as well as to toxicity. It should also be noted, however, that the effect on
respiration rate was dependent on the magnitude of the soil respiration rate in
plutonium-treated soil relative to untreated controls, which, in turn, was dependent on
the initial level of microorganisms in soil. In soils exhibiting a higher CO2 evolution rate,
the reduction of respiration rate due to plutonium amendment was more pronounced.
Studies of the toxicity of other transuranic elements to soil microflora have not been
conducted.

Mechanism of Effect. It is important to distinguish, where possible, chemical and
radiation effects of the transuranic elements on soil microorganisms to understand the
long-term effects of microorganisms on transuranic-element form. Pronounced initial
chemical toxicity can result in the development of special pathways of detoxification
leading to alteration of transuranic-element form. The lack of chemical toxicity may
imply chemical moditlcations of the transuranic elements through interaction with cell
metabolites. In contrast, radiation resistance is associated with an enhanced ability to
repair radiation damage to key macromolecules without development of new biochemical
pathways leading to alteration of transuranic-element form. However, the possibilities for
indirect alteration of transuranic-element form would be higher for a radiation-resistant
organism than for an organism that did not exhibit either radiation or chemical resistance
since, due to competitive advantage, these resistant organisms may be expected to be
present in larger numbers than less-resistant organisms where transuranic elements are
concentrated, such as in the vicinity of colloids.

The effects of plutonium on soil microorganisms may be due largely to radiation
damage. Schneiderman et al. (1974) measured the effects of plutonium form and
solubility on soil metabolic activity and on the types, numbers, and resistance of soil
fungi and actinomycetes in soil separately amended with ^.^^Pu (1 to 144jUg/g) and

318 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

^^^Pu (0.6/ig/g) in soluble nitrate and DTPA complex forms and with carbon, nitrogen,
and water to provide optimal microbial activity. Subsamples ot soil were removed over a
95-day aerobic incubation period to determine changes in the numbers of fungi and
actinomycetes and relative water solubilities (<0.01 ^m) of the plutonium forms.
Comparisons of soU fungal numbers in the presence of ^-'^Pu and ^■'^Pu at common
radioactivity levels, but at different mass concentrations, indicated that plutonium
toxicity was due to radiation rather than to chemical effects (Fig. 6). Solubility of

in
O

Z
Z)

UJ
CQ

D
2

2 -

1 -

O Control

A 239pu_Q-rpA, 10.0 juCi/g

A 238p^j_QjpA^ 10.0MCi/g
_ ■ 239pu (NOg)^, 10.0*iCi/g

□ 238pu (NOg)^, 10.0MCi/g

239p^

238p^

12 16 20 24

INCUBATION TIME, days

Fig. 6 Effect of different isotopes of plutonium on survival of soil fungi. [From
Schneiderman et al. (1975).]

plutonium in soil influenced plutonium toxicity to microorganisms with the more-soluble
Pu— DTPA forms resulting in the greatest reductions in numbers. Similar studies have not
been conducted with other transuranic elements.

Isolation of Resistant Organisms. Although much information is available regarding
organic ligands in soil (previous section), an organometal complex has never been isolated
intact from soils. A logical approach to the study of microbial transformations of the
transuranic elements is to isolate, from soil, resistant organisms most likely to alter
transuranic-element form, study the transformation in vitro, and validate the results in

RELATIONSHIP OF MICROBIAL PROCESSES 319

the soil system by using techniques specifically tailored to metabolites identified from the
simpler in vitro systems.

Application of enrichment techniques to the isolation of plutonium-resistant fungi,
which have been demonstrated (previous section) to be the most resistant class of
microorganisms, and actinomycetes from soil with the use of starch as a carbon source
(Schneiderman et al., 1974) resulted in the isolation of 14 fungal cultures and 13 cultures
of actinomycetes distinct in colonial morphology. Of these, 7 of the actinomycetes and 5
of the fungal isolates were capable of growth at 100|Ug/ml plutonium as the soluble
DTPA complex. There appeared to be a succession of actinomycete types in the soil
during incubation, as indicated by the different colony morphologies obtained from
enrichments after 4 and 25 days of incubation. Although this phenomenon may have
resulted from changes in the soil arising from the production of metabolites or chemical
degradation products, it may also have resulted from a response to the presence of
plutonium. Only one actinomycete isolate was found which was common to enrichments
from both incubation periods, and this organism was present at all plutonium
concentrations in the media. In contrast, the fungal isolates exhibited six common
morphological types regardless of incubation period.

Subsequent enrichment studies by R. A. Pelroy, Battelle-Northwest (unpublished
data. 1976), have resulted in the isolation of 30 distinct cultures of bacteria from soil. Of
these. 11 were resistant to plutonium at concentrations as high as lOO^L/g/ml. These
studies also indicate that carbon source as well as soil plutonium concentration will play a
role in determining the types and numbers of plutonium-resistant microorganisms present
in soil, which provides presumptive evidence that microbial metabolites, which will differ
with carbon source, may play a role in plutonium resistance. This subject is discussed in
the next section. The presence of plutonium-resistant organisms is apparently related to
factors that may be expected to vary with soil type and environmental conditions. Again
similar studies have not been conducted with other transuranic elements.

Microbial Transformations

Several means exist whereby microorganisms can transform trace metals in soil. These
may be generalized to (1) direct mechanisms, such as alteration in valence state or
alkylation; (2) indirect mechanisms, such as interactions with normal metabolites or
microbial alterations of the physicochemical environment: and (3) cycling mechanisms,
such as uptake during cell growth and release on cell decomposition. In the last case, any
combination of indirect and direct methods of alteration may be operational. Although
there have been no studies conducted to date that would allow the unequivocal
separation of these mechanisms, studies have been conducted that demonstrate the
alteration of plutonium form in vitro by soil microorganisms and provide evidence for
transformation of plutonium.

Direct Transformations. The potential for direct transformation of the transuranic
elements through alteration of valence state or alkylation is ditTicult to assess. Although
the transuranic elements have the potential for existing in aqueous solution in several
valence states, information is not available to assess the role of soil microflora in direct
alteration of valence. More information is available regarding the mechanism of metal
alkylation.

Alkylation of metals involving the alkyl donor methyl cobalamine and other alkyl
cobalamines has been clearly demonstrated for mercury, arsenic, and platinum (Wood,

320 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Kennedy, and Rosen, 1968; McBride and Wolfe, 1971; Taylor and Hanna, 1977). The
rr^thyl derivative of mercury may be present in significant quantities in soils (Beckert
et al., 1974). Wood (1974) suggested that methylated derivates of mercury and arsenic
are important in governing their behavior in the environment. McBride and Edwards
(1977) also suggested that these reactions occur abiotically. The process of biochemical
methylation of metals can be described as an overlap between the chemistry of methyl
cobalamine (an intermediate in methane synthesis by anaerobic bacteria and methionine
synthesis in aerobic bacteria) and the chemistry of the metals. In the case of the
transuranic elements, particularly plutonium, it is the complexity of the aqueous
chemistry that has limited research into alkylation phenomena.

It is unknown whether an ionic species of plutonium is capable of reacting in vitro
with an alkyl cobalamine. Further, if a mechanism for biological alkylation of plutonium,
similar to the mercury— arsenic— platinum alkylation reaction, did exist, it would be of
importance in influencing environmental behavior only if the alkylated molecule
exhibited stability (Wood, 1974), i.e., a half-life in soils and sediments of hours rather
than seconds. Considering the coordination chemistry of the actinides, Marks (1976)
noted that U— C and Th— C linkages are formed in organic solvents and that the
complexes are relatively stable thermally, although they are highly sensitive to oxygen.
Meaningful microbial studies await the development of an understanding of the chemical
speciation of transuranic elements in aqueous solutions at environmental concentration
levels.

Indirect Transformations. The potential for indirect transformation of the transuranic
elements may be greater than that for direct transformation. The potential for plutonium
interaction with microbial cells and metabolites has been demonstrated, and many of the
other transuranic elements form stable complexes with Icnown microbial metabolites.

Plutonium is taken up directly by microorganisms. Beckert and Au (1976)
demonstrated the uptake of ^^^Pu, applied initially to malt agar in nitrate, citrate, and
dioxide forms, by a common soil fungus, Aspergillus niger. By a specialized spore
collection method, the plutonium was shown to be present in the fruiting bodies.
Subsequent washing to remove external contamination indicated that the major portion
of the ^^^Pu was incorporated into the spores. The order of uptake (10^'') was related
to pH and expected solubility of the plutonium added; plutonium in the initially soluble
nitrate and citrate forms exhibited a factor of 2 to 3 greater uptake than the dioxide. The
availability to microorganisms of the plutonium in citrate and nitrate might be expected
to be considerably higher than that of the oxide from solubility considerations at the
picocurie per milliliter level. The relatively high microbial availability of plutonium as the
oxide is highly significant, and further studies are warranted to determine the mechanisms
of solubilization and uptake and the significance of microorganisms in recycling
processes.

The amount of literature on organic acids and bases, capable of complexing heavy
elements, which are produced directly or by secondary syntheses by a variety of
microorganisms, is increasing. Their concentration and form in soils will be dependent on
the environmental factors influencing microbial metabolism, such as carbon source, and
their residence time will be dependent on subsequent chemical and microbiological
stability.

In preliminary (unpublished) studies by ourselves and others, mixed cultures of soil
organisms, isolated from soil on the basis of carbon requirements and plutonium
resistance, were analyzed as to their ability to transport plutonium into cells and to alter

RELATIONSHIP OF MICROBIAL PROCESSES 321

plutonium form in the cellular and exocellular media. In addition, an experiment was
conducted to distinguish complexation reactions resulting from plutonium interactions
with metabolites arising from normal metabolic processes and plutonium interactions
with metabolites arising from plutonium resistance. For this distinction plutonium was
added at the stationary growth phase of soil microorganisms isolated from soil in the
absence of plutonium, and the transport and complexation were compared with microbial
cultures isolated from plutonium-containing soil and grown in the presence of plutonium.
After growth for 96 hr, the cuhures were separated into cellular and exocellular
fractions. The cell fraction was. in tum, homogenized into intracellular soluble and
cell-debris fractions. The results of studies in which plutonium was added at the
stationary growth phase of cultures of fungi or bacteria grown on mixed organic acids or
sugars are summarized in Table 5. These cultures, selected only on the basis of their
ability to grow on either of two carbon sources, differed to a first approximation in their

TABLE 5 Distribution of Plutonium in Mixed Microbial Cultures* Exposed to
Plutonium at Stationary Growth Phase and Grown on Different Carbon Sources

Plutonium

cultures, %

jngi

Bacteria

Fraction

Mixed sugars

Organic acids

Mixed sugars

Organic acids

Exocellular medium
Intracellular soluble
Cell debris

75
0.49
10

0.068
42

8.3
28

8.7

*Cultures were not replicated. Analytical precision was < ± 10% (1 a). Plutonium present in
cell washes before homogenization is not included.

interactions with plutonium. In general, the majority of the plutonium was associated
with the exocellular fraction, but significant quantities were insoluble and associated with
the cell wall and membrane fractions. However, the distribution of plutonium between
fractions was dependent on microorganism type and carbon source. In the case of fungi,
the exocellular fraction of organisms grown on the organic acid carbon source contained
less plutonium than when mixed sugars were used as a carbon source. Tlie reverse of this
relationship occurred with the bacteria.

Differences in plutonium distribution as a function of carbon source used in
enrichment were also found in cultures grown in the presence of plutonium throughout
incubation (Table 6). The fungal cultures grown on mixed organic acids exhibited larger
concentrations of plutonium both in the exocellular fraction and bound to the cell-debris
fraction; the cultures grown on mixed sugars contained a higher fraction of added
plutonium in the intracellular soluble fraction. In the bacterial cultures the situation was
somewhat different in that higher concentrations of plutonium occurred in the
exocellular fraction of the culture grown in organic acids; less plutonium was associated
with the cell-debris fraction as compared with cells grown on sugars.

In general, the continuous presence of plutonium during growth did not have
pronounced effects on the distribution of plutonium in the cultures (compare Tables 5
and 6). Rather, the metabolic properties of the mixed cultures as determined by carbon
source appeared to be the major factor resulting in the observed differences. Under both

322 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 6 Distribution of Plutonium in Mixed Microbial Cultures* Continuously
Exposed to Plutonium and Grown on Different Carbon Sources

Plutonium

cultures, %

Fungi

Bacteria

Fraction

Mixed sugars

Organic acids

Mixed sugars

Organic acids

Exocellular medium
Intracellular soluble
Cell debris

4.2
29

0.24
39

2.7
31

4
3.5

*Cultures were not replicated. Analytical precision was < ± 10% (1 a). Plutonium present in
cell washes before homogenization is not included.

sets of culture conditions, there was a high concentration of plutonium bound to the cell
wall and membrane fractions and thus was insoluble. As these materials are degraded by
lytic enzymes, e.g., proteases and chitinases, soluble plutonium compounds may be
formed.

Preliminary characterization, using gel permeation chromatography, of the mixed
culture of fungi isolated from soil and grown in sugars indicated that the plutonium form
was altered during fungal growth (Fig. 7). The exocellular and intracellular soluble
fractions obtained from organisms exposed to plutonium in a single exposure and in
continuous exposure contained a majority of plutonium in compounds of molecular size
greater than Pu— DTPA, which was used as the source of soluble plutonium. Furthermore,
there appeared to be a difference in plutonium chemical form when plutonium complexes
formed on simple interaction of plutonium with metabolites (single exposure) and
plutonium complexes formed on interaction after continuous plutonium exposure of the
culture were compared. This suggests either that the culture grown in the continuous
presence of plutonium contained metabolites capable of interacting with plutonium
which were different chemically from those produced by the culture grown in the
absence of plutonium or that the culture grown in the presence of plutonium contained
different organisms that were capable of adaptive response to the element leading to the
synthesis of compounds relatively specific to detoxification of plutonium.

Further chemical characterization with the use of thin-layer chromatography and
electrophoresis verified differences in plutonium form. Several solvents of different
polarities and pH values were used to provide a range of chemical conditions for
separation. Solvent systems included: A, butanol— pyridine, a system used in the
resolution of amino acids; D, pentanol— formic acid, a system used in the separation of
sugars and sugar acids; and G, water— acetic acid, a solvent used in the resolution of
keto-acids and sugars. These systems were used to resolve plutonium as Pu— DTPA and
plutonium in the soluble exocellular and soluble intracellular fractions of the above
cultures (Fig. 8). Thin-layer chromatography with the use of solvent A indicated that the
exocellular fraction contained one component of chromatographic mobility different
from the added Pu-DTPA, but the complex remained present in detectable quantities.
The intracellular soluble fraction contained a component of lesser chromatographic
mobility, but there was no evidence of Pu-DTPA. Solvents D and G did not provide good
resolution. Solvent D did not mobilize Pu-DTPA or other possible complexes; solvent G
mobilized Pu-DTPA and indicated the presence of immobile plutonium components in
the exocellular and intracellular fractions, but these were not resolved.

RELATIONSHIP OF MICROBIAL PROCESSES 323

I-
<

t-
z

o
z
o
o

D
_l
a.

50 70 90 110 130 150

VOLUME, ml

170 190 210 230

Fig. 7 Separation of soluble plutonium complexes in microbial cultures by gel
permeation chromatography.

Application of thin-layer electrophoresis (pH 6.6, pyridine— acetate buffer system;
cellulose support) indicated the presence (Fig. 9) of a relatively large amount of material
of greater negative charge than Pu— DTPA in the exocellular fraction along with
Pu— DTPA. The Pu— DTPA control contained a small quantity of plutonium, likely
hydrolysis products, that did not migrate from the origin. The plutonium ligands in the
intracellular fraction were either neutral in charge in this buffer system or were of a
molecular size too large to migrate under the conditions of electrophoresis. Similar
alterations of plutonium form by a single plutonium-resistant fungus exposed continu-
ously to plutonium during growth have also been reported (Robinson et al., 1977).

Several phenomena may have been responsible for the observed changes in the
chemical form of plutonium. The organism may have synthesized compounds that either
bind Pu— DTPA or bind plutonium more tightly than DTPA. thereby successfully
competing for plutonium in the presence of DTPA. Alternatively, the organism may

324 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TLC

Solvent A

Solvent D

Solvent G

Solvent front

" %

% *

•

• •

# « •

•

# 9

Pu-DTPA
CONTROL

INTRACELLULAR EXOCELLULAR '
SOLUBLE SOLUBLE

Pu-DTPA
CONTROL

INTRACELLL
SOLUBLE

EXOCELLULAR
SOLUBLE

Pu-DTPA
CONTROL

INTRACELLULAR
SOLUBLE

EXOCELLULAR
SOLUBLE

Fig. 8 Thin-layer chromatographic (TLC) behavior in three solvent systems of plutonium
complexes separated by gel permeation chromatography.

degrade or modify the DTPA moiety, allowing plutonium transfer to ligands arising from
microbial synthesis and degradation.

The number of known compounds with the potential to bind plutonium more
strongly than DTPA appears to be quite limited, although hydroxamate derivatives
(Emergy, 1974), catechol derivatives (Tait, 1975), and tetrapyrrole ring systems (Balker,
1969) may exhibit this property. If modification of the Pu-DTPA occurred prior to
hgand transfer, then a myriad of microbially produced compounds, e.g., phenolic acids,
peptides, and carboxylic acids, has potential for binding plutonium (see previous

Pu-DTPA control

•

Exocellular
soluble

•

Intracellular

soluble

Initial spotting

TLE, 20 min

Fig. 9 Thin-layer electrophoretic (TLE) behavior of plutonium separated by gel
permeation chromatography.

RELATIONSHIP OF MICROBIAL PROCESSES 325

section; also Alexander, 1971). In either case the plutonium was not in the form initially
added. Thus applications of gel-permeation chromatography, thin-layer chromatography,
and thin-layer electrophoresis indicate that soil microorganisms are capable, through
simple expressions of metabolic potential, of changing the chemical form of Pu— DTPA
with the resulting formation of plutonium complexes exhibiting a range in chemical
properties. Differences in plutonium distribution in microbial systems and in plutonium
form resulted from both simple interaction with metabolites and perhaps more specific
processes. These differences were dependent on organism type, metabolism, and
plutonium resistance. Investigations are presently under way with pure cultures of these
soil microorganisms to define complexation mechanisms. Detailed study is being directed
toward those organisms > producing exocellular metabolites which form plutonium
complexes that are soluble on elution through soil.

Although published information on the transformation of transuranic elements other
than plutonium is limited, it is likely that transformations similar to that of plutonium
will occur. The extent of these transformations will be dependent on the solubility of the
element, its availability to microorganisms, its toxicity to microorganisms, and its
potential for complexation. Investigations are currently under way with pure cultures of
soil microorganisms isolated in the same manner as the mixed cultures described above.
Exocellular complexes mobile in soil columns are being chemically characterized for
detailed study. Although microbial interactions remain to be elucidated, the solubility
and potential for complexation may be preliminarily assessed from known chemistry
(Table 7). It is evident that the transuranic elements form DTPA complexes with
stabihties similar in magnitude to Pu— DTPA over environmental pH ranges. It can be
concluded that complexation with organic ligands produced by soil microflora is highly
probable, and investigations to identify and characterize the indirect processes and the
ligands responsible for complexation of plutonium in soil are equally applicable to those
of other transuranic elements.

Cycling During Decomposition. A final process whereby the soil microflora may play a
role in transformation of the transuranic elements involves the biological uptake (plants
and microorganisms) of the elements and subsequent release on decomposition. Several
studies have demonstrated plant uptake of plutonium and americium and incorporation
into aboveground tissue. These tissues, deposited on soil either through Utter fall or
agricultural incorporation of crop residues, will be subject to microbial decomposition.
Furthermore, recent studies (Wildung and Garland, 1974) have indicated that barley roots
(uncontaminated with soil particles) contained three to eight times as much plutonium as
the shoots. The roots of plants are in intimate contact with the soil and can be expected
to decompose rapidly (weeks) under appropriate conditions of temperature and moisture
even in arid regions (Wildung, Garland, and Buschbom, 1975). Relatedly, microorganisms,
owing to their distribution in soil and large absorptive surface, compete efficiently with
plants for ions in soil (Alexander, 1961). Studies described in a previous section
demonstrated the association of plutonium with microbial cells. Growth of microbial
cells, a significant portion of the soil biomass, may therefore represent an important
mechanism for biological incorporation of the transuranic elements. Decomposition of
microbial cells generally proceeds at a more rapid rate than that of plant tissues.

Little is known of the form of the transuranic elements in plant or microbial tissues,
of the form, rate, and extent of the transuranics released on decomposition of these
tissues, or of the chemical reactions governing transuranic solubility after decomposition.

326 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Stability of DTPA Complexes with the Transuranic Elements*

Complext

Stability constant

Stable pH range

Neptunium

Np(III)

NpaV)-DTPA

■ 10'^

0.5 to 5.8

[Np(IV)],-DTPA3

10'"

>5.8

Plutonium

Pu(IV)^DTPA

10'^

1.0 to 5.8

[Pu(IV)j,-DTPA3

10"

5.8 to 8.5

Pu(IV)-DTPAj

10'^

>8.5

Americium

(Am(III)]2-DTPA

10'°

1.8 to 6

Am(III)-DTPA

10'^

*From Hafez (1969).

fCurium can be expected to form complexes of stabilities similar to
americium.

:(: Unstable in oxygenated solutions.

However, considering the known products of microbial metabolism of organic substances,
including a number of strong complexing agents and the susceptibility of a number of the
transuranic elements to complexation, it can be concluded that the transuranic elements,
initially immobilized through biological uptake, may be at least as soluble and perhaps
more soluble on decomposition.

In preliminary studies (R. E. Wildung and T. R. Garland, unpublished) plutonium-
amended soil containing largely undecomposed roots from a previous barley crop was
leached with water, and plutonium solubility was compared with a fallow soil containing
plutonium at similar levels. The results indicated that soluble plutonium was initially
immobilized by incorporation into roots, decreasing by a factor of 10 after root growth.
Root decomposition studies are in progress. Previously observed (Romney, Mock, and
Larson, 1970) increases in plutonium uptake from soils by plants with increased time,
generally attributed to increased root development, may have been due to increased
availability through a recycling process on the decomposition of plant roots. The
importance of the process will be dependent on transuranic-element availability to
different plants and microorganisms, the turnover rate of this tissue in soils under
different conditions, and the stability, chemistry, and biological availability of trans-
uranic-element metabolites. Studies are presently under way to provide this information
as a basis for establishing the long-term effects of recycling processes.

Microbial Influence on the Availability and Form of
Transuranic Elements in Plants and Animals

Plants

The results of investigations to physicochemically characterize the mobile forms of
plutonium in soils (<0.1% of total plutonium) suggested that mobile plutonium was
largely particulate (Garland and Wildung, 1977; previous section). The nonparticulate, or
soluble, fraction was present in insufficient quantities, at any single point in time, to
separate from soil and chemically characterize with the present methods. However, for

RELATIONSHIP OF MICROBIAL PROCESSES 327

plutonium to be available to plants, it must pass through the solution phase.
Furthermore, in studies where appropriate measurements have been made, the quantities
taken up by plants exceeded the quantity present in the soil solution. Thus plutonium
was being resupplied to solution and plants from the solid phase. Since Pu(IV) ions would
hydrolyze in solution, precipitating on soil surfaces, it is likely that the plant available
fraction was stabilized in soil solution by complexation with inorganic or organic ligands
and/or was present in a different, more soluble, valence state.

Circumstantial evidence suggested (Wildung and Garland, 1975; previous section) that
inorganic or organic ligands present in limited concentrations in soil stabilized plutonium
in soil solution. Furthermore, dissolution of plutonium from the solid phase has been
shown to be accelerated by complexing agents (Bondietti, Reynolds, and Shanks. 1976;
previous section). Organic ligands. which form the most stable plutonium complexes, are
generally derived from microbial processes in soil, and previous studies (Wildung, Garland,
and Drucker, 1973; previous section) have shown that plutonium solubility in soil and
availability and distribution in plants are influenced by niicrobial activity, although
mechanisms other than complexation may have been responsible. However, a synthetic
ligand (DTPA) was shown to maintain plutonium essentially soluble in soil for extended
periods. Thus it is likely that organic ligands of microbial origin, which differ markedly in
their form, concentration, and stability in soil (Keeney and Wildung, 1977), may play an
important role in stabilizing plutonium in solution for subsequent uptake by plants. A
key question is whether the relatively low uptake exhibited by plants from plutonium-
amended soils (reported concentration ratios of 10^^ to 10~^, Energy Research and
Development Administration, 1976) is due to limited solubility in soil as a result of
sorption on particulate surfaces or to discrimination by the plant. If discrimination is not
at the plant level, then the potential role of the soil microbiota in increasing plutonium
availability from the solid phase (oxide particles from the nuclear fuel cycle or soil
particles) becomes very important in influencing the long-term availability of plutonium
to plants.

The role played by organic ligands in facilitating plant uptake of ions, particularly
hydrolyzable ions, has long been a subject of controversy. The question has been whether
the complex simply .serves as a means of delivering the ion to the root membrane,
supplying the ion to the root by dissociation, or is taken up intact by the plant. Perhaps
the most meticulous investigation of this phenomena has been the work of Tiffm and
co-workers (summarized by Tiffin, 1972; 1977). At least in the case of iron, the evidence,
derived in part from direct analyses of xylem exudates, strongly supports the role of
complexors in increasing uptake by plants but indicates that the complex serves mainly to
deliver the metal to the root membrane and that the ligand is not taken up by the plant
stoichiometrically with the metal. Recently, Malzer and Barber (1976) concluded that
less than 16% of several calcium and strontium chelating ligands was removed from
nutrient solution by corn (Zea mays), whereas over 90% of the calcium and strontium
was taken up by the plant. Both studies used several physicochemical as well as
radiochemical measures of chelate concentration in aqueous solution. It should be noted,
however, that detailed, exhaustive procedures are required to purify metal chelates,
particularly in the case of the transuranic elements (Swanson, Garland, and Wildung,
1975). Without this effort it is possible that studies using only ^^C analysis as a measure
of chelate concentrations in plants would overestimate uptake since low-molecular
impurities containing ^^C might account for the ^'^C present in the plant, particularly
where chelate uptake rates are low.

328 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

For an aid in distinguishing soil sorption and plant root discrimination and for the
evaluation of the role of complexes in the plant uptake of plutonium, hydroponically
grown soybeans (Glycine max) were placed in micromolar level ^•^^Pu— DTPA solutions
and permitted to accumulate plutonium for up to 48 hr (T. R. Garland, D. A. Cataldo,
and R. E. Wildung, unpublished). Concentration ratios (microcuries per gram of plant per
microcuries per milliliter of nutrient solution) for shoot tissues were found to be
6x 10"^ and 0.3 after 1 and 24 hr, respectively. This would suggest that plants do
possess the potential to effectively accumulate plutonium, with much of the apparent
discrimination in soil— plant studies resulting from the effect of soil sorption in reducing
the quantity of plutonium available to the plant. In a preliminary effort to determine the
form of mobile plutonium in the plant, several plants were decapitated, xylem exudates
were collected in 1-ml aliquots at intervals over a 48-hr period, and the solutions were
subjected to thin-layer electrophoresis to resolve plutonium-containing components. The
electrophoretic mobilities of plutonium in the medium, in the control exudate spiked
with Pu(N03)4, and in the exudate collected from decapitated plants grown in 0.1/iM
Pu-DTPA solution over a 24-hr period are illustrated in Fig. 10. The plutonium-
containing components in the exudate from plants grown in the absence of Pu— DTP A
but spiked with Pu(IV) or Pu(Vl) indicated the presence of ligands capable of binding
plutonium and forming stable complexes. Similarly, several anionic and cationic
plutonium complexes were evident in exudates from plants grown in the presence of
Pu— DTPA. A major anionic component with an electrophoretic mobility similar to
Pu-DTPA reached maximum concentration in the second aliquot after decapitation and
then decreased in concentration with time. The application of several solvent systems to
separation subsequently indicated that this component was not the Pu-DTPA complex
suppHed in the growth media. These data suggest that plants do possess the ability to
effectively accumulate soluble plutonium and transport the plutonium to shoots in one or
more organic complexes. Furthermore, from the high concentration ratios for plutonium
supplied as Pu— DTPA and the lack of uptake of the complex, it can be concluded that
much of the apparent discrimination in soil-plant studies results from the effect of soil
sorption in reducing the quantity of soluble plutonium available to the plant. The form of
plutonium in alfalfa fed to animals was also characterized, and this provided insight into
observed differences in gastrointestinal absorption.

Animals

The accepted value for the gastrointestinal transfer ratio of plutonium from food to man
is 3 X 10~^ (U. S. Atomic Energy Commission, 1974). This value is based on the
gastrointestinal absorption of inorganic Pu(IV) in animals, administered by gavage (Weeks
et al., 1956), and the apparent assumption that foodstuffs would contain primarily
inorganic Pu(IV). However, the study demonstrated that gavaging Pu(VI) and complexed
forms of Pu(IV) resulted in higher absorption rates, e.g., 500 times as great as those for
Pu(VI). Until recently it was not possible to test the assumptions because plant tissues
with plutonium concentrations sufficient to measure uptake were not available. However,
studies with alfalfa indicated that tissues containing up to 400,000 d/min per gram could
be obtained under a sequential harvesting regime (T. R. Garland, D. A. Cataldo, and R. E.
Wildung, unpublished). These tissues, used in conjunction with a sensitive analytical
method (Wessman et al., 1971) for the measurement of plutonium at levels of <1 d/min,
allowed preHminary investigations of the availability to animals of plutonium in alfalfa

RELATIONSHIP OF MICROBIAL PROCESSES 329

<=> <=> O cr>:;- XYLEM EXUDATE 0 to 6.3 hr

15,200 (d/min) ml-1

6.3 to 9.8 hr 37,400 (d/min) m|-l

<=> '='* 9.8 to 14.7 hr 43,900 (d/min) ml-i

14.7 to 26.7 hr 60,100 (d/min) ml-i

® CONTROL, Pu-DTPA 50,000 (d/min) ml-i

c=:)C3 ^z^ CONTROL PLANT XYLEM 50,000

i EXUDATE PLUS Pu(N03)4 (d/min) ml

ORIGIN

(+) -*■ TLE, 40 mm ► Q

Fig. 10 Thin-layer electrophoretic behavior of plutonium in soybean xylem exudates.
[From Wildung, Drucker, and Au (1977).]

tissues grown on plutonium-containing soil. In these studies gut absorption of plutonium
gavaged in inorganic solution and plutonium fed in alfalfa tissue was compared for rats
(nonherbivorous) and guinea pigs (herbivorous). In both rats and guinea pigs, absorption
through the gastrointestinal tract of plutonium incorporated in alfalfa tissue was greater
(7.7 to 53 times) than that by gavaging inorganic Pu(IV) solutions (Sullivan and Garland,
1977). Uptake by rats fed alfalfa containing the plutonium exceeded that from gavaging
Pu(IV) nitrate or Pu(IV) citrate but was less than that resuhing from gavaging Pu(VI)
nitrate (Table 8). Uptake was higher when stems and leaves were fed than when leaves
only were fed. Several variables, including animal species, duration of feeding, and types
of plant tissues fed, were evaluated, but, as a result of the small number of animals that
could be used, the limited quantity of tissues, and variability, further studies are required
to evaluate the statistical significance of individual variables.

PreUminary chemical characterization (T. R. Garland, K. M. McFadden, and R. E.
Wildung, unpublished) of the form of plutonium present in the alfalfa tissue fed to the
rodents indicated that the plutonium was more soluble in stems than in leaves, perhaps by

330 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Gastrointestinal Transfer of Plutonium in Rats as a
Function of Plutonium Source

Fraction of

administered dose

Source

in bone plus liver*

Literature cited

Pu(IV) nitrate

0.000012

Weeks etal., 1956

Pu(VI) nitrate

0.0175

Weeks etal., 1956

Pu(IV) citrate

0.00026

Weeks etal., 1956

Plutonium in alfalfa

0.00086

Sullivan and Garland, 1977

*Analyzed 4 days after a single administration.

a factor of 2. Thus the differences in the gastrointestinal absorption between leaves and
stems appeared to reflect differences in relative solubility. The results of preliminary
studies indicated that more than 90% of the soluble plutonium in either stems or leaves
was complexed Pu(IV). These studies are continuing in an effort to evaluate the
importance of the ingestion pathway for transuranic-element uptake by man and to
determine the factors that influence the rate and extent of biological uptake.

Recommendations for Future Research

A broad range of soil— plant concentration ratios (10~^ to 10~^) has been reported for
the transuranic elements (Energy Research and Development Administration, 1976). If
this variability is attributed to the usual unexplained experimental and environmental
parameters, the information would be of little use in predicting the long-term behavior of
these elements. However, a closer examination shows that, although the ratios ranged
over many orders of magnitude, they encompassed, and were often dependent on,
different source terms, soil types, plant species, plant components, climatic conditions,
extent of foUar contamination, and kinetic factors. Consideration of these variables and
use of rapidly accumulating information on transuranic-element behavior at the chemical,
microbiological, and physiological levels should allow reduction of the level of
unexplained variabihty by many orders of magnitude and provide a valuable basis for the
prediction of transuranic-element behavior over a broad range of conditions.

There is a need to continue to develop a basic understanding of the processes
influencing the fate of transuranic elements in soils and plants over a range of soil types
and experimental conditions representing those likely to be encountered in the
environment. Research should emphasize those elements which are (1) expected to be
present in soil in the highest concentration, (2) most soluble in soil, (3) mobile in the
plant, and (4) transported to edible tissues. Source terms receiving future emphasis should
be those likely to result from the nuclear fuel cycle as opposed to fallout. From previous
discussions of the soil chemistry of the transuranic elements, which illustrate the marked
differences in soil behavior and plant availability resulting from different source terms,
sources due to fallout and local nuclear testing, useful for initial approximations, cannot
be taken as fully representative for validation purposes.

Research should emphasize determination of (1 ) source-term physicochemical
characteristics (composition, mineralogy, particle size, and valence) as a function of
source (fallout, reactor operation, reprocessing, and burial leachates); (2) physical
redistribution processes (erosion and resuspension); (3) biological redistribution processes

RELATIONSHIP OF MICROBIAL PROCESSES 331

(litter incorporation and root decomposition); (4) processes of solubilization and/or
transformation of refractory materials entering soil and the factors influencing the form
and equilibria between relatively insoluble forms and soluble chemical species (soil type,
soil solution composition, pH, temperature, redox conditions, diffusion, and microbial
action); (5) the capacity of representative plant species to assimilate soluble chemical
species, plant alteration of chemical form, and translocation to edible plant components
as a function of plant growth stage, form, and concentration in soil and the presence of
competitive ions in the soil solution; and (6) the nature and extent of the above processes
on a regional basis, as influenced by soil, plant, and climatic factors and land-use
practices. This information is currently being accumulated in laboratory studies, and
initial investigations are under way in several geographic regions to selectively validate the
findings in the tleld.

The role of the soil microflora must be viewed as only one contributory factor among
a number of highly important physicochemical and biological phenomena influencing the
overall behavior of the transuranic elements in terrestrial environments. However,
evidence is increasing that organic ligands resulting from microbial activity will play an
important part in influencing the behavior and plant availability of hydrolyzable species,
such as plutonium, in soils. Future studies in this area should involve a systematic
investigation of the major classes of soil organisms exhibiting highest transuranic-element
resistance and representing different metabolic types, determination of their ability to
alter transuranic-element form in soil, and evaluation of the soil and environmental
factors influencing the rate and extent of alteration. In view of the relatively high
concentrations of transuranic elements associated with roots and other organic
components in soil, particular emphasis should be placed on determining the role of the
microflora in recycling and redistribution processes. It is possible that microbial processes
responsible for alteration of metal form may function in a like manner for metals
exhibiting similar chemical properties, particularly for organisms exhibiting cross
resistance to these metals. Thus model systems may be available for those organisms
likely to be most responsible for alteration of metal form in soil over the long term, i.e.,
those organisms capable of growth and reproduction at higher metal concentrations in the
immediate vicinity of refractory oxides or soil colloids with surface deposits of the
transuranic elements.

A key integrating factor, in all studies of the transuranic elements and in
interpretation of environmental phenomena, is a knowledge of the chemical form of the
element. With this information studies conducted under broadly different conditions in
various substrates and biological media can be compared, and toxicological interpreta-
tions can be made on a common basis.

Investigations that are currently under way in several institutions across the country
should ultimately provide a realistic evaluation of the role of microbial processes in
influencing the long-term behavior of the transuranic elements in soil. Furthermore, the
studies should provide a basis for evaluation of the availabihty of transuranic metabohtes
to plants and insight into the potential for entrance of these elements into foodstuffs for
a broad geographical region over the long term.

Acknowledgments

Appreciation is extended to H. Drucker for his technical contributions and advice during
the conduct of the PNL microbiological studies and to F.H.F. Au for review of portions

332 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

of this manuscript. This chapter is based on work performed under U. S. Department of
Energy contract EY-76-C-06-1830.

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uptake of Transuranic Nuclides from Soil
by Plants Grown Under Controlled
Environmental Conditions

D. C. ADRIANO, A. WALLACE, and E. M. ROMNEY

Pla7it uptake of transuranic nuclides ranges through several orders of magnitude,
depending on plant, environmental, and edaphic conditions. Most information presently
available concerns root uptake of plutonium and americium. In environments where
resuspension prevails, direct deposition on plant foliage may exceed root uptake.
Atmospheric deposition is generally short lived, however, and the long-term assessment
precludes that root uptake, as in the case of surface land contamination and shallow
burial of nuclear wastes, will exceed that obtained from atmospheric deposition.
Concentration ratios for ^^^Am uptake generally ranged from I0~^ to 10^ ; those for
^^^Pu generally ranged from 10^'^ to 10~^ . Information for curium and neptunium is
scarce, but the range appeared to vary from 10^^ to I0~^ and from I0~^ to 10'^ ,
respectively, for these radionuclides.

Studies conducted using soils in pot culture showed that '^'^^Am uptake by crops
from southeastern U. S. soils was influenced by clay content and low cation exchange
capacity. Lime amendment suppressed ^'^^Am uptake, whereas organic matter amend-
ment appeared to temporarily reduce uptake from these soils. Commonly used
agricultural amendments generally were ineffective in altering ■^^^ Am and '^'^ Pu uptake
from western U. S. desert soils. However, the chelate diethylenetriamine pentaacetic acid
markedly and consistently increased root uptake of both plutonium and americium by
plants. Chelators and other chemical compounds that enhance complexation reactions
with transuranic elements appeared to be most effective in enhancing root uptake from
soils. Such compounds, which are usually present in shallow-burial waste-storage areas,
may accelerate plant uptake through deep-penetrating root systems.

Numerous studies on the root uptake and translocation of the transuranic elements have
been conducted which contributed to the understanding of some aspects of the processes
involved. Many investigators have demonstrated that transuranic elements entered plant
roots in trace quantities and were transported to aerial parts of plants (Jacobson and
Overstreet, 1948; Cline, 1968; Newbould, 1963; Newbould and Mercer, 1961; Rediske
and Selders, 1954; Romney, Mork, and Larson, 1970; Romney and Price, 1959; Wilson
and Cline, 1966; Rediske, Cline, and Selders, 1955). Adams et al. (1975) found that the
availability of plutonium to plants was very low from 100-/L/m ^•^^Pu02 particles
[concentration ratio (CR) of 10"^^ to 10""^ in ash] . In general, they found that plant
species differed in uptake, with about 25 times more ^'* ' Am taken up than ^ ^^Pu. Bean
seeds contained 200 times less plutonium than bean leaves, but radish roots contained 10
times as much ^^^Pu as did the tops. Peeling the radish roots, however, removed 99% of
the radionuclide, indicating that this radionuclide was mostly contained in or on the peel.

336

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 337

Plant uptake of transuranic elements via the root path has been the subject of recent
reviews (Francis, 1973; Price, 1972a; Brown. 1976; Bernhardt and Eadie, 1976; Mueller
and Mosley, 1976). There has been partial synthesis of the data, with most of the
emphasis on outlining the problems needing further study.

Recently, Adriano et al. (1977), Romney, Mork, and Larson (1970), and Wallace
(1969) found that the chelating agent DTPA (diethylenetriamine pentaacetic acid)
enhanced plant uptake of both ^ ^ ^ Am and ^ ^ ^ Pu up to at least one order of magnitude
under some conditions. These observations, confirmed by their subsequent studies, are
important for two reasons: (1) chelating agents are used in the nuclear industry and are
often present in wastes, and (2) chelating agents are also used in plant nutrition to supply
micronutrients to plants and therefore are likely to be present in agricultural soils.
Agricultural soils are usually treated with various types of amendments to optimize crop
production. Such amendments, like lime, organic matter, and fertilizers, change the
chemistry of the soil with a concomitant effect on the availability of the elements in
question and their subsequent translocation to the plant shoots.

Tlie main objective of this chapter is to evaluate how various soil factors, both
indigenous and introduced, affect uptake of the transuranic elements by various plant
species grown in potted soils obtained from two distinct regions of the United States: the
desert soils of the western United States and soils of the humid southeastern United
States.

Materials and Methods

Pot-Culture Experiments with Soils Representing the Humid
Environments of the Southeastern United States

The soils used for the Bahia grass and rice experiments were collected from the Savannah
River Plant (SRP), near Aiken, S. C. These were uncontaminated soils, higlily weathered
Ultisols, which were collected from the topsoil in a location that normally receives
approximately 130 cm of rainfall annually. These soils are either common in the coastal
plain area or similar to the soils found at the burial areas for nuclear wastes at Barnwell,
S. C, and at the SRP. The clay is usually reddish owing to ferric oxide and is dominated
by kaolin.

Bahia Grass Experiment. Collection, preparation, liming, spiking of soils, and potting
of soils (Troup sandy loam and Dothan sandy clay loam) are described in detail in a
previous paper (Adriano et al., 1977). Bermuda grass hay was ground to pass a 20-mesh
screen and mixed with both the limed and unlimed soils to give 0.0, 1.25, and 5.0%
organic matter (OM) by weight. Reagent-grade fertilizers (NH4NO3, KH2PO4, and
KNO3), lime, and OM were mixed well at the same time with each 2 kg of soil in plastic
bags.

A 500-g aliquot of the premixed soils was removed from each pot (top diameter,
1 5 cm; bottom diameter. 1 2.5 cm; height. 13.5 cm) and used for spiking. One microcurie
of ^"^^ Am. dissolved in 1 ml of 0.1 A^ HNO3. was placed in 125 ml of distilled water; then
10-ml aliquots of this solution were pipeted and added to a thin layer of soil
(approximately 40 g) placed on top of the remaining soil. This was repeated until the
total 500 g of soil was spiked.

Each treatment of the complete factorial (two soil types x two lime rates x three OM
rates) was replicated seven times, but only tlve replicates were spiked. The two unspiked

338 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

replicates were used to determine soil chemical changes at various times during the
180-day duration of the study.

After 50 days of equilibration, the pots were transferred into a water bath located in
a glasshouse. A 1 -cm-thick sheet of glass fiber was placed on top of the soil surface to
prevent the soil particles from adhering to lower plant portions during irrigation. Two
Bahia grass seeds were placed in each of three holes, cut in the fiber equidistant from
each other and the pot perimeter. On the 70th day, the plants were tliinned to three per
pot, one for each hole, and watered with deionized water as required. The plants were
chpped at the 100th, 130th, and 180th day after equiUbration. This gave a plant growth
time of 50, 30, and 50 days for the first, second, and tliird clipping periods, respectively.
The plants were clipped 2.5 cm from the glass-fiber surface. They were then cut into
shorter pieces, placed in paper bags, and dried to constant weights.

Rice Experiments. Only the Do than soil was used in these experiments. A total of
5 kg of soil was placed in each black plastic pot (top diameter, 22 cm; bottom diameter,
18 cm; height, 20 cm), and 2 /uCi of ^'''Am was added either by the soil-layering
technique earlier described or by injection to the ponded water used in the flooded
experiment.

In the first experiment ^'^^ Am, dissolved in O.IA^ HNO3 , was chelated by adding a
^^^ AmCNOaJa aliquot to 50 ml of DTPA solution. Chelated or nonchelated ^^ ' An. "'-'■
added to the ponded water at three various stages of growth: booting stage, flowering
stage, and dough ripening stage.

In the second experiment DTPA and OM (Bermuda grass hay) were premixed with
the whole soil to give 40 ppm DTPA and 5% OM by weight. The nonchelated
radionuclide was added to only the top 1 kg of soil by the layering technique. Two rice
varieties were grown, one under flooded condition and the other under nonfiooded
condition.

All pots were supplemented once with reagent-grade NH4NO3, KH2PO4, and KNO3.
All treatments were replicated five times. The plants were grown to maturity in a water
bath in a glasshouse, harvested, separated into various plant parts, cut into shorter pieces,
and dried.

The dried- plant tissues from all experiments were placed in counting tubes and
counted for at least 50 min for ^'^^ Am with a 7.6- by 7.6-cm Nal well crystal interphased
to a multichannel analyzer. Concentration ratios were calculated from the soil and plant
tissue radioactivity data.

Pot-Culture Experiments with Soils Representing the Desert
Environments of the Western United States

Some areas of the test-range complex in Nevada (NTS) were contaminated by fallout
debris during separate liigh-explosive (nonnuclear) detonations of devices containing
plutonium more than 20 yr ago. The ratios of plutonium to americium in soils and
vegetation collected in the field indicated that at least americium was taken up via roots
from the soil (Romney et al., 1976). Considerable amounts of plutonium and americium
have moved to the root zone in the soils involved. The soils in these areas are sandy desert
calcareous soils with pH values averaging about 8.0 (Leavitt, 1974). Sampling locations
were selected with a portable gamma spectrometer [FIDLER (field instrument for the
determination of low-energy radiation)] , which measures the 60-keV gamma radiation
emitted from ^"^^ Am. Radiochemical data of representative soil samples collected from

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 339

TABLE 1 ^ 3 ^ -2 ^ °Pu and ^ '^ ' Am Concentrations and
Ratios of Desert Soils from the Nevada Test Site and the
Tonopah Test Range Used in the Soybean Experiment*

Concentration, nCi/g
(dry weight)

Soil source 2 39,24opy ^^^\m Pu/Am

Area 11 B 2.8 ±0.41 0.44 ± 0.06 6.5 + 0.17

Area 11 C 9.6+1.3 1.8 ± 0.20 5.2 ±0.17

Area 11 D 4.6 ± 0.24 0.87 ± 0.06 5.2 ±0.12

Area 13 ' 6.0 ± 0.93 1.1 ±0.11 5.6 ± 0.27

Qean Slate 1 4.3 ± 0.57 0.23 ± 0.03 19.0 ± 0.33

aeanSlate2 15.0 ± 3.6 0.72 ±0.13 21.0+1.9

Qean Slate 3 11.0+1.6 0.59 ± 0.07 19.0 ±0.67

Double Track 5.9 ± 0.79 0.28 ± 0.03 21.0 ± 0.88

* Values are means ±1 standard error.

eight fallout areas at which FIDLER activity readings ranged from 20,000 to 30,000 cpm
are given in Table 1. These soils were used for the soybean experiment. The soUs used for
the alfalfa and barley experiments were collected at a greater distance from ground zero
in Area 13 and consequently had much lower ^ ^^Pu and ^"^^ Am contents.

Soils collected from an intermediate contamination zone in Area 13 were subdivided
into twelve 20-kg lots and mixed thoroughly with given amendments in a Patterson—
Kelley blender for 1 lir before subdividing the mixture into six 3.2-kg lots for potting.
The soil amendments consisted of nitrogen fertiUzer (at a rate equal to 200 kg N/ha as
NH4NO3), 2% agricultural-grade sulfur (to reduce pH from 7.6 to 5.4), and 57o OM (as
alfalfa meal). The treatments were done in three separate sets of three replications per
treatment and with and without DTPA (72 pots in total). Soil was potted in plastic pots
that were sleeved inside plastic buckets and covered with 5 cm of siUca sand to prevent
soil-particle resuspension. The soil activity levels turned out to be much lower than had
been anticipated. Consequently the plant materials from all rephcates grown on similar
treatments of the three sets had to be combined to produce an adequate sample size for
radioassay. Barley plants were grown first. They were harvested in the dough stage by
cutting at about 5 cm above the top of the sand layer and were divided into straw and
fruit head samples. Alfalfa was grown next. Tluee successive cuttings of foliage were
made in the quarter-bloom stage; then the plants were harvested and combined like the
barley plants.

Soils collected from eight different plutonium fallout areas on the NTS and the
Tonopah Test Range were used for the pot-culture experiments with soybeans. These
soils received only the DTPA chelate (200 ppm) amendment. The eight soils were again
arranged in three sets each, containing three replicates, with and without DTPA (144 pots
in total). Soil processing and potting and plant harvesting and preparation for radioassay
were done as indicated for the barley and alfalfa experiments. The soybean plants were
grown to maturity, harvested, and separated into fruit pods and foliage (leaf and stem).
All samples were radioassayed for ^"^^ Am and ^•'^Pu at LFE, Richmond, Cahf. (Majors
etal., 1973),

340 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Results and Discussion

Soils from the Humid Southeastern United States

Current field conditions at tlie SRP do not provide an environment suitable for the study
of the incorporation of transuranic elements in plant tissues through root uptake because
of a confounding effect from deposition of particles on the foliage following resuspension
or stack emission from reprocessing operation. Thus studies were conducted in a
glasshouse to evaluate the influence of some common soil amendments on the uptake and
translocation of ^"^ * Am by Bahia grass (Paspahmi notatum) and rice (Oryza sativa). Bahia
grass is a common pasture crop grown extensively in the southeast. Rice was included
because it is one of the most important food crops in the world and is widely distributed
throughout the tropical, subtropical, and temperate zones of all continents (Adair, Miller,
and Beachell, 1962; Harlan, 1976).

For brevity, only the CR* values are presented. In studies of this nature, the CR is a
convenient method of expressing the availability of an element from the soil and its
pattern of translocation to the plant parts. Since the soil '^^ ' Am concentrations are given
in the footnotes to the tables of results, the corresponding plant concentrations can be
calculated from the CR values.

Bahia Grass Experiment. The effect of soil type on ^'*'Am uptake in Bahia grass
(Table 2) was not so pronounced as that in the bush bean and corn seedlings (Adriano
et al., 1977). Nevertheless, the most striking differences in uptake were caused by soil
type and lime. On the average, ^^ ' Am concentrations in plant tissues from the unlimed
Dothan soil (pH 4.2) were approximately twice as high as those from the unlimed Troup
soil (pH 5.0). This wide disparity between these two soils was minimized when both soils
were limed. Consequently liming of both soUs (pH 7.1 for Dothan soil and pH 6.6 for
Troup soil) significantly (p < 0.01) reduced ^'*' Am availability to Baliia grass. Across the
OM treatments, the plant tissues from unlimed Troup soil had 12.0pCi/g (dry-weight
basis), compared to only 1.0 pCi/g from limed soil, on the average. On the average plants
from unlimed Dothan soil had 35.3 pCi/g vs. 0 pCi/g (below detection limit) from limed
Dothan soil.

The clipping period affected the ■^'*'Am concentration pattern in Bahia grass,
particularly in unlimed soils (Table 2). In unlimed Dothan soil, the concentration
progressively declined with clipping time. The peak occurred on the first clipping (50th
day of growth) and the minimum occurred on the last cHpping (130th day of growth),
irrespective of OM rate. The concentrations in the first clipping were always significantly
higher than those from subsequent clippings. In unlimed Troup soil, the concentrations
generally peaked during the second clipping, although at the 1.25% OM rate they were
not significantly different from the other clippings. No meaningful pattern can be
deduced when either soil was limed, and, in some cases, hming caused plant ■^^'Am
concentrations to be equal to background level.

The addition of OM affected ^^' Am plant concentrations to some extent. Plants in
unlimed Troup soil in wliich no OM had been added had the higliest concentrations. The
mean concentrations for all three chppings were 23.2, 9.5, and 5.6 pCi/g for 0.0, 1.25,

^^ ^ radioactivity/g (plant tissue)

*CR (concentration ratio) = —. — ; r-

radioactivity/g (soil)

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 341

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342 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

and 5.0% OM rates, respectively. Similarly, although less pronounced than in the Troup
soil, plant concentrations in the unlimed Dothan soil were highest at the 0% OM rate
(45.2pCi/g) and decreased to 36.9 and 14.5 pCi/g for the 1.25 and 5.0% OM rates,
respectively. In both unlimed soils, small differences in ^'''Am concentrations were
obtained from the 1.25% OM rate. However, the 5% OM rate decreased the ^"^'Am
concentrations in both unlimed soils somewhat substantially.

The CR values in Table 2 for limed treatments, in general, are 10 times lower than
those for the unlimed treatments and, in some cases, as much as two orders of magnitude
lower. The effects of OM rate and clipping period on ^"^^ Am availability can also be
easily deduced from Table 2 and further elaborate the effects of these treatments on
^^^ Am concentrations. The slight reduction in uptake with OM addition was possibly
caused by immobiUzation of ^^ ^ Am in the soil microbial biomass and the fixing capacity
of OM for metals.

The CR values were calculated on the basis of the total soil mass in the pot. These
values can also be calculated by using only the amount of soil that was spiked, in which
case the present CR values should be multipUed by a constant factor of 0.25.

Rice Experiment. Results (Table 3) indicate that, in some cases, americium applied in
water was detectable in the rice grain. However, these are low radioactivities compared
with other plant parts. No radioactivity was detected in the grain when americium was
applied to the soil. Americium increased in the following order: unshelled grain < green

TABLE 3 Influence of Chelate DTPA, Time of Spiking, and
Method of Placement on ^ '* ' Am Concentration Ratios for Rice
Grown in a Southeastern U.S. Soil Under Flooded Conditions*t

Unshelled
grain/soil

Green blades/soil

Old (dead) blades/soil

Applied to water

Period 1

Period 2

Period 3
Applied to soiH

(1.1 ±
(1.0 ±

0.36) X
0.18) X

10-'
10-'

With DTPAt

(3.8 ± 1.4) X 10-'
(1.3 ±0.18)x 10-'
(1.7 ±0.40)x 10-'
(0.2 ± 0.85) X 10-'

(2.7 ± 1.4) x 10-'
(6.5 ± 2.5) x 10-'
(1.6 ± 0.54) X 10°
(2.8 ± 0.27) X 10-

Applied to water

Period 1

Period 2

Period 3
Applied to soiH

(6.8 ± 2.3) X 10"

Without DTPA$

(5.3 ±2.8)x 10-'
(5.7 ± 1.3) X 10-'
(1.0 ± 0.80) X 10-'
(2.6 ± 0.80) X 10-'

(4.6+ 1.3) X 10-'
(4.8 ± 3.7) X 10-'
(4.2 ± 3.4) X 10-'
(2.5 ± 0.22) X 10-'

*Concentrations in plant materials can be calculated from the CR values and the
assumed concentrations (400 pCi/g dry soil) of the potted soil.

t Values are means of five replicates ±1 standard error.

^Chelated or nonchelated ''"Am (in 50 ml of 100 ppm DTPA as acid or water) was
added to the standing water when the rice plants were at booting (period 1), flowering
(period 2), and dough (period 3) stages.

§ Activities were below the detection Umit.

% Data taken from Table 4.

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 343

blade < dead blade. Morphologically, the green blades were located in the top portion of
the plant and the dead blades in the lower portion. Thus americium appeared to
accumulate in the older leaves to a much greater extent than in the newer leaves, which
developed in later stages of growth. The ratio of americium concentrations in dead and
green blades ranged from 48 to 140 with DTPA and from 8 to 44 without DTPA.
However, most of the americium accumulated in the sheath when this isotope was applied
in water and most notably when chelated. The accumulation in the sheath could be
attributed to physical absorption of the isotope rather than to physiological assimilation,
as in the case for leaf blades. Chelated americium applied to the ponded water was more
readily absorbed, as the leaf-blade data suggest, but it was not more readily absorbed
when applied to soil.

The relative magnitudes of americium in various parts are easily discernible from the
CR values. Plants that received chelated-americium through water application had
americium contents 10 times higher than plants supplied with nonchelated americiuni.
Dead blades had americium contents one to two orders of magnitude higher than green
blades. The CR values can be used to determine the relative availability of an element
from a substrate and the translocation pattern of this element within the plant. The CR
values indicate that americium was less available when added to soil and was not readUy
translocated to younger leaves.

In all CR calculations, 400 pCi/g dry soil was used, taking into account the total soil
mass per pot (5 kg). However, if based on only the top 1 kg of spiked soil, 2000 pCi/g
should be used. Thus, in the latter case, the reported CR values in Table 3 should be
multiplied by a factor of 0.20. With water application it is difficult to assign a conversion
factor. It should be pointed out that CRis probably not valid with water appUcation and
should be used with caution since traditionally it is used where the radionuclide was
applied to the soil and the total soil mass is taken into account in calculating the average
radionuclide concentration in the soil. Recently, however, conversion factors have been
introduced to consider also the fraction of the soil mass spiked (Lipton and Goldin,
1976). As would be expected, there is some question concerning the use of soil
concentration for determining the CR for water application, but the CR would
demonstrate the relative translocation or redistribution of ^"^^ Am in various rice parts
and could serve as a basis for calculating the plant concentration of '^^^ Am.

In the flood variety* the radioactivity in the grain was below the detection Umit
(Table 4). There was also Uttle translocation from old leaves to green leaves. The chelate
DTPA mixed with the soil slightly reduced ^^^ Am uptake. Apparently the chelate level
(40 ppm as acid) was harmful to the rice plants, retarding and reducing growth. Organic
matter did not have a clear-cut effect, although it tended to suppress the uptake by the
nonflood variety. It should be pointed out that the OM retarded growth of the rice
seedlings in early stages of growth, presumably because the organic acids inhibited root
development (Takijima, 1964). In general, the plant tissues of the nonflood variety had
higher '^^^ Am levels.

It appeared that the ^"^ ' Am content of the grain could be increased slightly by adding
chelated ^"^^ Am to the standing water. Thus the method of ^"^^Am placement would

*The flood variety was a dwarfed "miracle rice" variety from southeast Asia and was ponded with
water all the time. The nonflood variety, also from Asia, was not ponded and was taller than the flood
variety.

344 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 Influence of Chelate DTPA and Organic Matter on

^^ ^ Am Concentration Ratios for Rice Grown in a Southeastern

U.S. Soil Under Flooded and Nonflooded Conditions*!

Unshelled
grain/soil

Green blades/soil Old (dead) blades/soil

Control
+ DTPA
+ 5% OM

t
t
t

Control
+ DTPA

+ 5% OM

(7.1 ± 2.0) X 10^"
(2.5 ± 1.6) X 10-^
(2.5 ± 2.1) X 10-^

Flooded

(2.6 ± 0.80) X I0-'
(0.2 ± 0.85) X 10-'
(7.6 ± 2.4) X 10-^

Nonflooded

(2.7 ± 0.18) X 10~'
(2.9 ± 0.67) X 10"'
(1.4 ± 0.22) X 10"-'

(2.5 1 0.22) X 10-'
(2.8 ± 0.27) X 10"'
(1.5 ± 0.31) X 10"'

(5.9 ± 1.3) X 10-'
(2.2 i 0.76) X 10"'
(1.5 ± 0.10) X 10"'

* Values are means of five replicates ± standard error.

fConcentrations in plant materials can be calculated from the CR values and
the assumed ' " ' Am concentrations (400 pCi/g soil or 435 pCi/g soil + 5% OM) of
the potted soil.

:}: Activities were below the detection limit.

have an influence on its availability to the rice plants; i.e., application to the standing
water is more likely to result in higher uptake.

Rice has a peculiar uptake— translocation physiology (Chandrasekaran and Yosliida,
1973; Myttenaere and Marckwordt, 1967; Myttenaere, Bourdeau, and Masset, 1969). Its
only organ of economic importance is the grain. The straw is seldom used for animal feed.
Tlie "^' Am did not appear to be readily translocated to the grain; therefore its health
hazard to man is minimized.

Soils from tM Desert Environment of the Western
United States

Barley and Alfalfa Experiment. Results indicate that ^'^ ' Am generally was taken up
by barley two or more times as readily as was -^^^Pu (Table 5). The exception was for the
treatment acidified with sulfur and with DTPA. The reason was a relatively large uptake
of ^^^Pu with DTPA. The americium/plutonium ratios (Table 5) were obtained from the
respective CR values to normalize the levels in the soil. The CR values were generally in
the 10"^ to 10^ range (mean ^ '"^Pu^ 9.4 x lO""* and ^^' Am = 1.4 x IQ-^).
Without DTPA they were 1.3 X 10""* and 3.2 x 10"'*, respectively, for plutonium and
americium. Except for the control, which produced poor growth, DTPA enhanced the
uptake of both plutonium and americium. However, the increase, wliich was usually
greater than one order of magnitude, was equal for both elements.

The CR values for alfalfa were slightly less than those for barley (Table 5). The mean
CR for 2 3 9pu ^as ^ ^^ jq-^ ^^^ ^^^ ^"^^ Am ^Jvas 9.3 x 10"^. Without DTPA the CR
values were 7.6 x 10~^ and 6.6 x lO^'*, respectively. The preference for americium over
plutonium (9.9) was greater in alfalfa than in barley (4.2). The DTPA had much less
effec^t on the uptake of the two elements by alfalfa than by barley (ratio of about 2 for

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 345

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346 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

alfalfa and 17 for barley). This effect may indicate loss of DTPA from the soil with time,
possibly by metabolism and degradation.

Soybean Experiment. Results from the barley and alfalfa experiments showed the
necessity of using soils containing higher levels of contamination for better accuracy of
determination. In addition, the response from the soil amendments indicated a practical
influence from only the DTPA. Consequently the soybean experiments were conducted
using only chelate treatment of soils collected from eight of the Nevada Applied Ecology
Group study areas (Dunaway and White, 1974). Results of radiochemical analyses for
^•^^Pu and ^"^^ Am are shovm in Table 6 and indicate that the CR values were higlier for
soybean leaves and stems (Table 6) than for either barley or alfalfa (Table 5). Different
soil sources, in part, were involved, but this may not be the major factor. The mean CR
values. for soybean leaves and stems were 1.4 x 10~^ and 3.7 x 10~^ for plutonium and
americium, respectively; with DTPA they were 4.4 x 10"'* and 6.5 x 10~^ , respectively,
which are higher than equivalent values for barley and alfalfa.

The mean americium/plutonium ratio was 21.6, which is higher than that for either
barley (4.2) or alfalfa (9.9). It was not possible to determine if this ratio for soybeans
differed because of variability in soils.

The DTPA enhanced the uptake of both plutonium and americium from the Area 1 1
and Area 13 soils but only slightly over those from the Tonopah Test Range. For the
Area 11 soils, DTPA increased the uptake of americium more than of plutonium (about
2.5 times). This result did not appear to be significant since it was not observed for the
other soils described in Table 6 except for one. Different chemical and physical properties
of americium and plutonium in the soils from different locations may be involved. The
soil plutonium/americium ratios were highest (Table 1) for the soils with least response to
DTPA.

The mean CR for fruit pods was 3.7 x 10"^ for ^^^Pu and 8.0 x lO"'* for ^"^^ Am.
Without DTPA the values were 1.5 x 10~^ and 3.1 x 10"'', respectively. The mean
americium/plutonium ratios were slightly higher for fruit pods than for vegetative
material (32.8 vs. 21.6). The mean americium/plutonium ratio for fruit pods was 23,6
without DTPA and 42.0 with DTPA. The difference, however, was not due to
DTPA-induced transport from leaves to fruit because the americium/plutonium ratios
with and without DTPA were really not different (Table 6). Also, the mean CR for
vegetative parts for plutonium was 9.9 without DTPA and 11.5 with DTPA, which
difference was not significant. For americium the means were 15.9 and 10.1, the lower
value being with DTPA. It appears that DTPA caused more plutonium and americium to
be translocated to the fruit pods because plutonium and americium were liigher in the
leaves when DTPA was added. This resulted in correlation coefficients of +0.988 for
plutonium and +0.983 for americium between these two plant parts. The CR values for
fruit pods vs. leaves and stems from the eight soils were calculated for plutonium and
americium and show that, on the average, the ratios were 0.035 ± 0.003 SE for ^^^Pu
and 0.058 ± 0.009 SE for ^"^^ Am. The americium/plutonium ratio was 1.7 ± 0.24 SE for
the eight soils. The ratio of transport with DTPA to that without DTPA was
1,0±0.21SE for plutonium and 1.3 ± 0.36 SE for americium. This method of
calculation indicates that DTPA did not directly increase plutonium and americium
transport to fruits, nor did DTPA influence the two elements differentially in transport
from shoots to fruits. It appears tlierefore tliat there was a mass-action effect for
transport from leaves to fruit.

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 347

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>48 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

ATMOSPHERIC

ENVIRONMENTAL
CONDITIONS

PLANT
PROPERTIES

SOIL
PROCESSES

SOIL
PROPERTIES

SOURCE

Fig. 1 Schematic diagram of various factors influencing the bioavailability of the
transuranic nuclides in the soil-plant system.

Discussion

Several factors influence the availability of the transuranic nuclides to plants. A
generalized outline is shown in Fig. 1. Plant uptake is influenced by soil pH, Eh
(oxidation state), cation exchange capacity, texture (particularly percent clay), fertilizers
and other amendments, and soil OM.

Lime significantly suppressed ■^'^'Am uptake by crops grown in southeastern soils
owing either to a lower solubility of ^'*' Am at higher pH, an increased cation exchange
capacity caused by liming (Fiskell, 1970; Helyar and Anderson, 1974), or to
calcium— magnesium and ^'*' Am antagonism. The first two processes could have caused
high Kd* values. The latter could have resulted in Ca^^ and Mg^"" ions suppressing the
uptake of ^'^^ Am^'' ions. Chelates have been known to make insoluble cations available
to plants. Such chelates increase the diffusion and mass tlow of cations to roots by
replenishing those taken up by the plants. It has been shown that chelates, including
DTPA, decreased the Kd values of plutonium; i.e., less plutonium was retained by the soil
(Relyea and Brown, 1978). Wallace (1972a; 1972b) observed that ^^'Am-DTPA was
most stable at about pH 7.7, where plant uptake was greatest.

*Kd (distribution coefficient) =

concentrution ot" ^ "' Am/g (soil)
concentration of ^ " ' Am/ml (solution)

UPTAKE OF TRANS URANIC NUCLIDES FROM SOIL BY PLANTS 349

Soil processes involving soil OM decomposition (Cataldo et al., 1976; Wildung and
Garland, tliis volume); microbial, particularly fungal, growth (Wildung and Garland, this
volume); chelation; hydrolysis; and oxidation likewise influence the phytoavailability of
the nuclides.

The long-time phytoavailability of the transuranic nuclides, once on the soil, will
depend on the long-time soil processes. Leaching and capillary rise will cause the nuclides
to move in the profile. Animal activity will also do the same.

The slow process of sheet erosion likewise can cause the nuclides to move in the
profile. Particularly with plowing, which moves nuclides to greater depths, sheet erosion
can result in movement of the nuclides up in the profile because of a decrease in the level
of the soil surface. Hundreds of years may be involved in this process, but it must be
considered in the long-time availability of the nucHdes.

Plant properties influence the uptake of the transuranic elements. It is well known
that plant roots excrete protons, organic and amino acids, chelators, and other
substances. These liave profound effects on tlie uptake translocation of many metals,
including transuranic elements. Also, roots have a reducing capacity that is exceptionally
important in the physiology of iron uptake by plants. At least in some species, Fe^"" must
be reduced to Fe^"" before absorption can occur. This process is believed to be of
importance in the various oxidation states of some transuranic elements. Decomposition
of plant residues influences both uptake and recycling.

The depth of rooting of various plant species, a characteristic of both the plant and
the soil in which it is growing, is a factor in transuranium-element uptake. This is
especially important in areas that are or will be plowed or in areas where wastes have been
buried. Nuclides in lower horizons of soil may be mobilized by deep roots.

The aboveground contamination of plants and subsequent leaf absorption—
translocation is covered elsewhere in this volume. Leaf uptake does depend on the
physiological and anatomical characteristics of the leaves.

Environmental conditions other than soil characteristics may be more important in
leaf uptake of transuranic elements tlian in root uptake of them. The effect of soil
moisture is as yet unknown for root uptake, but soil moisture is a factor in wind
resuspension and subsequent deposition of contaminated materials on leaf surfaces.

The sources of transuranic elements are of much importance to phytoavailability. The
particle size greatly determines the availability. Small particles generally are more subject
to weathering and release nuchdes faster than do large particles. Oxides are less available
than other forms. Transuranic elements in wastes containing acids and chelator chemicals
probably are more available than those from other sources because of the complexation
processes involved. Oxidation state has been referred to previously. The transuranic
elements themselves differ in phytoavailability (Pu < Am < Cm < Np) (Price, 1972b).

The various factors that influence transuranic -element uptake by plants result in a
very wide range covering several orders of magnitude in the CR. Data have been compiled
from the literature and are shown in Table 7. One contributing factor is the very low level
of the nuclides that move from soil to roots and then to shoots of plants. The root uptake
is generally believed to be very much lower than that which comes from the atmosphere.
Tliis is correct only so long as the nuclides can be resuspended or released to the air. On a
long-term basis, root uptake would exceed that from the atmosphere.

(Text continues on page 357.)

350 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Concentration Ratios of the Transuranic Elements for Plants Grown in

Pot Cultures or Field Environments

Plant species

Range of CR

Conditions and comments

Reference

Pot culture
Alfalfa

Alfalfa

Barley
Grain
Foliage

Barley

Barley (shoots)

Barley (plant)

10-= to lO-''

10-

Plutonium

Grown in 3 kg of NTS soil; 0.6 nCi/g This study
(soil) of 2'^'^ ""Pu; high-
fired PuOj ; liighest CR
caused by chelate treat-
ment; used mainly for
livestock feed.

Grown in 1.5 to 5 kg of contaminated Adams et al.,
soU from Palomares, Spain; 1975

60 nCi/g soil of ' " Pu;
used mainly for live-
stock feed.

10-* to 10-'
10-= to 10-'

10-'
10"=

Same as for alfalfa above;
used mainly for human
food and an ingredient
of livestock feed.

Grown in 3 kg of NTS soil;
high-fired PuO 2 ; 10 to

50nCi/g(soa)of

2 3 9 ,2 4 0

Pu.

Shoots

lO-'' to 10-'

Grown in split-root (soil-

Roots

10-" to 10-'

nutrient solution) plant
culture; 0.05 to
10MCi/g(soU)
of total plutonium
asPu(N03)4.

Barley

Grain

10-'

10MCi/g(soii)of

Foliage

10-= to 10-'

^"'Puand^''*Pu
asPu(N03),.

Barley (shoots)

10-=

Grown in 1.6 kg of soil

10-' (av.)

— 4

spiked with 0.1 AiCi/g

of ^"PuasPu(NO,)^;

uptake similar from three

soil types.
Grown in 1.2 kg of Los

Alamos mountain

meadow soil spiked

with 300 MQ/g of

^"Pu asPuOj.
Same as for alfalfa above.

This study

Schulz et al.,
1976b

Wildung and
Garland, 1974

Energy Research
and Development
Administration,
1976

Wilson and Qine,
1966

Adams et al.,
1975

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 351

TABLE 7 (Continued)

Plant species

Range ofCR

Conditions and comments

Reference

Barley (shoots)

10"

Beans (shoots)
Beans (shoots)

Qover, Ladino

Lettuce (plants)
Lettuce (plants)
Oats (plants)

10"

10-' to 10"

Corn

Grain

10-'

Leaves

10-*

10-' (av.)*
10-' (av.)*

Grown in Cinebar (pH 4.5)
and Ephrata (pH 7.5)
soils using the Neubauer
technique; soils spiked
with lOMCi/gof "'Pu
as Pu(N03)4 : uptake
similar from two soils.

Same as for barley above;

seeds mainly used for

human food.
Grown in standard Hoagland

solution spiked with 1.0

MCi/literof "'Pu

asPu(N03)^.

Grown in 1 20 kg of NTS soil;
high-fired PuO^ ; 70
nCi/g(soil)of ^'''^^''Pu;
CR increased sevenfold
in 5 yr; mainly used
for livestock feed.

Grown in 5 to 7 kg of soil
collected from top layer
of a field adjacent to
a reprocessing facihty
at SRP: homogenized soil
contained 2 pCi/g total
plutonium; used mainly
for livestock feed plus
some for human food.

Same as for alfalfa; mainly

for human food.
Same as for barley.

Same as for barley; grain
mainly used for human
food plus ingredient for
animal feed; plants
used as forage.

Qine, 1968

Wilson and Gine,
1966

Qine, 1968

Romney, Mork, and
Larson, 1970

Adriano, Corey, and
Dahlman, this
volume

Adams et al.,

1975
Adams et al.,

1975

Peas (plants)

10-' to 10°

Grown in about 1.5 kg of

Lipton and Goldin,

sand spiked with ^ '' PuO^

1976

at 20 mCI per container;

highest CR caused by

chelate ; lesser effects

by colloid size and

placement depth; fruits

Radish

mainly for human food.

Bulb

10-^

Same as for alfalfa; mainly

Adams et al..

Shoots

10-'

for human food, including

1975

the shoots as greens.

(Table continues on the next page.)

352 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 (Continued)

Plant species

Range of CR

Conditions and comments

Reference

Soybean

Fruit

10-*

to 10-'

Grown in 3 kg of NTS soil;

This study

Forage

10-"

to 10-^

5 nCi/g (soU) of
2 "'='°Pu: high-fired
PuOj ; highest CR caused
by chelate treatment.

Soybean (foliage)

10-^

Same as for corn; fruit
used mainly for human
food plus ingredient
for animal feed.

Adriano, Corey,
and Dahlman,
this volume

Tomato

Fruit

10-=

Same as for alfalfa; fruit

Adams et al.,

Plant

10-'

used mainly for human
food.

1975

Wheat

Grain

10-'

to 10-*

Grown in 3 kg of soil

Schulz, Tompkins,

Leaf

10-*

to 10-^

spiked with 2 3%240py
in either the chloride

and Babcock,
1976a

Wheat (plants)

10-

Cheatgrass (plant)

10-= to 10"

Tumbleweed (plant)

Crops and vegetables
Fruit
Foliage
Subterranean

10-= to 10-'

10-= to 10-'
10"' to 10-^
10-' to 10-'

or nitrate form; 10 ^Ci/g
(soil); highest CR
occurred when plutonium
was added in nitrate
form to an alkaline,
calcareous soil.
Same as for corn; grain
used mainly for human
food plus ingredient for
animal feed; plants when
green can be used for forage.

Grown in 1 kg of soil
spiked with 50 nCi/g of
''''PuasPu(N03)^;high
CR values caused by
organic acids; native
species with no apparent
food value to man and
livestock animals.

Same as for cheatgrass.

Bush beans, beets, carrots,
lettuce, millet, potatoes,
radishes, soybeans, and
tomatoes grown in a flood-
plain in Oak Ridge
contaminated for
30 yr from weapon
development; soil had
25-100 pCi/g total
plutonium; nominal
surface contamination;
peeling the skins of

Adriano, Corey,
and Dahlman,
this volume

Price, 1972b;
1973

Price, 1973

Adriano, Corey,
and Dahlman,
this volume

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 353

TABLE 7 (Continued)

Plant species

Range of CR

Conditions and comments

Reference

subterranean crops and

vegetables removed

Wheat

most of the plutonium.

Grain

10-^

Grown on a field adjacent to

Adriano, Corey,

Foliage

10-' to 10"'

a reprocessing facility

and Dahlman,

at SRP which aerially

this volume

Soybean
Grain
Foliage

Native vegetation

10-"
10-^

released <3 mCi total
plutonium per yr; 1 to
3 pCi/g (soil) of total
plutonium; 90-977o of
contamination was external.

Grown on a field adjacent to
a reprocessing facility
at SRP which aerially
released <3 mCi total
plutonium per yr; 1 to
3 pCi/g (soU) of
total plutonium; 90-
97% of contamination
was external.

Corn

Grain

10-"

Grown on a field adjacent to

Leaves

10-'

a reprocessing facility
at SRP which aerially
released <3 mCi total
plutonium per yr; 1 to
3 pCi/g (soil) of total
plutonium; 90-97% of
contamination was external

Corn + cobs

10-'

Corn, potatoes, and peas

Potatoes

10-'

grown on garden plots

Peas (shelled)

10-^

in North Eastham,

10-' to 10"

Mass., on Cape Cod;
top 30 cm of soil
had6fCi/g2^''"'°Pu;
crops shielded from
direct deposition or
resuspension.

Tree foliage, shrubs, and
herbaceous samples collected
from an area adjacent to a
reprocessing plant at SRP.

10-" to 10-'

Adriano, Corey,
and Dahlman,
this volume

Hardy, Bennett,
and Alexander,
1977

Unpublished SRP
data.

Koranda et al.,
1973

Leaves of nonedible
perennials, messerschmidia,
scavola, and pandanas
native to Enewetak AtoU;
coral sandy soil.

(Table continues on the next page.)

354 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 (Continued)

Plant species

Range of CR

Conditions and comments

Reference

Native vegetation

10"' to 10''

Native vegetation

10"^ to 10"

Native vegetation

10"

Coconut

10"

Native perennials and shrubs
at the NTS environments
contaminated by safety
shots: resuspension and
deposition suspected on
hirsute plants.

Mixed native grasses grown
in contaminated soil
blocks at Rocky Flats,
Colo.; metallic pluto-
nium in oil at 1 nCi/g
(soil); root and fobar
surface contamination
contributed about equally.

Tree foliage and herbaceous
samples collected from
a floodplain in Oak
Ridge contaminated for
30 yr; evidence for
plutonium-organic
complex; plutonium in
monomeric forms.

Same as for native vegetation;
coconut meat used for
human food in the
tropics.

Romney et al.,
1976

Whicker, 1976

Energy Research
and Development
Administration,
1976

Koranda et al.,
1973

Pot culture
Alfalfa

Alfalfa

Bahia grass

Barley

Americium

lO"'' to 10"^ Grown in 3 kg of NTS soil;
1 nCi/g(soU)of ' Am;
chelate increased CR;
OM tended to decrease
CR.

10""' Grown in 1.2 kg of Los

Alamos mountain
meadow soil spiked
with l8nCi to 0.5
juCi/g (soil) of '■* ' Am.

10""" to 10~' Grown in 2 kg soil from
SRP; 0.5 kg of top
pot soil spiked with
Am(N03)3 to give
500 pCi/g (soil); lowest
CR caused by lime.

10~' to 10° Grown in 500 g of soil

spiked with 1 nCi/g
of "'" Amas Am(N03)3;

This study

Adams et al.,
1975

This study

Wallace et al.,
1976

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 355

TABLE 7 (Continued)

Plant species

Range ofCR

Conditions and comments

Reference

Barley

highest CR caused
by chelate applied
to a calcareous soil.

Grain

lO-*"

Same as for barley

Schulz et al..

Foliage

10-"

Plutonium study; 1 to

11 nCi/g(soil)of ="' Am.

1976b

Barley

Grain

10-' to 10"'

Grown in 3 kg of NTS

Energy Research

FoUage

10-' to 10-'

soil; 0.6 nCi/g
(soil)of = "'Am;
highest CR produced
by chelate treatment.

and Development

Administration,

1976

Barley (plants)

10-"

Grown in 1.2 kg of Los

Adams et al.,

Alamos mountain meadow

1975

10"

Bean, bush (shoots) 10 ' to 10^
Bean, bush (shoots) 10-' to 10'

Beans

10"

soil spiked with 18 nCi to

0.5 iuCi/g(soil)of '"' Am.

Grown in Cinebar (pH 4.5)
and Ephrata (pH 7.5)
soils using the
Neubauer technique;
soils spiked with 1 .8
/uCi/g of * " ' Am
as Am(N03)3 ; uptake
similar from two soils.

Same as for barley.

Grown in 500 g of soil
spiked with 2 nCi/g
(soil) of '" ' Am as
Am(N03)3; lowest
CR occurred on limed
clay soil; highest CR
occurred on limed and
chelated treatment.

Grown in standard Hoagland
solution spiked with
0.9 AiCi/liter of
'"'Am as Am(NO,),.

Corn (shoots)

10-' to 10"

Same as for barley.

Corn (shoots)

10-^ to 10"

Same as for beans.

Rice

Grain

BGt to 10-'

Grown under flooded or

Leaves

10-' to 10"

nontlooded conditions
in 5 kg of soil from

aine, 1968

Wallace et al.,

1976
Adriano et al.,

1977

aine, 1968

Wallace et al.,

1976
Adriano et aL,

1977

This study

(Tabic continues on the next page.)

356 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 (Continued)

Plant species

Range of CR

Conditions and comments

Reference

Soybean

Fruit

10-' to 10-^

Foliage

10-' to 10-'

Wheat (grain)

Qieatgrass
Tumbleweed

Field studies
Corn + cobs

Potatoes

Peas (shelled)
Native vegetation

Native vegetation

10-' to 10- =

10-' to 10-"
10-' to 10""

10"

10"' to 10

10-^ to 10

3t culture

Cheatgrass

10-'

Tumbleweed

10-'

SRP; spiked top 1 kg
with Am(N03)3 to
give 400 pCi/g; highest
CR caused by chelate
treatment.

Grown in 3 kg of NTS
soil; 0.2 to 2 nCi/g
(soil); highest CR
caused by chelate.

Same as for wheat plu-
tonium study; ^ " ' Am
added as chloride or
nitrate to give
45 nCi/g(soU);
highest CR obtained
when added as chloride
to a neutral soil.

Cheatgrass and tumbleweed
grown in 1 kg of soil
spiked with 25 juCi/g
of ^" Am as Am(N03)3;
some organics suppressed
uptake.

Same as for plutonium
study; 1 fCi/g (soil)
of-'"Am

Same as for plutonium
study; 1 fCi/g (soil)
of'"' Am.

Same as for plutonium
study; 1 fCi/g (soil)
of 'Am.

Native perennial shrubs
inhabiting the NTS
environs; mainly sur-
face contamination.

Same as for plutonium
study.

Curium

Cheatgrass and tumbleweed
grown in 1 kg of soil
spiked with 25 nCi/g
of ""'CmasCm(N03)3;
uptake not affected by
organics.

This study

Schulz, Tompkins,
and Babcock,
1976a

Price, 1973

Hardy, Bennett,
and Alexander,

1977

Energy Research
and Development
Administration,
1976

Koranda et al.,
1973

Price, 1973

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 357

TABLE 7 (Continued)

Plant species

Range of CR

Conditions and comments

Reference

Pot culture

Cheatgrass
Tumbleweed

10
10

Neptunium

Cheatgrass and tumbleweed
grown in 1 kg of
soil spiked with
50nCi/gof ^"'Np
as NpCNOj)^ ; organics
tended to enhance
uptake.

Price, 1973

^_ _ radioactivity/g (plant ash)

radioactivity/g (soil)
t Background level of activity.

Summary

Experiments on plant uptake of the transuranic nuclides conducted under controlled
environmental conditions using spiked soils from the Savannah River Plant and
contaminated soils from the Nevada Test Site revealed the following:

1. In general, the uptake of '■*' Am by crop plants (Baliia grass and rice) grown on
soils from the humid southeastern United States was influenced by soil amendments and
indigenous soil factors. Lime generally immobilized ^"^^Am in the soil and decreased
uptake by Baliia grass. The DTPA chelate somewhat enlianced the uptake by other crops
tested in these acidic soils, but its greatest effect occurred where DTPA was supplied in
Hmed soils.

The addition to soil of up to 5.0% OM appeared to demote ^^^ Am uptake in Bahia
grass, but its effect was not so great as that of lime alone. This was possibly caused by
temporary immobilization of ^"^'Am in microbial bio mass, by an increase in cation-
exchange capacity, or production of organic ligands from the OM.

The addition of ^ "* ^ Am to soil resulted in almost no translocation of this radionuclide
to the rice grain. However, when ^^'Am was introduced in a chelated form to the
ponded water, it appeared that there was relatively more absorption and translocation to
the grain.

2. Crop plants (barley, alfalfa, and soybeans) grown on Nevada Test Site soils had an
average CR of 10^"^ for plutonium in the vegetative parts. Americium appeared more
available than plutonium; the average americium/plutonium ratio was about 4 for barley,
10 for alfalfa, and 22 for soybeans.

Of the various soil amendments (nitrogen, sulfur, OM, and DTPA) used, only DTPA
markedly and consistently increased both plutonium and americium uptake by plants; the
increase with barley and soybeans was usually more than one order of magnitude but
only a factor of 2 with alfalfa.

Americium appeared to be more mobile for transport (70% greater) than plutonium
from shoots to fruits in soybeans. Concentrations of both plutonium and americium in
shoots were liighly correlated compared to those in fruits. The chelator DTPA did not
differentially influence the transport of plutonium and americium from shoots to fruits.

358 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Acknowledgments

Tliis work was supported by the U. S. Department of Energy through contract
Nos. EY-76-C-09-0819 (to Savannah River Ecology Laboratory, University of Georgia)
and EY-76-C-03-0012 (to University of CaUfornia, Los Angeles).

References

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States, /tc?i. Agron., 14: 61-108.
Adams, W. H., J. R. Buchholz, C. W. Christenson, G. L. Johnson, and E. B. Fowler, 1975, Studies of
Plutonium, Americium, and Uranium in Environmental Matrices, ERDA Report LA-5661, Los
Alamos Scientific Laboratory, NTIS.
Adriano, D. C, M. S. Delaney, G. D. Hoyt, and D. Paine, 1977, Availability to Plants and Soil
Extraction of Americium- 241 as Influenced by Qielating Agent, Lime and Soil Type, Environ.
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Bernhardt, D. E., and G. G. Eadie, 1976, Parameters for Estimating the Uptake of Transuranic
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Brown, K. W., 1976, Americium — Its Behavior in Soil-Plant Systems, Report EPA-600/3-76-005,

U. S. Environmental Protection Agency, Las Vegas, Nev.
Cataldo, D. A., R. C. Routson, R. E. Wildung, and T. R. Garland, 1976, Uptake and Distribution of
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BNWL-2000(Pt. 2), Battelle Pacific Northwest Laboratories, NTIS.
Chandrasekaran, W., and T. Yoshida, 1973, Effect of Organic Acid Transformations in Submerged

Soils on Growth of the Rice Plant, Soil Sci. Plant Nutr. (Tokyo), 19(1): 3945.
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Report for 1977 to the USAEC Division of Biology and Medicine. USAEC Report BNWL-714,
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USAEC Report NVO-142, Nevada Operafions Office, NTIS.
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Elements, Proceedings of the Workshop, Battelle Seattle Research Center, Seattle, Wash.,
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FiskeU, J. G. A., 1970, Cation Exchange Capacity and Component Variations of Soils of Southeastern

USA, Soil Sci. Sac. .Am., Proc, 34: 723-727.
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2(1): 67-70.
Hardy, E. P., B. G. Bennett, and L. T. Alexander, 1977, Radionuclide Uptake by Cultivated Crops,
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Harlan, J. R., 1976, The Plants and Animals that Nourish Man, Sci. Am., 235(3): 89-98.
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Nutrients in Acid Soil, Soil Sci. Sac. Am., Proc, 38: 341-346.
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Nuclear Fission Adsorbed on SoU Colloids, 5o// 5a.. 65: 129-134.
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Terrestrial Biota Survey, in Enewetak Radiological Survey, USAEC Report NVO-140,
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in Nevada, USAEC Report NERC-LV-539-Z8, Nafional Environmental Research Center, NTIS.
Lipton, W. v., and A. S. Goldin, 1976, Some Factors Influencing the Uptake of Plutonium-239 by Pea

Plants, Health Phys., 31: 425430.
Majors, W. J., K. D. Lee, R. A. Wessman, and R. Melgard, 1973, Determination of ^ ^ ' Pu and ^ "' Am
in Large NAEG Vegetation Samples, in The Dynamics of Plutonium in Desert Environments, P. B.

UPTAKE OF TRANSURANIC NUCLIDES FROM SOIL BY PLANTS 359

Dunaway and M. G. White (Eds.), USAEC Report NVO-142, pp. 1U7-118, Nevada Operations

Office, NTIS.
Mueller, Anita A., and R. F. Mosley, 1916, Availability, Uptake and Translocation of Plutonium

Within Biological Systems: A Review of the Literature, Report EPA-600/3-76-943, U. S.

Environmental Protection Agency, Las Vegas, Nev.
Myttenaere, C, and U. Marckwordt, 1967, Indirect Contamination of Lowland Rice with

Manganese-54, in Symposium on Radioecology , Proceedings of the Second National Symposium,

Ann Arbor, Mich., May 15-17, 1967, D.J. Nelson and F. C. Evans (Eds.), USAEC Report

CONF-670503, pp. 561-570, NTIS.
, P. Bourdeau, and M. Masset, 1969, Relative Importance of Soil and Water in the Indirect

Contamination of Flooded Rice with Radiocesium, Health Pliys.. 16: 701-707.
Newbould, P., 1963, Absorption of Plutonium 239 by Plants, Annual Report on Radiobiology,

Agricultural Research Council, p. 86.
^— , and E. R. Mercer, 1961, Absorption of Plutonium-239 by Plants, in Agricultural Research

Council Radiobiological laboratory Annual Report, 1961-1 962, Report ARCRL-8, pp. 81-82,

Agricultural Research Council.
Price, K. R., 1972a, The Behavior of Waste Radionuclides in Soil-Plant Systems, in Pacific Northwest

Laboratory Annual Report for 1972 to the USAEC Division of Biomedical and Environmental

Research, USAEC Report BNWL-1750(Pt. 2), BatteUe Pacific Northwest Laboratories, NTIS.
, 1972b, Uptake of'-^'' Np,^ ^"^Pu. ^ * Mm, and ' " " Cm from Soil by Tumbleweed and Cheatgrass,

USAEC Report BNWL-1688, Battelle Pacific Northwest Laboratories, NTIS.
— -, 1973, Tumbleweed and Cheatgrass Uptake of Transuranium Elements Applied to Soil as Organic

Acid Complexes, USAEC Report BNWL-1755, Battelle Pacific Northwest Laboratories, NTIS.
Rediske, J. H., and A. A. Selders, 1954, The Absorption and Translocation of Plutonium-239 and

Cerium-144 by Plants, in Biology Research— Annual Report for 1953. USAEC Report HW-30437,

Hanford Atomic Products Operation, NTIS.
, J. F. Qine, and A. A. Selders, 1955, The Absorption of Fission Products by Plants, USAEC

Report HW-36734, Hanford Atomic Products Operation, NTIS.
Relyea, J. F., and D. A. Brown, 1978, Adsorption and Diffusion of Plutonium in Soil, in

Environmental Chemistry and Cycling Processes, DOE Symposium Series, Augusta, Ga.,

Apr. 28-May 1. 1976, D. C. Adriano and I. L. Brisbin (Eds.), pp. 479-494, CONF-760429, NTIS.
Romney, E. M., and K. R. Price, 1959, Availability of Pu-239 to Plants from Contamination, in

Semiannual [^ogress Report for the Period Ending December 31, 1969, USAEC Report

UCLA-457, University of California at Los Angeles, NTIS.
- — , H. M. Mork, and K. H. Larson, 1970, Persistence of Plutonium in Soil, Plants, and Small

Mammals, Health Phys., 19: 487-491.
— , A. WaUace, R. O. Gilbert, and J. E. Kinnear, 1976, 239-24opy ^^^ ^4 1^,^ Contamination of

Vegetation in Aged Fallout Areas, in Transuranium Nuclides in the Environment, Symposium

Proceedings, San Francisco, 1975, pp. 470-491, STI/PUB/410, International Atomic Energy

Agency, Vienna.
Schulz, R. K., G. A. Tompkins, and K. L. Babcock, 1976a, Uptake of Plutonium and Americium by

Plants from Soils: Uptake by Wheat from Various Soils and Effect of Oxidation State of

Plutonium Added in Soil, in Transuranium Nuclides in the Environment, Symposium Proceedings,

San Francisco, 1975, pp. 303-310, STI/PUB/410, International Atomic Energy Agency, Vienna.
— , G. A. Tompkins, L. Leventhal, and K. L. Babcock, 1976b, Uptake of Plutonium and Americium

by Barley from Two Contaminated Nevada Test Site Soils,/ Environ. Qual, 5(4): 406410.
Takijima, Y., 1964, Studies in Organic Acids in Paddy Field Soils with Reference to Their" Inhibiting

Effects on the Growth of Rice Plants 2: Relations Between Production of Organic Acids in

Water-Logged Soils and the Root Growth Inliibition, Soil Sci. Plant Nutr. [Tokyo), 10(5): 22-29.
Wallace, A., 1969, Behavior of Certain Synthetic Chelating Agents in Biological Soil Systems, Annual

Progress Report for the Fiscal Year, USAEC Report UCLA-34P-51-35, University of California at

Los Angeles, NTIS.
, 1972a, Increased Uptake of ^"'Am by Plants Caused by tlie Chelating Agent DTP A, Health

Phys., 22: 559-562.
, 1972b, Effect of Soil pH and Chelating Agent (DTPA) on Uptake by and Distribution of ^ "' Am

in Plant Parts of Bush Beans, Radiat. Bot., 12: 433435.

360 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

, E. M. Romney, R. T. Mueller, and P. M. Patel, 1976, ^^ ' Am Availability to Plants as Influenced

by Chelating Agents, in Radioecology & Energy Resources, Proceedings of Fourth Symposium on
Radioecology, Oregon State University, May 12-14, 1975, pp. 104-107, C. E. Gushing (Ed.),
Ecological Society of America Special Publication Series No. 1, Academic Press, Inc., New York.

Whicker, F. W., 1976, Radioecology of Natural Systems in Colorado, Fourteenth Annual Progress
Report, May 1, 1975-July 31, 1976, USAEC Report COO-1 156-84, Colorado State University,
NTIS.

Wildung, R. E., and T. R. Garland, 1974, Influence of Soil Plutonium Concentration on Plutonium
Uptake and Distribution in Shoots and Roots of Barley,/ Agric. Food Chem., 22: 836-838.

Wilson, D. O., and J. F. Qine, 1966, Removal of Plutonium 239, Tungsten 185, and Lead 210 from
Soih,Nature. 209: 941-942.

Comparative Uptake and Distribution
of Plutonium, Americium, Curium,
and Neptunium in Four Plant Species

R. G. SCHRECKHISE and J. F. CLINE

The uptake of the nitrate forms of^^^Pu, '^^^Pii, '^'^^Am, ^ "* "* Cm, and ^ ^ "^ Np from soil
into selected parts of four different plant species grown under field conditions was
compared. Alfalfa, barley, peas, and cheatgrass were grown outdoors in small weighing
lysimeters filled with soil containing these contaminants. The plants were harvested at
maturity, divided into selected components, and radiochemically analyzed by alpha-
energy analysis. Soil concentration did not appear to affect the plant uptake of ^^^Pu,
^^^Pu, ^^^ Am, or ^^^Cm for the two levels used. The relative uptake values of ^^^Pu
and ^^^Pu were not significantly different from each other and the ^'^ ^ Am uptake values
were not significantly different from the ^^'^Cm values. The relative plant uptake of the
four different transuranium elements was Np > Cm —Am> Pu. Relative uptake values of
neptunium into various plant parts ranged from 2,200 to 45,000 times as great as those of
plutonium, whereas americium and curium values were 10 to 20 times as great. The seeds
were significantly lower than the rest of the aboveground plant parts for all four
transuranics. The legumes accumulated approximately 10 times as much as the grasses. A
hypothetical comparison of the radionuclide content of plants grown in soil contami-
nated with Liquid Metal Fast Breeder Reactor fuels indicates that concentrations of
isotopes of americium, curium, and neptunium would exceed ^^^ Pu values.

The release of transuranium nuclides to environmental systems, whether planned or not,
poses potential hazards^, especially if the biologically toxic materials enter food chains
leading to man. Quantitative information on transport parameters is required for an
assessment of the potential health hazards from such releases. One parameter that
warrants close attention is the plant uptake of transuranics from contaminated soil. It
used to be assumed that all transuranium elements behaved like plutonium and were
equally discriminated against by plants. However, studies by Cline (1968) and Schulz et
al. (1976) indicate a difference in the phytoavailability of ■^'^ ' Am and ^^^Pu.

Variations in the relative uptake of plutonium by plants, as summarized by the
Energy Research and Development Administration (1975), can be explained by
differences in plant species or fragments examined. Other factors that affect uptake are
edaphic parameters, environmental conditions, and differences in the chemical form or
valence state of the plutonium initially added to the soil. This chapter reports the relative
plant uptake of ^^^Pu, ^^^Pu, '^^ 'Am, ^"^"^Cm, and ^^''Np from soil into selected parts
of four different plant species grown under field conditions. The data presented here are
first-year results of a long-term study of the effects of aging, weathering, and associated
biological processes in soil on the phytoavailability of transuranium elements.

56/

362 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Materials and Methods

The radiological safety requirements for experimentally placing transuranic elements
ill the field are stringent and precise. The radionuclides must be securely contained,
readily retrievable, and isolated to eliminate biological and physical transport away from
the study site. The plants in this study were grown outdoors in transuranic-contaminated
soil contained in small weighing lysimeters. The containers were isolated from biota by a
wire-mesh exclosure designed to exclude mammals and birds (Hinds et al., 1976). The
exclosure was situated on the Arid Lands Ecology Reserve located on the Department of
Energy's Hanford Site in South Central Washington.

The containers were constructed from 1-m lengths of polyvinyl chloride (PVC) pipe
measuring 13.2 cm in inside diameter. Metal bale handles were attached to the open end,
and the other end was enclosed with a watertight end cap. These containers were placed
inside a slightly larger (15.7-cm-diameter) PVC pipe buried vertically in the ground so
that the upper end was level with the ground surface. This arrangement facilitated
retrieval of the contaminated soil and exposed the soil profile in the containers to realistic
outdoor conditions of temperature and precipitation (Hinds, 1975).

Treatment containers were initially filled to within 35 cm from the top with 1 1.1 kg
of oven-dried soil. Nitrate forms of ^^^Pu, ^^^Pu, ^'^'Am, ^^"^Cm, and ^'^^Np were
individually added to a 3.4-kg aliquot of oven-dried soil, which was then placed in
separate containers in a layer 20 cm thick. An additional 1.7 kg (10 cm) of clean soil was
added to the top of the contaminated soil. This brought the level of soil to within 5 cm of
the top of the container. The surface layer of clean soil was intended to prevent the
spread of radionuclides by wind to the surrounding environment or their deposition on
the surface of the experimental plants, which would have produced erroneous uptake
values.

The soil used for this study was a silt loam of the Ritzville series. The soil has a pH of
6.2 and a cation-exchange capacity of 22.5 meq/100 g at pH 7 (Wildung, 1977).
RadionucUdes were added to the soil by pipetting I ml of the 4M HNO3 solutions
directly onto the soil, which had been adjusted to 5% moisture content. The oxidation
state of the radionuclides when added to the soil was +4 for the plutonium isotopes, +3
for ^"^^ Am aiTd ''^'^Cm, and +5 for the ^-^''Np. Enough CaCOj was added to the soil to
neutralize the HNO3. The amended soil was stored for 24 hr and then thoroughly mixed
in a V-blender before it was transferred to the containers. Two different amounts, 1 .0 and
0.1 mCi/3.4 kg soil, of "^Pu, ^^^Pu, ^^'Am, and ^^^Cm were added to the soil.
Neptunium-237 was added to the soil only at a concentration of 0.1 inCi/3.4 kg soil.

Control containers containing only uncontaminated soil were also prepared so that
the levels of contamination in the treatment vegetation attributable to external
deposition or root uptake of radionuclides present in the soil from fallout or other
sources could be determined.

Cheatgrass {Bronius tectorum L.), an annual grass, was planted in some of the
containers. The only water the cheatgrass containers received came from natural
precipitation, which averages 16 cm/yr for the study site (Hinds and Thorp, 1971 ). Peas
(Pisuni sativum, var. Blue Bonnet), barley (Honlcum viilgare, var. U. Cal. Briggs), and
alfalfa {Medicago sativa, var. Ranger) were planted in tlie spring. The crop plant
containers were irrigated and weighed so that the soil moisture content was maintained at
about 20% by weight throughout the growing season. The peas, barley, and alfalfa plants
were fertihzed with NH4NO3, at the rate of 300 kg/ha, approximately halfway through

UPTAKE AND DISTRIBUTION OF Pit, Am, Cm, AND Np 363

the growing seasons. Super phosphate (P2O5) was also added to the alfalfa containers at a
rate of about 250 kg/ha.

Aboveground plant parts were hand harvested at maturity and divided into selected
components. The entire cheatgrass plant, separated from other plant species that had
invaded the containers, was analyzed. Barley seeds were analyzed separately from the rest
of the plant. Peas, harvested at the dry-seed stage, were divided into seeds, leaves, and
stem and pod fragments for analysis. The entire alfalfa plant (three separate harvests) was
analyzed. Radiochemical analyses (Major et al., 1973; Wessman et al., 1978) of the plant
materials were conducted by the LFE Environmental Analysis Laboratory, Richmond,
Calif.

Radiochemical analysis of the control plants was used to determine the net uptake of
transuranics by the plants in the amended soil. The radionuclide concentration observed
in the treatment plants was corrected by subtracting the corresponding values of the
control plant parts.

Results

Results of the radiochemical analyses are summarized in Tables 1 and 2. The data are
presented as a ratio of the concentration of the radionuclide in the vegetative part to the
total amount of that radionucUde added to the 3.4 kg of soil. The ratio can be used to
compare the relative uptake values of the five different radionuclides into the various
parts of the four plant species. Since the contaminated soil in this study was covered with
a 10-cm layer of clean soil, these values are not to be regarded as concentration ratios
(CR). Concentration ratio values are normally calculated by dividing the concentration of
the vegetation (activity per unit dry weight) by the concentration of the top 10 to 20 cm
of soil (also in activity per unit dry weiglit).

TABLE 1 Relative Uptake of Transuranium Elements by Cheatgrass and Alfalfa

Millicuries

per
container

Average (pCi/g dry vegetation)

(± standard error)*

mCi/container

Cheatgrass

Alfalfa

Isotope

1st harvest

2nd harvest

3rd harvest

238p^j

0.108

1.3 ±

0.3

30 ± 7

1 3 ± 3t

58 ±42$

238 p^

1.06

2.7 ±

0.6

16 ± 2±

22±6±

34 ± 13$

239p^j

0.103

12±

31:

80 ± 21

17 ±5

45 ± 20

239p^,

0.989

5.0 ±

0.8$

24 ± 7

14 ±2i

16 ± 2$

^^'Am

0.0946

(1.4 ±

0.2) X 10'

(60 ± 14) x 10'

(1.7 ± 0.2) X 10'

(2.7 ± 0.4) x 10'

^^'Am

0.983

(0.48 ±

0.11) X 10'

(7.0 ± 1.1) X 10'

(1.7 ±0.5)x 10'

(3.5 ± 0.5) X 10'

^^^Cm

0.103

(0.65 ±

0.25) X 10'$

(7.8 ± 1.9) X 10'

(2.3 ±0.5)x 10'

(3.8 ± 0.6) X 10'

^^^Cm

1.04

(0.80 ±

0.17)x 10'

(3.5 ±0.8)x 10'

(1.5 ±0.3) x 10'

(2.5 ± 0.6) X 10'

23 7 ^p

0.101

(1.1 ±

O.Dx lO't

(21 ± 7)x 10'

(5.5 ± 1.3) x 10"

(4.7 ± 1.2) X 10"

Grams of dry tissue

per container

3.9 ±

0.3

7.3 ± 0.8

10.0 ± 0.3

6.3 ± 0.2

*n =

= 5, except as noted.

tn =

= 3.

$n = 4.

364 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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UPTAKE AND DISTRIBUTION OF Pu, Am, Cm, AND Np 365

As illustrated in Fig. 1 , the relative uptake of either plutonium isotope was not
statistically different (a = 0.01) for the two soil concentrations used (approximately 0.03
and 0.3 juCi/g soil). Soil concentration did not appear to have any effect on the uptake of
■^"^'Am or ^'^'*Cm, which were also added at the same two levels. Also, as shown in
Fig. 1, the relative uptake of ^^^Pu was not statistically different from ^•'^Pu (a = 0.01).

Since soil concentration did not appear to affect phytoavailability or differences in
^^^Pu vs. ^^^Pu, the relative plant-uptake data were combined for each element so that
further statistical comparisons could be made. Most interesting was a comparison of the
uptake values for the four different elements (Fig. 2). The relative uptake of ^^^Np
ranged from 2,200 to 45,000 times as great as that of plutonium, depending on the plant
part compared. Neptunium-237 was accumulated 35,000 and 45,000 times as great as
plutonium in barley and pea seeds, respectively, and averaged a factor of 4,700 more in
the remaining plant tissues. There was no significant difference (a = 0.01) in the uptakes
of ^'^^ Am and ^^'^Cm, which were both 10 to 20 times as great as plutonium.

The relative uptakes of the transuranics by the four different plant species were
noticeably different. Generally, the legumes (peas and alfalfa) accumulated approxi-
mately 10 times as much as the grasses (cheatgrass and barley). Concentrations of
transuranics in various plant tissues examined were also different. The values were
considerably lower in the seeds than in other aboveground plant parts. The concentra-
tions in barley seed were lower by a factor of 30 to 50 than those in the entire combined
plant parts for plutonium, americium, and curium and about a factor of 5 lower for
neptunium. For peas the ratio of concentration in the seeds compared to the rest of the
plant was 230, 150, 70, and 30 for plutonium, americium, curium, and neptunium,
respectively.

Discussion

Our results showed soil concentration to have no observable effect on the uptake of
either plutonium isotope. This differs somewhat from results reported by Wildung and
Garland (1974). They observed an increase in the relative uptake of plutonium as the soil
concentration decreased. However, they noted little effect on uptake at soil concentra-
tions of 0.5 /iCi/g or less, which exceeded the maximum level used in this study (0.3
A/Ci/g).

Plutonium-238 has been reported to be more available than ^^^Pu in a grassland
ecosystem (Little, 1976) and in the southeastern United States (McLendon et al., 1976).
As noted previously, this difference was not observed in this plant uptake study.

The results of this study showed that the relative plant uptake of the four different
transuranium elements was Np > Cm ^ Am > Pu. These trends are consistent with data
reported by Price (1972) on the uptake of "^Pu, ^^'Am, ^^"^Cm, and ^^^Np by
cheatgrass and tumbleweeds (Salsola kali). The same differences in the relative uptake of
2"^^ Am and "^Pu have been reported by Cline (1968) and Schulz et al. (1976). A
significant aspect of this trend is that, if the CR of plutonium is approximately 0.0001 as
summarized by Price (1973) and ^^''Np was taken up into the entire plant some 3900
times as great as plutonium, one can infer that the CR value for neptunium would be
about 0.4; i.e., under usual agronomic conditions, the concentration of vegetation
growing on soil contaminated with ^^^Np in the upper 15 to 20 cm would be equal to
approximately one-half the soil concentration on a dry-weight basis. Likewise, the CR
values for ^'^^ Am and ^'^'^Cm would be expected to be about 0.002 since the relative

366 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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UPTAKE AND DISTRIBUTION OF Pu, Am, Cm, AND Np 361

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368 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

uptake of these two transuranics was approximately a factor of 16 times that of
plutonium.

The significance of these CR values can be related to transuranics associated with
plutonium breeder reactor fuels. Thomas and Healy (1976) reviewed the uptake of
neptunium and transplutonium elements by plants. They calculated the relative
abundance of a number of long-lived transuranics relative to ^"'^Pu from data reported
earher by Bell (1970) on nuclide levels in spent Liquid Metal Fast Breeder Reactor
(LMFBR) fuels. As shown in Table 3, the concentration of ^"^ • Am, ^''^ Cm, ^-^ ^Np, and
^■^^Np in vegetation would be higher than that of ^^^Pu for various time periods
following environmental releases of spent LMFBR fuels. A difference is noted in the
relative plant uptake values summarized by Thomas and Healy for americium and
neptunium when compared with the values reported in this study. The discrepancy in the
americium values is probably due to the fact that they included uptake data in which the
soil had been amended v^th a chelating agent, such as diethylenetriaminepentaacetic acid,
which significantly increases plant uptake (Hale and Wallace, 1970; Wallace, 1972). Their
relative neptunium uptake value was taken from Price (1972) and was approximately
one-tenth of the results presented here. Price noted some toxicity symptoms in the
neptunium-contaminated seedlings which may have caused a reduction in neptunium
uptake.

Another interesting finding in this study was the low concentrations of transuranics
in pea and barley seeds when compared with those of the entire plant. This is important
because in many dose-assessment models the CR values used are often calculated from the
entire aboveground plant parts. As shown in this study, the levels of plutonium,
americium, and curium in barley seeds were lower by a factor of 30 to 50 of those in the
entire plant. For neptunium the seeds were one-fifth of the entire plant values. Pea seeds
were lower by a factor of 70 to 230 for plutonium, americium, and curium and a factor
of 30 lower for neptunium. Differences in plant-part concentrations must be considered
when CR values are used in dose-assessment models and can also describe some of the
discrepancy in CR values reported in the literature.

Acknowledgments

We thank W. T.- Hinds for his technical guidance during the initial stages of this study
and H. A. Sweany, M. J. Harris, L. F. Nelson, M. A. Combs, and V. D. Charles for their
teclinical assistance througliout the study. This research was funded by the U. S.
Department of Energy, Office of Health and Environmental Research, under contract
EY-76-C-06-1830.

References

Bell, M. J., 1970, Heavy Element Composition of Spent Power Reactor Fuels, USAEC Report

ORNL-TM-2897, Oak Ridiie National Laboratory, NTIS.
Qine, J. F., 1968, Uptake of ^'" Am and "'Pu by Plants, in Pacific Northwest Laboratory Annual

Report for 1967 to the USAEC Division of Biology and Mcdicinf. USAEC Report BNWL-714. pp.

8.24-8.25, Battelle, Pacific Northwest Laboratories, NTIS.
Energy Research and Development Administration, 1975, Workshop on Environmental Research for

Transuranium Elements, Proceedings of the Workshop, Battelle Seattle Research Center, Seattle,

Wash., Nov. 12-14, 1975, ERDA Report ERDA-76/134, NTIS.
Hale, V. Q.. and A. Wallace, 1970, Effect of Chelates on Uptake of Some Heavy Metal Radionuclides

from Soil by Bush Beans. Soil Sci., 109: 262-263.

UPTAKE AND DISTRIBUTION OF Pu, Am. Cm, AND Np 369

(SO

£3

•«

"OJ

-4-^

-S

O h

^^H

■4— •

r<^

.2 «
55:

c oa
< u-

"^ ^

1* *-

« a.

I °

C3 ^

•a
c

■^■J

•^-t

>
■0

o
o
o

3 *-
c >

a. J2

ff'

b«

■X

kri

o
o
o

re s

3
<

t^.

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■te-'

«
~

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o
o
o

rr,

00
VO

0
0

00000

<N (N CN O O

o o

o
o

0000

O 'd-

q
d

O lO m O o
O — I — " <^> <~o
00 c^ t^

\q q
— d

— d

o o

o o o
^ ^ —

ON 00

."2

^w-

L--

iz:

■^

3 10 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Hinds, W. T., 1975, Energy and Carbon Balances in Cheatgrass: An Essay in Autecology, Ecol.

Monogr., 45: 367-388 (1975).
, and J. M. Thorp, 1971, Annual Summaries of Microclimatological Data from the Arid Lands

Ecology Reserve: 1968 Through 1970, USAEC Report BNWL-1629, Battelle, Pacific Northwest

Laboratories, NTIS.
, J. F. Cline, H. A. Sweany, and V. D. Charles, 1976, Weather and Aging Effects on Uptake of

Transuranics by Plants, in Pacific Northwest Laboratory Annual Report for 1975 to the USLRDA

Division of Ecological Science, ERDA Report BNWL-2000 (Pt.2), pp. 176-177, Battelle, Pacific

Northwest Laboratories, NTIS.
Little, C. A., 1976, Plutonium in a Grassland Ecosystem, Ph.D. Thesis, Colorado State University, I'ort

Colhns, Colorado.
Major, W. J., K. D. Lee, R. A. Wessman, and R. Melgard, 1973, Determination of Plutonium-239 and

Americium-241 in Large NAEG Vegetation Samples, in Nevada Applied Ecology Group

Procedures Handbook for Environmental Transuranics. M. G. White and P. B. Dunaway (Eds.),

pp. 271-282, ERDA Report NVO-166, Nevada Operations Office, NTIS.
McLendon, H. R., O. M. Stewart, A. L. Boni, J. C. Corey, K. W. McLeod, and J. E. Pinder, 1976,

Relationships Among Plutonium Contents of Soil, Vegetation and Animals Collected on and

Adjacent to an Integrated Nuclear Complex in the Humid Southeastern United States, in

Transuranium Nuclides in the Environment, Symposium Proceedings, San Francisco, 1975,

pp. 347-363, STI/PUB/410, International Atomic Energy Agency, Vienna.
Price, K. R., 1972, Uptake of ^'■'Np, ^^"Pu, ^^'Am and ^""Cm from Soil by Tumbleweed and

Cheatgrass, USAEC Report BNWL-1688, Battelle, Pacific Northwest Laboratories, NTIS.
, 1973, A Review of Transuranic Elements in Soils, Plants, and Animals, J. Environ. Quai, 2(1):

62-66.
^Schulz, R. K., G. A. Tompkins, L. Lerenthal, and K. L. Babcock, 1976, Uptake of Plutonium and

Americium by Barley from Two Contaminated Nevada Test Site Soils, J. Environ. Quai, 5(4):

406-410.
Thomas, R. L., and J. W. Healy, 1976, Appraisal of Available Information on Uptake by Plants of

Transplutonium Elements and Neptunium, ERDA Report LA-6460-MS, Los Alamos Scientific

Laboratory, NTIS.
Wallace, A., 1972, Increased Uptake of ^'"Am by Plants Caused by the Chelating Agent DTPA,

Health Phys., 22(6): 559-562.
Wessman, R. A., K. D. Lee, B. Curry, and L. Leventhal, 1978, Transuranium Analysis Methodologies

for Biological and Environmental Samples, in Environmental Chemistry and Cycling Processes,

DOE Symposium Series, Augusta, Ga., Apr. 28-May 1, 1976, D. C. Adriano and I. Lehr Brisbin,

Jr. (Eds.), pp. 275-289, CONI'-760429, NTIS.
Wildung, R. E., 1977, Soils of the Pacific Northwest Shrub-Steppe. Occurrence and Properties of Soils

on the Arid Lands Ecology Reserve, Hanford Reservation, ERDA Report BNWL-2272, Battelle,

Pacific Northwest Laboratories, NTIS.
, and T. R. Garland, 1974, Infiuence of Soil Plutonium Concentration on Plutonium Uptake and

Distribution in Shoots and Roots of Barley,/. Agric. Eood Chem., 22(5): 836-838.

Comparative Distribution of Plutonium
in Contaminated Ecosystems at Oak Ridge,
Tennessee, and Los Alamos, New Mexico

ROGER C. DAHLMAN, CHARLES T. GARTEN, JR., and THOMAS E. HAKONSON

The distribution of plutonium was compared in portions of forest ecosystems at Oak
Ridge, Tenn., and Los Alamos, N.Mex., which were contaminated by liquid effluents.
Inventories of plutonium in soil at the two sites were generally similar, but a larger
fraction of the plutonium was associated with biota at Los Alamos than at Oak Ridge.
Most (99.7 to 99.9%) of the plutonium was present in the soil, and very little (0.1 to
0.3%) was in biotic components. Comparative differences in distributions within the two
ecosystems appeared to be related to individual contamination histories and greater
physical transport of plutonium in soil to biotic surfaces at Los Alamos.

Currently most of the plutonium present in terrestrial ecosystems of the United States
originates from nuclear weapons testing and from the reentry burnup of the SNAP-9A
satellite power source (Hanson, 1975). In the future, however, local ecosystems will
receive small quantities of plutonium released from nuclear fuel reprocessing and
fabrication facilities. The purpose of this chapter is to compare and contrast the
distribution of plutonium in two contaminated ecosystems that are representative of
humid and semiarid environments of the United States. Current plutonium inventories for
ecosystems several decades after initial contamination can help ecologists forecast the fate
of plutonium in the environment. One important question is whether the availability of
this element to plants and other organisms will change after it has been subjected to
weathering and ecological processes of the environment. Potential radiological toxicity
and long physical half-lives of plutonium dictate that its behavior in ecosystems be
understood.

Although the ecosystems at Oak Ridge, Tenn., and Los Alamos, N.Mex., are
dissimilar owing to differences in geology, climate, and ecology, there are certain similar
features of these contaminated environments. A forested floodplain at Oak Ridge and a
canyon at Los Alamos were contaminated by treated hquid waste effluents which have
resulted in detectable levels of plutonium in most ecosystem components. However, the
discharge and chemical characteristics of plutonium were not similar at Oak Ridge and
Los Alamos; therefore it is not possible to conduct a rigorous intercomparison of
plutonium behavior in the different ecosystems. Althougli many environmental variables
are uncontrolled, a comparative analysis of plutonium in both ecosystems provides insight
on patterns of plutonium behavior in the environment.

371

312 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Plutonium in Oak Ridge and Los Alamos Environments

Rates of plutonium releases to the Oak Ridge and Los Alamos environs during the study
period and resulting concentrations in abiotic components are summarized in Table 1 .
The release of plutonium to the atmosphere at Oak Ridge was considerably less than that
at Los Alamos, and ambient plutonium in air reflects the higher release rate at Los
Alamos. However, these concentrations in air are not easily distinguished from plutonium
in air originating from global fallout (Bennett, 1976).

The quantity of plutonium released to surface water is similar at both sites, but the
source of plutonium entering the drainage systems is different. At Oak Ridge release from
routine operations is negligible; most of the plutonium in surface water originates from
contaminated locations in the White Oak Drainage (WOD). At Los Alamos routine release
of plutonium in treated liquid waste accounts for nearly all the plutonium entering
Mortandad Canyon (Hakonson, Johnson, and Purtymun, 1973).

TABLE 1 Plutonium in the Oak Ridge and Los Alamos Environments*!

Oak Ridge

Los Alamos

Plant

Mortandad

Laboratory

Component

Floodpiain

area

perimeter

Canyon

area

Off site

Release, mCi/yr

Atmosphere

NAt

0.004

NAt

NAJ

Surface water

NM§

2011

8.8

NM§

Air concentration, fCi/m^

0.11**

0.02

0.01

NM§

0.05

Soil concentration, pCi/g

0 to 1 cm

NM§

0.04

NM§

0 to 20 cm

10 to 150

NM§

9 to 250

0.05

0.02

Surface-water concentra-

tion, pCi/liter

ND«

NM§

0.2tt

7.7

NM§

0.12

*Total plutonium includes the 238, 239, and 240 nuclides. Prior to 1968 the plutonium released
to Mortandad Canyon was 2 3 9,24opy ^-^^q jj^at time ^^^Pu has been the dominant isotope. The

present

Pu/

2 3 9 >24 0

Pu activity ratio is 3 : 1. The predominant isotope released to the Oak Ridge

floodpiain is ^ 3' ''""Pu.

fData for Oak Ridge obtained from Union Carbide Corporation, Nuclear Division (1976), Oakes
and Shank (1977), Dahlman and McLeod (1977), and Bondietti and Sweeton (1977); for Los Alamos
from Herceg (1973), and Schiager and Apt (1974). The data matrix is obviously incomplete in terms
of measurements for certain components and in terms of estimates of errors. The values presented
represent single measurements, averages of a few measurements (accompanied by high variances), and
summations or products of several measurements. Accordingly, the inclusion of error estimates was
not considered practical. The reader is referred to original data sources where additional information
on variability is provided.

JNA, not applicable.

§NM, not measured.

H Represents total release from three plant areas to three different surface streams; value is based
on total alpha analysis but excludes uranium and thorium and includes other transuranium elements.

**Not detectable in 350-m^ air samples where the minimum detectable level is 40 fCi per sample;
air sampled in the 0- to 10-cm zone contained 0.14 fCi/m^ under ambient conditions; air samples at
1-m height during soil cultivation contained 26.3 fCi/m^ .

tfWater collected from White Oak Lake at the point of discharge, which is approximately 2 km
below the floodpiain site.

i|:$ND, not determined.

DISTRIBUTION OF PLUTONIUM IN ECOSYSTEMS 373

Plutonium in liquid effluents released from Oak Ridge National Laboratory (ORNL)
during the Manhattan Project contaminated environments in the White Oak Drainage
Basin. The Oak Ridge site, approximately 0.5 km downstream from ORNL, received
effluents containing plutonium when it served, for 6 months in 1944, as a temporary
settling basin for radioactive wastes. The impoundment drained in late 1944, and since
then a forest has developed on the floodplain.

Mortandad Canyon has received liquid waste since 1963, and from 1972 to 1973 the
release rate was approximately 9 mCi ^^^?\\jyi. Soil has become contaminated with both
^^^Pu and ^•^^Pu; concentrations range from 250 pCi/g at the waste outfall to 9 pCi/g
about 2 km down the canyon. Surface water from the outfall completely infiltrates into
the alluvium within 1.2 km of the eftluent outfall. Downstream transport of plutonium
into dry portions of the streambed occurs only during storm runoff events (Hakonson,
Nyhan, and Purtymun, 1976).

Chemical and isotopic characteristics of plutonium released from Oak Ridge and Los
Alamos are also different. The Oak Ridge method of treating radioactive liquid waste in
1944 involved coprecipitation with carbonate to remove radionuclides, primarily ^°Sr. At
Los Alamos plutonium may be associated with laundry and laboratory chelating agents
[e.g., nitrilotri (methylene phosphoric) acid-ATMP; 1 -hydroxyethyhdene 1,1 di-
phosphoric acid-HEDP] . The environmental stability of these complexes is unknown,
but plutonium is presently associated with the soil-sediment component of the canyon.

Soil and Biotic Characteristics of the Oak Ridge Floodplain

The floodplain is representative of bottomlands of the East Tennessee valley. The soil
profile is azonal because of periodic erosion and deposition of sediments related to
flooding. An accumulation of humus is evident from the dark-brown appearance of soil in
the 0- to 3-cm zone. Soil texture is a loamy clay (72% silt and 24% clay). The soil
reaction is neutral to slightly alkaline (pH = 7.1 to 7.6), which is atypical of regional
forest soils. The sliglitly alkaline condition is attributed to the alkaline coprecipitation of
wastes during the Manliattan Project.

The forest ecosystem of the floodplain occupies a 3-ha area. The present successional
stage of the forest is dominated by sycamore (Platanus occidentalis L.) and white ash
{Fraximis americana L.). Ground vegetation is chiefly wild rye grass (Elvmus virginicus
L.), Microstegium vimineum [(Trinus) A. Canus] , jewelweed (Impatiens capensis Meerb.j,
and Japanese honeysuckle {Lonicera japonica Thunberg). The resident small-mammal
population includes the white-footed mouse (Peromyscus leucopus), the rice rat
(Oryzomys palustris), and the short-tailed shrew (Blarina brevicaiida). Earthworms
( Lumbricus rubellus) and crayfish {Cambams sp.) are important soil invertebrates.

Biomass [grams (dry weight)] of the major compartments of the floodplain
ecosystem is given in Table 2. Biomass for arborescent species was estimated from
mensuration data (Van Voris and Dalilman, 1976) and from regression equations (Harris,
Goldstein, and Henderson, 1973). The arborescent component (leaf, root, and wood)
contains 95% of the total forest biomass, whereas animals comprise only 0.02%. These
estimated biomass standing crops compare favorably witli average values compiled for
different forest stands of the eastern deciduous forest biome (EDFB) (Burgess and
O'Neill, 1975) with the exception of litter (550 g/m), which was considerably less than
the average for EDFB (~2000 g/m^). The low estimate for Utter was probably due to (1)
the young age of the floodplain forest (approximately 30 yr); (2) the effect of periodic

374 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Estimates of Mass of the Major

Components of the Oak Ridge Floodplain and the

Los Alamos Mortandad Canyon

Oak Ridge

Los Alamos

■t

floodplain

Mortandad Canyon

Component

(g/m^)

(g/mM

Soil*

2.6 X 10'

2.4 X 10=

Tree wood

10,500t

15,000$

Tree root

3,000t

Tree leaf

400t

600 1

Litter

550§

1,000

Ground vegetation

110§

Not determined

Grass

Not determined

Forb

Not determined

Soil fauna

3§

Not determined

Consumer

0.03 §

0.07

*Mass is based on a profile depth of 20 cm and on
densities of 1.3 and 1.2 g/cm^ for Oak Ridge and Los
Alamos, respectively.

t Estimated from mensuration data and regression
equations (Harris, Goldstein, and Henderson, 1973).

^Estimated from mensuration data and regression
equations (Wheeler, Smith, and Gallegos, 1977).

§ Estimated from field measurements of populations and
biomass.

floods on litter accumulation; and (3) moist-mesic conditions, which favor rapid
decomposition.

Soil and Biotic Characteristics of Mortandad Canyon

The canyon soil consists of an alluvial deposit (<30 cm) derived from volcanic tuff. The
coarse soil is less than 3% by weight silt and clay (Nyhan, Miera, and Peters, 1976a).
Cation-exchange capacity is low (2 to 20 meq per 100 g) (Schiager and Apt, 1974).
Organic matter ranges from 0.1 to 0.2%. Calcium as high as 3.7% and soil pH up to 9.2
are measurable (Scliiager and Apt, 1974), which reflects the carbonate contribution from
liquid waste. Uncontaminated soil from the canyon floor has a pH of 5.7.

The dominant arborescent species of the canyon are ponderosa pine (Pinus
ponderosa), Douglas fir (Pseudotsuga menzeziij, and Gambels oak (Quercus gambelii)
(Miera et al., 1977). Dominant forb and grass species are wheat grass (Elymus sp.),
bluegrass (Poa pratensis), wild strawberry (Fragaria americana), and dandelion (Taraxa-
cum officinale). The most common mammal residents are pinon mouse (Peromyscus
tniei), deer mouse (P. maniculatus), and the least chipmunk (Eutamias minimus).

Standing-crop biomass [grams (dry weight)] for arborescent, herbaceous, and animal
components of the floodplain and canyon ecosystems is shown in Table 2. Arborescent
species contribute more than 80% of the total mass at the site.

Characteristics of Plutonium in the Floodplain Soil

Concentration of plutonium in the floodplain soil ranges from about 10 to 150 pCi/g over
a 3-ha area (Fig. 1). The highest concentrations were found behind the former dike, along

DISTRIBUTION OF PLUTONIUM IN ECOSYSTEMS 375

WHITE

OAK

CREEK

S30

S60

S90

PPZ^ 150 pCi/g
^^ 100 pCi/g
^3 25 pCi/g

[ ■ .] 10 pCi/g

S300

Fig. 1 Approximation of areal distribution of plutonium at the contaminated Oak
Ridge floodplain. Grid size is 30 by 30 m for an area of 900 m^ . Distribution is generally
estimated from 0- to 10-cm soil samples plus a few samples from the 10- to 20-cm depth.

White Oak Creek, and in the upper part of the floodplain. The maximum concentration
of plutonium occurred where the creek is beheved to have entered the historic
impoundment between coordinates E60-S120 and E120-S30 (Fig. 1). Sediment-borne
plutonium probably settled from the water column as the stream velocity decreased on
entering the impoundment. Downstream from the site of initial deposition the higher
plutonium concentrations tended to follow the watercourse of White Oak Creek (WOC).

Plutonium is not distributed uniformly between the loam and clay fractions of the
soil. The loam fraction (>0.002 mm) contains 60% of the plutonium; whereas 24% of the
<0.002-mm size class contains 40% of the plutonium. The high affinity of plutonium for
colloids may be responsible for plutonium enrichment in clay. The distribution
coefficient (K^), determined by desorption in 0.01MNaHCO3,is 5 X 10^ (Bondietti and
Tamura, this volume).

The highest concentration of plutonium is observed in the 0- to 10-cm zone of the
soil profile. Occasionally a higher concentration of plutonium occurs below 10 cm. This
atypical distribution is attributed to deposition of sediments over the initial plutonium
deposit rather than to movement of the element by biogeochemical processes.

516 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Soil Plutonium Characteristics in the Canyon Ecosystem

Concentrations of plutonium in alluvial soil from the canyon are strongly a function of
distance from the waste-effluent outfall (Hakonson and Bostick, 1975; Nyhan, Miera, and
Peters, 1976b). Highest concentrations, averaging about 250 pCi/g, occur near the outfall,
whereas successively lower concentrations are measured with increasing distance
downstream. Downstream transport of waste has occurred to about 3500 m below the
outfall where fallout levels of plutonium are measured. Stream-bank soils to 1 m either
side of the channel are contaminated to the same level as adjacent channel soils.

Plutonium is rather uniformly distributed within the canyon soil profile to depths of
30 cm, wliich reflects the effect of turbulent mixing processes during storm runoff events.
The mixing of plutonium with depth has been rapid (<years) as inferred from the
distribution of ^^^Pu with depth (Hakonson and Bostick, 1975).

Over 85% of the plutonium in canyon alluvium is associated with sand particles of
from 0.05 to 23 mm in diameter. About 14% of the plutonium inventory is present in the
silt— clay fraction (<0.05 mm), which comprises only 2% of the soil mass (Nyhan, Miera,
and Peters, 1976a). Significant correlations with solid-phase constituents (i.e., carbonates
and the colloidal exchange complex) suggest that plutonium is sorbed to particles and
minerals.

Concentration Ratios

The availability of soil plutonium to biota can be inferred from the concentration ratios
(CR's) in Table 3. The low CR's for biota at both sites demonstrate the small fraction of
plutonium that has moved from soil to biota. Vegetative components (root, litter, and
herbaceous species) most intimately associated with the soil exliibited the highest CR's at
both sites. Arborescent components generally exhibited the lowest CR's. The fact that all
CR's are less than 1 confirms that there is no evidence of biomagnification of plutonium
in these terrestrial ecosystems two to three decades after the environments were
contaminated.

TABLE 3 Concentration and Inventory of Plutonium in Major Components of the Oak Ridge
Floodplain and the Los Alamos Mortandad Canyon

Oak

Ridge*

Los Alamost

Component

pCi Pu/g

pCi Pu/m'

CRt

IR§

pCi Pu/gf

pCi Pu/m'

CRJ

IR§

Soil

1.6 X 10'

0.999

1.2 X 10'

0.997

Tree wood

0.003

5 X 10~'

2 X 10"'

0.05

7.5 X 10'

1 X 10-'

6 X 10-'

Tree root

3.6

1.1 X 10*

6 X 10-'

7 X 10"*

Tree leaf

0.003

1.2

5 X 10"'

8 X 10-'

0.05

1 X 10-'

3 X 10-'

Litter

6.0

3.3 X 10'

1 X 10"'

2 X 10"*

3.2 X 10*

6x 10-"

3 X 10"'

Soil fauna

1.0

3.0

2 X 10"'

Ground vegetation

0.05

5.5

8x10-*

3x 10-'

Grass

4.8 X 10'

1 X 10°

4 X 10-'

Forb

4.0

1 X 10'

8 X IQ-'

8x 10-'

Consumer

0.04

0.001

6 X 10-*

6x 10-"

0.27

0.02

5 X 10-'

2 X 10-'

*Concentrations and inventories averaged for the 3-ha floodplain.

t Concentrations and inventories based on the 0- to 1500-m segment of Mortandad Canyon.
JCR = [PuJcomponent/lPulsoil-

§1R = [Pu] receptor/(P") source- The IR's for all components except soil are based on soil as the plutonium source.
The soil IR represents the fraction of total plutonium of the ecosystem that is present in soil,
fl Weighted average for intensive study sites 1 and II.

DISTRIBUTION OF PLUTONIUM IN ECOSYSTEMS 377

Plutonium Inventories in the Floodplain and Canyon Ecosystems

Total plutonium present in different ecosystem components is referred to as a static
inventory or budget. For each ecosystem the inventory was calculated by multiplying the
mass values of Table 2 and the plutonium concentrations (picocuries per gram) in the
respective components (Table 3). Results on areal distribution of soil plutonium (Fig. 1)
were integrated to provide an estimate of the plutonium inventory for the entire 3-ha
floodplain. Total soil plutonium inventory in the top 20 cm, based on the summation of
subinventories of four different concentration zones, was 0.5 Ci (8 g). The 100- and
150-pCi/g zones contained 88% of the plutonium, but they occupied only 46% of the
area of the 3-ha floodplain*. Because only 50% of the soil determinations included both
the topsoil (0 to 10 cm) and subsoil (10 to 20 cm) and because the zones of
concentration indicated by the isopleths in Fig. 1 are interpolated between sampling
stations, the soil inventory is provisional and does not represent high resolution of
plutonium distribution in the floodplain.

The inventory of plutonium for the 0- to 20-cm soil depth in Mortandad Canyon was
calculated for the 3 X 1500-m segment (intensive study sites I and II) of stream channel
immediately below the effluent outfall. Nearly all the plutonium inventory is present in
this segment of the canyon (Hakonson and Bostick, 1975). The estimated plutonium
inventory was 0.054 Ci as of 1974, which corresponded closely to the estimated input of
0.051 Ci based on treatment-plant release records.

A summation of component inventories shows that, by far, the majority of the
plutonium resides in soil. More than 30 yr after the initial deposit, less than 0.1% of the
total plutonium is present in living components of the Oak Ridge floodplain. For
Mortandad Canyon, the inventory of plutonium in biotic components of the canyon
ecosystem was 0.00015 Ci in the 0- to 1500-m segment, or 0.28% of the total plutonium
present in the canyon ecosystem.

Inventory Ratios for Plutonium in the Floodplain and Canyon Ecosystems

One approach used to understand the significance of relative distributions of plutonium
in ecosystems is to relate the plutonium inventory of the receptor to the plutonium
content of the donor or source component in terms of inventory ratios (IR's):

P _ Activity/unit area in receptor
Activity/unit area in source

Contaminated soil is the major source of plutonium for all biotic components of both the
floodplain and canyon ecosystems; consequently soil has been used as the denominator to
calculate IR's (Table 3). Components other than soil may contribute plutonium to certain
other components, e.g., wood is a likely source of plutonium in tree leaves. The reader
may choose data from Table 3 to derive other IR's not provided by our analysis.

The IR results (Table 3) demonstrate the importance of soils as the reservoir for
environmental plutonium. Over 99% of the plutonium of both study sites was associated
with soils. Biota serve as an incidental, although potentially important, receptor for
plutonium in the environment, but only a small fraction of the total plutonium is present
in the biotic components. Root and Utter components of the floodplain had an IR of
approximately lO""^ compared with a value of 10~^ for litter at Los Alamos. Inventory
ratios for tree and aboveground vegetative components ranged from 10~^ to 10~^ on the

318 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

floodplain. Inventory ratios for grass, forb, and tree components of the canyon ecosystem
were about 10~^ . The higher IR's of the canyon biota reflect both higher concentrations
of plutonium and greater standing crops of biomass.

Discussion

There are Hmitations in the use of CR and IR data to describe the distribution of
plutonium among components of the ecosystem. There is the difficulty of distinguishing
between contributions resulting from several simultaneous transfers to a receptor.
Sometimes the relationships between sources and receptors are not clear. Another
limitation is that IR's and CR's are static indexes of plutonium distribution in the
environment. Both indexes are based on an assumption of equilibrium conditions for the
ecosystem; i.e., that biomass and concentrations are constant. The equilibrium assump-
tion is not strictly met in terrestrial communities that are dynamic. In spite of these
limitations, both CR and IR values can give order-of-magnitude estimates on the
availabiUty of plutonium in the environment.

More than 30 yr after deposition in floodplain sediments, plutonium in the Oak Ridge
soil appears to be a monomeric species associated with endemic organic and mineral
constituents (Bondietti, Reynolds, and Shanks, 1976). There have been no recent
amendments of plutonium to the floodplain ecosystem. In contrast, there are annual
additions of plutonium in the waste effluents released to Mortandad Canyon. Tliese
effluents contain diverse industrial wastes, including chelating agents; consequently
plutonium in the canyon environment can be complexed by chelators. Enlianced mobiUty
of plutonium would be expected in the canyon ecosystem if a chelated form of
plutonium were a stable and dominant species in the environment. This factor could
increase the biochemical assimilation of plutonium by plants and animals; thus, along
with surface contamination, it may be partly responsible for the higlier CR's and IR's
observed for the canyon ecosystem.

Plutonium uptake by the root pathway yields a plant : soil concentration ratio of
about 10"^ for floodplain species (Table 3). Tliis ratio is about one to two orders of
magnitude greater than CR's determined from short-term experiments when plants are
grown in soil contaminated with plutonium solutions (Daltlman, Bondietti, and Eyman,
1976). Root uptake is the main mechanism of plutonium incorporation by plants in the
floodplain ecosystem because the negligible contribution of plutonium to the Oak Ridge
atmosphere from resuspension and industrial release would create minimal contamination
of vegetative surfaces. As mentioned previously, external contamination of vegetation is
considered an important mechanism of uptake at the canyon.

The fraction of physiologically available plutonium is largely determined by
environmental chemistry and reactions of plutonium with soil. The great affinity of
plutonium for soil particles results in distribution coefficients of the order of 10^ to 10^ .
Sorption of plutonium to colloids is a surface reaction that occurs predominantly with
the clay constituents of soil because this component possesses the greatest specific surface
area. Enrichment of plutonium in the clay fraction has been observed in several
contaminated environments that contain appreciable clay. The differences in percentage
clay in canyon and floodplain soil (2% vs. 24%, respectively) may be responsible for
diminished sorption of plutonium to canyon soil, and this could account for the increased
incorporation of plutonium into biotic components of the canyon ecosystem. Althougli
plutonium enrichment in soil clay apparently occurs at both sites, and, althougli

DIS TR IB UTION OF PL UTONIUM IN ECOS Y STEMS 379

incorporation of plutonium by biota of the canyon appears inversely related to clay
content, the minimum quantity of clay in soil that is required to sorb plutonium and
restrict its movement to biota is unknown.

Comparative studies of the-biogeochemical behavior of plutonium in ecosystems can
facilitate the application of plutonium data to assessments of future environmental
impact. Relative distributions and concentrations of plutonium in components of two
different ecosystems confirm that the element is not readily incorporated by biota after it
has been in the terrestrial environment for 20 to 30 yr. However, in the absence of data
over many decades, it is difficult to forecast with certainty the future biological
availability of this element. Yet the small inventory in biota and the absence of any
evidence of biomagnification indicate limited environmental mobility of the element.

Currently available indexes of mobility in forest ecosystems 20 to 30 yr after initial
contamination suggest that the properties of plutonium have not been modified in a way
that would affect its long-term biogeochemical behavior.

Acknowledgments

Research for the ORNL study was sponsored by the U. S. Department of Energy under
contract with Union Carbide Corporation, Publication No. 1347. The LASL study was
funded under contract No. W-7405-Eng.36 between the U. S. Department of Energy and
LASL.

References

Bennett, B. G., 1976, Transuranic Element Pathways to Man, in Transuranium Elements in the
Environment, Symposium Proceedings, San Francisco, 1975, pp. 367-383, STI/PUB/410,
International Atomic Energy Agency, Vienna.

Bondietti, E. A., and 1'. H. Sweeton, 1977, Transuranic Speciation in the Environment, in
Transuranics in Natural Environments, Symposium Proceedings, Gatlinburg, Tenn., October 1976,
M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178, pp. 449-476, Nevada Operations
Office, NTIS.

, S. A. Reynolds, and M. A. Shanks, 1976, Interaction of Plutonium with Complexing Substances

in Soils and Natural Waters, in Transuranium Elements in the Environment, Symposium
Proceedings, San Irancisco, 1975, pp. 273-287, STI/PUB/410, International Atomic Energy
Agency, Vienna.

Burgess, R. L., and R. V. O'Neill (Eds.), 1975, Eastern Deciduous Forest Biome Progress Report,
Sept. I-Aug. 31, 1974, Odk Ridge National Laboratory, NTIS.

Dahlman, R. C, and K. W. McLeod, 1977, loliar and Root Pathways of Plutonium Contamination of
Vegetation, in Transuranics in Natural Environments. Symposium Proceedings, Gatlinburg, Tenn.,
October 1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178, pp. 303-320,
Nevada Operations Office, NTIS.

, E. A. Bondietti, and L. D. Eyman, 1976, Biological Pathways and Chemical Behavior of

Plutonium and Other Activities in the Environment, in .Actinides in the Environment. A. M.
I'riedman (Ed.), ACS Symposium Series, No. 35, pp. 47-80, American Chemical Society.

Hakonson, T. E., and K. V. Bostick, 1975, Cesium-137 and Plutonium in Liquid Waste Discharge
Areas at Los Alamos, in Radioecology and Energy Resources, Proceedings o{' the Symposium on
Radioecology, Oregon State Univ., May 13-14, 1975, The Ecological Society of America, Special
Publications No. 1 , C. E. Gushing, Jr. (Ed.), pp. 40-48, Halsted Press, New York.

"~~', L. J. Johnson, and W. D. Purtymun, 1973, The Distribution of Plutonium in Liquid Waste
Disposal Areas at Los Alamos, in Proceedings of the Third International Radiation Protection
Association Meeting. Washington, D. C, Sept. 9-14, 1973, W. S. Snyder (Ed.), USAEC Report
CONF-730901 . pp. 248-252. International Radiation Protection Association, NTIS.

380 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

, J. W. Nyhan, and W. D. Purtymun, 1976, Accumulation and Transport of Soil Plutonium in

Liquid Waste Discharge Areas at Los Alamos, in Transuranium Nuclides in the Environment,

Symposium Proceedings, San Francisco, 1975, pp. 175-189, STI/PUB/410, International Atomic

Energy Agency, Vienna.
^, J. W. Nyhan, L. J. Johnson, and K. O. Bostick, 1973, Ecological Investigation of Radioactive

Materials in Waste Discharge Areas at Los Alamos, USAEC Report LA-5282-MS, Los Alamos

Scientific Laboratory, NTIS.
Hanson, Wayne C, 1975, Ecological Considerations of the Behavior of Plutonium in the Environment,

Health Phys., 28: 529-537.
Harris, W. F., R. A. Goldstein, and G. S. Henderson, 1973, Analysis of Forest Biomass Pools, Annual

Primary Production and Turnover of Biomass for a Mixed Deciduous Forest Watershed, in lUFRO

Biomass Studies, Mensuration, Growth and Yield, pp. 41-64, H. Young (Ed.), Univ. of Maine Press,

Orono.
Herceg, Joseph E. (Comp,), 1973, Environmental Monitoring in the Vicinity of Los Alamos Scientific

Laboratory, Calendar Year 1972, USAEC Report LA-5184, Los Alamos Scientific Laboratory,

NTIS.
Miera, F. R., Jr., K. V. Bostick, T. E. Hakonson, and J. W. Nyhan, 1977, Biotic Survey of Los Alamos

Radioactive Liquid-Effluent Receiving Areas, ERDA Report LA-6503-MS, Los Alamos Scientific

Laboratory, NTIS.
Nyhan, J. W., F. R. Miera, Jr., and R. J. Peters, 1976a, Distribution of Plutonium in Soil Particle Size

Fractions of Liquid Effluent-Receiving Area at Los Alamos, 7. Environ. QuaL, 5(1): 50-59.
, F. R. Miera, Jr., and R. J. Peters, 1976b, The Distribution of Plutonium and Cesium in Alluvial

Soils of the Los Alamos Environs, in Radioecology and Energy Resources. Proceedings of the

Symposium on Radioecology, Oregon State Univ., May 13-14, 1976, The Ecological Society of

America, Special Publications No. 1, C. E. Cashing, Jr. (Ed.), pp. 49-57, Halsted Press, New York.
Oakes, T. W., and K. E. Shank, 1977, Surface Investigations of Energy Systems Research Laboratory

Site at Oak Ridge National Laboratory, ERDA Report ORNL-TM-5695, Oak Ridge National

Laboratory, NTIS.
Schiager, Keith J., and Kenneth E. Apt (Comps.), 1974, Environmental Surveillance at Los Alamos

During 1973, USAEC Report LA-5586, Los Alamos Scientific Laboratory, NTIS.
Union Carbide Corporation, Nuclear Division, 1976, Environmental Monitoring Report, U. S. Energy

Research and Development Administration, Oak Ridge Facilities, Calendar Year 1975, ERDA

Report Y/UB-4, Union Carbide Corporation, NTIS.
Van Voris, P., and R. C. Dahlman, 1976, Floodplain Data: Ecosystem Characteristics and '^''C?

Concentrations in Biota and Soil, USAEC Report ORNL-TM-5526, Oak Ridge National

Laboratory, NTIS.
Wheeler, M. L., W. J. Smith, and A. F. Gallegos, 1977, A PreUminary Evaluation of the Potential for

Plutonium Release from Burial Grounds at Los Alamos Scientific Laboratory, ERDA Report

LA-6694-MS, Los Alamos Scientific Laboratory, NTIS.

Plutonium Contents of Field Crops
in the Southeastern United States

D. C. ADRIANO, J. C. COREY, and R. C. DAHLMAN

Agricultural crops were grown at the U. S. Department of Energy Savannah River Plant
(SRP) and at Oak Ridge National Laboratory (ORNL) on soils at field sites containing
Plutonium concentrations above background levels from nuclear weapon tests. Major
U. S. grain crops were grown adjacent to a reprocessing faciUty at SRP, which releases low
chronic levels of plutonium through an emission stack. Major vegetable crops were grown
at the ORNL White Oak Creek floodplain, which received plutonium effluent wastes in
1944 from the Manliattan Project weapon development.

The plutonium contents of grain crops (wheat, soybeans, and corn) at SRP were
affected by distance from the emission stack, plant height, and grain-processing method.
In general, vegetative materials growing close to the stack liad higher plutonium
concentrations than those growing in an adjacent field. Plutonium concentrations of
portions of plants, such as wheat and corn, collected highest from the ground level
indicate tliat plutonium contamination of these plant parts from soil resuspendible matter
was minimal. The plutonium content of the grain when harvested by combine was
elevated because the grain was mixed with extraneous matter and straw, which had
relatively higher plutonium concentrations. Results from glasshouse studies using the
same field-grown crops indicate tliat root uptake contributed insignificantly to the total
plutonium contents of the field-grown crops.

Plutonium contents of vegetable crops grown at the ORNL White Oak Creek
floodplain were influenced by part of plant, stage of maturity, and method of processing
for the edible portions of the subterranean crops. Plutonium concentrations of fruits were
at least one order of magnitude lower than those of the foliage. The plutonium content of
the vegetable foliage was maximum when the foliage biomass was at maximum. Peeling
the skins from potatoes and beets removed approximately 99% of the residual plutonium.

In general, the concentration ratios of vegetative parts of crops at SRP were
approximately one order of magnitude higher than those at ORNL, which indicates the
influence of aerial deposition of plutonium at the SRP site.

Research on transuranic nuclides in the environment has gained momentum in recent
years as a result of proposed increases in production and use of plutonium in the nuclear
fuel cycle. Because of the toxicity and long half-life of plutonium (specifically
2 3 9,2 4 0p|j^^ its health hazard to man is being evaluated. Although inhalation has been
considered the major pathway by which plutonium reaches man (Bennett, 1976),
ingestion of plutonium-contaminated foodstuff through the soil-to-plant pathway should
be critically evaluated because of the long persistence and general immobility of
plutonium in the environment (Francis, 1973). Numerous studies have been conducted to

3S1

382 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

determine plant uptake of plutonium under both field and controlled conditions (Francis,
1973; Price, 1972; Hanson, 1975; Bennett, 1976), However, most of these experiments
were done in pot culture; experiments done in the field involved only nonagricultural
vegetation. This general lack of information on field-grown crops is apparent from the
proceedings of the 1976 international symposium on the Transuranium Nuclides in the
Environment {InXQxmiiondX Atomic Energy Agency, 1976).

This chapter describes plutonium contents of crops grown on fields at the U. S.
Department of Energy Savannah River Plant (SRP) and at Oak Ridge National
Laboratory (ORNL) in soils containing plutonium concentrations at levels above those
attributable to fallout from nuclear weapon tests. The fields at SRP, near Aiken, S. C,
have been receiving plutonium at low chronic levels from the emission stack of a
reprocessing facility since 1955; the field plots at ORNL are located on the White Oak
Creek floodplain, which received plutonium from Manhattan Project operations in 1944.
We compared the plutonium contents of major grain crops (wheat, soybeans, and corn) at
SRP, where the major mode of contamination is through deposition from the emission
stack, with those of major vegetable crops at Oak Ridge, where the major pathway of
contamination is via root uptake.

Description and History of the Study Sites

Savannah River Plant

The SRP is on a reservation of 77,830 ha. Public access to the reservation has been
controlled since its acquisition in 1951. The reservation consists of freshwater streams,
old fields, and forest; most of the old fields are in the upper Coastal Terraces.

For over 20 yr this integrated nuclear complex has included nuclear reactors (three of
the original five are operating at present), two nuclear-fuel reprocessing plants, a fuel
fabrication facihty, a heavy-water production unit, and nuclear and environmental
research laboratories (Fig. 1). It also includes an ecological research laboratory to assess
the effects of nuclear technology on the environment and its biota. The reprocessing
plants (F and H) and global fallout are the sources of the transuranic elements that enter
the SRP environs. Each source releases plutonium of unique isotopic composition: 25 and
95 a % ^^^Pu.* from the F- and H-area reprocessing facilities, respectively, and 10 a %
^^^Pu from global fallout. These isotopic differences provide a convenient basis for
studying the origin and transport of plutonium in various SRP ecosystems.

In 1974 two crop fields were estabhshed adjacent to the H-area reprocessing facility.
Low-level atmospheric releases of plutonium have occurred from H area since the start of
operations in July 1955. Approximately 440 mCi of plutonium had been released from H
area before the installation of high-efficiency filters on the exhaust-air systems in
December 1955. These releases contained 2 3 9,24 0p^ From 1956 through 1966,
2 3 9,2 4 0pjj re[easgs averaged 4 mCi/yr. From 1967 through 1974, normal releases
averaged 12 mCi ^^^Pu/yr and 4 mCi ^^^'^'^'^Pu/yr. An accidental failure of the filtering
mechanism in 1969 released an additional 560 mCi of ^^^Pu and 58 mCi of "^'^"^^Pu.
Total releases through 1974 were 640 mCi of " ^Pu and 570 mCi of "^ '^"^^Pu.

* ~238t, ^^*Pu alpha activity ,^„

*a % ^^*Pu = --; r-^- — — I -^-. — X 100.

total plutonium alpha activity

PLUTONIUM CONTENTS OF FIELD CROPS 383

^s. c

—7

\
Columbia

-^^'^

\^ Augusta /______

"Tr^

r^^^^^^v^

^^X^fy ^

J-^^

-<;

V\i^ ]

\\ /

i^lv

/♦ ^"^"^'v/l

GEORGIA \/ly

0 5 \\

1 1 v>.

SOUTH k\
CAROLINA I

km ^\,

F]
Hj

C I

> - Reprocessing Plants
■ - Nuclear Reactors

^4°"

D - Heavy Water Plant

M
A

- Fuel Fabrication Plant

- Administration and
Research Facilities

SAVANNAH RIVER
PLANT

Fig. 1 Lxjcations of the two reprocessing plants (H and F) at the Savannah River Plant.
The agricultural study area is near the H area.

The South Field was 145 m by 30 m, and its long axis was oriented to the northwest
away from the point of release, a 62-m stack. The North Field shared the same long axis
as the South Field but was smaller ( 105 m by 30 m). The centers of the South and North
fields were approximately 230 and 420 m, respectively, from the stack. The A soil
horizon and parts of the B soil horizon had been removed from the South Field during
construction of the reprocessing facility. In addition, some fill dirt had been deposited on
the South Field. The North Field had been disturbed less and had a soil profile that was
typical for the area. Before being tilled, both fields supported a herbaceous plant
community dominated by Andropogon spp., Lespedeza cuneata (Dumont) G. Don,
Panicum spp., and Smilax spp. with scattered loblolly pines {Pinus taeda L.). The plant

384 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

communities were typical of those occurring on abandoned fields of similar soil types in
the southeastern United States. A 5-m-wide strip along the southwestern margin of each
field has been mowed regularly for several years. The soil in both fields is classified as
Vaucluse, a liighly leached Ultisol characterized by sandy A horizon and sandy clay B
horizon. Both fields were acidic with a similar pH (~4.6).

The climate of the SRP area consists of mild winters and long, warm, humid summers.
Temperatures average about 9°C in the winter and 30°C in the summer. The average
annual temperature is 18°C. The average annual relative humidity and rainfall are 70%
and 120 cm, respectively. The maximum annual precipitation recorded was 187 cm in
1929, and the minimum was 71 cm in 1933 (Langley and Marter, 1973).

White Oak Creek Floodplain, Oak Ridge

Manhattan Project operations in 1944 produced treated wastes containing plutonium,
americium, and curium. Following soda-ash treatment to precipitate various ions, the
liquid effluents were released to White Oak Creek (WOC). Trace quantities of ^^^Pu,
"^^^Pu, ^^' Am, and ■^'*'*Cm were deposited along the water course, in an intermediate
retention pond, and, finally, in White Oak Lake. These elements were deposited in
sediments of a retention pond over a 6-month period in 1944. No additional radioactivity
is believed to have been deposited at this location since the pond was drained in late
1944. The former retention-pond site currently constitutes the first floodplain terrace of
WOC.

The alluvial soil of the floodplain is representative of bottomlands of the Tennessee
Valley and Ridge province. Azonal characteristics predominate because of concurrent
erosion and deposition of materials during periodic floods. The soil profile remains
relatively undeveloped, although the accumulation of humus is evident from a
dark-brown coloration in the 0- to 3-cm zone.

Sedimentation during the 6-month impoundment in the temporary holding basin 30
yr ago contributed new sedimentary materials. A Tennessee Valley Authority (1975)
survey in 1951 reported that approximately 2100 m"^ of sediment had been deposited at
this locality. This volume, distributed uniformly over the research site, represented an
increment of approximately 9 cm of new sediment. However, equal deposition was
unlikely, and the exact history of sedimentation associated with the 1944 impoundment
is unknown.

Textural analyses indicate that soils are silty loam (72% silt and 24% clay) and
contain almost no sand or gravel (Tamura, 1976). Although this texture is representative
of the floodplain, isolated gravel lenses occur irregularly across the floodplain and within
the soil profile.

The soil reaction is mildly alkaline since pH ranges from 7.1 to 7.6. The mild
alkalinity is probably caused by the soda-ash method of waste treatment used before
effluents were released from laboratory operations. Data have not been obtained on base
exchange or soil fertility, althougli, on the basis of plant growth and crop performance,
the site possesses high fertility potential.

The climate at Oak Ridge, Tenn., is characterized by mild winters and warm, humid
summers. Average temperatures of the continental climate are 24°C in summer and

PLUTONIUM CONTENTS OF FIELD CROPS 385

5°C in winter; average precipitation is 140 cm, with 10 cm of snowfall. Winter and
summer temperature variations tend to be greater at Oak Ridge than at SRP.

Methods

Savannah River Plant

Pine trees were bulldozed from the old fields and bushes and herbaceous vegetation were
cut with a tractor-drawn rotar>' mower to prepare the land for crops. All debris, including
tree roots, was removed before any operation. The fields were disked with a bush and
bog disk harrow, then with a standard disk harrow, subsoiled, redisked, limed, redisked,
fertilized, and, finally, redisked. The disking was done to a depth of about 20 cm.
Agricultural Hme (consisting of 60% CaCOg, 25% MgCOg, and 15% H.O by weight) was
added at tlie rates of 908 kg on the South Field and 590 kg on the North Field.
Mixed-grade fertihzers were added at rates of 318 kg of 3-9-18* and 227 kg of
5—10—15 for the South and North fields, respectively. The South Field was divided into
18 equal plots, and the North Field, into 12 equal plots.

Wheat (variety Coker 68-19) was sown on the fields by hand in November 1974 at the
rate of approximately 53 kg of seed per field. The seeds were then covered by disking the
fields at a very shallow depth. Foliage samples were collected in March and April. At
harvest in June, plants were separated into grain and straw. Wheat-grain samples were
obtained by two techniques. The first technique was to carefully collect wheat heads by
hand from each of the 30 plots and separate the grain with a laboratory thrashing
machine. This machine was carefully cleaned between samples from each plot to
minimize the amount o{ dust that would adhere to the grain. The second technique was
to harvest the grain with a tractor-pulled combine. Grain collected by this method was
exposed to the usual soil and dust stirred up by harvest activities. Total biomass for the
crop (grain plus hull plus straw) was estimated at 3615 and 3545 kg/ha for the South and
North fields, respectively. Wlieat was also grown 37 km from the reprocessing plants as a
control.

The fields were prepared for soybeans by disking them several times and adding
mixed-grade fertilizer (3—9—18) at the rate of 227 kg for each field. Inoculated soybeans
(variety Bragg) were planted in July 1975 with a two-row planter; about 27 kg of seeds
was used for the two fields. The crop was harvested with a combine in November 1975.

The fertilizer-addition and field-preparation techniques for the corn crop were similar
to those for soybeans. Field corn (variety Coker) was planted in May 1976 with a
two-row planter and harvested in October with a combine.

The chronic releases of plutonium-bearing particles at low levels from the emission
stack make it impossible to determine the plutonium uptake by the crops from the soil
through the root pathway. Whatever amount of plutonium is translocated to the plant
foliage from the soil would be obscured by external deposition and retention from stack
fallout. Therefore glasshouse studies were conducted to determine the amount of
plutonium translocated to the aerial portions of the plants from the soil.

'3-9-18 refers to 3% elemental nitrogen, 97o P^O. , and 1 87r K^O, respectively.

386 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Soils used in the glasshouse studies were obtained from the surface layer (0 to 20 cm)
of the South Field, where crops were grown. The soil had an average total plutonium
content of 1.96 pCi/g (dry weight) with 21 a % ^^^Pu. The soils (~8 kg) were potted and
fertilized, and the same varieties of wheat, soybeans, and corn were grown to maturity.

In the selection of locations for sampling soil and resuspendible particles, 18 grid
points were located in the South Field and 12 in the North Field. Grids were composed
of samipling blocks placed 3, 15, and 27 m from the southwestern margin of the fields on
transects across the short axis of the fields. The transects originated at 30.4-m intervals on
the long axis of each field. Each sampling block was 3 m by 10 m and contained ten 1- by
3-m plots. A randomly chosen plot was permanently marked in each block. Thus there
were 18 sampling locations in the South Field and 12 in the North Field.

Soil cores of 3.8-cm diameter were taken with a split-barrel sampler and divided into
0- to 5-cm, 5- to 15-cm, and 15- to 30-cm fractions for plutonium analysis. For
unexplained reasons plutonium concentrations in the 0- to 5-cm samples collected before
tillage were 50% lower than in samples collected after tillage; therefore these samples
were replaced with samples that had been collected at the same depth before tillage with
a hand soil auger.

The resuspendible particles on the soil surface from areas where aboveground
vegetation had been removed were collected by drawing a nearly laminar flow of air
(velocity = 6 m/sec) across a 232-cm^ area under a 1 -cm-tall stainless-steel hood. The
resuspended materials were collected in the paper bag of a small electric vacuum cleaner.
The interior of the plastic compartment holding the paper bag was wiped to collect
materials passing through the bag, and these materials were included in the sample.

Samples of soil and of resuspendible particles were collected at three different times:
(1) before tillage, (2) after soil preparation for wheat, and (3) after wheat harvest.

All samples were ashed before deterniination of plutonium contents. Plant, grain, and
resuspension samples were ashed at 550°C. Soil samples were ground to a particle size of
<500 Mm and ashed at 500°C.

Actinide elements were leached from a <10-g aliquot of the sample ash with hot 8M
HCl, and valences were adjusted to ensure formation of Pu(IV) and Np(lV). Plutonium,
neptunium, and uranium were extracted into 10% triisooctylamine in xylene, and the
plutonium was separated from the neptunium and uranium by reducing Pu(IV) to Pu(III)
with NH4I and extracting into 8Af HCl. This solution was evaporated to dryness and
oxidized to destroy residual organic matter. The plutonium was taken up in 8M HNO3,
and the valence was adjusted to Pu(IV). Final purification was accomplished by adsorbing
the Pu(IV) onto an anion-exchange column and removing any residual iron, uranium, or
other contaminants from the column with 4MHNO3. The Pu(IV) was reduced to Pu(lII)
and eluted from the column with H2SO3. Following purification, the plutonium was
electrodeposited on platinum plates, and the amounts of '^"'^Pu and 239,240p^ ^^^^
determined by alpha spectrometry with low-background high-resolution surface-barrier
detectors. Counting times ranged from 2 to 7 days, depending on the plutonium
concentrafion of the sample. An internal standard of ^^^Pu was used to determine
recovery efficiencies.

After the concentrations of ^^^Pu and 23 9,24 0pjj j^^ ^ sample had been determined,
the rafio of the ^^^Pu concentration to the 2 3 9,24 0pu concentrafion was computed.

PLUTONIUM CONTENTS OF FIELD CROPS 387

This ratio was used to evaluate the relative importance of different pathways of
Plutonium movement.

White Oak Creek Floodplain, Oak Ridge

A successional floodplain forest characterizes the vegetation of the study site (Dalilman
and Van Voris, 1976). A small area was cleared of native vegetation to provide adequate
sunlight for growing vegetables and field crops on the contaminated site. Soil was
cultivated by tilling in the early spring of 1975 and 1976. Common varieties of vegetable
and forage seed stock were 6btained from local vendors. All plants were grown in parallel
rows randomly placed in each of four 5- by 5-m subplots. Each species (except potato
and tomato) was grown in two replicated rows per subplot. The entire 100-m^ plot
contained eight row replications of all species except potato and tomato. Single rows of
potatoes and tomatoes were grown in each subplot for a total of four rows per 100-m^
plot. Analysis of variance among subplots showed no significant difference in plutonium
concentration for a given species among subplots where sufficient plant material was
collected for plutonium analysis.

After the seedUngs emerged, black plastic mulch was placed on the soil surface to
prevent weed growth and resuspension of soil-borne plutonium. This method proved
effective because the ambient air concentration of plutonium 10 cm above the soil
surface did not exceed 0.14 x 10"^ pCi/m^ (Dahlman and McLeod, 1977). Air samples
were collected over mulched soil containing 63 ± 0.4 (standard error) pCi Pu/g.
Monitoring results for the Oak Ridge area showed that ambient plutonium was
0.016 X 10"^ pCi/m^ at a height of 1 m above ground (Oakes and Shank. 1977). It was
not determined if the higher value at 10 cm represented normal plutonium in ambient air
near the floodplain soil or if it represented radioactivity induced by air currents frorh the
high-volume samples (approximately 30 cfm). We suspect the latter caused it.

Vegetative samples were cleaned before radiochemical analysis. Bush bean, soybean,
and tomato leaves were washed and rinsed in an ultrasonic cleaning device, where fresh
samples (about 5 to 10% by volume) were placed in a 14iter-capacity cleaning tray
containing 700 ml of distilled water. The sonifier was tuned to give an optimal frequency
for cleaning each sample for a 2-min period. This was determined by the maximum
agitation delivered to the loaded cleaning tray by the wave generator. The cleaned
vegetation was carefully separated from residues, which settled in the bottom of the
sonifier tray. After the residue had been discarded and the tray rinsed with acid, the
vegetation was cleaned a second time with fresh distilled water. Samples were finally
removed from the tray, dried, chopped, and prepared for radiochemical analysis.

Other garden vegetables were cleaned according to kitchen food-preparation
techniques. Lettuce was cleaned under running tap water by gently rubbing the leaves.
The samples were drained, dried, chopped, and prepared for radiochemical analysis. Root
samples (radish, carrot, and potato) were vigorously hand scrubbed under running tap
water. The samples were dried, chopped, and prepared for radiochemical analysis.

The samples were sent to LFE, Richmond, Calif, for radiochemical analysis. The
analytical procedure has been discussed previously by Wessman et al. (1978).

388 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Results and Discussion

Field Crops at SRP

Wheat. Data for wheat are summarized in Table 1. The results for the South and North
fields are presented separately because of the differences in proximity to the stack and
hence the possibly different deposition patterns. There appears to be no significant
change in the plutonium concentrations of wheat foliage with time on either the South or
the North Field. The plutonium content of foliage or straw for both fields averaged
4 X 10"^ pCi/g(dry weight).

The isotopic composition of the foliage, as indicated by the a % ^^^Pu, changed with
time (Table 1). In March the wheat plants were short, and the foliage was not yet dense.
At this stage contamination of the foliage appeared to originate primarily from the soil, as
in rain splash or resuspendible matter. As the plants grew taller and their foliage became
denser, they were able to intercept fallout particles from the stack more efficiently. Also,
the dense foliage minimized rainfall energy on impact with the ground. The foUage had
much higher ^'^^Pu values in April and June than in March. This discrepancy indicated
that contamination originated primarily from fallout particles from the emission stack
because the ^^^Pu values were closer to those of the deposifion particles than to those of
the soil samples (Dahlman and McLeod, 1977). The plutonium concentration of straw
was approximately 300 times greater than that of the laboratory-thrashed grain and 40
times greater than that of the field-combined grain. That the plutonium values for the
grain were lower than those for the straw indicates that the grain was possibly shielded
from atmospheric deposition. The plutonium concentrations in foliage, straw, and grain
samples were slightly higher in the South Field.

The straw from the off-plant control site had only (8 ± 2) x lO"'* pCi/g (data not
shown), about two orders of magnitude lower than SRP field samples. The thrashed grain
from the control site, however, had the same plutonium contents as the SRP thrashed
samples. However, the high ^^*Pu percentage of 62 ± 15 for thrashed grain from the
control site, which is disparate from the straw value, should be noted. This liigh value
was apparently caused by contamination from thrashing of the SRP samples since all
tlirashed samples had similar plutonium values.

The concentration ratios (CR) for the June straw were 5 X 10"^ and 1 X 10"^' for
the South and North fields, respectively. Similar values (1 X 10~') were obtained for
samples from the control site. The CR for the grain collected from the combine averaged
2 X 10~^ compared with an average of 2 X lO"""* for the thrashed grain from the two
fields.

In Table 2 plutonium contents of field- and glasshouse-grown wheat plants are
compared. The glasshouse-grown plants had total plutonium contents one order of
magnitude lower than the plants from the South Field. The combined grain from the
fields was sifted with a soil screen to evaluate the effects of removing extraneous matter
on the plutonium content. Data from Tables 1 and 2 for the unsifted grain harvested with
a combine are remarkably similar. The sifted combined grain had a factor of 2 less total
plutonium content than the unsifted grain, which indicates that the extraneous matter
had higher plutonium contents than the grain. The extraneous matter included unshelled
grain and chopped straw.

PLUTONIUM CONTENTS OF FIELD CROPS 3H9

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PLUTONIUM CONTENTS OF FIELD CROPS 391

Soybeans. Data for soybeans from the field and glasshouse are given in Table 2. Total
plutonium concentrations in glasshouse-grown plants were an order of magnitude less
than those in field-grown plants. The CR for plants grown in the glasshouse was a factor
of 10 lower than that for plants grown in the field. The glasshouse plants (whole
vegetative plants) were sampled when the bean pods were still green and before the plants
started defoliating. The field-grown plants were sampled approximately 1 month before
the field was combined, when the plants had the maximum biomass. The plutonium
content of the combined grain was two orders of magnitude lower than that of the
vegetative parts.

The bean plants had plutonium concentrations [10~^ pCi/g (dry weight)] and ^■^^Pu
percentages similar to those of the wheat straw (whole plants without the grain).
Consequently the two crops had similar CR values. However, the soybean grain had
slightly lower plutonium contents than the wheat grain.

Com. Plutonium contents of field- and glasshouse-grown corn are shown in Table 2. So
that the extent of interception of fallout particles from the stack by foliage could be
determined, corn leaves were sampled from the field when the plants had their maximum
biomass. The leaves were sampled from two heights: 0 to 1 m from the ground and 1 to 2
m (or 1 m to the top). Plants in the glasshouse were sampled when the ears had matured.
The total plutonium content of leaves from glasshouse-grown corn was one order of
magnitude lower than that from field-grown corn, and the CR was a factor of 30 lower.
Leaves from the 1- to 2-m section of the corn plants had plutonium contents almost a
factor of 2 higher than leaves from the lower section (0 to 1 m). This indicates that the
upper foliage partially filtered and retained plutonium-bearing particles from atmospheric
deposition.

Shelled grain from the glasshouse had sliglitly lower plutonium concentrations than
the field-combined grain (Table 2). However, the CR of the grain from the glasshouse was
an order of magnitude lower than that of the field-combined grain. The unsifted
combined grain had total plutonium content a factor of 2 higher than that of the sifted.
Apparently the extraneous matter separated from the grain contained approximately 50%
of the plutonium in the sample.

Relative Contribution of Root Uptake to Plutonium Contents of Field Crops. A
comparison (Table 3) of the total plutonium contents of the vegetated material of the
crops grown under field and glasshouse conditions indicates that plutonium contamina-
tion of the field crops was primarily external in nature. For wheat and corn samples,
approximately 97% of the total contamination was external. For soybeans contribution
from root uptake appeared greater than 10%. No plausible explanation can be offered at
this time for this discrepancy.

Forage and Vegetable Crops at WOC Floodplain, Oak Ridge

Forage and vegetable crops harvested from a field plot located on the contaminated
floodplain contained measurable quantities of plutonium. The field experiments were
designed to minimize resuspension of dust and subsequent contamination of leaf surfaces
by plutonium-contaminated soil from the field experimental plots. Five factors aided this
objective: (1) The plastic mulch that covered the plot served as a barrier to the transport

392 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Comparison of Total Plutonium Contents of

Vegetative Materials from Crops Grown on the South

Field and in a Glasshouse at SRP*

Plutonium contents, fCi/g

(dry weight)
Vegetative contaniuiation,f

External
tamin:
material South Field Glasshouse %

Wheat straw 88.0 3.0 97
Soybeans,

whole plants 52.2 5.6 89

Corn leaves t 33.5 1.1 97

*The soils used in the glasshouse were from the top layer of the
South Field.

fCalculated from the equation: [(field content - glasshouse
content)/field content] x 100.

t Fiberglass mats were placed on top of the soil in the corn pots in
the glasshouse to prevent plant contamination by resuspension.

of particles from the soil surface to the air. (2) The floodplain soil moisture was
maintained continuously near field capacity owing to capillary conductivity from a
shallow water table. The cohesive force of a thin film of water surrounding the soil
prevented particles from becoming airborne. (3) Owing to protection provided by the
surrounding forest, the floodplain site was not exposed to wind, and particles did not
become airborne by aeoUan mechanisms. (4) The plastic mulch used for weed control
eliminated the need to cultivate; therefore airborne dust particles were not created by
mechanical operations. (5) The surrounding floodplain soil was covered by moist
decomposing litter and by a dense cover of herbaceous understory vegetation. The
boundary layer afforded by such conditions prevented soil-borne plutonium from
becoming airborne. These five factors reduced the likelihood that soil-borne plutonium
would become airborne and deposited on vegetation; indeed plutonium was not
detectable in 346-m^ air samples collected next to the test plot at 1-mheiglit. Attempts
to measure plutonium in air were unsuccessful because of liinited sample size over a 6- to
8-hr collection period, but the air concentration would be less than 0.3 X 10~^ pCi/m^
on the basis of a minimum detectable level of 0.1 pCi per sample (Dahlman and McLeod,
1977).

Concentrations of 239,240p^ -j^ foliage of bush beans, soybeans, tomatoes, lettuce,
radishes, and millet ranged from 0.01 to 0.33 pCi/g (dry weiglit) (Table 4). Concentra-
tions of plutonium in the fruit of these species were lower by at least an order of
magnitude tlian those in the foliage. Liinited observations on root crops (carrots,
potatoes, and beets) indicated that the plutonium concentrations of the edible portion
were similar to those of the foliage.

The plutonium concentrations of foliage and fruit appeared to represent plutonium
assimilated by the root pathway because surface contamination was removed before
radiochemical analysis. Samples of bush beans, soybeans, and tomatoes were washed and
rinsed in a sonic bath. This procedure effectively removed surface-bound plutonium

PLUTONIUM CONTENTS OF FIELD CROPS 393

TABLE 4 Plutonium Concentration and Concentration
Ratio of Field Crops from the Oak Ridge WOC Floodplain

Number

Total plutonium,*

Concentration

Crop

samples

pci/g

ratiof

Foliage

Bush bean

(1.1 ± 0.3) X

10-'

2x 10-'

Soybean

(1.0 + 0.2) X

10-'

2x 10-'

Tomato

(3.3 ± 0.7) X

10-'

5 X 10-'

Lettuce

(8.0 ± 2.0) X

10-^

1 X 10-'

Radish

(1.0 ±0.5) X

10-'

2x 10-'

MiUet

(1.0 ± 0.3) X

10-^

2x 10-"

Fruit

Bush bean

Whole bean

(7.0 ± 2.0) X

10-'

1 X 10-'

Shelled bean

(1.0 ± 0.4) X

10-^

2x 10-"

Tomato

(4.0 ± 2.0) X

10-'

6x 10-'

Soybean

Whole bean

(1.5 ±0.3)x

10-^

2 X 10-"

Shelled bean

(1.7 ± 1.2) X

10-^

3 X 10-"

Subterranean

Radish

(4.1 ± 0.6) X

10-'

6x 10-'

Carrot

3.1 X

10-'

5 X 10-'

Irish potato

Whole

(8.0 ± 1.0) X

10-^

1 X 10-'

Peeled

(6.0 ± 1.0) X

10-'

1 X 10-"

Skin only

(5.3 ± 1.2) X

io-»

8 X 10-'

Beet

Peeled

(5.0 ± 1.0) X

10-'

8x 10-'

Skin only

(1.3 ±0.5)x

10"

2x 10-'

* Values are means ± 1 standard error, which includes analytical
error of approximately 10%.

fConeentration ratio values are based on a soil concentration of
63 + 0.4 (standard error) pCi/g.

because no particles were observed in microscopic examination on any sample except
soybean pods where hirsute structures effectively retained surface contaminants.

The plutonium concentration of soybean fruit, which included the pod, was similar to
that of the whole bean. Concentrations of whole snap beans and shelled beans were also
similar. For both species the bean pod protected bean seeds from possible surface
contamination while young bean seeds matured on the vine. Because plutonium
concentrations were similar when bean seeds were analyzed separately and when seeds
and pods were analyzed together, these observations reinforced the argument that any
residual surface soil contaminant was removed from vegetative surfaces by the cleaning
process. Thus the plutonium content of these vegetables is attributed to assimilation by
the root pathway. Other results in support of root assimilation are discussed elsewhere
(Dahlman, Bo.ndietti, and Eyman, 1976; Dahlman and McLeod, 1977).

394 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Concentration of plutonium in root-vegetable crops reflected contamination by
residual soil plutonium. Concentrations ranged from 0.08 to 0.41 pCi/g for whole radish,
carrot, and potato samples that had been hand scrubbed before radiochemical analysis.
These samples were not examined microscopically for the presence of residual particles,
as had been done for sonically cleaned vegetation; yet concentrations of plutonium in
storage organs of roots were similar to those in cleaned vegetation. Peeling the skins from
potatoes and beets removed most of the residual plutonium. Potato skins averaged 0.53
pCi/g, as compared with 0.006 pCi/g for peeled potatoes. This indicates that about 99%
of the plutonium was removed by peeling. Beet skins had even greater concentrations of
plutonium, 1.3 pCi/g for skins vs. 0.005 pCi/g for peeled beets. Thus peeling the beet
removed about 99.5% of the total plutonium in the beet.

TABLE 5 Plutonium Concentration and Concentration

Ratio for Soybean Foliage Harvested at Different Stages

of Growth at Oak Ridge WOC Floodplain

Number
Growth of Total plutonium,* Concentration

stage samples pCi/g ratiof

8 weeks 8 (2.0 ± 0.4) x IQ-^* 3 x 10""

12 weeks 8 (1.0 ± 0.3) x lO"' 2x10"^

16 weeks 8 (1.9 ± 0.4) x lO"' 3 x 10"'

*Values are means ± 1 standard error.

f Concentration ratio values are based on a soil concentration of
63 + 0.4 (standard error) pCi/g.

The uptake pattern for the soybean plants at the WOC floodplain is shown in Table 5.
The plutonium concentration in the foliage after 8 weeks of growth was only 0.02 pCi/g
(dry weight). The plutonium concentration had increased five times (0.10 pCi/g) by the
12th week and ten times (0.19 pCi/g) by the 16th week. The fohage biomass was at
maximum at around the 16th week. Correspondingly, the CR of the foliage increased by
an order of magnitude from the 8th to the 16th week.

Dose to Man from Ingestion of Plutonium- Contaminated Foodstuff

The dose to man from ingesting wheat or other foods (Tables 2 and 4) can be estimated
from the nomograms in Fig. 2. For example, if a person ate 2 x 10^ g of wheat grain
containing 2 fCi Pu/g of dry grain, the 70-yr dose to bone would be ~0.5 mrem.
Likewise, if a person ate 2 X 10^ g of peeled Irish potatoes (approximately equivalent to
2000 kg of wet potatoes) containing 6 fCi Pu/g (dry weight), the 70-yr dose to bone
would be ~1.0 mrem. These doses are low compared with the 130-mrem dose from
background radiation (Klement et al., 1972). The basic data used for these calculations
came from ICRP publications (International Commission on Radiological Protection,
1959; 1972a; 1972b).

PLUTONIUM CONTENTS OF FIELD CROPS 395

1 X 10^--
5 --

2 --

10--
5 --

2 --

5 --

2--

1 X 10"^

5--

2--
1 X 10"2 —

--1 X 10^
-- 5

-- 2

--10
--5

-- 2

-- 5

-- 2

--1 X 10"''
-- 5

-- 2

— 1 X 10"2
-- 5

-- 2

— 1 X 10"^
-- 5

-- 2

— 1 X 10"''
-- 5

--2

--1 X 10'^

70-yr
DOSE,

BONE
mrem

-ri X lo'^

-- 5

-- 2

'1 X 10^

-- 5

--2

--1 X 10^

--5

--2

-^1 X 10^

FOODSTUFF
CONSUMED, g

PLUTONIUM

CONCENTRATION,

fCi/g

Fig. 2 Nomograph for calculating dose to bone from consumption of plutonium-
contaminated foodstuff. The dose is that which will be received by the bone during a
70-yr life-span. The following assumptions from ICRP publications were used: effective
absorbed energy per disintegration, 270 MeV (International Commission on Radiological
Protection, 1959); fraction from gastrointestinal tract to blood, 3 x 10~^ ; fraction from
blood to bone, 0.45; half-Ufe in bone, 100 yr (International Commission on Radiological
Protection, 1972); mass of bone, 5 x 10^ g (International Commission on Radiological
Protection, 1972).

Soils at SRP

As with plant materials, the resuspendible matter and soils from the South Field (closer
to the stack) had higlier plutonium concentrations than those from the North Field
(Table 6). The differences in concentrations were more pronounced in the resuspendible
matter where the South Field had plutonium levels one order of magnitude higher than
the North Field. Soils (0 to 5 cm) from both fields had plutonium levels about three
orders of magnitude higlier than soils from the control site. Cultivation of the fields
resulted in more uniform plutonium concentrations in the soil profile, increasing the
plutonium at the deeper depths (5 to 15 cm). Cultivation also altered the isotopic

396 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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PLUTONIUM CONTENTS OF FIELD CROPS 39 7

composition of the resuspendible portion to approach that of the soil values. After the
initial cultivation, the ^^^Pu percentage for the resuspendible matter decreased from an
average of 50 to an average of 27 for the two fields.

Soils at woe Floodplain, Oak Ridge

Concentrations of plutonium in the floodplain soil generally ranged from 10 to 150 pCi/g
(dry weiglit) over a 3-ha area (Fig. 3). The highest concentrations occurred behind the
former dike, along WOC, and at the upper end of the floodplain. Concentrations of
plutonium were less than 25 pCi/g near the margins of the floodplain. The ^'^^Pu
concentration of soil at a garden plot site (W35, S295) was 63 ±0.4 (standard error)
pCi/g.

A typical profile distribution of plutonium showed that the highest concentration of
plutonium was observed in the uppermost zone of the profile (Duguid, 1975).
Occasionally, however, where higher concentrations of plutonium occurred in the subsoil.

WHITE

OAK

CREEK

S30

S60

S90

150 pCi/g
100 pCi/g
^S 25 pCi/g
I. \ 10 pCi/g

Fig. 3 Areal distribution of plutonium at the contaminated Oak Ridge WOC floodplain.
Grid size is 30 m by 30 m (900 m' ).

398 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

this atypical distribution was attributed to variable deposition of sediments over an initial
plutonium deposit rather than to plutonium cycling by leaching processes.

Plutonium was not distributed uniformly between the silt and clay fractions of the
soil. Although 72% of the soil was silt, this fraction contained 60% of the plutonium,
whereas the 24% clay contained 40% of the plutonium. In this case the affinity of
plutonium for colloids, the clay fraction, may be responsible for plutonium enrichment in
clay.

The results of various attempts to extract plutonium from the soil indicated that
more than 95% of the plutonium can be removed by hot 8M HNO3 ' milder HNO3
treatments removed smaller amounts of plutonium (Auerbach, 1975). These treatments
indicated that plutonium recovery from the floodplain soil did not require rigorous
treatment, such as HNO3— HF or carbonate fusion. Milder treatment with O.IM citric acid
removed 16 to 24% of the soil plutonium. Contact with a CHELEX resin removed 11%.
Humic acid solubilized by the presence of Na from a sodium-saturated CHELEX resin
contained 5% of the total plutonium after 4 weeks. The carbonate treatment removed
54% of the soil plutonium. Further analysis of the carbonate extract showed that at least
90% of the plutonium behaved as Pu(lV) (Bondietti and Sweeton, 1977; Bondietti,
Reynolds, and Shanks, 1976).

General Discussion

The various CR values for the agricultural vegetation at SRP are generally higher than
those obtained for indoor-grown plants. Schulz et al. (1976a; 1976b) obtained values for
barley on the order of 10~^ for vegetative material; grain was a factor of 20 to 100
lower. In wheat grain they obtained CR values ranging from 1.1 X 10~^ to 3.8 x 10~^.
Lipton and Goldin (1976) obtained CR values for pea plants on the order of 10"'*.
Natural plant species not subject to contamination from atmospheric releases or
resuspension were observed to have CR values on the order of 10~^ to 10"'' (Francis,
1973; Hanson, 1975), much lower tlian values obtained at SRP. However, in arid, windy
environments, dust and soil particles can become airborne and can be deposited and
retained in leaves, causing plutonium CR values to approach 10° (Romney et al., 1975).
Earher studies at SRP indicated that deposition on the surfaces of tlie leaves and stems
was the principal mechanism of plutonium contamination of natural vegetation (Adriano
and Pinder, 1977; McLendon et al., 1976). The plutonium concentrations of all
ecosystem components decreased as the distance from reprocessing plants increased
(McLendon, 1975; McLendon et al., 1976). Thus considerable external contamination of
plants from atmospheric releases and resuspension is a complicating factor in field data
interpretation.

The CR values from the glasshouse studies at SRP are on the order of 10"'* to 10~^
for the vegetative materials and, in general, are similar or sUghtly higher than the values
obtained by Schulz et al. (1976a; 1976b) and Lipton and Goldin (1976). Our glasshouse
results suggest that ^^^Pu is more available than ^'^^Pu, as indicated by relatively higher

Pu percentage values in the vegetative samples than in the soil. The percent Pu
ranged from 33 ± 6 for soybeans to 55 ± 1 1 for corn. The soils used for growing these
crops had only 21% ^^^Pu. This difference in availability between the two radionuclides
has been observed also at the Trinity Site ecosystem, where changing ratios of

PLUTONIUM CONTENTS OF FIELD CROPS 399

^^^Pu/^^^Pu in soils, vegetation, and animal components were obtained (Hanson, 1975).
Although the ^■^^Pu data in the various ecosystem components were not conclusive,
earlier studies by McLendon et al. (1976) at SRP support evidence presented in other
studies that there was an apparent increase in the bioavailability of ^^^Pu relative to that
of ^^^Pu in tlie environment. Hanson (1975) has already discussed the possible
mechanisms causing this difference.

In general, plutonium concentrations in vegetative parts of agricultural plants at SRP,
where tlie primary mode of contamination was external in nature, were similar to those in
the forage and vegetable foliage cultured at ORNL floodplain ecosystem, where
contamination was caused primarily by root assimilation. However, the CR values of
vegetative parts of crops from SRP were about one order of magnitude higher than those
from ORNL.

Plutonium data on the edible portions of root crops are almost nonexistent. Potatoes
grown on soils receiving only global fallout plutonium which had been peeled had a CR of
3 X 10"^ (Bennett, 1976). This is similar to the ORNL data, which indicate that peeling
subterranean crops removed most of the plutonium, as high as 99.5% in the case of beets.
Whether plutonium in these vegetative tissues occurred as a free ion is still unknown.

Data on plutonium in a variety of species from the ORNL floodplain suggest that CR
values in the range of 10""'* to 10~^ are related to plutonium assimilation by the root
pathway. The order of CR values was 2.4 x 10~^ (foliage av.) > 1 .7 x 10""* (fruit
av.)>0.9 X lO""* (peeled root). The results for foliage compare favorably with CR values
for plants grown in the glasshouse at SRP where aboveground tissues were protected from
airborne sources of plutonium. Comparative results described in tliis chapter and
elsewhere (Dahlman, Bondietti, and Eyman, 1976; Dahlman and McLeod, 1977) clearly
ascribe high plutonium CR values (10~^ to 10*^) to contamination of foliage surfaces.
Assimilation of plutonium by the root pathway was responsible for CR values in foliage
of 10~^ or less. There was no evidence that assimilation of plutonium by the root
pathway had been especially enhanced as a result of weathering, complexation, or other
soil processes in the 30-yr period since the site was contaminated in 1944.

Summary

Agricultural crop experiments conducted adjacent to a reprocessing facility at SRP, which
releases low chronic levels of plutonium per year through an emission stack, and at the
woe floodplain at Oak Ridge, where plutonium was deposited in 1944 from the
Manhattan Project weapon development, revealed the following:

I.Plutonium concentrations and isotopic compositions in winter wheat grown in
1975 at SRP were affected by distance from the stack, plant heiglit, and method of
processing the grain. In general, vegetative materials closer to the stack (South Field) had
plutonium concentrations a factor of 2 higlier than samples farther away (North Field).
The wheat straw had approximately two orders of magnitude higher plutonium content
than the straw from the control site. Grain harvested by a combine had an order of
magnitude higher plutonium content than laboratory-thrashed grain. When the plants
were short, as in March, the sources of contamination were stack emission and
resuspendible and soil materials, as indicated by an average 38 a % ^^^Pu. Later in the

400 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

season, as tlie plants matured and grew taller, emission became the principal source of
contamination, as indicated by a higher '^^^Pu percentage.

2. Plutonium concentrations of the foliage of soybeans and corn grown in the South
Field at SRP were similar to those of tlie June wheat straw [on the order of 10^^ pCi/g
(dry weiglit)] collected from the same field. However, the corn foliage had higlier a %
^^^Pii than did the wheat and soybean foliage, which reflects the minimal contribution
of resuspension to the corn canopy. Plutonium contents of the combine-harvested grain
of tlie three crops were basically similar. Sifting the combined grain separated about 50%
of the plutonium-bearing particles and/or extraneous matter from the grain and resulted
in plutonium concentrations a factor of 2 lower in the sifted grain than in the unsifted
grain.

3. Preliminary results from the SRP glasshouse studies indicated tliat only about 3%
of the total contamination of field-grown crops adjacent to a reprocessing facility was
contributed by root uptake. Also, it appeared from glasshouse results that ■^■^^Pu was
more available than ^^^Pu.

4. Cultivation of the fields at SRP before planting wheat caused the plutonium
concentrations in the top soil layer to become more uniform and to increase the
plutonium concentrations at the deeper depths (5 to 15 cm). It also caused the ^^^Pu
percentage for the resuspendible portion to approach the soil values. Subsequent
cultivations caused greater uniformity in plutonium concentration in the top soil.

5. Concentrations of 239,240p|j ^^ ^^^ foliage of forage and vegetable crops grown at
the woe floodplain ranged from 0.01 to 0.33 pCi/g. This represented the amounts of
plutonium taken up by plants exclusively via the root pathway from soils with plutonium
concentrations ranging from 10 to 150 pCi/g and averaging 63 pCi/g. Plutonium
concentrations in the fruit of these species, however, were lower by at least an order of
magnitude than those in the foliage.

6. Peeling the skins of potatoes and beets grown at the WOC floodplain removed
approximately 99% of the residual plutonium.

7. Plutonium contents of soybean foliage were related to the stage of maturity and
were maximum when the foliage biomass was maximum.

8. In general, the CR values of vegetative parts of crops at SRP were approximately
one order of magnitude higher than those at Oak Ridge, which suggests the influence of
aerial deposition of plutonium at the SRP site.

Acknowledgments

This work was supported by the U. S. Department of Energy through contract Nos.
EY-76-C-09-0819 (to Savannah River Ecology Laboratory, University of Georgia),
AT(07-2)-l (to Savannah River Laboratory, E. I. du Pont de Nemours and Co.), and
W-7405-eng-26 (to Oak Ridge National Laboratory, Union Carbide Corp.).

References

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from Nuclear luel Reprocessing, 7. Environ. Qual , 6: 303-307.
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the Hnvironment, ERDA Report ORNL-5 193, pp. 6-7, Oak Ridge National Laboratory, NTIS.

PLUTONIUM CONTENTS OF FIELD CROPS 401

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Enviro)iment , Symposium Proceedings, San Francisco, 1975, pp. 367-383, STI/PUB/410,

International Atomic Energy Agency, Vienna.
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Substances in Soils and Natural Waters, in Transuranium Nuclides in the Environment, Symposium

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1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178, pp. 303-320, Nevada

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Symposium on Radioecology, Oregon State University, May 12-14, 1975, C. E. Cashing (Ed.),

pp. 291-298, Ecological Society of America Special Publication Series No. 1, Academic Press, Inc.,

New York.
Duguid, J. O., 1975, Status Report on Radioactivity Movement from Burial Grounds in Melton and

Bethel Valleys. ERDA Report ORNL-5017, p. 55, Oak Ridge National Laboratory, NTIS.
Francis, C. W., 1973, Plutonium Mobility in Soil and Uptake in Plants: A Review, /. Environ. Qual . 2:

67-70.
Hanson, W. C, 1975, Ecological Considerations of the Behavior of Plutonium in the Environment,

Health Phys..2S: 529-5 37.
International Atomic Energy Agency, 1976, Transuranium Nuclides in the Environment. Symposium

Proceedings, San Francisco, 1975, STI/PUB/410.
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Dose for Internal Radiation. ICRP Pubhcation 2, Pergamon Press, Inc., New York.
, (Task Group of Committee 2), 1972a, The Metabolism of Compounds of Plutonium and Other

Actinides, ICRP Pubhcation 19, Pergamon Press, Inc., New York.
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Publication 20, Pergamon Press, Inc., New York.
Klement, A. W., C. R. Miller, R. P. Minx, and B. Shlelen, 1972, Estimates of Ionizing Radiation Doses

m the United States, 1960-2000, Report EPA-ORP/CSD-72-1, U. S. Environmental Protection

Agency, Office of Radiation Programs, Washington, D. C.
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du Pont de Nemours & Co., Savannah River Laboratory, NTIS.
Lipton, W. v., and A. S. Goldin, 1976, Some lactors Influencing the Uptake of Plutonium-239 by Pea

?hnts. Health Phys.. 31: 425^30.
McLendon, H. R., 1975, Soil Monitoring for Plutonium at the Savannah River Plant, Health Phys.. 28:

347-354.
, O. M. Stewart, A. L. Bom, J. C. Corey, K. W. McLeod, and J. E. Pinder III, 1976, Relationships

Among Plutonium Contents of Soil, Vegetation and Animals Collected on and Adjacent to an

Integrated Nuclear Complex in the Humid Southeastern United States of America, in

Transuranium Nuclides in the Environmoit. Symposium Proceedings, San Francisco, 1975, pp.

347-363, STI/PUB/410, International Atomic Energy Agency, Vienna.
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laboratory Site at Oak Ridge National laboratory, ERDA Report ORNL-TM-5695. Oak Ridge

National Laboratory, NTIS.
Price, K. R., 1972, Uptake of ^'''Np, ""^^Pu. ^"Mw, and '""' Cm from Soil by Tumbleweed and

Cheatgrass, USAEC Report BN\VL-1688, Battelle Pacific Northwest Laboratories, NTIS.

402 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Romney, E. M., A.Wallace, R. O. Gilbert, and J. E. Kinnear, 1975, "'''""Pu and ' "' Am
Contamination of Vegetation in Aged Plutonium l-'allout Areas, in Radioecology of Plutonium and
Other Transuranics in Desert Environments, Nevada Applied Ecology Group Progress Report as of
January 1975, M.G. Wliite and P. B. Dunaway (Eds.), ERDA Report NVO-15 3, pp. 43-88, Nevada
Operations Office, NTIS.

Schulz, R. K., G. A. Tomkins, and K. L. Babcock, 1976, Uptake of Plutonium and Americium by
Plants from Soils: Uptake by Wheat from Various Soils and Effect of Oxidation State of
Plutonium Added to Soil, inTransuranium Nuclides in the Environment, Symposium PToctadings,
San Francisco, 1975, pp. 303-310, STI/PUB/410, International Atomic Energy Agency, Vienna.

, G. A. Tomkins, L. Leventhal, and K. L. Babcock, 1976, Uptake of Plutonium and Americium by

Barley from Two Contaminated Nevada Test Site Soils, /. Environ. Qual , 5 : 406-410.

Tamura, T., 1976, Physical and Chemical Characteristics of Plutonium in Existing Contaminated Soils
and Sediments, in Transuranium Nuclides in the Environment, Symposium Proceedings, San
Francisco, 1975, pp. 218-220, STI/PUB/410, International Atomic Energy Agency, Vienna.

Tennessee Valley Authority, 1975, personal communication from T. C. Bounds to R. C. Dahlman,
Sediment Investigation, 195 1, White Oak Creek and Lake.

Wessman, R. A., K. D. Lee, B. Curry, and L. Leventhal, 1978, Transuranium Analysis Methodologies
for Biological and Environmental Samples, in Environmental Chemistry and Cycling Processes,
DOE Symposium Series, Augusta, Ga., Apr. 28-May 1, 1976, D. C. Adriano and I. L. Brisbin, Jr.
(Eds.), pp. 275-289, CONF-7 60429, NTIS.

Ecological Relationships of Plutonium
in Southwest Ecosystems

T. E. HAKONSON and J. W. NYHAN

A comprehensive summary of results was prepared on plutonium distribution and
transport in Los Alamos and Trinity Site study areas. Despite differences in ecosystems
and plutonium source, there are several similarities in plutonium distribution between Los
Alamos and Trinity Site study areas. First, the soils I sediment component contains
virtually all the plutoniufn inventory, with vegetation and rodents containing less than
0.1% of the total in all cases.

Plutonium has penetrated to considerable soil depths at both locations, although it
has occurred much more rapidly and to a greater degree in the alluvial soil at Los Alamos
than in the arid terrestrial system at Trinity Site. However, in all cases less than 50% of
soil-column plutonium inventories was found in the surface 2.5 cm. The plutonium
penetration depth appears to correspond to the moisture penetration depth at Trinity
Site. This is probably the governing factor at Los Alamos, although storm runoff and
accompanying turbulent mixing processes complicate the process. In Acid-Pueblo
Canyon, the bulk of the soil column inventory lies in the lower profiles, an indication of
the loss of the plutonium from surface layers due to sediment transport.

Soil plutonium, in most cases, was associated with relatively coarse-size fractions. The
silt-clay (<53 ym) fraction contained relatively little (<15%) of the plutonium: this
reflects the small amounts of this size fraction in study area soils. An exception was in
Area 21 at Trinity, where the <53-im72 soil-size fraction contained about 73% of soil
plutonium inventories. The importance of these distributional differences was demon-
strated for Trinity Site, where Bagnold dust samples from Area 21 contained 54%
silt-clay material and samples from Area Ground Zero (GZJ contained less than 10% of
this material

Concentrations in herbaceous vegetation were generally related to those in soils from
all sites. Our belief, although unsubstantiated, is that external contamination of the plant
surfaces is the major contaminating meclmnism in these arid systems. Vie plutonium
concentrations in certain rodent tissues from all study areas were related to corresponding
soil concentrations. Over 95% of the plutonium body burden in rodents was associated
with pelt and gastrointestinal tract samples, indicating the dominance of physical
processes as the contaminating mechanism

Horizontal transport of soil plutonium is dominated by physical processes. At Los
Alamos water governs the downstream transport of soil plutonium, and indications are
that wind is a relatively more important transport vector at Trinity Site.

In no case was there evidence for trophic-level increase due to physiological processes
as plutonium passes from the soil to vegetation to the rodents, although food habits of
rodents are not hiown sufficiently to preclude a trophic-level increase. We believe,
however, that rodents most likely come into contact with environmental plutonium
directly from the soil and not through a food-web intermediary.

403

404 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Several reviews on environmental plutonium distribution and transport indicated a
general lack of published field data from representative areas of the United States
(Francis, 1973; Price, 1973; Roniney, 1977; Hanson, 1975; Hakonson, 1975). Several
field studies of plutonium have been initiated in the last few years to address
informational needs at a number of locations which encompass a wide spectrum of
climatic conditions ranging from deserts to humid forests and contain plutonium from
industrial, weapons, or accidental-release sources.

The comparison of plutonium data from two southwest ecosystems in this chapter is
one step in the total synthesis of information from various regions of the United States
where types of ecosystems and sources of plutonium differ. The southwest United States
is an important study locale because of the energy activities that may develop and the
lack of understanding of the processes in arid systems which govern distribution and
transport of contaminants. In this regard studies on environmental plutonium are useful
to develop an understanding of patterns that are applicable to the transport and fate of
other materials.

The objective of this chapter is to use existing plutonium contamination in the
canyon waste areas at Los Alamos and in the grasslands in the fallout zone at Trinity Site

• To evaluate the role of environmental transport processes in distributing and
redistributing surface inputs of plutonium.

• To evaluate the transport of environmental plutonium to the biosphere and the
relationships that lead to the potential for human exposure.

• To compare plutonium behavior in these two major southwest ecotypes.

The tasks in this study were to (1) document plutonium inputs where possible,
(2) develop an understanding of distributions by inventory of major environmental
components, and (3) evaluate transfers as functions of ecological variables. Plutonium, as
used in this chapter, denotes ^^^Pu and/or ^^^'^'**^Pu.

Site Descriptions

Los Alamos

The canyons at Los Alamos, in north central New Mexico (Fig. 1), are typical of those
in the southwest plateau region of New Mexico, Arizona, Colorado, and Utah. They vary
from 10 m to over 200 m in depth and were formed by water erosion of the volcanic
substrate of the Pajarito Plateau. The area has a semiarid continental mountain climate
(Table 1) with annual precipitation ranging from about 20 to 50 cm as elevation increases
from 1650 to 2200 m; rainfall accounts for about 75% of the annual precipitation.
Drainage from the 113-km^ Laboratory site is via the many canyons that bisect the
plateau. The biotic resources of the canyons are diverse (Miera et al., 1977); total
vegetative ground cover is variable but generally high and approaches 100% in some areas
owing to the dense overstory, which is partly due to the industrial liquid effluents.

Nearly all the liquid wastes generated by the Laboratory since 1943 have been
collected by industrial waste lines, treated (since 1951), and released into one of three
canyons (Fig. 1 ). The resuUs of studies in two of these canyons are emphasized in this
chapter since they represent the extremes in temporal use history. The oldest
waste-receiving area is Acid -Pueblo Canyon, which was used from 1943 to 1963 and

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 405

LEGEND

/':

■ RADIOECOUOGICAL STUDY AREAS

/ AREA 21

SCALE (km) 0

5 10

/ \ chupad'era I

f I

/mesa

BINGHAM ^

-'■'' "^

\~ whiTe sands"/

\ MISSLE range/

\ MILITARY f

RESERVATION

AREA 16 \

1
1
/

380

\ (^^^^^*>'H'perimeter6f .-,:: 1

TRiNirXqiTF l«^^<^ : FALLOUT ZONE i
rRnlwnVyJpnAS f \s ' B'^SED ON SURVEY
GROUND \ZERO-tQ_J \* -, ||^ 1948 1

V ■ V !

WSMR -^CONTROL '.,^'. '
ROUTE 7 \ \p . 1

Fig. 1 Plutonium study areas at Los Alamos and Trinity Site, New Mexico.

406 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Some Characteristics of Plutonium Study Areas at Los Alamos

and Trinity, New Mexico

Mortandad

Acid-Pueblo

AreaGZ

Area 21

Annual precipitation,

43 to 52

39 to 54

12 to 25

30 to 40

cm
Average annual

7.5

7.1

temperature, °C

Range

-26 to 36

-23 to 31

-5 to 39

-4 to 38

Sou

Sandy alluvium

Sandy loam

Loam

SoUpH

8.6 to 9.2

7.1 to 7.9

7.5 to 8.4

8.2 to 8.4

Soil cation exchange capacity.

0.06 to 0.09

0.05 to 0.10

0.02 to 0.02

0.02 to 0.03

equivalents/kg soil

Soil organic carbon, %

0.10 to 0.40

0.04 to 0.54

0.40 to 0.70

0.75 to 1.5

Clay mineralogy

Amorphous

Mixed

Plutonium source

Industrial
liquid effluent

Weapon fallout

So^239,240py/238pu

0.35*

100*

12t

21t

concentration ratio.

(pCi/g)/(pCi/g)

Weathering time of pluto-

Oto 11

14 to 30

nium in environment (yr) as

of 1973, the year most of

the data in this paper

were collected

*See Miera et al. ( 1977); Nyhan, Miera, and Peters (1976a).
tSee Neher and Bailey (1976).

received an estimated 173 mCi of plutonium; Mortandad Canyon has been used for the
least amount of time (from 1963 to present) and currently receives most of the
Laboratory's liquid waste plutonium. As of 1973 and 1974, the years from which data in
this chapter were collected, Mortandad Canyon had received about 61 mCi of plutonium.
Surface water exists in the upper reaches of both canyons as a result of Laboratory
effluents and/or domestic sewage-treatment effluent; the lower portions of the canyons
are normally dry. Surface water, including the pulse releases of plutonium-contaminated
liquid effluent, rapidly percolates into the alluvium and generally disappears about 1 km
downstream. Relatively large flows occur in both canyons during storm runoff events.
Storm runoff reaches the Rio Grande via Acid— Pueblo and Los Alamos Canyons (Fig. 1 ),
but the runoff water in Mortandad Canyon rapidly soaks into the thick alluvial deposits
and seldom reaches postoutfall distances beyond 3 km. Many rainstorms at Los Alamos
are intense, of short duration, and result in dramatic flash floods in the canyons.

Trinity

Trinity Site and the associated fallout zone is located in the northern end of the
Tularosa basin in south central New Mexico (Fig. 1 ). The region is characterized (Table 1)
by low rainfall (12 to 40 cm), high summer temperatures (commonly greater than 37°C),
and severe wind and water erosion on exposed and disturbed ground surfaces. Rainfall

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 407

accounts for about 90% of the annual precipitation. Tlie area supports a relatively dense
vegetation cover, considering the region; total vegetative ground cover ranges from about
15 to 25%(Neher and Bailey, 1976).

On July 16, 1945. a 20-kt atomic bomb was detonated 31 m above the ground
surface at Trinity Site during a relatively unstable climatic regime when winds were to the
northeast and were accompanied by intermittent thundershowers. Fallout from the cloud
deposited in a northeast direction in the general pattern outlined in Fig. 1 (Larson et al.,
1951 ). Relatively high fallout deposition occurred on Chupadera Mesa about 35 to 55 km
from the crater. The reasons for the heavy deposition in this zone are unknown but may
be related to weather or topographic factors. Tlie elevation increases from about 1 500 m
at the crater to 2100 m on Chupadera Mesa. The fallout zone within 15 km of the crater
is on the Wliite Sands Missile Range, which is under U. S. Army jurisdiction. Beyond 15
km the fallout zone is on mixed private and public (Bureau of Land Management, State;
and U. S. Forest Service) lands that are used heavily for domestic livestock grazing.

Plutonium Distribution

General

The chronic release of treated liquid effluents to the Los Alamos canyons has resulted
in soil plutonium concentrations that are generally much higher than those at Trinity
Site. Concentrations of a few hundred picocuries per gram (dry weight) are found in soils
from the canyons, whereas those in Trinity soils average less than 1 pCi/g (Table 2).
Worldwide fallout concentrations of 2 3 9,2 4 op^^ ^ ^^ Alamos and Trinity Site soils
average about 0.01 pCi/g (Apt and Lee. 1976; Nyhan, Miera, and Neher, 1976b).

TABLE 2 Ranges in Plutonium Concentration and Variability

Estimates in Some Los Alamos and Trinity Ecosystem

Components in 1973 and 1974

Component*

Los Alamos

Trinity

Soil (0 to 15 cm)

pCi Pu/g

l-290t

0.02-0.32

cvt

0.32-2

0.52-0.88

nCi Pu/m'

190-80,000

2.8-63

Vegetation

pCi Pu/g

0.08-76

0.002-0.37

cvt

0.65-2.2

0.38-1.1

pCi Pu/m^

0.7-600

0.07-5

Rodents

fCi Pu/g

7-300

3-100

CV$

0.16-1.3

0.52-1.3

fCi Pu/m^

0.2-20

0.03-2

*Dry-weight concentrations.

tlncludes^^'Puand 2 3 9,2 4 op^,

^Coefficient of variation (CV = standard deviation/mean).

408 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Vegetation at both study locations contains the highest plutonium concentrations of
any biotic component yet examined (Hakonson and Bostick, 1976). Plutonium
concentrations in native grasses and forbs ranged from 0.08 to 76 pCi/g (dry weight) at
Los Alamos and from 0.002 to 0.37 pCi/g in the Trinity fallout zone; levels in vegetation
generally do not exceed those in corresponding soil samples. Additionally, the highest
plutonium concentrations were associated with plants growing closest to the ground
surface; taller growth forms, such as shrubs and trees, contained the lowest concentra-
tions (Hakonson and Bostick, 1976; Hakonson and Johnson, 1974).

Plutonium concentrations in rodents, as representatives of the primary consumer
trophic level, reflect the low physiological availability of the element. Pooled samples of
internal organs from rodents generally do not contain measurable levels of plutonium,
even though habitat soils may contain up to a few hundred picocuries per gram.
Plutonium concentrations in whole rodents ranged from analytical detection limits of
about 5 fCi/g to a few hundred femtocuries per gram; most of this radioactivity was
associated with samples of pelt and gastrointestinal (GI) tract and contents.

Plutonium concentration variability, as characterized by the coefficient of variation in
soils, plants, and animals, was uniformly higli at all study sites. It commonly varied up to
2.0, with extreme values approaching 3.0 (Hakonson and Bostick, 1976; Nyhan, Miera,
and Neher, 1976b). Variability of this magnitude has been observed at several
environmental plutonium study sites in the United States (Little, 1976; Gilbert and
Eberhardt, 1976) and results in the need for very large sample sizes in field experiments
(Gilbert and Eberhardt, 1976; White and Hakonson, 1978).

Soils

Horizontal Distribution. Horizontal plutonium concentration gradients are evident in
both study areas, reflecting the dispersion from point sources of plutonium. Concentra-
tions in the Los Alamos stream channels decrease one to two orders of magnitude in a
predictable fashion (Nyhan, Miera, and Peters, 1976a; Hakonson and Bostick, 1976)
within 10 km of th€ effluent sources, whereas similar differences occur over much greater
distances at Trinity and do not necessarily decrease with distance. For example,
plutonium concentrations in Trinity soils gradually increase from a minimum just outside
the crater to a maximum at about 50 km from the crater (Nyhan, Miera, and Neher,
1976b; Larson etal., 1951).

Liquid effluent radionuclides at Los Alamos have been transported laterally into the
stream banks as well as to downstream areas. Stream-bank soils accumulate radionuclides
to levels equivalent to adjacent channel soil (Anonymous, 1977), and they serve as a
long-term source of these materials to stream-bank biota. The stream banks, wliich are
heavily vegetated, retard the downstream movement of radionuclides since they are not
subject to the severe erosion encountered in the channel.

Although plutonium concentrations average much higher in the canyons than at
Trinity, the extent of the contamination in the canyons is confined to less than 0.1 km^,
whereas the low-level contamination at Trinity Site covers several thousand square
kilometers. Consequently the ecosystems at risk at Los Alamos are exposed to higher
concentrations of plutonium than those at Trinity; however, the areas involved are
smaller, and corrective action could be taken more easily should it ever be necessary.

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 409

Vertical Distribution. Some data from Area 21 (see Fig. 1) at Trinity Site indicate that
the plutonium originally deposited on those environs in 1945 has been depleted from the
soil surface over a 23-yr period (Table 3). Area 21 soils contained about 700 nCi/m^ in
1950 (Olafson. Nishita, and Larson, 1957) and 18 nCi/m^ in 1973 (Nyhan. Miera, and
Neher, 1976b).

The depletion of plutonium from the soil surface is primarily due to the vertical
transport of the element into the soil profile rather than to horizontal transport away
from the study site by wind or water. Evidence that plutonium has migrated into the soil
profile at the two Trinity Site locations is illustrated in Table 4 and is presented in detail
by Nyhan, Miera, and Neher (1976b). In 1973 plutonium was detected at the 28- and
35-cm depths at Areas GZ and 21, respectively, whereas in 1950 plutonium was confined
exclusively to the surface 2.5 cm (Olafson, Nishita, and Larson, 1957). The patterns of
distribution with depth were typical of those observed in terrestrial soils in that
plutonium concentrations decreased with depth.

TABLE 3 Comparison of Plutonium Concentrations

in Surface (0 to 2.5 cm) Soils from Chupadera Mesa

as a Function of Time After the Atomic Bomb Test

at Trinity Site in 1 945

Plutonium concentration, nCi/m-
1950* 1951* 1973t

746{0.31)t 341(0.82)$ 18(0.48)*
n=6 n=3 n=8

*Data for 1950 and 1951 from Larson et al. (1951),
and Olafson, Nishita, and Larson (1957).

+ Data for 1973 from Nyhan. Miera, and Neher (1976b).

^Parenthetic value is coefficient of variation
(CV = standard deviation/mean).

TABLE 4 Mean Percent Plutonium Inventory in Soil Profiles from
Los Alamos and Trinity Site Study Locations in New Mexico

Trinity Site*

Los Alamos*

Depth, cm

Area GZ

Area 21

Depth, cm

Mortandad

Acid-Pueblo

0-2.5

29(0.78)t

41(0.46)1

0-2.5

20(0.44)i

4.0(0.76)t

2.5-5.0

18(0.72)

19(0.63)

2.5-7.5

36(0.23)

10(0.48)

5-10

21(0.81)

6.0(0.88)

7.5-12.5

21(0.55)

20(1.3)

10-15

15(0.67)

8.0(0.92)

12.5-30

24(0.79)

67(0.18)

20-25

17(1.3)

16(1.0)

25-33

NDt

10(1.2)

*n = 8 for Trinity Site data; n = 10 for Los Alamos data.

fParenthetic value is coefficient of variation (CV = standard deviation/mean).

$Not detectable.

410 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

The depth of plutonium transport into channel and bank soil profUes in the Los
Alamos canyons is much greater than that at Trinity. In areas where permanent surface
water exists (i.e., Mortandad Canyon), elevated plutonium concentrations are found at
depths of 100 cm in the channel and at depths of 50 cm in the stream bank. Plutonium
concentrations in channel soils do not show any consistent patterns with sampling depth,
whereas decreasing concentrations with depth are evident in bank soils. In downstream
areas, which are dry except during periods of storm runoff, plutonium occurs at depths of
at least 30 cm (Nyhan, Miera, and Peters, 1 976a).

The transport of plutonium into the channel alluvium and stream-bank soil has been
rapid, as shown by the presence of elevated ■^ ^^Pu at the lower sampling depths. Elevated
^^^Pu was observed at soil depths of 30 cm in Mortandad Canyon in 1972, about 4 yr
after the first significant release of this element to the canyon (Hakonson and Bostick,
1976). In contrast, fallout 239,240p^ -^ Trinity soils Syr after the bomb test was
confmed to the upper 2.5 cm of soU (Olafson, Nishita, and Larson, 1957).

A common feature of plutonium distribution in soils from both locations was that in
1974 less than one-half the total plutonium in the soil column was present in the surface
2.5 cm (Table 4) despite differences in soils and source of plutonium. In Acid— Pueblo
Canyon lOyr after the decommissioning of those facilities for waste disposal, an average
of 67% of the soil column inventory was below the 12.5-cm depth, which reflects
depletion of plutonium from the surface layers by vertical and horizontal transport
processes. Previous studies in the canyons have shown that horizontal transport of soil
during storm runoff events is an important mechanism in the downstream transport of
plutonium (Purtymun, 1974; Hakonson, Nyhan, and Purtymun, 1976).

The depletion of plutonium from the soil surface decreases the probability of
horizontal transport by wind and water but may increase the probability of uptake by
vegetation during the time that the element is distributed within the plant rooting zone.
However, over long periods of time, continued movement of plutonium into the soil
profile may remove the element from the biologically active zone of the soil.

Particle Size Relationships. Tlie highest concentrations of plutonium in soil at the Los
Alamos and Area 21 locations were associated with the silt-clay fraction, whereas this
fraction at Area GZ, 1 km from the crater, contained the lowest concentrations of
plutonium (Table 5) (Nyhan, Miera, and Neher, 1976b; Nyhan, Miera, and Peters, 1976c).
At the GZ location, the highest concentrations were measured in the 1- to 2-mm soil
particles, which perhaps reflects the physical characteristics of the fallout debris near the
detonation site and/or depletion of the plutonium from smaller size fractions by wind or
water transport vectors. Decreasing plutonium particle sizes with increasing distance from
the crater were also noted at weapons test sites in Nevada (Romney, 1977).

The inventory of plutonium among the various soil size fractions in surface soils at
the Los Alamos and Area GZ Trinity study sites was similar in that the silt— clay size
fraction (<53 jim) comprised less than 10% of the soil mass and contained less than 15%
of the plutonium (Table 5), whereas over 80% of the plutonium was associated with soil
particles greater than 53 /am (Nyhan, Miera, and Neher, 1976b; Nyhan, Miera, and Peters,
1976c). The reverse was true at Area 21, Trinity Site, where the <53-/im fraction
comprised 36% of the soil mass and contained over 70% of the soil plutonium inventory.

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 411

TABLE 5 Comparative Distribution of Plutonium in Surface Soil
(0 to 2.5 cm) Size Fractions at the Los Alamos and Trinity Study Areas

Soil size fraction*

<53ium*

53-105 Mm 105-500 /im

500-1000 Mm

1-2 mm

2-23 mm

Mortandad Canyon

pCi Pu/gt

1500.0

1300.0 610.0

310.0

87.0

69.0

Soil weight, %

2.2

1.8 14.0

21.0

26.0

35.0

Pu in fraction, %

14.0

6.0 27.0
Acid-Pueblo Canyon

21.0

16.0

pCi Pu/gt

85.0

60.0 25.0

8.8

7.9

25.0

Soil weight, %

3.0

3.0 16.0

26.0

28.0

24.0

Pu in fraction, %

7.0

7.0 31.0
Trinity Site, Area GZ

19.0

17.0

19.0

pCi Pu/gt

0.07

0.05 0.92

2.1

5.3

0.01

Soil weight, %

8.9

11.0 49.0

23.0

6.1

2.0

Pu in fraction, %

0.78

0.43 36.0
Trinity Site, Area 21

38.0

25.0

0.01

pCi Pu/gt

3.8

1.7 0.42

0.64

1.6

0.23

Soil weight, %

36.0

18.0 25.0

4.2

2.9

14.0

Pu in fraction, %

73.0

16.0 5.5

1.4

2.4

1.8

*Size fraction data based on composite samples.

tPu denotes primarily ^^'Pu in Mortandad Canyon and 2 3 9,2 4opjj .^^ all other study locations.

Vegetation

Plant- Soil Relationships. The concentrations of plutonium in the study area vegetation
were related to the levels of plutonium in associated soils (Fig. 2). The relationship
between plutonium concentrations in vegetation and in soils was predictable over a range
of five orders of magnitude in concentrations; this relationship is similar to relationships
that were observed in the Rocky Flats environs (Little, 1976).

Plant -soil plutonium concentration ratios (CR = picocuries per gram of vegetation/
picocuries per gram of soil) are a convenient means of estimating the plutonium levels in
vegetation growing on contaminated soils. Ratio estimates for native grasses in the Los
Alamos and Trinity Site study areas (Table 6) ranged from 0.05 to 1.2, whereas the values
for forbs ranged from 0.04 to 1.1. All these ratios are higli relative to those derived from
experimental studies where root uptake was the contamination mechanism (Romney and
Davis, 1972; Wilson and Cline, 1966), which indicates that either plutonium is much
more available to plants under field conditions or that mechanisms other than root
uptake are responsible for the plutonium measured in plant samples from the field.

The relative amounts of plutonium associated with the internal and external portions
of the vegetation are difficult to assess under field conditions, although we contend that

412 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

100 —

10 —

c
o

0.1

0.01

0.001

1 1 1 1 1

1 1 1

1 1

1 M 1

1 1

A Trinity

O y

—

0 Los Alamos

Y = 0.25x1-^

r^ = 0.85

—

0 o y

—

1 1 1 1 1

1 1 1

1 1

\ 1 1

1 1

0.01

0.1

1 10

pCI Pu/g soil (0 to 15 cm)

100

1000

Fig. 2 Relationship of average plutonium concentration in herbaceous vegetation
(grasses and forbs combined) and in corresponding soils in Los Alamos and Trinity Site
study areas.

TABLE 6 Plutonium Concentration Ratios for Vegetation and Associated Soils
from Los Alamos and Trinity Study Sites

Plutonium concentration ratio*

Los Alamos

Trinity Site

Component

Mortandad Canyon n Acid-Pueblo Canyon

n Area GZ n

Area 21

Grass
Forb

24
16

0.93(0.94! 19 0.13(1.2)
0.31(1.3) 11 0.23(0.28)

13 0.05(0.54) 16
17 0.04(0.97) 21

1.2(0.74)
1.1(0.92)

*Ratio calculated as [pCi Pu/g (dry weight) plant] /[pCi/g (dry weight) soil (0 to 15 cm depth)].
Parenthetic value is coefficient of variation (CV = standard deviation/mean).

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 413

most of the plutonium in our study areas is externally deposited on plant surfaces.
Information supporting this conclusion includes:

• The high plant /soil plutonium concentration ratios compared to greenhouse studies.

• The obvious presence of soil in vegetation samples.

In addition, other investigators have shown that some of the plutonium associated with
native vegetation samples can be removed by a wasli treatment (Alldredge, Arthur, and
Hiatt, 1977).

Rodents

Rodent-Soil Relationships. Plutonium in internal organs (i.e.. liver, bone, and muscle)
of rodents sampled within our study areas generally could not be measured. However,
concentrations of plutonium in pelt and Gl tract samples were readily measured and were

TABLE 7 Inventory of Plutonium in Small
Mammal Tissues from Mortandad Canyon

Percent of total

Total

plutonium.*

Percent total

Tissue

body we

ight

pCi/g

plutonium

Pelt

0.85

50.0

GI tract

46.0

Lung

0.034

0.02

Liver

0.035

0.5

Carcass

0.018

2.8

* Based on six pooled samples.

directly correlated with levels in the study area soils (r^ = 0.90). Over 95% of the
plutonium body burdens in rodents was associated with these two tissues, as shown by
the data for Mortandad Canyon in Table 7. Thus we conclude that, in our study areas,
physical and biological processes (i.e., contamination of the pelt or ingestion of soil)
dominate in the transport of plutonium to rodents.

Plutonium Inventories

The fractional distribution of plutonium in Los Alamos and Trinity ecosystem
components (Table 8) is based on quantitative estimates of ecosystem component mass
(grams per square meter) and corresponding plutonium concentrations (picocuries per
square meter) in those compartments. The distribution of plutonium among five
components was generally quite similar between sites in that over 99% of the plutonium
was associated with soil and less than 1% with biota. Live vegetation contained 10"^ to
10~*% of the plutonium inventory. We conclude that very little of the environmental
plutonium present in our study areas has appeared in the biological components of the
ecosystem even after 30 yr of exposure. These results are essentially the same as those
observed at Rocky Flats and Oak Ridge (Little, 1976; Dahlman, Garten, and Hakonson,
this volume).

414 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Plutonium Inventory Ratios for Some Components of Los Alamos
and Trinity Study Areas in New Mexico

Plutonium inven

tory ratio*

Los

Alamos

Trinity

Component*

Mortandad

Canyon

Acid

-Pueblo Canyon

Area GZ

Area 21

Grass

4.1 )

< 10-

'(0.90)

5.6

X 10"

""(1.6)

2.0 X

10"

-'(0.99)

1.3x 10"

"(0.76)

Forb

4.8 X 10-

'(1.2)

1.7

X 10-

""(1.4)

1.7 X

10"

""(1.0)

3.5 X 10"

-=(0.77)

Litter

1.6 X

10"

""(2.0)

1.1 X 10"

-"(0.81)

Rodents

1.5 X 10-

'(0.77)

4.5

X 10"

"'"(0.99)

3.7 X

10"

-'(1.7)

2.3 X 10"

-'(0.47)

Soil

0.99(0.00009)

0.99(0.001)

0.99(0.0003)

0.99(0.00008)

*Inventory ratio = (pCi Pu/m^ in component)/(total pCi Pu/mM. All plutonium values are "''24opy except
Mortandad Canyon which is " ' Pu ; parenthetic value is coefficient of variation (CV = standard deviation/mean ).

Tlie relative inventory of plutonium within all our study ecosystems is governed
primarily by component mass relationsliips since differences in mass of the various
ecosystem components are much greater than the differences in plutonium concentrations
between the same components. The data in Table 9 demonstrate that mass inventory
ratios for Mortandad Canyon provide a good approximation of the plutonium inventory
ratio.

TABLE 9 Mass and Plutonium Inventory Ratios
in Mortandad Canyon at Los Alamos

Component

Mass inventory

Plutonium

Component

mass, g/m^

ratio

inventory ratio*

Sou

(0 to 15 cm)

2.3 X 10=

0.999

Grass

20.0

9.0 X 10- =

4.1 X 10"'

Forb

10.0

4.4 X 10"'

4.8 X 10"'

Rodents

0.03

1.3 X 10""

1.5 X 10"'

*Data from Table 8.

Plutonium Transport

Soils

Rainstorm runoff in the intermittent streams receiving wastes was identified over 30 yr
ago in the environmental transport of plutonium (Kingsley, 1947). Additional studies
were begun to determine the relationships of rainfall, runoff, and suspended sediments
with radionuchde transport (Purtymun, 1974; Purtymun, Johnson, and John, 1966;
Hakonson, Nyhan, and Purtymun, 1976).

Results of these studies demonstrate that runoff from snow melt and summer
rainstorms serves as a radionucUde transport vector in Los Alamos intermittent streams
and that the magnitude of this transport is higlily dependent on the hydrologic

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 415

characteristics of the watershed and the intensity of runoff flow (Purtymun, 1974;
Hakonson, Nyhan, and Purtymun, 1976). The dependency of concentrations of
suspended sediments and plutonium in runoff on flow rate is indicated in Fig. 3 for one
rainstorm runoff event in Mortandad Canyon. Tlie nonlinearity in the curve is due to the
relationship of flow rate with the particle size of resuspended material. At flows less than
0.25 m^/sec, only the silt— clay size materials were in suspension in the runoff. However,
at flows greater than 0.25 m^/sec, coarser sands containing most of the sediment
plutonium inventory (Table 5) were resuspended, which resulted in increased suspended
sediment and radionucUde concentrations. High flow rates typically occur during the
early phases of runoff events at Los Alamos owing to the intense nature and short
duration of area rainstorms. We found that nearly 80% of the sediment and 70% of the
radioactivity was transported within the first half of such events.

Additionally, there was a highly significant (P < 0.01 ) relationship between sus-
pended sediment and radionuclide concentrations in runoff water. About 99% of the
radioactivity in runoff was associated with suspended sediments greater than 0.45 /um in
diameter, whereas only 1% of the radioactivity in the liquid phase was associated with
sediments less than 0.45 jum in diameter.

6600

0.10

0.15 0.20 0.25

FLOW RATE, m^/sec

0.35

Fig. 3 Concentration of sediment and radioactivity in unfiltered runoff water from
Mortandad Canyon as a function of runoff flow rate.

416 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Studies were recently begun on wind transport of plutonium in the Trinity fallout
zone, where evidence of wind erosion of soil is readily apparent. Although these studies
are not complete, several important observations have been made. First, soil flux by
surface creep and saltation processes is highly seasonal and has peaked in the months of
July and August for two consecutive years of observation. Second, soil particle size
analyses on dust-collector samples show major differences in the amount of silt -clay
material between study sites. About half the dust material at Area 21 is in the silt— clay
size range, whereas less than 1% of collected dust at Area GZ is in the silt— clay size range.
These differences become important when coupled with the plutonium concentrations in
the various soil particle size fractions (from Table 5). For example, silt— clay material in
dust collectors at Area 21 contains over 200 times as much plutonium as the silt— clay
fraction of dust samples at Area GZ.

Summary and Conclusions

Despite differences in ecosystems and plutonium source, there are several similarities in
plutonium distribution between the Los Alamos and Trinity study areas. First, the
soils— sediment component contains virtually all the plutonium, with vegetation and
rodents containing less than 0.1% of the total. Plutonium has penetrated to considerable
soil depths at both locations, although it has occurred much more rapidly and to a greater
degree in the alluvial soil at Los Alamos than in the arid terrestrial soils at Trinity. At
both locations less than 50% of soil column plutonium inventories was found in the
surface 2.5 cm.

The plutonium penetration depth appears to correspond to the moisture penetration
depth in the Trinity fallout zone. This is probably the governing factor at Los Alamos,
although storm runoff and accompanying turbulent mixing complicate the process. In
Acid— Pueblo Canyon, the bulk of the soil column inventory lies in the lower profiles, an
indication of the loss of plutonium from surface layers due to sediment transport.

The plutonium in most cases was associated with relatively coarse soil size fractions.
The silt— clay (<53 /jm) fraction contained relatively little (<15%) of the plutonium, a
reflection of the small amounts of this size fraction in study area soils. An exception was
in Area 21 at Trinity, where the <53-iLtm soil size fraction contained about 73% of soil
plutonium inventories. The importance of these distributional differences stems from the
fact that silt— clay soil particles can be transported farther and are more likely to adhere
to biological surfaces than larger size fractions.

Concentrations in herbaceous ground vegetation were generally related to those in
soils from all sites. Our data strongly indicate that external contamination of plant
surfaces is the major soil-to-plant transport mechanism in these arid systems. The
plutonium concentrations in pelt and Gl tissues were related to corresponding soil
concentrations at all sites. Over 95% of the plutonium body burden in rodents was
associated with pelt and GI tract samples, an indication of the dominance of physical
and/or biological processes as the contaminating mechanism.

Horizontal transport in both areas is dominated by wind- and water-driven processes.
At Los Alamos surface runoff water governs the downstream transport of plutonium;
indications are that wind is a relatively more important transport vector at Trinity,

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 417

although splash-up of soil by raindrops may be an important transport mechanism in
these arid, sparsely vegetated study locations.

There was no evidence for a trophic-level increase of plutonium from soil to
vegetation to rodents. We believe that rodents come into contact with environmental
plutonium directly from the soil and to a lesser extent througli a food-web intermediary.

Research Needs

The importance of the soils component as a receptor of plutonium released to the Los
Alamos and Trinity Site study areas coupled with the direct role these soils play in
contamination of biota emphasizes the importance of understanding soil formation and
transport processes. Factors governing these processes will be instrumental in determining
plutonium distribution and transport as a function of time. Hydrologic and wind
transport processes discriminate against certain soil particle sizes; therefore studies on the
relationsliip of plutonium to soil separates will be useful in evaluating the potential
importance of a transport pathway and the resuhing hazard. We know, for example, that
wind transport of silt -clay material at Area GZ, Trinity Site, would represent a relatively
smaller inhalation hazard than corresponding transport at Area 21 simply because the
silt— clay fraction of Area GZ soil contains very little of the plutonium inventory.

Factors affecting migration of plutonium into the soil profile require understanding
since depletion of plutonium from the soil surface will likely reduce the horizontal
transport potential and may alter the availability of the element to vegetation.

Field studies should be conducted to quantify the relative importance of the root
pathway for contaminating vegetation to serve as a basis for judging changes in
physiological availability of environmental plutonium with time. As yet few field studies
have been able to show conclusively the relative importance of the two contamination
mechanisms.

in our opinion studies should be continued on the availability of plutonium to native
animals in our study ecosystems; however, on the basis of present concentrations and the
high variability associated with these measurements, we believe that the frequency of
sampling should be drastically reduced. Perhaps sampling at intervals of 5 to 10 yr would
be adequate to judge whether significant changes in plutonium availability have occurred.

Acknowledgments

We wish to recognize the following individuals for their valuable contributions to this
research: J. L. Martinez, G. Trujillo, E. Trujillo, K. Bostick, T. Schofield, G. Martinez,
K. Baig, P. Baldwin, R. Peters, W. Schwietzer, and S. Lombard. We also wish to thank
R. 0. Gilbert, D. Adriano, J. Corey, and G. Matlack for their efforts in reviewing this
manuscript. This research was performed under U. S. Department of Energy contract No.
W-7405-ENG-36.

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of America, Special Publications: No. 1, C. E. Gushing, Jr. (Ed.), pp. 49-57, Halsted Press, New

York.
, F. R. Miera, Jr., and R. E. Neher, 1976b, The Distribution of Plutonium in Trinity Site Soils

After 28 Years,/. Environ. Qual, 5: 431-437.
, F. R. Miera, Jr., and R. J. Peters, 1976c, Distribution of Plutonium in the Soil Particle Size

Lractions in Liquid Effluent-Receiving Areas at Los Alamos,/ Environ. Qual., 5: 50-56.
Olafson, J. H., H. Nishita, and K. H. Larson, 1957, The Distribution of Plutonium in the Soils of

Central and Northeastern New Mexico As a Result of the Atomic Bomb Test of July 16, 1945,

USAEC Report UCLA-406, pp. 1-25, University of California at Los Angeles, NTIS.
Price, K. R., 1973, A Review of Transuranic Elements in Soils, Plants and Animals,/ Environ. Qual.,

2: 62-66.

ECOLOGICAL RELATIONSHIPS OF PLUTONIUM 419

Purtymun, W . D., 1974, Storm Runoff and Transport of Radionuclides '■' DP Canyon. Los Alamos

County, New Mexico, USAEC Report LA-5744. Los Alamos Scientific i^aboratory, NTIS.
, G. L. Johnson, and E. C. John, 1966, Distribution of Radioactivity in the Alluvium of a Disposal

Area at Los Alamos, New Mexico, U. S. Geological Survey, Professional Paper 550-D,

pp. D250-D252.
Romney, E. M., 1977, Plutonium Contamination of Vegetation in Dusty Field Environments, in

Transuranics in Natural Environments, M. G. White and P. B. Dunaway (Eds.), USAEC Report

NVO-178, pp. 287-302, Nevada Operations Office, NTIS.
, and J. J. Davis, 1972, Ecological Aspects of Plutonium Dissemination in the Environment, Health

Phys., 22: 551.
White, G. C, and T. E. Hakonson, 1978, Statistical Considerations and Survey of Plutonium

Concentration Variability in Some Terrestrial Ecosystem Components, /. Environ. Qual, 8:

176-182.
Wilson, D. O., and J. F. Cline, 1966, Removal of Plutonium-239, Tungsten-185, and Lead-210 from

'S,o'\[s, Nature, 209: 941-942.

Plutonium in a Grassland Ecosystem

CRAIG A. LITTLE

Ttiis chapter is primarily concerned with plutonium contamination of grassland at the
U. S. Department of Energy Rocky Flats plant, which is located northwest of Denver,
Colo. Major topics include the definition of major plutonium-containing ecosystem
compartments; the relative amounts in those compartments: whether or not the
predominant isotopes, ^^^Pu and ^^^Pu, behaved differently; and what mechanisms
might have allowed for the observed patterns of contamination.

Samples of soil, Utter, vegetation, arthropods, and small mammals were collected for
plutonium analysis and mass determination. Small aliquot s (5 g or less) were analyzed by
a rapid scintillation technique and by alpha spectrometry.

Of the compartments sampled, greater than 99% of the total plutonium was
contained in the soil. The concentrations of plutonium in soil were significantly inversely
correlated with distance from the contamination source, depth of sample, and particle
size of the sieved soil samples. The soil data suggested that the distribution of
contamination largely resulted from physical transport processes.

Concentrations of plutonium in litter and vegetation were inversely correlated to
distance from the source and directly correlated to soil concentrations at the same
location. Comparatively high concentration ratios of vegetation to soil suggested wind
resuspension of contamination as an important transport mechanism.

Arthropod and small-mammal tissue samples were highly skewed, kurtotic, and quite
variable. Plutonium concentrations were lower in bone than in other tissues. Hide,
gastrointestinal tract, and lung were generally not higher in plutonium concentration than
kidney, liver, and muscle. All data tended to indicate that physical transport processes
were the most important.

Median isotopic ratios of ^ ^ ^ Pu to^^ ^Pu by activity concentration in soil were 40 to
50. Litter and vegetation isotopic ratios were similar to those of soil. Arthropod and
small-mammal isotopic ratios were lower than those of soil, which implied that the two
isotopes were differentially incorporated into the animal bodies and ^^^Pu was taken up
at a higher rate. However, further investigations suggested that statistical bias may have
spuriously contributed to the lower isotopic ratios in small animals.

Most of the world's agriculture occurs on land that, before tilling, was once covered by
stands of grasses. An important untilled tract of land contaminated with plutonium* is
the grassland immediately adjacent to and contained within the Rocky Flats plutonium
processing plant and associated buffer zone about 12 km northwest of Denver, Colo.,
metropohtan area. Because Rocky Flats is a prime example of plutonium-contaminated

*The word "plutonium" indicates ^ ^ ' '^ '^ " Pu in this chapter, unless otherwise noted.

420

PLUTONIUM IN A GRASSLAND ECOSYSTEM 421

grassland, this chapter will dwell primarily on data from environmental sampling at
Rocky Flats.

The Rocky Flats installation uses nearly 30 km^ as a buffer zone to separate the
pubhc from plutonium-handUng operations. The climate at the installation is typified by
occasional strong WNW winds exceeding 40 m/sec and moderate precipitation, i.e., 40
cm/yr average (Rocky Flats 1975 annual weather summary, unpublished). The Rocky
Flats grassland has been modified by the activities of humans and includes plant species
typical of short-grass plains (Buuteloua gracilis and Biichloe dactyloides) as well as
tall-grass prairie (Agropyron spp. and Andropogon spp.) and ponderosa pine (Pimts
ponderosa) woodland (Web'er. Kunkel, and Shultz, 1974). Mule deer {Odocoileus
hemionus) are found on the site along with grassland species of reptiles, rodents, and
birds (Whicker, 1974).

Source of the Contamination

Investigations by Krey and Hardy (1970) of DOE's Environmental Measurements
Laboratory (EML, formerly Health and Safety Laboratory) suggested that the most likely
contamination source was a storage area of stacked 55 -gal barrels that leaked
plutonium-poUuted oil. Data supporting the conclusion of Krey and Hardy and a
description of the nature of the stored oil— plutonium mixture are delineated.

Air-sampling data from Rocky Flats link the barrel storage area to the east -southeast
contamination pattern discovered by Krey and Hardy (1970). Air-samphng station S-8,
one of many such stations maintained and sampled regularly by Rocky Flats personnel, is
located about 75 m east and slightly south of the barrel storage area. Except for a brief
period during 1961, montlily averages of daily airborne contamination values have been
kept since at least 1960 to the present (Fig. 1).

The S-8 air-sampling data indicated that contamination peaks in the air were
associated with dates of perturbation of the contaminated surface (Table 1 and Fig. 1).
Except for periods of disturbance, the gross alpha concentrations in ambient air were near
0.01 pCi/m^ . However, during excavation and paving of the barrel storage area, the alpha
concentration in air markedly increased (Table 1).

The plutonium-contaminated cutting oil, about as viscous as lightweight motor oil but
thinned by carbon tetrachloride, was stored in the 55-gal barrels for periods of up to 7 yr.
The interactions between the oil, air, CCU , and plutonium within the barrels were
probably quite important in determining the eventual fate of the element.

The oil was filtered through 2- to 3-jum filters before being placed in the barrels. The
discard Umit at the time of storing was 1 x 10~^ g of plutonium per liter of oil. If the
limit and the filtering had been observed and performed faithfully, each of the
approximately 3570 plutonium-containing barrels would have had no more than 2.1 g of
plutonium (0.13 Ci) (M. A. Thompson, Environmental Sciences, Rocky Flats Plant, and
F. J. Miner, Chemical Resources, Rocky Flats Plant, personal communications).

It is difficult to assess what occurred once the oil was inside the barrels. The presence
of carbon tetrachloride in the drums allows the possibility that hydrochloric acid was
formed, which, in turn, may have reacted with the plutonium metal to form very low
concentrations of plutonium chloride, a more-soluble form of the element (J.M.
Cleveland, Environmental Studies, Rocky Flats Plant, personal communication). This
possibility is given credibility by the work of J. Navratil of Rocky Flats Chemical
Research Division, who has studied contaminated cutting oil in recent years.

422 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

1966 I 1967
DATE

Fig. 1 Monthly means of daily gross alpha activity in ambient air at station S-8 (75 m
east of the oil-barrel storage area). Aliquots of Gelman AH filter material were counted
in a gas-flow proportional detector. Data adapted from D. C. Hunt, Environmental
Sciences, Rockwell International, personal communication.

TABLE 1 Total Monthly Gross Alpha Activity in Ambient Air at Station S-8

(75 m East of Oil-Barrel Storage Area) During Disturbances

of the Storage-Area Surface*!

Dates

Event

Alpha activity,
pCi/m'

7/59-9/63 No large-scale leaking.

1/64-1/65 Large-scale leaking.

1/65 Contaminated soil covered with fill.

1/66 Small building added to filter contaminated oil from

leaking to new drums.

1/67 Drum-removal activity begun.

6/68 Last drums removed but high winds spread some activity.

2/69 Weeds burned and area graded for paving.

9/69 Asphalt pad completed.

11/69 Four sampling wells dug through pad.

4/71 Drainage ditch dug on west side of asphalt pad.

0.009
0.025
0.01
0.014

0.038

0.188

0.34

0.013

0.033

*Air-filter material was counted directly in a gas-flow proportional counter.

fAdapted from D. C. Hunt, Environmental Sciences, Rockwell International, personal

communication.

PLUTONIUM IN A GRASSLAND ECOSYSTEM 423

Navratil and Baldwin (1976) found that filtering the contaminated oil through a
O.Ol-jum filter removed only about 50% of the plutonium. This result strongly suggested
that about half the plutonium was in a relatively large particulate form whereas the other
half was associated with very small particles. Fission-track analysis of the filtered oil
confirmed that the remaining plutonium was monomeric. It is doubtful that the barrels
consistently held the above proportions of particulate and nonparticulate plutonium
oxide, but each probably contained some plutonium cliloride.

J. M. Cleveland (personal communication) also suggested that the filtered 3-ium
plutonium particles might combine to form larger aggregates of the metal. Of course, the
size and binding tenacity of 1,hese conglomerates are unknown.

Methods

Two macroplots were chosen for intensive sampling of plutonium in soil, vegetation, and
litter. The locations of these macroplots relative to the supposed plutonium source, the
barrel storage area, and the prevailing wind are shown in Fig. 2. A sampling grid was
superimposed over each macroplot. The macroplot 1 grid was approximately 0.75 ha

1 km

ACROPLOT 1

ASPHALT BARREL
STORAGE PAD

SECURITY
FENCE

Fig. 2 Schematic representation of the southeast comer of the Rocky Flats Plant
indicating the location of study macroplots and sampling transects. The wind rose
indicates the relative magnitudes of wind velocities during 1974.

424 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

(1 ha = 2.5 acres) and contained 100 sampling markers. Plutonium data in this chapter are
from macroplot 1 unless specifically noted otherwise.

Depth-profile soil samples were taken by hand with a trowel. After vegetation and
litter had been clipped and bagged separately, four 5- by 5- by 3-cm samples were
removed and bagged separately for each of seven 3-cm-depth layers to 21 cm deep. If
rocks precluded sampling at a given depth, the column was resumed below the blockage.
Soil samples were transported to the laboratory and air-dried. Rocks or debris greater
than 0.5 cm in diameter were removed from the sample. After oven-drying and weighing,
samples were mechanically shaken on brass soil sieves. The accumulation on each sieve
was weighed and placed in a small paper envelope, and the envelope was sealed with tape.

Litter and standing vegetation were sampled from 0.25-m^ and 0.5-m^ areas,
respectively. Vegetation was clipped and bagged, and litter was gathered by hand and
bagged. In the laboratory litter and vegetation samples were air-dried and weighed. Soil
was separated from the Utter samples by a flotation process (Little, 1976). The net
vegetation or the litter dry weiglit was divided by the microplot size, 0.5 m^ or 0.25 m^ ,
respectively, to calculate mass per unit area. For plutonium analysis vegetation and Utter
samples were ground on a Wiley mill with an SSO-jLim opening screen, and 5-g aliquots
were taken.

Arthropods were sampled by a combination of sweep netting, pitfall trapping, and
drop trapping at random sites on established grids. At the laboratory animals were
separated into generic groups that were weighed separately. These generic totals were
combined for an estimated weight per 0.5-m^ microplot. Arthropods obtained by the
drop-trap method were not analyzed for plutonium owing to fear of cross contamination
from soil during the vacuuming process. Samples for plutonium analysis and a species
inventory list resulted from the sweep netting and pitfall traps. For plutonium analysis a
representative composite was analyzed for each sampling period. Arthropods were not
cleaned prior to plutonium analysis.

Small-mammal trapping grids were superimposed over each macroplot in a manner
resembUng that used by the U. S. International Biological Program Grassland Biome
(Packard, 1971). Animals were trapped about six times yearly. Sherman live traps were
used for cricetid and sciurid rodents. Geomyid rodents were trapped less regularly in
homemade live traps placed in burrows systems. Approximately 15% of the estimated
population was removed from each macroplot during each trapping session. These animals
were coUected by removing dead-in-t^ap individuals during the regular session and the
remainder randomly in one extra trapping night. Small mammals were either dissected
immediately or frozen for dissection later. Special precautions were taken during
dissection to minimize cross contamination between tissues. Approximately 10 cm^ of
hide was used as an aUquot. Lungs, liver, and gastrointestinal (GI) tract were removed
intact. Muscle samples were taken from the legs in most cases. Bone samples consisted of
the whole skeleton, which had been cleaned of flesh by a dermestid beetle colony. All
samples except bone were placed on tared, ashless filter papers, oven-dried at 50 to 60°C,
and weighed. The sample was placed in a snap-cap vial for storage or transport to a
commercial laboratory.

Some soil-sample aliquots (5 g) were analyzed for plutonium content by commercial
laboratories (LFE, Richmond, Calif., and Eberline Instrument Corp., Albuquerque,
N. Mex.). Most soil samples were analyzed in our laboratory, as were most litter and
vegetation samples. Small-mammal tissues and arthropods were commerciaUy analyzed.
The LFE method used concentrated hydrofluoric acid to dissolve samples (Wessman

PLUTONIUM IN A GRASSLAND ECOSYSTEM 425

et al., 1971); Eberline modified a pyrosulfate fusion technique for the same purpose (Sill,
1969). Ion-exchange columns removed interfering elements and isolated plutonium from
the samples before alpha spectrometry analysis. Chemical recovery was measured by
adding ^^^Pu tracer to each sample. Agreement between homogenized split samples sent
to these laboratories was good (Little, 1976). In our laboratory a procedure was used that
incorporated harsh digestion of the sample by nitric and hydrofluoric acids, ion exchange,
organic extraction, and liquid scintillation spectrometry (Little, 1976). This method had
an estimated minimum detectable activity of 0.42 pCi (P < 0.05). Plutonium data in this
chapter are 2 39,240p^ unless otherwise noted. Plutonium-240 contributed about 20% on
the average to the alpha activity of ^^^'^'*°Pu.

Plutonium Compartmentalization

The inventories of plutonium in the principal compartments of the grassland ecosystem
were calculated. Compartments investigated were soil, in 3-cm increments from 0 to
21 cm, litter, standing vegetation, arthropods, and small mammals.

The compartmental inventories of plutonium were calculated by multiplying the
mean mass of each compartment (g/m^) by the mean plutonium concentration of the
compartment (|UCi/g). A total ecosystem inventory was calculated by summing over all
compartments. The compartmental fraction (unitless) of the total plutonium inventory
was calculated by dividing each compartmental inventory (^Ci/m^) by the total inventory
(AfCi/m^).

The soil compartment had vastly the largest fraction of the total plutonium, 99.69%
(Table 2). As expected, the fraction of the total plutonium contained within a soil layer
decreased as the depth increased. The litter compartment comprised less than 1% of the
total plutonium (53 nCi/m^) in the study areas, and the vegetation represented only
about 0.01% of the total plutonium. By virtue of representing both low biomasses and
plutonium concentration, the animal compartments, arthropods and small mammals, had
extremely small fractions of the total plutonium, 6.8 X 10~^ and 33 x 10~^,
respectively.

In summary, the .compartmentalization data indicated that greater than 99% of the
plutonium in the study area was located in the soil. At the time of sampling, nearly
one-half (49.7%) the total plutonium was in the top 3 cm of soil. In decreasing order,
smaller plutonium-inventory fractions were found in Htter, vegetation, arthropods, and
small mammals. The implication of these results is that, in the present state, transport of
plutonium is closely Unked to soil movement or erosion. Therefore efforts to prevent
plutonium transport off contaminated grasslands should be directed primarily at
minimizing soil transport rather than mobihzation by biota.

Plutonium in Soil

Analysis of the soil plutonium data suggested that two primary generalizations about
plutonium in soil could be stated. First, the plutonium concentrations in the soil samples
were highly variable. Second, the plutonium concentration in a soil sample was a function
of sample location, sample depth, and the soil particle composition of the sample.
Rationales for both of these conclusions are examined in some detail.

Frequency distributions of plutonium in soil samples were positively skewed
(P < 0.05) with coefficients of variation (CV = standard deviation ^ mean) ranging to

426 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Distribution of ^ ^ ''Pu in Samples
from the Rocky Fiats Study Macroplot*

Compartment

Mean nf 95% confidence interval:]:

Plutonium concentrations, pCi/g

Soil, 0-3 cm

835 72

554-1120

Soil, 3-21 cm

105 309

69-141

Litter

412 29

314-509

Vegetation

28.6 76

15.7-41.4

Artiuopods§

5.48 23

3.13-7.83

Small mammals

6.50 304
Fraction of total plutonium

2.38-10.6

Soil, 0-3 cm

5.0x10-' 2.5 X

10-' -7.4 X 10-'

Soil, 3-21 cm

5.0x10"' 2.5 X

10-' -7.5 X 10"'

Litter

2.9x10"^ 1.6 x

10~'-4.2 X 10"'

Vegetation

1.0 X 10"" 4.1 X

10-5-1.6 X 10"^

Arthropods §

1.2 xlO"" 4.6 X

10-^-2.0 X 10-"

Small mammals

3.3 x 10-' 6.6 x 10-'" -6.0 x 10"'

*Compartmental ^^'Pu inventory (pCi/m^ ) equals mean biomass
[g(dry)/m^ ] times mean concentration [pCi/g(dry)] . Fraction of total
equals mean compartmental inventory (pCi/m^) divided by total
inventory.

■[Number of samples for which the mean is calculated: For
arthropods and vegetation n is the number of groups of individuals
analyzed; for small mammals n is the number of tissue samples, not
individual animals.

±95% confidence interval equals mean + (1.96 standard error of the
mean).

§Includes data from Bly (1977).

greater than 2.0. Although positive skewness is a characteristic of iognormal distributions,
the natural-log transformation of soil data did not result in normal distributions
(Kolmogorov— Smirnov one-sample test, P > 0.05) but did reduce the skewness for the
seven depth groups tested.

Three adjacent soil columns (5 by 5 cm) from a 5- by 15-cm area on macroplot 2
exhibited the extreme spatial variability that sometimes occurred in plutonium
concentrations in the soil. The mean plutonium concentrations in the 5-g aliquots from
each column were 480 (column A), 5.4 (column B), and 0.57 pCi/g (column C) at the 0-
to 3-cm depth. Virtually all the plutonium in column A was in the top 3 cm, the other
depths (in 3-cm increments to 21 cm) being at or near background. In columns B and C,
the majority of the plutonium was found at lower depths. No other cases of such extreme
spatial variation in soil plutonium concentrations were detected during the sampling at
Rocky Flats.

As expected, surface soil samples (0 to 3 cm) had a higher mean plutonium
concentration than subsurface samples (Table 3). This result agreed with data from
Rocky Flats soil sampling reported by Krey and Hardy (1970). Plutonium concentrations
were also a function of the size range of soil practices comprising the aliquot (Tables 3
and 4).

PLUTONIUM IN A GRASSLAND ECOSYSTEM 427

TABLE 3 Mean Plutonium Concentrations of Soil Samples from Rocky Flats

Soil particle
size range,

Plutonium concentration, pCi/g

p.xn

0-3 cm

3-6 cm

6-9 cm

-12 cm

-15 cm

15-18cm

18-21 cm

850-2000

740

140

5.4

1.8

425-850

460

120

100

7.0

5.5

250-425

440

130

5.7

180-250

460

120

130

8.9

150-180

770

130

100

140

6.5

75-150

870

210

100

45-75

1400

310

210

100

160

0-45

1500

180

810

190

220

0-2000

830

170

200

TABLE 4 Regression Parameters of Soil Plutonium Concentration

(pCi/g) Adjusted for the Sample Location as a Function of Soil

Particle Diameter at Various Depths*

Correlation

Significant

Depth, cm

Intercept (bg)

Slope (b,)

coefficient (r)

at a =

0-3

4.8

-0.336

-0.312

0.01

3-6

3.6

-0.270

-0.291

0.05

6-9

0.89

-0.753

-0.471

0.001

9-12

1.6

-0.544

-0.564

0.001

12-15

0.67

-0.799

-0.719

0.001

15-18

-0.21

-0.775

-0.706

0.001

18-21

-0.42

-0.572

-0.358

*The model

used was in

Pu = bf, +b, In

Least-squares regressions were calculated with linear, exponential, and power-function
models of plutonium concentration in surface soil as functions of the distance east or
south from the asphalt pad. The power-function model gave the best fits of the data for
both curves (Figs. 3 and 4). A t-test indicated that the slope of the distance-south curve
(Fig. 4) was steeper (P < 0.05) than that of the distance-east curve (Fig. 3). These results
conform to the concept of wind-distributed plutonium; the more effective, stronger
winds were to the east, and hence the slope of that curve was smaller.

Several multiple linear-regression models were calculated. The model that accounted
for the largest fraction of the total variance (0.868) had the following form:

In Pu = 1 1 .1 5 - 0.0535 In E - 1 .628 In S

where Pu is the plutonium concentration (pCi/g), E is the distance east of the asphalt pad
centerline (m), and S is the distance south of the asphalt pad centerline (m).

With this model plutonium concentrations oi samples in the soil depth profile were
adjusted to estimate the concentration expected at a common location. The adjusted
values were then regressed as a function of sample depth (Fig. 5). As with the distance

428 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

2 5 10^ 2 5 io3 2 5

DISTANCE EAST OF ASPHALT PAD (E), m

Fig. 3 Plutonium concentration in 0- to 3-cm-deep Rocky Flats macroplot 1 soil as a
function of distance east of the center of the asphalt oil-barrel storage pad.

relationships, a power-function regression model had the highest correlation of plutonium
concentration with depth of the models attempted and was significant (P < 0.01).

The relationship of plutonium concentration in soil as a function of soil particle
diameter (as represented by the opening of the final passage sieve) was examined for each
depth layer (Table 4). The model resulted in a significant (P < 0.05) regression for each
depth group except the 18- to 21 -cm group. The steepest slope (—0.799), at 12 to 15 cm,
was significantly different (P < 0.05) from the flattest slope (-0.270), at the 3- to 6-cm
level. However, there was no obvious trend in slope vs. soil particle-size curves with depth.
Because the amount of surface area represented by the soil particle spheres in a constant
mass of soil is inversely related to soil particle diameter, it followed that the plutonium
concentration in a soil sample should be inversely related to the surface area of the
particles in the sample. A tabulation of the fractions of the total soil-sample mass

PLUTONIUM IN A GRASSLAND ECOSYSTEM 429

1U-'

■ 1 1 1

■^~

■^

—

\
\
\

Cs =

2. OB

X 10'" X S-"

—

r =

-0.93

d.f. = 44
P < 0.01

— ^

,—

•

I •

^-"

Q-

LU 2

•

—

H 5

•

^^"

•

Z 2

)

^ 102

_^^

2
LU

U 5

^—

•\

J?.

•^

i 2

•

^ 10^

•\

10°

1 III

10^

10-^

IC*

DISTANCE SOUTH OF ASPHALT PAD (S), m

Fig. 4 Plutonium concentration in 0- to 3-cm-deep Rocky Flats macroplot 1 soil as a
function of distance south of the asphalt oil-barrel storage pad.

represented by each sieve size organized by depths did not produce any obvious patterns
with either depth or particle size range. Consequently regressions of the soil mass fraction
per sample as a function of depth were not significant for most sieve fractions. However,
these last results would not preclude a surface-attachment mechanism.

The data on plutonium in soil at Rocky Flats can be summarized by several
statements. First, the variance in the plutonium concentrations of the soil samples was
large; CV's within groups of like samples (same depth and particle size) ranged to over
2.0. Frequency distributions for soil samples were positively skewed. Spatial variation was
also large; in one instance the plutonium concentrations of aliquots taken less than 15 cm
apart varied by nearly three orders of magnitude.

Second, in spite of the large degree of variance in the data, the plutonium
concentrations in soil were significantly correlated with the location and soil particle

430 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10^

5 10^ 2 5 102 2

DEPTH (D), cm

5 10-^

Fig. 5 Rutonium concentration in Rocky Flats macroplot 1 soil as a function of depth
of sample. Sample concentrations adjusted for distance east and south of center of
asphalt oil-barrel storage pad.

composition of the soil sample. The spatial distribution of plutonium (i.e., more
plutonium downwind than downslope) implicated wind as the prime mechanism of
plutonium transport onto the studied areas. Such factors as resuspension with or without
added mechanical disturbances by humans or fauna undoubtedly contributed to the wind
transport of plutonium but to a presently unknown degree. The data also indicated that
plutonium was found to a depth of 21 cm in most samples from downwind of the barrel
storage area but that about two-thirds of the contamination was in the top 5 cm. The
relationship between plutonium concentration and soil particle size suggested a
surface-attachment mechanism of plutonium attacliment to soil particles. However, the
lack of any pattern of soil mass fraction with depth for the various particle sizes probably
indicates that plutonium transport with depth is not simply a case of transport of the
various plutonium— soil particles downward.

Plutonium in Plant Compartments

The vegetation community of the study area was composed mostly of grasses and
members of the sunflower family. Members of the sedge, pea, and mustard families were

PLUTONIUM IN A GRASSLAND ECOSYSTEM 431

TABLE 5 Plutonium Concentrations

in Rocky Flats Vegetation

and Litter Samples

Plutonium concentra-

tion,

, pCi/g

Coefficient

Mean

variation

Litter

412

0.65

Vegetation

28.6

2.02

also present but in much lower numbers of individuals. Rather than study numerous plant
types individually, two plant-derived compartments were studied, litter and detritus and
standing vegetation. Although these compartments accounted for only a small fraction of
the total plutonium (about 0.2%). the study of those compartments helped derive some
concepts of plutonium transport.

As with the soil, frequency distributions for vegetation samples were positively
skewed. Further, the hypothesis that plutonium concentrations in vegetation were
lognormally distributed could not be rejected (P > 0.05). Unexpectedly, the hypothesis
that plutonium concentrations in litter were normally distributed could not be rejected
(P> 0.05).

Mean plutonium concentrations in Utter were liigher than those in vegetation
(Table 5). Concentrations of plutonium in Utter and vegetation were each inversely
correlated with distance east or south from the asphah pad (P < 0.01).

The fact that litter had a higher mean concentration of plutonium than standing
vegetation is not surprising. This result reinforces the suggestion made above that soU
transport was the primary mechanism of plutonium transport.

Plutonium in Animal Compartments

Two animal compartments were studied, arthropods and small mammals. These
compartments together contained about 2 x 10~^ of the total plutonium estimated to be
in the studied areas. Nevertheless, the mobiUty of the animals makes them potential
transporters of plutonium, albeit relatively small amounts, off the site.

As expected from the soU and vegetative sampling, the frequency distributions of
plutonium concentrations in smaU mammals were positively skewed, as indicated by the
histogram in Fig. 6. Not only were there many samples that had plutonium concentra-
tions below the detection Umit but also much of the total activity was supplied by
relatively few samples. Frequency distributions of plutonium concentrations in arthro-
pods were also positively skewed (Bly, 1977). Bly (1977) further indicated that
logarithmic transformations were useful in aUeviating the skewness. Therefore the
plutonium concentrations of the arthropod samples were probably lognormaUy distrib-
uted.

Concentrations of plutonium in 23 groups of individual arthropods and in
small-mammal tissues were of comparable magnitude (Table 6). The smaU-mammal

432 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

0 1

2 3 4 5 6 7 8 9 10 11 12 15 16

MULTIPLES OF THE MEAN 239pu CONCENTRATION

Fig. 6 Representative histogram of small-mammal tissue samples from Rocky Flats.

TABLE 6 Mean Plutonium Concentrations

of Arthropods* and Small Mammals Sampled

from Rocky Flats Macroplot 1

Plutonium concentration,

pCi/g

Coefficient

Sample type

Mean

variation

Arthropods

5.48

1.05

Small mammals

Bone

0.288

2.28

GI tract

7.03

2.50

Hide

1.51

1.84

Kidney

13.6

4.39

Liver

8.38

5.45

Lung

3.57

1.90

Muscle

8.92

5.85

External tissues

134

3.88

2.76

Internal tissues

169

8.59

5.58

*Includes data adapted from Bly (1977).

PLUTONIUM IN A GRASSLAND ECOSYSTEM 433

tissue-sample means ranged from 0.288 pCi/g for bone to 13.6 pCi/g for kidney, and the
mean of whole arthropods was 5.48 pCi/g.

The patterns, or rather, lack of patterns, in the small-mammal data were puzzling. The
tissues were arbitrarily classed either external or internal, depending on whether or not
the tissue had a direct contact with the animal's environment. External tissues included
GI tract, hide, and lung; internal tissues included bone, kidney, Uver, and muscle. By
virtue of the supposed low biological availability of plutonium and the proximity of the
external tissues to the contaminated soil, external tissues were expected to have larger
plutonium concentrations than internal tissues. Inexplicably, this was not the case. The
three highest plutonium concentrations were found in internal tissues, i.e., kidney,
muscle, and liver; hide and lung comprised two of the three lowest means. Additionally,
the amount of variation in samples within a given tissue was quite high. The minimum
tissue variation was in hide samples (CV = 1.84), and the maximum was in muscle
(CV = 5.85).

Only two explanations for the high degree of variability are at hand. First, the
possibility of cross contamination always exists no matter how carefully one removes
tissues during dissection. Second, the extremely small sample mass of a few samples (a
dry kidney may be as small as 0.05 g) may have had a tendency to magnify the relative
plutonium concentrations. However, a plot of plutonium concentration in small mammals
vs. sample mass indicated that about as many samples had large mass and small plutonium
concentrations as had small mass and large concentrations. Beyond this, however, the
tendency for small-mass samples to skew the distribution has not been investigated.

The nonparametric Kruskal-Wallis technique (Siegel, 1956) was used to test whether
or not the seven tissue means were from the same population. The resulting chi-square
value of about 44 indicated that the difference between the tissue groups was highly
significant (P < 0.001). Although no test was performed, it was intuitively obvious that
the mean plutonium concentration of the bone samples (0.29 pCi/g, n = 28) was lower
than that of other tissues.

Plutonium Concentration Ratios

The concentration ratio (CR) is a potential indicator of plutonium redistribution by
wind, water, or plant uptake. Concentration ratio is defined as the concentration in
activity per unit mass or volume divided by the concentration of the same nuclide in the
same units in another material. In this section the CR will have 0- to 3-cm-deep soil as the
material in the denominator [e.g., CR of Utter = (pCi Pu/g litter)^ (pCi Pu/g 0- to
3-cm-deep soil)] .

The CR's of litter, vegetation, arthropods, and small mammals are listed in Table 7.
Litter had the largest CR followed in descending order by vegetation, small mammals, and
arthropods. Regressions of litter and vegetation CR's vs. distances east and south of the
asphalt pad did not achieve significant correlation coefficients (P > 0.05).

The plutonium concentrations in litter and in vegetation were plotted vs. soil
plutonium concentrations from the same locations. Only the litter curve is shown here
(Fig. 7). The litter regression was interesting because of its high correlation (r = 0.975)
and near-unit slope (1.001). Although the number of samples here was limited, the data
comprising Fig. 7 suggested that litter may be an excellent estimator of soil plutonium
concentration in the grassland. The regression of plutonium concentration in vegeta-

434 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Plutonium Concentration Ratios and
95% Confidence Intervals of Ecosystem Compartments
in Rocky Flats Macroplot 1 with 0- to 3-cm-Deep Soil*

Compartment Concentration ratio 95% confidence interval

Litter
Vegetation
Arthropods
Small mammals

4.9x10-'
3.4 X 10-'
6.8 X 10-'
7.8 X 10"'

2.9 X 10"' -7.0 X 10-'
1.5 X 10-' -5.4 X 10-'
3.1 X lO-'-l.l X 10-'
2.1 X 10-^-1.3 X 10-'

*Mean plutonium concentration in 0- to 3-cm-deep soil equals
835 pCi/g. Concentration ratio equals mean pCi/g compartment
divided by mean pCi/g in 0- to 3-cm-deep soil.

10'

I-

5 —

10-^ —

5 —

2 —

z
o

_J
a.

10'

— 1

1 1 1

Mil 1

1 1 1 nil 1

1 11 1 ■ -

—

= 1.09 C^°°^

= 0.97

—

d.f.

= 8

—

P < 0.01

• ~/-

^r ^

—

/ -_

;^-

• X

—

•

—

/•

—

1 1 1

Mill 1

\ \ Mini 1

1 1 lllll

5 —

2 —

10° 2

5 IqI 2

102 2

5 103

PLUTONIUM CONCENTRATION IN 0- TO 3-CM-DEEP SOIL (C,), pCi/g

Fig. 7 Plutonium concentration in litter vs. plutonium concentration in soil at the same
sample location.

PLUTONIUM IN A GRASSLAND ECOSYSTEM 435

tion vs. plutonium concentration in underlying soil was also statistically significant
(P < 0.01) but was less conclusive than the litter vs. soil curve and is not shown here.

The CR's of the vegetation were higher than those produced in greenhouse studies.
Typically, uptake of plutonium under laboratory conditions has been on the order of
10"^ to lO"'* of the soil concentrations (Newbould, 1963; Wilson and Cline, 1966;
Romney, Mork, and Larson, 1970; Schulz, Tompkins, and Babcock, 1976). The Rocky
Flats CR of 3.4 X lO"'^ suggests either increased root uptake by grassland species or
another method of contamination, such as aerial deposition of resuspended soil particles.
The high surface-to-volume ratio of grasses and the hairy nature of the leaves of many
members of the sunflower family would be amenable to a high rate of impaction and
attachment of small soil particles. Given wind-redistributed plutonium at Rocky Flats,
surficial attachment of contaminated soil particles to plants is the likely mechanism of
contaminating the vegetation.

Plutonium Isotopic Ratios

Ratios of plutonium isotopes or ratios of ^^^Pu and "^' Am have been reported from
several sites (Emeiy et al., 1976; Gilbert et al., 1975; Hakonson and Jolinson, 1974;
Markham, 1976). In the hope that the examination of the isotopic ratios of ^^^Pu and
^^^Pu in the grassland would give some insight into the relative ecological availability of
these two nuclides, isotopic ratios were calculated for samples analyzed by alpha
spectrometry [isotopic ratio (IR) = ^^^Pu pCi/g of sample -^ ^ ^ ^ Pu pCi/g of same
sample] . Ratios were not calculated for samples where either isotope was below the
detection limit. Ratios were tabulated according to sample type and tested for goodness
of fit to a normal distribution. The distribution of IRin the various soil depths was either
lognormal or marginally normal. Small-mammal tissues appeared to be lognormal with
respect to the IR.

As suggested by Doctor and Gilbert (1977), the concentration of ^^^Pu was plotted
vs. ■^ ^ ^ Pu for each of the seventeen sample types. Seventy percent of these groups exhibit
a zero intercept, based on a t-test. These results implied that the ratios were constant
within the tested sample groups and that ^^^Pu/'^^^Pu would be an unbiased estimator.
However, because of the likelihood that the IR is lognormally distributed within most of
the sample groups, median IR's are reported here (Table 8). The R4 method of
calculation discussed by Doctor and Gilbert (1977) was used to calculate these values.

At first glance the median isotopic ratio in soil appeared to decrease as depth
increased. However, the overlapping 95% confidence intervals for the listed medians
suggested that the ratio is relatively constant. As expected, neither linear, exponential,
nor power-function regressions of the raw IR data vs. soil depth were statistically
significant (P > 0.05).

Despite the limited number of htter and vegetation samples analyzed for both ^^ Pu
and ^"^"Pu, the median IR's of these two compartments were very similar to IR's of the
soil. These results tended to indicate that the litter and vegetation were closely linked to
the soil.

The IR's in the animal compartments raised some very interesting questions (Table 8).
Only two sample types (GI tract and muscle) exhibited 95% confidence intervals that
overlapped with soil IR confidence intervals. Therefore, it appeared that the small-
mammal and arthropod compartments had lower IR's than soil. A lower IR would imply

436 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Median Isotopic Ratios

(" ^ Pu pCi/g ^ " « Pu pCi/g) in

Rocky Flats Environmental Samples*

Isotopic ratiof

95% confidence

Compartment

Median

interval

Soil depth, cm

0-3

54.70

37.04-80.78

3-6

43.70

35.09-54.43

6-9

46.41

35.23-61.14

9-12

46.45

40.67-53.06

12-15

48.76

42.81-55.54

15-18

46.94

32.27-68.26

18-21

38.67

34.16-43.78

Litter

55.47

51.62-59.61

Vegetation

59.98

39.92-90.12

Arthropods

9.88

5.69-17.15

SmaU-mammai tissues

Bone

7.49

2.99-18.71

GI tract

24.82

17.18-35.90

Hide

19.94

13.99-28.43

Kidney

11.07

3.98-30.80

Liver

17.55

11.62-26.50

Lung

7.42

3.97-13.87

Muscle

13.20

4.66-37.41

*Only data in which both ^^'Pu and ^"*Pu were
above detectable Umits were included.

tThe median and confidence limits were calculated by
method R^ of Doctor and Gilbert (1977).

relatively enhanced assimilation of ^^^Pu, compared to ■^^^Pu,into these compartments
than into soil.

Obviously, there are some physical reasons for skepticism regarding data wliich
suggest that two isotopes of the same element behave differently in biological systems.
The difference in mass between '^■'^Pu and ^^^Pu is less than that between ^^"^U and
^^*U, on which millions of dollars have been spent for enrichment. Alpha-recoil energy
from ^^^Pu and ■^^^Pu could displace other atoms from near the surface of a particle of
plutonium metal. However, unless the particle is composed of either pure ^^^Pu or pure
'^^Pu, there would probably be no preferential displacement of either nuclide relative to
their ratio in the original metal. Rocky Flats plutonium metal probably did not contain
either pure ^ -^ ^ Pu or ^ ■^ ^ Pu particles .

However, if a particle of pure "•'^Pu were in some way introduced into an organism,
autoradiolysis by this high-specific-activity nuclide might allow relatively fast biological
transport compared to ^•'^Pu. This idea is not unprecedented. Rats that inhaled ^^*Pu02
and ^"'^Pu02 of the same particle size and crystalline form translocated up to seven times
as much ^^^Pu as ^"'^Pu to systemic organs at times up to a year postinhalation (Stuart,
1970). Ballou et al. (1973) allowed rats and beagle dogs to inhale Pu02 aerosols of

PLUTONIUM IN A GRASSLAND ECOSYSTEM 437

identical size and preparation. According to these workers, "The much greater
translocation of ^^^Pu ... suggests that solubilization of the "^^^PuOt occurs to a
significant degree within the dog ... ."

The previous two paragraphs do little to help explain the animal IRdata. A possible
explanation may be had in statistical bias that heretofore has gone undetected. Basically,
the bias has to do with the fact that both ^^^Pu and ^^*Pu are probably lognormally
distributed in environmental compartments. Therefore the ratio of ^^^Pu to •^^^Pu
should also be lognormally distributed (Aitchison and Brown, 1969, p. 11). Unfortu-
nately, the distribution of both ^^^Pu and ^^^Pu was censored; i.e., some proportion of
the data points was below a detectable limit (Aitchison and Brown, 1969). Shaeffer and
Little (1978) have shown that both the mean ratio and the variance of the ratio of two
censored lognormal variates will be decreased relative to ratios of uncensored variates if
the denominator (^^^Pu) has a lower magnitude than the numerator (^"'^Pu). The
magnitude of the decrease in mean ratio and variance is influenced by the relative
closeness to the detection limit of the two variates.

This appears to be essentially the case with the IR data presented herein. The soil,
vegetation, and litter compartments had relatively high plutonium concentrations and
also relatively large IR"s. As the plutonium concentration began to approach the
detection limit, e.g., in arthropods and small mammals, the IR also decreased. Therefore,
if the censoring is large, an estimate of the mean or median of the uncensored ratios will
be in error because of the effect of censoring.

A solution for the problem of ratios of two censored distributions is to try to
estimate the population parameters for each distribution and then use method R2 , i.e.,
mean ratio equals mean ^^^Pu divided by mean '^^Pu, as suggested by Doctor and
Gilbert (1977). Kushner (1976) discusses two methods of estimating such parameters.

Lognormality was assumed, and the methods of Hald (1949) as modified by Kushner
(1976) were used to calculate population parameters. Then, a method of Aitchison and
Brown (1969, p. 45) was used to calculate the "minimum variance unbiased estimator" of
the arithmetic mean isotopic ratio tor hide. The mean ratio of hide by these methods was
found to be 37. The median ratio published in this chapter was 20, and the mean ratio
calculated by summing all hide ratios and dividing by the number of ratios (method R3 in
Doctor and Gilbert, 1977) was 29. Therefore, although no confidence interval was
calculated, the mean IR in hide calculated by Kushner's (1976) method would be little
different from the mean IR in soil. Unfortunately, some of the small-mammal tissue data
are censored to such a degree that some of the functional values are extreme enough that
they were not tabulated by Hald (1949), one of Kushner's (1976) prime references.
Therefore the parameters of most of the censored small-mammal data cannot be
estimated by the methods of Kushner (1976) and Hald (1949).

In summary, the median IR was constant in soil and vegetation compartments.
However, the median IR's also suggest that *^'^^Pu is preferentially mobile in animal
compartments of the grassland relative to '^^^Pu and soil. There is reason to believe that
the IR data are biased toward lower magnitudes as influenced by their nearness to the
detection limit. The mean IR for hide estimated with procedures of Kushner (1976) and
Hald (1949) suggested that these data may be similar to soil IR's. Other small-mammal
tissues were not compatible with these estimation procedures. Further field sampling to
eliminate the censoring difficulties is probably necessary if the question of differential
concentration of "^ ^"^Pu and ^ ''^^Pu in small mammals is to be resolved.

438 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Summary

The soil to a depth of 21 cm contained more than 99% of the plutonium estimated to be
in the studied areas of the Rocky Flats grassland. Litter contained a larger fraction of the
total plutonium (~10~^) than vegetation (~10~'*), arthropods (~10~^), or small
mammals (~I0~^). These results implied that soil— plutonium relationships and
soil-management practices are very important at contaminated sites.

Plutonium-concentration frequency distributions for soil samples were positively
skewed and characterized by CV's that were generally greater than 100%. Plutonium
concentrations in surface (0 to 3 cm) soil were inversely related to distance from the
plutonium source, the former oil -barrel storage area. Soil— plutonium concentrations
tended to decrease as depth increased and tended to increase as the soil particle size
decreased. This latter result suggested that plutonium— soil interaction was a surface-
attachment mechanism.

Mean concentrations of plutonium were higher in litter than in vegetation. Frequency
distributions of plutonium concentration were normal in litter and lognormal in
vegetation. In a manner similar to soil, plutonium concentration both in litter and in
vegetation was also inversely related to distance from the barrel storage area. Plutonium
concentrations in plant-derived compartment samples were also significantly correlated to
plutonium concentration in surface soil at the same locations.

Plutonium frequency distributions in arthropods and small mammals were also
positively skewed. Plutonium concentrations in bone samples were lower than those in
the other tissues sampled, namely, GI tract, hide, kidney, liver, lung, and muscle.

Concentration ratios of litter, arthropods, and small mammals relative to soil
indicated that litter had the highest value. The other compartments, in descending order,
were vegetation (3.4 X 10"^), small mammals (7.8 x 10~^), and arthropods
(6.8 X 10'^). The relatively liigh CR's suggested that most of the contamination of
vegetation resulted from surficially attached plutonium— soil particles as opposed to root
uptake. All the above data strongly indicate that in the grassland soil is by far the most
important compartment insofar as plutonium content and transport are concerned. The
primary conclusion is that, if transport of plutonium is to be avoided, then transport of
soil should be avoided. Therefore soil stabilization should be promoted by maximizing
vegetative cover growth and minimizing mechanical disturbances.

Isotopic ratios of ^^^Pu to ^^^Pu were calculated for soil, litter, vegetation,
arthropod, and small-mammal samples processed by commercial laboratories. The soil
results indicated that the median ratio was about 50. Litter and vegetation IR's were
similar to IR's in soil. The IR's of small-mammal tissues and arthropods were likely lower
than those of soil.

The meaning of the lower IR's in animal compartments was clouded by the fact that
the frequency distributions of the ^^^Pu and ■^''^Pu concentrations, from which the
ratios were formed, were censored. Further, the '^^^Pu concentration distribution was
censored to a much larger degree than was the ^^^Pu distribution. This situation may
have the effect of spuriously decreasing the mean or median ratio if the ratios are formed
before the average is calculated. An estimation procedure was used to calculate the mean
of both ^^^Pu and ^"'^Pu by taking into account the degree of censorship. Although
most small-mammal compartments may not be amenable to such a procedure, the ratio in
hide was calculated to be about 37. This value was within the 95% confidence interval of
most of the soil IR's. Without further analysis, the hide data suggested that the IR may

PLUTONIUM IN A GRASSLAND ECOSYSTEM 439

not be changing between environmental plutonium compartments, as previously
suggested (Little, 1976), but may indeed be constant.

Acknowledgments

Most of the Rocky Flats data presented here were collected by my associates and me in the
Department of Radiology and Radiation Biology of Colorado State University while
under contract with the Energy Research and Development Administration [now the
U. S. Department of Energy (DOE)] . Substantial work was contributed by T. F. Winsor,
F. W. Whicker, and J. A. R. Bly.

References

Aitchison, J., and J. A. C. Brown, 1969, The Lognormal Distribution, Cambridge University Press,
New York.

Ballou, J. E., D. K. Craig, J. F. Park, H. A. Ragan, and C. L. Sanders, 1973, Dose-Effect Studies with
Inhaled Plutonium in Rats and Dogs, in Pacific Northwest Laboratory Annual Report for 1972 to
the USAEC Division of Biological and Environmental Research, USAEC Report BNWL-1750(Pt.
1), pp. 21-27, Battelle, Pacific Northwest Laboratories, NTIS.

Bly, J. A.,- 1977, Plutonium Concentrations in Arthropods at a Nuclear Facility, Colorado State
University, unpublished.

Doctor, P. G., and R. O. Gilbert, 1977, Ratio Estimation Techniques in the Analysis of Environmental
Transuranic Data, in Transuranics in Natural Environments, Symposium Proceedings, Gatlinburg,
Tenn., October 1976, M. G. White and P. B. Dunaway (Eds.), ERDA Report NVO-178, pp.
601-619, Nevada Applied Ecology' Group, NTIS.

Emery, R. M., D. C. Klopfer, T. R. Garland, and W. C. Weimer, 1976, The Ecological Behavior of
Plutonium and Americium in a Freshwater Pond, in Radioecology and Energy Resources,
Proceedings of the Fourth National Radioecology Symposium, Oregon State University, May
12-14, 1975, C. E. Gushing (Ed.), Ecological Society of America Special PubUcation Series:
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Gilbert, R. O., L. L. Eberhardt, E. B. Fowler, E. M. Romney, E. H. Essington, and J. E. Kinnear, 1975,
Statistical Analysis of ^ 39-24 opy ^^j 24 1 ^^^ Contamination of Soil and Vegetation on NAEG
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Hakonson, T. E., and L. J. Johnson, 1974, Distribution of Environmental Plutonium in the Trinity
Site Ecosystem After 27 Years, in Proceedings of the Tfiird International Radiation Protection
Association Meeting. Wash.. D. C, Sept. 9-14, 1973, W. S< Snyder (Ed.), USAEC Report
CONF-730907-P1, pp. 242-247, International Radiation Protection Association, NTIS.

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Truncated at a Known Point, Skand. Aktuartidskrift, 32: 1 19-134.

Krey, P. W., and E. P. Hardy, 1970, Plutonium in Soil Around the Rocky Flats Plant, ERDA Report
HASL-235, New York Operations Office, NTIS.

Kushner, E. J., 1976, On Determining the Statistical Parameters for Pollution Concentration from a
Truncated Data Set, Atmos. Environ., 10: 975-979.

Little, C. A., 1976, Plutonium in a Grassland Ecosystem, Ph.D. Thesis, Colorado State University, Ft.
Collins, Colo.

Markham, O. D., 1976, Plutonium and Americium in Soils Near the INEL Radioactive Waste
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Radioecology -Ecology Programs, ERDA Report IDO-12080, pp. 35-46, Idaho National Engi-
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Navratil, J. D., and C. E. Baldwin, 1976, Recovery Studies for Plutonium Machining Oil Coolant,
ERDA Report RFP-2579, Rockwell International, NTIS.

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440 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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Biological Program Grassland Biome Technical Report 114.
Romney, E. M., H. M. Mork, and K. H. Larson, 1970, Persistance of Plutonium in Soil, Plants and

Small Mammals, Health Phys., 19: 487.
Schulz, R. K., G. A. Tompkins, and K. L. Babcock, 1976, Uptake of Plutonium and Americium by

Plants from Soils, in Transuranium Nuclides in the Environment, Symposium Proceedings, San

Francisco, 1975, pp. 303-310, STI/PUB/410, International Atomic Energy Agency, Vienna.
Sliaeffer, D. L., and C. A. Little, 1978, unpublished data.
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Sill, C. W., 1969, Separation and Radiochemical Determination of Uranium and the Transuranium

Elements Using Barium Sulfate, Health Phys., 17: 89.
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the Oxide, in Pacific Northwest Laboratory Annual Report for 1968 to the USAEC Division of

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Pacific Northwest Laboratory, NTIS.
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Final Report, USAEC Report COO-237 1-2, University of Colorado, NTIS.
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Transuranic Elements in Arctic Tundra
Ecosystems

WAYNE C. HANSON

Concentrations and inventories of worldwide fallout of^^'^Cs, ^ ^ ^Pu, and ^ ^ ^ '^ '* '^pu ifj
soils, lichens, and animals from northern Alaska and Greenland during the period
1968-1976 are discussed. Cumulative ^^^Cs fallout deposition at the soil surface at
Anaktuvuk Pass, Alaska, during the period 1959-1976 was estimated to be 43 mCi/km^ ,
compared to 16 mCi/km^ at Thule, Greenland. Measured ^^^Cs values in surface (top
5 cm) soil were 7.9 mCi/km^ at Anaktuvuk Pass and 21.5 mCijkm^ at Thule. The
discrepancy is presumably due to measuring difficulties and to rapid movement of
radionuclides into the soil profiles. An effective half-time of 0.4 to 0.5 yr was estimated
for Plutonium isotopes in surface soil at Anaktuvuk Pass. Average concentrations and
inventories of 239,240^^^ -^^ uncontaminated Thule lichen communities were, re-
spectively, 0.25 pCi/gand 0.21 nCi/m^ in 1968 and 0.33 pCi/g and 0.25 nCi/m^ in 1974;
however, these values were not significantly different. Inventories of ^^^Pu and
23 9,240^ w Alaskan lichen carpets were 0.019 and 0.28 nCijm^ , respectively, in 1968
and 0.040 and 0.67 nCi/m^, respectively, in 1974. Concentrations of^^ '' Cs, ^^^Pu, and
2 3 9,2 4 0p^^ were significantly higher in the upper 6-cm stratum than in the lower 6-cm
stratum o/Cladonia— Cetraria lichen carpets at Anaktuvuk Pass; concentrations of ^^S'r
were less consistent.

Radionuclides in arctic ecosystems have been investigated for nearly 20 yr because of the
efficient transfer of worldwide fallout materials through arctic food chains. Initial
investigations were .concerned primarily with the radiological health aspects of the
appreciable body burdens of ^*^Sr and '^^Cs obtained by circumpolar populations that
were involved in the lichen— reindeer/caribou-man/carnivore food webs. The discovery of
measurable amounts of ^ ^ ^Cs in Nearctic Eskimos and Indians and Palearctic
reindeer-herding peoples in 1961 and 1962 coincided with the advent of a second major
period of nuclear weapons tests, which resulted in appreciable radioactive fallout
deposition and increased the efforts of several investigators.

Several ecological aspects of arctic tundra ecosystems recommended them for study
of the transfer of worldwide fallout radionuclides. Although their geographical location,
mostly beyond 60°N latitude, is a region of appreciably less fallout deposition than other
more-populated areas of the world, the plant and animal communities are so related that
^^^Cs body burdens of the native peoples in the Arctic regions, for example, are often
100 to 1000 times greater than those of Temperate Zone residents. This has resulted from
the effective accumulation of atmospheric materials, radioactive or otherwise, by the
hchen communities, which provide a reservoir of such materials at the base of northern
food webs. Transfer from this relatively rich soufte is enhanced by (1) the low

441

442 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

concentration of potassium, the chemical analog of cesium, and other nutrients; (2) the
arctic climate, which enhances food intake by the native animals; (3) the simple and
direct food webs that are clearly defined; and (4) the few interfering factors that expedite
evaluation of small changes over a sufficient time span.

This chapter summarizes information on '^^Cs and plutonium isotopes which was
gathered as a part of mtensive studies of worldwide fallout in soils, lichens, and animal
samples from northern Alaska during the period 1969 to 1976. Soil and lichen samples
from Thule, Greenland, were obtained during participation in Danish expeditions of 1968
and 1974 which investigated ecological consequences of plutonium that had been released
to those environs by the nonnuclear explosion of four unarmed nuclear weapons during
the crash of a U.S. Air Force B-52 bomber on Jan. 21, 1968 (Aarkrog, 1971a; 1971b;
1977; Hanson, 1971 ; 1972; 1975). Results from these studies were compared with results
from radiological studies of plutonium isotopes in the Scandinavian lichen-reindeer—
Lapp food web.

Most of the data are from studies centered at the inland Eskimo village of Anaktuvuk
Pass, located in the central Brooks Range (Fig. 1), where annual precipitation is about
20 cm. Soil samples were collected from undisturbed locations at Settles, some 125 km
south of Anaktuvuk Pass, with annual precipitation of about 32 cm, and from Fairbanks,
about 480 km southeast of Anaktuvuk Pass, with annual precipitation of 28 cm.

ARCTIC yv. OCEAN

GULF OF ALASKA

Fig. 1 Map of northern Alaska showing samphng locations.

TRANSURANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 443

Methods

Sample Collection and Processing

Soil samples were collected at eight locations near Thule (Fig. 2) during 1974 to estimate
the amounts of '^"^Cs and 2 3 8,239,240p|j jgpQsited on the landscapes. Seven of the
locations (numbers 5,7, 13, 14, 14A, 18, and 21 A) were chosen in the downwind vector
of the debris cloud that drifted from the 1968 crash site, and one location (number 3)
was located 20 km upwind from that site. Five 1-dm^ by 0.5-dm-deep samples were
collected at 0.2- to 0.4-km intervals along transects over landscapes selected for

'21A

WOLSTENHOLME
ISLAND

KILOMETERS

70 W

68 W

Fig. 2 Map of Thule, Greenland, environs showing 1968 and 1974 sampling sites for soils,
alluvium, and lichens.

uniformity of slope, direction, orientation to the crash site and relationship to lichen
sampling sites. The variable intervals were selected to best represent the landscape unit to
be sampled. Soil samples were taken from sites that were free of vegetation and large
rocks. The five samples were composited, yielding 0.05 m^ of surface area, for inventory
of the radionuclide deposition. The method was similar to the template method adopted
by the Environmental Measurements Laboratory (Harley, 1972). Alluvium samples were
collected in the same manner as soil samples from seasonal streambeds that drained the
landscapes across which the soils were collected.

Alaskan soil samples were collected from three locations at Anaktuvuk Pass and at
single locations near Bettles and Fairbanks in the same manner as at Greenland locations
except that single 1-dm^ by 0.5-dm-deep samples (0.01 m^) or three composited samples
(0.03 m^) were analyzed for radionuclides. The different methods were used to better
define analytical variability and to remain within the analytical capability of our
laboratory.

444 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

All soil samples were screened to separate rocks greater than 0.6 cm in diameter from
the fine-soil component. Both size fractions were dried, and 100-g aliquots of the
fine-size (<0.6 cm) fraction and the entire large-size fraction were leached for about 16 hr
with a heated mixture of HCl and HF acids to dissolve the plutonium in the samples.

Lichens were collected on an areal basis whenever possible to provide estimates of
biomass, community composition, and compartmental analysis for the various radio-
nuclides. Alaskan samples were 0.25-m^ blocks cut from the Cladonia Cetraria and
Alectoria-Cladonia-Cetraria fruticose lichen communities that form carpets on the
Anaktuvuk Pass (Fig. 1) landscapes. Greenland samples were obtained from several
discontinuous but representative "islands" of each community type which result from
microhabitat differences. The term "community" included all lichen or other plant
species (populations) within a sample and was designated by the dominant lichen species
at a specific location. The various dominant lichen species contributed an average of 90%
of the total community biomass in Greenland samples and 80% in the Alaskan samples.
Other populations separated from the community samples consisted of subordinate lichen
species, vascular plants, lichen and vascular plant debris, and fine soil, which normally
comprise a lichen community. The component samples were dried at 100°C for 24 hr to
determine dry weights and then dry-ashed at 425°C and dissolved in HCl and HF acids for
radiochemical analyses. An average of 1 to 5 g of ash resulted from minor sample
components, and up to 100 g of ash from major components of the populations was used
for plutonium determination.

Eskimo residents of Anaktuvuk Pass provided the animal samples. Emphasis was
placed on sampling caribou (Rangifer arcticus tarandus) because of its importance as a
food base for the entire carnivore (including human) population in northern Alaska.
Major sampUng efforts were made in autumn (September— October) and spring
(May— June) months when the caribou were intercepted by Eskimo hunters during their
migrations to and from wintering grounds and summer ranges. Standard samples consisted
of the upper femur and attached muscle. Appreciable numbers of red foxes (Viilpes
fulva), tundra wolves (Canis lupus), and wolverines (Gulo gulo) and lesser numbers of
arctic foxes (Alopex lagopus) and lynx (Lynx canadensis) were taken each winter by
Anaktuvuk Pass Eskimo hunters; entire hindquarters of each animal type were obtained
for separation into muscle and bone samples, drying at 100°C. ashing at 425°C, and
subsequent radionuclide analyses.

Cesium- 137 was measured by counting the 0.661 -MeV gamma-ray emissions from
dried soil, plant, and animal samples in a calibrated plastic container atop a 7.6- by 7.6-cm
sodium iodide crystal connected to a 400-channel analyzer spectrometer. Most counting
times for lichen samples were in the range of 30 to 40 min; counting times were longer
for small samples. Spectra were corrected to individual radionuclide amounts by
comparison with National Bureau of Standards sources and corrected for background-
radiation contributions.

Plutonium-242 tracer was added to sample solutions to determine recovery of the
plutonium isotopes, and the mixture was deposited as the nitrate on an anion-exchange
resin column. The plutonium was eluted from the column with a nitric acid-ammonium
iodide solution and electrodeposited on stainless-steel planchets; the planchets were
counted on a silicon surface-barrier alpha spectrometer for 165 to 1330 min. Recovery of
plutonium isotopes, as measured by recovery of the ^^^Pu tracer, was usually in the 60
to 80% range. Isotopic exchange was considered to be uniform within the samples on the
basis of standards and interiaboratory comparisons. Counting efficiency for this particular

TRANS URANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 445

measurement system was 30% with an average background of 3 counts per 1330 min.
Counting data were reduced by a computer program that expressed results in picocuries
of ^^^Pu or ^^^•^'*°Pu per gram of sample with one standard deviation for the counting
statistics.

An improved analytical procedure for the determination of ^"^'Am in large (up to
100 g of ash) samples became available during 1977 (Knab, 1977) and yielded the first
realistic results for that radionuclide in a limited number of Alaskan samples. This
procedure consists of DEHPP [phosphorus pentoxide, bis(2-ethylhexyl) phosphoric acid,
and cyclohexane] extraction of both plutonium and americium from the sample residue,
separation of plutonium frorfi americium by anion exchange onto a nitric acid prepared
column, and purification of americium by ion exchange in methanol-nitric acid and
ammonium thiocyanate anion columns. The eluted americium is then electrodeposited on
a stainless-steel planchet and counted on the same alpha-spectrometer system used for
plutonium. Yields were monitored by ^'^^ Am tracer.

Several samples, particularly the animal tissues, which were analyzed for transuranic
elements yielded net values that were lower than the minimum detection limits (MDL) of
the system; for ^^^Pu extracted from 10 g of soil samples and counted for 1333 min, the
MDL was 0.003 pCi/g, and for "''■^^"Pu it was 0.002 pCi/g. Values of zero or negative
numbers are a common occurrence in environmental sampling owing to statistical
fluctuations in the measurements. Although a negative value for a measurement does not
represent a physical reality, a valid long-tenn average of many measurements can be
obtained only if very small or negative values are included in the population. The data
reported here are often averages of several samples, including those below the minimum
detection limit or negative numbers. Zero values were considered to represent <0.0061
dpm at the 95% confidence level. This procedure is consistent with data treatment at
other laboratories (Harley and Fisenne, 1976). Unless specifically stated, data reported as
^^^Pu include the minor radioactive contribution of ^'^"^Pu.

Results and Discussion

Radionuclide Deposition Estimates

Studies of transuranic elements in ecosystems are greatly aided by relating their behavior
to that of fallout '^"^Cs, which is easily measured and is consistently near a ratio
(^^^Pu/'^'^Cs) of 0.016 (Hardy, 1975). Fallout deposition on the Alaskan and Greenland
landscapes was calculated from data published by the U. S. Department of Energy
Environmental Measurements Laboratory (Hardy, 1975) with ^^ ^Cs deposition estimated
from the ratio ^^ "^Cs/^^Sr = 1.6 ± 0.2 (Hardy and Chu, 1967). Values were directly
available for Thule, Greenland, from 1959; however, the fallout deposition at Anaktuvuk
Pass was estimated by extrapolating the measured deposition at Fairbanks, some 500 km
southeast of Anaktuvuk Pass, to the study area by multiplying by 0.67, the ratio of the
annual precipitation rates at the two locations (21 and 32 cm/yr, respectively) (Volchok
and Kleinman, 1971). Similarly, fallout deposition at Bettles, about 300 km northwest of
Fairbanks, was estimated by multiplying by 1 .34, the ratio of their annual precipitation
rates (43 and 32 cm/yr, respectively).

The correlation of worldwide fallout deposition with precipitation and the prac-
ticality of estimating the integrated fallout deposited in a geographic region by careful
soil sampling have been demonstrated by Hardy (1974; 1975) and Hardy and Krey
(1971). Our data indicated that ^ ^ "^Cs inventories in Greenland lichen communities were

446 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

about 60% as great as those in Alaska, or about proportional to their differences in annual
precipitation. However, the estimated '^^Cs deposition at Thule was only 37% of that
estimated for Anaktuvuk Pass, which suggests that "dry fallout" must account for an
appreciable part of the worldwide fallout deposited on Thule lichen communities. The
close agreement between most of the ^^''Cs inventories in the soils and adjacent lichen
communities of that region suggested that the two components were balanced by some
physical transport mechanism, probably the substantial winds of the region.

Fallout collections at both Thule and Fairbanks began near the end of a previous
major period of fallout deposition that resulted from an extensive series of atmospheric
nuclear weapons tests during 1952 to 1958 by Great Britain, Russia, and the United
States. Therefore the '^^Cs estimate of 43 mCi/km^ at Anaktuvuk Pass at this time and
of 16 mCi/km^ at Thule at the time of sample collection in August 1974 represent
conservative deposition estimates. Cornparison of fallout collection data at New York
during periods before 1960 and after 1960 suggests that the preceding deposition
estimates for the northern Alaska and Greenland areas should be about doubled to
account for the total deposition of fallout since the beginning of nuclear weapons testing.

The estimated deposition of ^^^Cs from fallout in the Thule environs (annual
precipitation, 13 cm/yr) determined from ^^ ^Cs concentrations in soil samples (Table 1)

TABLE 1 ' ^ ^Cs Concentrations and Areal Inventories

in Soils* of the Thule, Greenland, Environs

During August 1974

'^^Cs

Concentration,

Inventory,

Location/ numberf

pCi/g

; (dry weight)

nCi/m^

Saunders Island/5

1.05

32.1

Narssarssuk/14

0.29

21.6

Wolstenholme Island/ 18

0.45

21.4

Narssarssuk/7

0.27

21.0

Cape Abernathy/3

0.22

18.2

Narssarssuk/13

0.22

14.7

Wolstenholme Island/21 A

<0.01

Narssarssuk/14 A

<0.01

*Each value is derived from measurement of a 100-g aliquot of dry
soil taken from a composite of five 0.01-m^ samples taken along a
transect of 0.2 to 0.4 km in the various locations.

t Refer to Fig. 2.

ranged from 14.7 to 32.1 nCi/m^ with an average of 21.5 ±2.4 (SE). The maximum value
occurred on the southwest side of Saunders Island, where large snowdrifts accumulate
during winter periods; the minimum values (<0.01 pCi/g) were obtained in two samples
from the windswept headlands at the southern edge of the study area. The validity of this
estimate was substantiated by a value of 20.6 nCi/m^ calculated from a 2 3 9,24 0p|j
inventory of 0.33 mCi/km^ measured in a large (1-kg) aliquot of 622 cm^ of Thule soils
during 1970 and 1971 (Hardy, Krey, and Volchok, 1973) and a ^^^'^'^^Pu/'^ ^Cs ratio
of 0.016. Fallout collections at Thule indicated that an additional 0.008 mCi of
23 9,240pjj pgj. 3q^2^.g kilometer was deposited between July 1971 and August 1974,
bringing the two estimates even closer to agreement.

TRANS URANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 44 7

Radionuclide Concentrations in Soils

Soil samples o\^ the Thule region during 1974 contained an average of 13.0 ± 6.3 (SE) fCi
(10' '^ Ci) of 2 3 9,240pu pgj. gj.gj^ I'^jj.y v^eiglit) in those areas considered to be
uncontaminated by the January 1968 accident debris (Table 2). The 239,240p^
inventory was estimated to be 0.35 ±0.10 nCi/m^ from these samples. Two sampling
locations (numbers 7 and 13 in Fig. 2) within an area of about 16 km'^ near the small
habitation of Narssarssuk contained 20 to 100 times that inventory, which reflects
contamination from the accident. Alluvium samples from seasonal streambeds that
drained snowmelt from the landscapes across which the soil transects were taken
corroborated the soil measurements. The nature of the variability in aliquots taken from
replicate soil samples from the Narssarssuk area indicated that small particles of
indetenninate size contributed most of the radioactivity. Plutonium oxide particles with a
calculated mass median diameter of 4 jum were determined by nuclear track auto-

TABLE 2 ^^^'^^°Pu Concentrations in Replicate 100-g

Aliquots* of Soil and Alluvium Samples of the Thule,

Greenland, Environs During August 1974

Sample

2 3 <3 , 2 4 0 py concentration

aliquot

fCi/g

Location/numberf

(replicates)

(dry weight)

nCi/m^

SoUs

Narssarssuk/7

929

232

Narssarssuk/ 13

217

379

Narssarssuk/14A

3.5

0.2

Narssarssuk/14

1.2

1.0

Wolstenholme Islan

id/18

0.8

0.9

Saunders lsland/5

0.6

Cape Abernathy/3

1.2

0.5

Wolstenholme Islar

id/21A

A
B

2
2

Alluvium

Narssarssuk/7

A
B

622
197

Narssarssuk/14

130

Wolstenholme islar

id/ 18

A
B

20
16

Cape Abernath\73

A
B

18
10

*Aliquot values above the dashed lines are considered to be contaminated by
the 1968 accident debris.
T Refer to liij. 2.

448 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

radiographic and microscopic studies of debris, and geometric mean particle diameters of
2 to 5.6 //m were reported in snow/ice samples obtained shortly after the accident
(Langham, 1970; Gjc^rup, 1970).

Isotopic ratios in the Thule soil and alluvium samples from uncontaminated sites were
in the range of 0.011 to 0.065 for ^^^'^'^^Pu/'^ ''Cs and 0.15 to >1.0 for
2 38p^/2 3 9,240p^ jj^g 2 3 8 p^/2 3 9 ,2 4 0 p^ ^^^-^^ -^ ^-^^^^ communities, by comparlsoH,

were usually within the range of 0.020 to 0.024, which was reported for global fallout
(Harley, 1975).

Radionuclide concentrations in soil samples collected at the Alaskan sites during
summer periods of 1975 and 1976 are shown in Table 3. Inventories of ' "^ ''Cs, expressed
as nanocuries per square meter, were generally proportional to average annual rainfall
regimes at the three locations. The greatest amounts were in Fairbanks samples, although
Fairbanks receives only about three-quarters as much precipitation as Bettles; however,
these values were not significantly different (t-test value, 1.54; df, 2;P<0.2 to 0.3), nor
were the Anaktuvuk Pass and Bettles values (t value, 1.43; df, 6;P<0.2). Fairbanks
values were significantly different from those for Anaktuvuk Pass (t value, 4.01; df, 6;
P<0.01). There was a significant difference between the areal inventories of '^''Cs
measured at Anaktuvuk Pass in 1975 and 1976 (t value, 1.81; df, 7; P<0.10) but not
between those in July and September 1976 (t value, -0.88; df 8; P<0.40). This was
apparently due to slightly greater fallout deposition during 1974 and 1975 after the large
atmospheric nuclear weapons tests conducted by the People's Republic of China in 1973
and 1974 (Carter and Moghissi, 1977). The ^^^'^'^"Pu/' ^ ^Cs ratio was 0.016 in 1975
and then decreased to 0.004 to 0.007 at all sites in 1976 as greater amounts of ^^"^Cs
were apparently deposited.

Concentrations and inventories of ^^^'^'*°Pu in Alaskan soils during 1975 and 1976
were substantially less than those in Greenland during 1974; in most cases they were

TABLE 3 Radionuclide Concentrations (Mean ± Standard Error)

in Soil Samples* Collected at Anaktuvuk Pass, Bettles, and Fairbanks,

Alaska, During Summer Periods of 1975 and 1976

Date

1 3

'Cs

2 3 8

2 3 9,2.

4 0py

2 4 1

Location

pCi/g

nCi/m-

pCi/g

nCi/m^

pCi/g

nCi/m'

pCi/g

nCi/m^

Anaktuvuk

7/75

0.49

9.6

-0.0018

0.0078

0.152

0.0005

0.0092

Pass

±0.22

±4.3

±0.0018

±0.0054

±0.101

±0.0003

±0.0061

Anaktuvuk

7/76

0.26

7.5

0.0004

0.0185

0.0012

0.042

Pass

±0.16

±3.7

±0.0001

±0.0050

±0.0012

±0.030

Bettles

7/76

0.47

22.0

0.0004

0.0287

0.0016

0.115

±0.27

±11.8

±0.0000

±0.0130

±0.0006

±0.081

I'airbanks

7/76

1.15

±0.57

34.1
±13.2

0.0022

0.0858

0.0057

0.222

Anaktuvuk

9/76

0.36

7.9

0.0003

0.0040

0.0010

0.032

Pass

±0.20

±2.1

±0.0001

±0.0023

±0.0010

±0.017

*Kach sample consisted ofa 10-g aliquot taken from a 0.1- by 0.1- by 0.05-ni core after drying and
sieving to remove rocks greater than 6.35 mm in diameter,
t Number of samples.

TRANSURANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 449

about one-tenth the Thule values. This difference could be due to the short residence
time of ■^"'^Pu in the upper 5 cm of soil at both the Alaska and Greenland sites.
Plutonium inventories in surface soil at Anaktuvuk Pass decreased between 1975 and
1976 sampling periods at an effective half-time of 0.4 to 0.5 yr. A tentative application of
these values to the Greenland situation suggests that the decline in Thule soils is of a
similar rate.

Radionuclides in Lichen Communities

The ability of lichens to retain and recycle fallout radionuclides has been observed by
several northern investigators. The radiological health aspects of the lichen— caribou— man
food web have been the dominant theme of the studies carried out in European nations.
Similarly, the ecosystem studies at Thule, Greenland, were mainly oriented toward
defining the consequences of the accidentally released plutonium in the marine food webs
of that area which were of importance to the local Eskimos (Aarkrog, 1971a; 1971b;
1977). Most of the plutonium contamination (~30 Ci) resulting from the accident was
associated with the sea ice and other Bylot Sound marine components. Approximately 1
to 5 Ci of plutonium was estimated to have been contained in the cloud of smoke and
debris that drifted west-southwesterly from the crash site and deposited in uncertain
amounts on the sea ice and landscape of the area (Langham, 1970). This uncertainty was
enhanced by the discontinuous distribution of lichens in the Thule region, the arid
climate and light character of the soils, and the appreciable winds that redistributed the
plutonium particles that originated from worldwide fallout from nuclear weapons tests,
the April 1964 burnup of the SNAP-9A satellite power source, and the aircraft accident.

Lichen samples collected from several Thule locations during 1968 (Hanson, 1972)
and 1974 (Table 4) illustrated the highly variable nature of plutonium concentrations
compared with more uniform ' ^ ''Cs concentrations in the lichen communities exposed to
the 1968 accident debris. During 1968 most lichen samples from uncontaminated areas
contained a mean 239,240p^^j concentration of 0.25 ± 0.07 (SE) pCi/g (standard dry
weight) and a total inventory of 0.21 nCi/m^ ; comparable values during 1974 were
0.33 ±0.09 pCi/g and 0.25 nCi/m^ which presumably included the 0.016 nCi of
2 3 9,2 4 0pjjyj^2 ^j.|^^ Yi^^ been deposited on Thule landscapes in the 6-yr interval between
collections. Those two sets of data were not significantly different (t value, -1.47; df. 12;
P<0.1 to 0.2) nor were the "^Pu/^^^-^'^^Pu ratios of the uncontaminated 1974 lichen
samples (last eight values in Table 4) significantly different from those which contained
appreciably greater amounts of 2 3 9,2 4 0p|j ^j^^^ apparently originated from the 1968
accident. However, the ^^^'■^'**^Pu/^ ^ ^Cs ratios of uncontaminated samples varied over
a 200-fold range and showed coefficients of variation (CV = standard deviation ^mean)
that averaged >2.0. By comparison, the ■^^^•^'*°Pu/' ■^ ^Cs ratios in uncontaminated
lichen samples were relatively stable at 0.02 ± 0.01. This greater variation in radionuclide
ratios in lichen samples collected near the accident site suggests that the rigorous climatic
and edaphic factors of the Thule region probably had a major influence on the
redistribution of radionuclides and led to a balancing of concentrations in lichens and
soil.

The ^^^Cs inventory in the Ciadonia-Cetraria lichen carpet at Anaktuvuk Pass
increased steadily from 6.2 nCi/m^ in the initial sampling in 1962 to maximum values of
about 50 nCi/m^ in 1965 and has subsequently fluctuated near 35 nCi/m^ (Table 5). The
estimated ' ^ ^Cs deposition and the amount in the lichen carpet were in close agreement,
althougli the lichens had also been exposed to an undetermined amount of fallout during

430 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 Estimated Inventories of "^Pu, ^^^^^°Pu, and '^^Csin
Areal Samples* of Greenland Lichen Communities During August 1974

Sample type

Radionuclidi

e inventory,

nCi/m^

Location/ nu mberf

2 3 8 py 2 :

)9,2 4 0py

'^'Cs

Saunders Island/ 12

Cetraria nivalis

0.068

3.97

Saunders Island/ 12

Cetraria nivalis

0.071

3.79

Saunders Island/ 12

Cetraria delis ei

0.078

5.50

Saunders Island/ 12

Alectoria ochroleuca

0.058

0.86

Saunders Island/ 12

Alectoria ochroleuca

0.077

3.79

Kap Atoll/ 19

Cetraria delisei

0.022

0.68

Narssarssuk/7

Cetraria nivalis

0.046

9.38

Narssarssuk/14

Cetraria nivalis

0.154

8.24

Narssarssuk/13

Cetraria nivalis

1.452

81.61

Narssarssuk/I3

Cetraria nivalis

0.010

0.28

Wolstenholme Island/ 18

Cetraria nivalis

0.006

0.12

Wolstenholme Island/ 18

Cetraria nivalis

0.009

0.28

Wolstenholme Island/ 18

Cetraria nivalis

0.003

0.17

Saunders IsIand/5

Cetraria nivalis

0.009

0.21

Saunders Island/5

Cetraria nivalis

0.006

0.19

Saunders Island/5

Cetraria nivalis

0.006

0.24

Cape Abernathy/3

Cetraria nivalis

0.015

0.49

*Samples above dashed line are considered contaminated by 1968 accident debris,
t Refer to I ig. 2.

the pre-Health and Safety Laboratory (HASL) measurement period. For example, values
in lichens sampled in July 1964 reached 41 ± 2.6 nCi/m^ compared to a calculated fallout
deposition at Anaktuvuk Pass during the period 1960 to 1964 of 26 nCi/m^ . This
maintenance of high '^^Cs inventory prompted the experiments on effective half-times
of radionuclides in lichens and the modeling of Arctic food chains which demonstrated
the significant differences in radionuclide behavior within lichen communities and the
important implications to Arctic ecosystems (Ebeihardt and Hanson. 1969; Hanson and
Eberhardt, 1967; Hanson, 1973). A salient feature of those data was the longer effective
half-time of '^"^Cs (10 yr) compared with that of ''°Sr ( 1 .0 to 1.6 yr) in lichens. This
was due primarily to the greater mobility and recycling of '^^Cs and the impedance of
^°Sr translocation by cation-exchange phenomena (Tuominen, 1967; 1968). Similar
mechanisms may be operative in the relatively rapid loss (Ti^ , 6.1 yr) of ^^^'~'*'^Pu from
lichen carpets reported from Scandinavia (Holm and Persson, 1975). Similar results were
obtained from Alaskan liclien carpets at Anaktuvuk Pass (Table 6 and Fig. 3), which
showed (1) a general increase of both ^^^Pu and 239,240p^j jy^^^^ ,q^^g ^^ 1071 ; (2) a
decline during 1972 and 1973, periods of low fallout deposition; (3) a sudden increase
during 1974 that correlated with increased fallout deposition presumably due to the
Chinese nuclear weapons tests of 1973 and 1974; and (4) another decline in 1975 and
then an increase in 1976 samples. The lower 1975 values for plutonium are unexplained
but also occurred to a lesser degree in the '^^Cs values; the decrease was more
pronounced when compared with similar decreases that occurred in 1972. The
probability that this was due to either sampling or analytical error is considered to be

TRANSURANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 451

TABLE 5 Worldwide Fallout ' ^ ''Cs Inventory*

in the Cladonia-Cetraria Lichen Carpet at

Long-Term Sampling Sites Near Anaktuvuk Pass,

Alaska, During the Period 1962 to 1976

•^^Cs

SamplLng

Biomass,

inventory,t

Deposition, §

date

kg/m^

nCi/m^

3/7/62

0.52

6.2 ± 2.5

5.14

7/7/63

0.61

14 ± 1.4

9.82

12/7/64

1.53

41 ± 2.6

11.01

26/7/65

1.45

48 ±4.3

10.97

2/8/66

1.26

34 ± 1.2

2.13

2/8/67

1.20

30 ± 1.2

0.63

30/6/68

1.16

30 ± 1.3

0.60

29/7/69

2.95

44 ± 1.6

0.66

2/8/70

1.66

24 ± 5.3

0.91

9/8/71

2.04

0.67

25/7/72

2.82

44 ± 1.0

0.00

2/7/73

2.42

23 ± 0.8

0.06

26/9/74

1.74

44 ± 1.3

0.32

21/7/75

3.33

33 ±4.2

0.06

8/7/76

2.76

35 ± 3.0

0.05

17/9/76

2.78

33 ± 1.0

0.00

43.03

*Vaiues are based on samples of 0.25-m^ to 0.5-m^
replicates from contiguous sampling areas.

t Number of samples.

$Mean ± standard error.

§Cumulative fallout deposition at Anaktuvuk Pass
between successive sampling periods based on 0.67 of
monthly measurement at 1 airbanks (Hardy, 1975).

very sliglit because the decline occurred in both lichen communities sampled and the
analytical data have been verified. One-way ANOVA tests performed on untransformed
and log-transformed lichen radionuclide data for the years 1974 to 1976 showed
significant (P<0.05) differences between 1974 and 1975 and between 1975 and 1976
which were identified by multicomparison procedures. These differences were confirmed
by such nonparametric procedures as the Kruskal— Wallis and Kolmogorov— Smirnov tests
(Hollander and Wolfe, 1973). The ^^^•^'^^Pu/^ ^ ^Cs ratios in Alaskan lichens usually
were stable near 0.013, but they decreased to 0.006 during 1975. Statistical analysis (t
statistic for two means) of the 1971 to 1976 plutonium concentrations (Table 7) revealed
that the 1975 values were significantly lower than the other years and that the
2 3 8pjj^2 3 9.2 4 0pjj ra^JQg j^i the Alectona-Claclonia-Cetraria lichen community samples
during 1976 were significantly greater than those in the Cladonia-Cetraria community
samples.

During 1969 and 1970 the Cladonia-Cetraria lichen carpet samples were fractionated
into upper 6-cm (Cu) and lower 6-cm iC\) components to test the hypothesis that there
were no significant differences between their radionuclide concentrations (Hq : Cy = Ci)

452 TRANSURANJC ELEMENTS IN THE ENVIRONMENT

TABLE 6 Estimated Inventories of "^Pu, "^'^^"Pu, ^^ ' Am, and * ^ ''Cs in
Areal Samples of Alaskan Lichen Communities During Summers of 1968—1976

Taxon

Radionuclide inver

itory, nCi/m^

Year

238p„

2 3 9 ,2 4 0 p

'^^Cs

^"'Am

1968

Cladonia- Cetraria

0.019

0.28

30.0

1969

Cladonia-Cctraria

0.018

0.36

44.5

1970

Cladonia- Cetraria

0.027

0.41

50.2

1971

Cladonia alpestris

0.030

0.47

42.1

Cetraria delisei

0.013

0.18

40.4

Alectoria ochroleuca

0.002

0.04

9.2

1972

Cladonia- Cetraria

0.024

0.28

44.0

Cladonia alpetris

0.029

0.30

27.9

Cetraria delisei

0.028

0.30

40.2

1973

Cladonia- Cetraria

0.024

0.33

23.4

Cladonia alpestris

0.038

0.30

26.8

Cetraria delisei

0.007

0.10

27.7

Stereocaulon paschale

0.069

0.32

57.4

1974

Cladonia- Cetraria

0.040

0.67

44.3

Alectoria ochroleuca

0.023

0.40

38.2

1975

Cladonia- Cetraria

0.008

0.14

29.2

Alectoria ochroleuca

0.009

0.16

34.3

1976

Cladonia- Cetraria

0.030

0.43

34.6

0.122

Alectoria ochroleuca

0.034

0.34

21.8

0.078

^Number of samples.

at the P<0.05 level. Subsequent statistical analyses for four common fallout
radionuclides were performed using the untransformed data in a two-tailed t test which
allowed separate variance estimates (Nie et al., 1975), and the log-transformed data were
tested by Kruskal-Wallis and Kolmogorov- Smirnov procedures (Hollander and Wolfe,
1973) (Table 8). The upper 6 cm usually contained significantly greater concentrations of
^^Sr, '^^Cs, "^Pu, and 239,240p^ ^^^^ ^^g j^^^j. 6 cm, except for ^°Sr during 1970.
Cesium-137 showed the greatest differences in concentrations in the layers, apparently
owing to its greater mobility and concentration in the more rapidly photosynthesizing
upper portion of the lichens (Moser, 1977). These data are consistent with similar studies
in Sweden (Holm and Persson, 1975) in which Cladonia alpestris carpets were
fractionated into several vertical layers. Considering the variation of radionuclide
depositions, sampling, and analytical differences, the values reported for the ^■'^Pu and
2 3 9,2 4 0pj^j concentrations in lichen samples from central Sweden are similar to those
from northern Alaska.

Concentration ratios of ^^^Pu, ^^^'^'*'^Pu. and ^^''Cs in Greenland and Alaska
(lichens/soil) (shown in Table 9) were generally consistent; the exception occurred in
■^^^Pu measured in Greenland samples. The values for the plutonium isotopes were
considerably higher than the values in the range of 10~^ to 10"'* reported for most
Temperate Zone plants (Francis, 1973). Resuspension of radionuclides from soil to
lichens was assumed to be a strong possibility in the Greenland sites and very minor in

TRANSURANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 453

0.7

0.6

0.5

CM 0.4

E
o

^ 0.3

0.2

0.1

0.07

0.06

— 0.05

— 0.04 -

— 0.03 3

Q-
00
n

0.02

— 0.01

1968

1970

1972
YEAR

1974

1976

Fig. 3 Inventories of ^ ^ *Pu and ^ 3' '^ 4o p^ -^^ ^^i^ Cladonia-Cetraria lichen carpet at
Anaktuvuk Pass, Alaska, during the period 1968-1976.

TABLE 7 " ^ Pu and 2 ' "^ -^ ^ ^ Pu Concentrations in

Northern Alaskan Lichen Communities

During 1971-1976

Radionuclide concentration,t pCi/

g (dry weight)

Year

.3 8py

239,240py 2

38py/239,240py

1971

0.012 + 0.002

0.201 ±0.100

0.065 ± 0.010

1972

0.026 ± 0.003

0.280 ± 0.026

0.091 ± 0.003

1973

0.013 ± 0.003

0.146 ± 0.026

0.089 ± 0.010

1974

0.017 ±0.004

0.280 ± 0.006

0.064 ± 0.005

1975

0.003 ±0.001

0.058 ± 0.002

0.050 ± 0.004

1976

0.014 ± 0.003

0.176 ±0.010

0.078 ±0.012

*Nuniber

of samples.

fMean ± s

tandard error.

Alaska; however, that mechanism did not appear to be of major significance in this study,
or possibly it was masked by other parameters.

Correlations between radionuclide isotopic ratios obtained from soil and lichen
samples (Table 10) indicated that Unear relationships between ^^^Pu/^^^Pu and
2 3 9 p^j^i 3 7 (->g \fjQXQ more constant in Alaskan samples than in Greenland samples and that
ratios in lichen samples were more strongly related than ratios in soil samples. There was

454 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Statistical Analyses (t Statistic for Two

Means) of the Null Hypothesis of Equal Radionuclide

Concentrations in Upper and Lower 6-cm Segments

oi Cladonia-Cetraria Lichen Carpet Samples

(Ho : Cu = Ci) During 1969 and 1970

Degrees of

Year

Radionuclide

t value

freedom

Probability

1969

'"Sr

5.23

0.002

'^^Cs

10.55

0.000

2 38py

3.07

0.016

2 39,2 40p^j

3.10

0.010

1970

'°Sr

0.63

0.482

'^•'Cs

27.45

0.000

"«Pu

11.83

0.008

2 39,2 "Opy

27.33

0.016

TABLE 9 Concentration Ratios of

Greenland and Alaska Lichen Samples

Concentration ratio,
(pCi/g lichen)/(pCi/g soil)

Location

2 3 9,240

'Cs

Alaska 140 180 36

Greenland 3 140 52

TABLE 10 Correlation Coefficients of "^Pu/"^Pu and ^^^Pu/' ^ ''Cs Concentration
Ratios in Soil and Lichen Samples from Alaska and Greenland

2 3 8py/2

3 9pu

''»Pu/'

'^Cs

Sample type

Degrees of
freedom

Conclusion about
hypothesis, p = 0

Degrees of
freedom

Conclusion about
hypothesis, p = 0

Alaska soil
Alaska lichens
Greenland soil
Greenland lichens

7
24

8
29

0.4690
0.6481
0.4369
0.9695

Not rejected
Rejected at 1%
Not rejected
Rejected at 1%

24
6

0.7280
0.5292
0.6164
0.1000

Rejected at 1%
Rejected at 1%
Not rejected
Not rejected

also an indication that sample sizes were often too small to validate the hypothesis that
p = 0, i.e., that the sampled populations were normally distributed about the regression of
YonX.

Radionuclides in Caribou Tissues

Flesh and bone samples of 69 caribou taken by the Anaktuvuk Pass Eskimos during the
period 1964 to 1976 contained highly variable concentrations of^'^^Puand 239,240p^^

mANSUHANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 455

Highest values were found in samples taken during early 1965; values then declined to 0
to 04 tri ^^^•^'*^Pu/g (dry weight) and 0 to 0.1 fCi "^Pu/g. A relatively constant
isotopic ratio of 1/4 (^^^Pu/^-'^'^'*°Pu) prevailed throughout the sample series analyzed.
There was little or no correlation between plutonium and ^^^Cs concentrations in flesh
On a seasonal basis, and '^^Cs concentrations were generally 2.5 x 10^ times as great as
2 3 9,2 4op|^j concentrations. Concentrations of ^"^^Am in 35 of the preceding flesh
samples were of the same general values as the plutonium concentrations but were often
more variable.

Radionuclides in Carnivore T^issues

A limited number of carnivore tissues have thus far been analyzed for their transuranic
nuclide content. The results were so variable and near the minimal detectable amounts of
radionuclides that larger sample volumes (up to 100 g of ash) of most mammalian species
were analyzed to provide more positive values for meaningful comparison and
interpretation. The results (Table 11) showed that (1) only 2 3 9,2 4 0pjj (.Q^l^J ]jq reliably

TABLE 11 Concentrations of^3^Pu,"''-^'*°Pu, and '* ^ Am in Large

(38- to 100-g ash) Muscle and Bone Samples Composited from Several Animals

Collected at Anaktuvuk Pass, Alaska, During April and May 1976

Species
sample type

Sample

ash
weight, g

Radionuclide concentration,!
fCi/g ash

238pu

2 3 9 ,2 1 0 pu

^^'Am

Caribou (Rangifer

arcticusj
Muscle

60.5

-0.19 ± 0.17

0.90 ± 0.20

-0.1 ± 0.3

Bone

83.4

0.03 ± 0.07

0.23 ±0.11

0.7 ±0.2

Bone

100.2

0.33 ± 0.11

0.53 ±0.14

1.0 ± 0.3

Wolf (Canis lupus)
Muscle

43.0

-0.40 ± 0.20

0.80 ± 0.40

0.3 ± 0.4

Bone

71.7

-0.09 + 0.11

0.30 ± 0.13

0.2 ±0.3

Bone

92.2

0.07 ± 0.15

0.05 ± 0.10

0.1 ±0.2

Fox ( Vulpes fulvaj
Muscle

38.0

-0.30 ± 0.30

0.40 ± 0.40

-0.1 ± 0.5

Bone

85.6

-0.07 ± 0.12

0.09 ±0.13

0.1 ±0.2

* Number of animals.

fMean ± standard error (counting statistics).

reported from these composited animal samples. (2) caribou tlesh contained slightly
higher 2 3 9,24 0p|j concentrations than wolf and fox flesh, and (3) there was no indication
of biomagnification in the upper trophic levels of the food web. Concentration ratios of
the plutonium isotopes were generally similar; the lichen/soil ratio was 140—180; the
caribou/hchen flesh ratio was 0.004-0.005; and the wolf/caribou ratio was 0-0.9.

Conclusions

Two major periods of worldwide fallout deposition on northern Alaskan and Greenland
ecosystems have occurred as a result of atmospheric nuclear weapons tests of Great

456 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Britain, Russia, and the United States, the first and most sustained during 1953 to 1959
and the second during 1961 to 1964. Recent additions of fallout have been made by
atmospheric tests conducted by France and the People's Republic of China; tests
conducted during the summers of 1973 and 1974 by China have apparently made
important contributions of radionuclides of interest in this chapter.

Estimates of ^^^'^''"Pu fallout inventories in the Arctic landscapes discussed in this
chapter began during 1959 and 1960 and have recently been estimated to be 0.33
mCi/km^ at Thule, Greenland, and 0.40 to 0.60 mCi/km^ in northern Alaska. If an
average ^^^'^^^Pu/^^'^Cs, ratio of 0.016 is assumed, these values translate to 20.6 mCi
^^''Cs/km^ at Thule and 35.6 mCi '^^Cs/km^ at Anaktuvuk Pass. This is reasonably
close to the amounts calculated from HASL fallout deposition data for the period after
1959 and 1960, when the measurements began. A large increment of the pre-1959 fallout
is therefore unaccounted for in both the soils (to a depth of 5 cm) and lichen
communities studied in this chapter. The lichen communities of northern Alaska during
1964 contained 41 nCi ' ^ ^Cs/m^ (mCi/km^), and values since that time have fluctuated
near 35 to 40 nCi/m^ . This is equivalent to a predicted level of 0.56 to 0.64 nCi
239,240p^/j^2 j^^^g^ Qj^ ^j^g assumption of a ^^^'^"^^Pu/'^ ^Cs ratio of 0.016. However,
measurements of ^^^Pu and ^^^'^'^^Pu in lichen samples were about one-half the
predicted concentration. This lower value was confirmed by consistently lower
239,240py|i 3 iQ^ ratios (near 0.006 to 0.012) in Hchens during recent years.

Isotopic ratios in Greenland soil and alluvium samples during 1974 were in the range
of 0.011 to 0.014 for "^'^^°Pu/^^^Cs and 0.010 to 0.019 for ^^^Pu/^^^'^^^Pu.
Although these values do not strongly indicate the presence of plutonium released by the
1968 aircraft accident that deposited an estimated 1 to 5 Ci of ^^ ^'^'^'^Pu on the Thule
landscapes, the ratios are substantially lower than the 0.020 to 0.024 values usually
found. The presence of 2 3 9,24 0p|j pa^^j^-igs j^ soils, alluvium, and lichens of the Thule
environs in both 1968 and 1974 samples from sites south and southwest of the 1968
crash site was inferred from the extreme variation in sample aliquots from those areas.
Uncontaminated areas of Thule showed a more balanced distribution of plutonium
concentrations in soil and lichens, probably because of the edaphic and climatic factors of
the region which resuspended soils to a considerable degree.

Samples of soils and lichens from northern Alaska contained lower plutonium
concentrations in relation to fallout deposition than were noted in Thule samples.
Lichen/soil ratios during 1975 and 1976 were 0.92 and 13.0, respectively; this contrasts
with the plant/soil ratios that are often of the order of 10""* to lO"'' in Temperate Zone
environments (Francis, 1973). Caribou/lichen ratios of ^^^'■^'^^Pu concentrations were in
the range of 10"^ to 10^'* , and carnivores contained transuranic nuclide concentrations
that were equal to or less than those in caribou, which were undoubtedly their major
food source.

Acknowledgments

I am grateful to Mary Ann Hanson for long hours spent in the meticulous separation of
the lichen sample components; to Eliza Trujillo, William Goode, and Kenneth Bostick
for laboratory assistance; and to Daryl Knab, Richard Peters, and David Curtis for
plutonium analyses. Gary White assisted in statistical analyses of the data. Ludi Kupinski
and Karen Tallent rendered secretarial and editorial assistance. This work was performed
under U.S. Department of Energy contract No. W-7405-ENG-36.

TRANSURANIC ELEMENTS IN ARCTIC TUNDRA ECOSYSTEMS 457

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No. 1, Proceedings of Symposium on Radioecology, Oregon State University, May 12-14, 1975,

pp. 29-39, Academic Press, Inc., New York.
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Eskimos, in Symposium on Radioecology, Proceedings of the Second National Symposium, Ann

Arbor, Mich., May 15-17, 1967, D.J. Nelson and F. C. Evans (Eds.), USAEC Report

CONF-670503, pp. 627-634, NTIS.
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Report HASL-286, pp. 1-2 to I-IO, Health and Safety Laboratory, NTIS.
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Summary Report, USAEC Report H.ASL-288. pp. 1-2 to 1-9, Heahh and Safety Laboratory, New

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, and N. Chu, 1967, The Ratio of ' ^ 'Cs to '" Sr in Global Fallout, in Fallout Program Quarterly

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HASL-300, pp. B-05-01 through B-05-11, Health and Safety Laboratory, NTIS.
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Laboratory, NTIS.
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(Ed.), Health and Safety Laboratory, NTIS.
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438 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

IialrrU.E.-i and R. B. R. Persson, 1975, Fail-Out Plutonium in Swedish Reindeer Lichens, Health Phys.,

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NTIS.

Nevada Applied Ecology Group Model

for Estimating Plutonium Transport

and Dose to Man / ^ " ^ *

DOCUMENl

COLLECTION

W. E. MARTIN and S. G. BLOOM

A Standard Man is assumed to live in and obtain most of his food from a
plutonium-contaminatcd area at the Nevada Test Site (NTS). A plutonium-transport
model, based on the results of other Nevada Applied Ecology Group plutonium studies, is
used to estimate potential chronic rates of ^ ^ '^Pu inhalation and ingestion as functions of
the average concentration of ^^^Pu (C^, picocuries per gram) in the surface soil (0- to
5-cm depth) of the reference area. A dose-estimation model, based on parameter values
recommended in publications of the International Commission on Radiological Protec-
tion (ICRP), is used to estimate organ burdens, accumulated doses, and dose
commitments as functions of exposure time. Tliese estimates are combined with ICRP
recommendations for allowable public exposure to radiation to arrive at acceptable soil
concentrations at NTS.

The plutoniimi-transport model is based on a relatively simple ecosystem that was
used as a preliminary model to guide data-acquisition studies at NTS. Tlie preliminary
model provides a framework for developing more detailed dynamic models of the
ecosystem, but at present there are insufficient data to implement these dynamic models;
so the estimates of inhalation and ingestion rates are based on simpler steady-state
models. If we assume the transport system to be in steady state, the estimated inhalation
and ingestion rates (picocuries per day) are 0.002 Cs and 0.2 Q, respectively.

A number of dose-estimation models were examined, and calculations were made for
comparison. Tlie results of these calculations indicated that the dose estimates to the
most sensitive organs were comparable. The model recommended by the Task Group on
Lung Dynamics of ICRP was used for dose estimates at NTS because it is the model most
widely accepted. Estimated doses (rem) due to chronic inlialation and ingestion of^^'^Pu
for 50 vr at the rates indicated above are: thoracic lymph nodes, 0.610 C^: lungs, 0.025
Cs; bone, 0.014 C^; liver, 0.009 Cs; kidney, 0.003 Cs; total body, 0.0007 Cs; and
gastrointestinal tract (lower large intestine), 0.0002 Cs- Inhalation accounts for 100% of
the estimated dose to the lungs and thoracic lymph nodes and for about 95% of the
estimated dose to bone, liver, kidney, and total body. Ingestion accounts for >99%o of the
dose to the gastrointestinal tract.

According to the ICRP (International Commission on Radiological Protection, 1966)
recommendations for individual members of the public, the dose rate to the lungs after
50 yr exposure should not exceed 1.5 rem/yr. Tlie plutonium-transport and dose-
estimation models described in this chapter indicate that the average concentration of
^^^Pu in the surface (0 to 5 cm) soils of contaminated areas at NTS which could residt in
a maximum dose rate of 1.5 rem/yr to the lungs is approximately 2.8 nCi/g, or about 140
IdG/m'^ for soils weighing 1 g/cm^.

459

460 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

An important goal of the Nevada Applied Ecology Group (NAEG) plutonium program is
to evaluate the potential radiological hazard to man due to the presence of plutonium in
various nuclear safety test areas at the Nevada Test Site (NTS). Since the contaminated
areas of interest are uninhabited, we have based our analysis on the assumption that a
Standard Man resides in and obtains most of his food from a plutonium-contaminated
area at NTS.

In this chapter we use information provided by other NAEG studies to develop a
plutonium-transport model that attempts to characterize the general behavior of
plutonium in a typical NTS ecosystem and to provide a basis for estimating potential
rates of plutonium ingestion and inhalation by the hypothetical Standard Man. We discuss
the mechanisms involved in the transport processes and, in most cases, include
appropriate mathematical expressions for these mechanisms. However, the final form of
the transport model is determined by the available data, which often limits us to using
only the simplest mathematical expressions.

The estimates of inhalation and ingestion rates provide the input for a dose-estimation
model that is used to calculate potential organ burdens, cumulative organ doses, and dose
commitments due to chronic inhalation and ingestion of ^^^Pu. Although several models
are considered, the preferred dose-estimation model is based entirely on the recommenda-
tions and publications of the International Commission on Radiological Protection (1959;
1964; 1966; 1972).

Finally, a procedure is described whereby the combined results of the transport
model and the dose-estimation model can be applied to the practical problem of deciding
whether and to what extent environmental decontamination might be required to limit or
reduce potential health hazards due to plutonium.

A preliminary model of potential plutonium transport from the environment to man
was introduced during the planning stage of the NAEG plutonium program to ensure
consideration of laboratory and field studies that would provide the data and parameter
estimates required for implementation of more detailed transport and dose-estimation
models to be developed later in the program. This model forms the basis for discussing
the various transport mechanisms in this chapter. Some of the parameters sought at the
outset have proved to be elusive or impossible to measure accurately, and consequently
the proposed dynamic model has not been fully implemented. This chapter represents our
best effort to ^udge and interpret the information currently available and to select the
best available methods for estimating potential intake rates and doses. Tlie design of the
transport and dose-estimation models plus the assumptions and parameter values selected
for their implementation comprise what we believe to be a reasonable and conservative
working hypothesis that provides a method for evaluating the potential health hazards
associated with plutonium-contaminated areas at the NTS. As a working hypothesis, it is
subject to continuing reappraisal, and the results or conclusions derived from it are
subject to unavoidable uncertainties. To a considerable extent, however, these uncer-
tainties are compensated for by conservative assumptions, which tend to result in
overestimates of potential intake rates, organ burdens, and doses rather than underesti-
mates.

Plutonium-Transport Model (Preliminary Model)

Figure 1 is a diagram of the potential transport pathways considered in the preliminary
planning model. The large square represents an arbitrary boundary of a contaminated
area. Boxes represent the principal ecosystem compartments of interest, and arrows

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 461

represent net transport via the pathways indicated. Arrows that cross the arbitrary
boundary represent net transport out of the system.

The distribution of plutonium in the contaminated areas of principal interest at the
NTS has been described by Gilbert et al. (1975). Present levels of soil contamination in
the areas of interest range from about 1 .0 ^Ci/m^ to >6000 idCi/rn^ . Because these levels
of soil contamination resulted from nuclear safety tests carried out from 1954 through
1963 and because current worldwide fallout rates are insignificant compared with existing
levels of contamination, Fig. 1 shows no current plutonium input to the system.

ARBITRARY BOUNDARY

(RESUSPENSION)

- ^C)\ 1

AIR

1 *

VEGETATION

HERBIVORES

MAN

Fig. 1 Principal pathways of plutonium transport to man.

Under these conditions the plutonium concentration in soil is the principal factor
forcing the transport system. Air is contaminated by resuspension of plutonium-bearing
soil particles. Vegetation is contaminated internally by root uptake from soil and
externally by deposition of resuspended particles. Plutonium input to herbivores is due to
ingestion of soil and vegetation and to inhalation. Plutonium could reach man by
inhalation of contaminated air, by accidental ingestion of contaminated soil, by ingestion
of contaminated vegetation, and by ingestion of milk or meat (skeletal muscle or internal
organs) from animals raised in the contaminated area. Drinking water for herbivores and
man is assumed to come from deep wells or from sources outside the contaminated area
and to contribute nothing to plutonium intakes by herbivores or by man. Numerous
other pathways, most of them trivial and unsubstantiated, could be postulated, but we
have tried to limit our consideration to genuinely important pathways.

If it is assumed that (1 ) the major ecosystem compartments and important transport
pathways are as indicated in Fig. 1,(2) the plutonium in each compartment is well mixed
with the other contents of the compartment, and (3) the net rate of transfer from one
compartment to another is the product of a transfer coefficient and the quantity of
plutonium in the transmitting compartment, then the intercompartmental flux of

462 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

plutonium is represented by a system of linear, first-order, ordinjiry differential
equations, the general formula for which is

n 11

^=X^..Y,-Y,^Xj. 0 = 1,2,3... n) (I)

i=i i=i

where j is the compartment of reference and all other compartments are designated i, Yi
is the amount (picocuries) of plutonium in compartment j at time t (days), and Xy and Xjj
are transfer coefficients (day" ' ) for flows into and out of compartment j. The positive
expression on the right side of Eq. 1 represents the flow rate into compartment j, and the
negative expression represents the flow rate out of compartment j. The amount present in
a given compartment at a given time, Yj(t), is therefore dependent on the rates of input
and output.

In a general way, Fig. 1 and Eq. 1 identify the principal kinds of information needed
to estimate the transport of plutonium to man. The compartments of Fig. 1 indicate the
principal ecosystem components, and the arrows indicate the pathways of transport from
environment to man via inhalation and ingestion. Equation 1 suggests that intercompart-
mental How rates might be expressed as the product of a transfer coefficient and the
quantity of plutonium in the transmitting compartment. It was recognized, however, that
some parts of the transport system (Fig. 1) might not behave in accordance with the
first-order kinetics model suggested by Eq. 1. Consequently the objectives of the NAEG
plutonium studies were stated in broader terms. The general objectives related to the
estimation of potential human ingestion and inhalation rates were simply to (1)
determine plutonium concentrations in ecosystem components and (2) quantify the rates
of plutonium transfer among ecosystem components.

In the remainder of this chapter, we discuss in turn each compartment of the
preliminary model (Fig. 1), soil, air, vegetation, herbivores (cattle), and man. In the
sections on soil, air, vegetation, and cattle, we describe what is known about the
compartment and discuss the processes that involve it in the transport of plutonium to
man. In the section on man, we provide methods for estimating plutonium inhalation and
ingestion rates, based on concentration in soil. These rates are used in the section on
Dose-Estimation Models to estimate organ burdens, cumulative doses, and dose
commitments by alternative methodologies. In the final section. Practical Applications,
we show how the results of the plutonium-transport and dose-estimation models can be
used to determine an "acceptable soil concentration."

We wish to emphasize at the outset that this is not a definitive study of the behavior
of plutonium in desert ecosystems. It is merely an inquiry that asks how we can best use
the theory and data presently available to obtain a reasonable assessment of potential
hazards and a credible criterion on which to base preliminary consideration of
countermeasures that may or may not be planned and executed in the future. Our study
identifies some of the obstacles between present knowledge and a workable cleanup
criterion and recommends a pro tem path around these obstacles. In plotting this
sometimes tortuous path, we have encountered theory that cannot be applied for lack of
data, and we have encountered data that cannot be used because they are too scanty to
be fitted into the present theoretical framework. The result is a compromise between
knowledge and ignorance. We make use of the knowledge we have, but we are made
uneasy by the awareness that there are other paths, perhaps equally defensible, which

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 463

may lead to far different conclusions. Or, to put the matter more bluntly, the present
state of the transport- and dose-modeling art is such that, by careful selection of
published parameter values and model equations, one could obtain a preselected result.
We have made every effort to avoid doing this, but we feel obliged to offer this comment
to warn the reader that such efforts are necessary.

Soil

Plutonium Concentration in Soil. Various soil surveys have been conducted to delineate
highly contaminated areas at NTS, to determine the horizontal and vertical distribution
of plutonium in contaminated soils, and for various other purposes (see several papers in
reports by Dunaway and White, 1974; White and Dunaway, 1975; 1976; 1977).
Inventories of 239,240p^ -^^ ^l^g surface soils (0- to 5-cm depth) of NAEG study areas
were reported by Gilbert et al. (1975, p. 379) and revised by Gilbert (1977, p. 425).

As mentioned earlier, soil is the principal reservoir for plutonium at NTS, and soil
concentration (picocuries per gram) is the factor that drives or forces the transport
system. In developing equations to estimate potential plutonium inhalation and ingestion
rates for the hypothetical Standard Man, we shall attempt to relate the concentrations in
air and foods to the average concentration in soil. Soil concentrations based on data
provided by Gilbert et al. (1975) are given in Table 1. The estimated inventories (as
revised) are given in Table 2.

TABLE 1 Average Concentrations of ^ ^ ^ '^ "* ° Pu in Surface Soils
(0- to 5-cm Depth) of NAEG Study Areas*

Study area

Strata

239,240pu^|

nCi/g

Study area

Strata

239,240py^.^

nCi/g

0.036 ±0.0078

Clean Slate 2

0.086 ± 0.028

0.10 ±0.025

1.8 ± 0.74

0.40 ± 0.075

6.2 ±2.5

1.1 ±0.15

5.4 ± 1.4

5
6

20
47

2.4 ± 0.43
14 ± 6.4

Clean Slate 3

1
2

28
12

0.24 ± 0.046
1.2 ±0.33

5 (GMX)

0.059 ±0.013

4.6 ± 1.3

0.73 ±0.15

7.9 ± 3.9

4.5 ± 1.2

Area 1 1 sites

7.3 ± 1.6

<5000-cpm region

0,021 ± 0.0066

0.084 ± 0.03

CD overlap

0.30 ±0.13

Double Track

1
2
3
4

0.12 ±0.057
5,7 ±4.0
2.9 ± 0.97
44 ± 15

B site
C site

2
3
4
2
3

12
14
23
12
10

0.73 ± 0.45

5.5 ± 1.4

33 ± 7.0

0.85 ± 0.44
2.2 ± 0.65

Clean Slate 1

0.36 ±0,17

26 ± 7.5

1.6 ±0.57

120 ±52

2,7 ± 1.1

D site

1.0 ± 0.20

2.9 ± 0.97

3
4

12
18

4.3 ± 1.6
18 ±6.5
49 ± 15

*Bascd on data from Gilbert et al. (1975, pp. 393-395).

jMean ± standard error (SE). SE = s/(n) '2, where n is the number of samples.

464 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Estimates of Inventory of ^ ^ '^ '^ "* ° Pu in Surface Soil (0- to 5-cm Depth)

of NAEG Study Areas*

Size of

Estimated

Percent

area

239,240py^

inventory,!
Ci

of total

Area

Strata

Percent

MCi/m=

inventory

1,245,000

31.0

1.9 + 0.34

2.4 ± 0.42

2,547,000

63.4

5.8 ± 1.4

15 ± 3.6

108,000

2.7

23 ±4.3

2.5 ± 0.46

74,000

1.8

54 ±8.8

4.0 ± 0.65

19,000

0.5

110± 19

2.1 ± 0.36

24,000

0.6

820 + 340

20 ± 8.2

Total

4,017,000

100.0

169

46 ± 9.0

100

5 (GMX)

111,300

88.8

3.1 +0.66

0.35 ± 0.073

8,400

6.7

42 ±9.1

0.35 ± 0.076

800

0.6

270 ± 64

0.22 ± 0.051

1,000

0.8

530 ± 150

0.53 ± 0.15

3,800

3.0

4.6 ± 1.6

0.02 ± 0.006

Total

125,300

99.9

111

1.5 ± 0.19

100

Double Track

176,000

98.3

6.7 ± 3.5

1.2 ± 0.62

1,600

0.9

350 ± 250

0.56 ± 0.40

800

0.4

190 ±59

0.15 ± 0.047

600

0.3

2,800 ± 1,000

1.7 ± 0.60

Total

179,000

99.9

3.6 ± 0.95

100

Clean Slate 1

157,000

88.9

15 ± 7.0

2.4+ 1.1

10,000

5.7

64 ± 22

0.64 ± 0.22

8,400

4.5

110± 35

0.92 ± 0.29

1,700

1.0

120 ± 39

0.20 ± 0.066

Total

177,100

100.1

4.2 ± 1.2

100

Clean Slate 2

351,000

74.7

4.1 ± 1.3

1.4 ± 0.46

82,300

17.4

73 ± 30

6.0 ± 2.5

26,200

5.5

270 + 99

7.1 + 2.6

1 1 ,000

2.3

260 ± 65

2.9+ 0.72

Total

470,500

99.9

17± 3.7

100

Clean Slate 3

1,615,000

93.2

12± 2.2

19.4 ± 3.6

61,000

3.5

58 + 16

3.5 ± 0.98

40,000

2.3

210± 63

8.4 ± 2.5

16,000

0.9

370 ± 190

5.9 ± 3.0

Total

1,732,000

99.9

37 ± 5.4

100

Area 1 1 sites

<5000-cpm

region

4,672,000

96.7

0.97 ± 0.30

4.5 ± 1.4

12.4

CD overlap

62,200

1.3

12 ±5.2

0 75 ± 0.32

2.1

B site

8,200

30± 18

0.25 + 0.15

6,000

220 ± 55

1.3 ± 0.33

3,300

1,400 ± 300

4.6 ± 0.99

Total

17,500

0.4

6.2 ± 1.1

17.0

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 465

TABLE 2 (Continued)

Size of

area

:2 3 9 ,2 4 Opu

Estimated
inventory ,t

Percent
of total

Area

Strata

Percent

nCilxn'

inventory

Csite

16,400

34 ± 22

0.56 ± 0.36

5,600

88 ± 25

0.49 ±0.14

3,500

1,400 ±390

4.9 ± 1.4

300

6,200 ± 2,800

1.9 ± 0.84

Total

25,800

0.5

7.8 ± 1.7

21.6

Dsite

32,300

46 ± 9.5

1.5 + 0.31

13,300

220 ± 86

2.9 ± 1.1

4,900

990 ±370

4.9 ± 1.8

2,900

2,700 ± 840

7.8 ± 2.4

Total

53,400

1.1

17.1 ± 3.2

47.0

\rea 11 Total

4,830,900

100.0

36 ±4

100.1

*Based on data from Gilbert (1977, p. 425).

fMean ± standard error (SE). SE = s/{n)'2= standard error.

Losses from Soil Compartment. As suggested by Fig. 1, plutonium can be transferred
from the soil compartment to compartments representing other ecosystem components.
It can also be removed from the soil of a given area by water or wind erosion. Percolation
into the profile could remove plutonium from the surface, where it is most susceptible to
resuspension, and could, if the soil were plowed and rainfall were plentiful, transport
some plutonium below the root zones of crop plants.

Owing to the extreme variability of plutonium concentrations in soil samples taken
from the same general area and to the arbitrary nature of soil-compartment boundaries
(usually specified by a depth measurement), it would be difficult to design field studies to
estimate the overall rate of plutonium loss from the soil compartment. In fact, no such
studies have been undertaken in the field or in the laboratory, and we have no basis for
assuming that the average concentrations of plutonium in the soils of large contaminated
areas will decrease significantly in the next lOOyr or so. Consequently the soil
concentrations given in Table 1 will be treated as constants for the areas indicated; i.e.,
the soil compartment is assumed to be a continuous and constant source for plutonium
transfer to other compartments. In the absence of any evidence that the rate of
plutonium loss is, in fact, significantly greater than the rate of loss due to radioactive
decay, the equation for the soil compartment is

Cs = Cs(0)exp(-XAt)

(2)

where Cs = average concentration of plutonium in the surface soil of a contaminated area
at time t (pCi/g)

Cs(0) = initial concentration as given in Table 1

Xa = radioactive decay rate of ^^''Pu (7.7829 x 10"^ day^')

t = time (days)

466 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Air

As indicated in Fig. 1 , plutonium contained in surface soils can be resuspended and
transported to vegetation via external deposition or to herbivores and man via inhalation,
and some of it can be carried by wind and redeposited beyond the arbitrary boundary. In
the absence of data to the contrary, we have assumed that deposition and resuspension
processes in contaminated areas at NTS are in approximate steady state, althougti data
presented by Anspaugli and Phelps (1974, pp. 292—294) suggest that resuspension may
exceed deposition, at least to a small degree. Several methods have been suggested for
analyzing and modeling deposition and resuspension processes. These are discussed in the
following paragraphs. Of these, the mass-loading approach requires the least information
for implementation and was used in the present model owing to the absence of data to
implement the other methods at NTS.

Deposition Velocity. The rate at which resuspended plutonium is deposited on soil
could be estimated as the product of a deposition velocity (centimeters per day) and
concentration in air (microcuries per cubic centimeter) to yield a rate that has dimensions
of juCi cm"^ day" ^ . Deposition velocities are functions of meteorological factors and the
aerodynamic properties of plutonium-bearing soil particles and soil surfaces.

Deposition velocities measured under field conditions have been reported by Van der
Hoven (1968), Sehmel. Sutter, and Dana (1973), and Healy (1974). Measurements under
controlled conditions in a wind tunnel have been reported by Sehmel, Sutter, and Dana
(1973) and Sehmel (1973: 1975). These data indicate that the deposition velocity
increases with increasing air velocity, increases with increasing particle size for sizes
greater than about 1 jum, increases with decreasing particle size for sizes less than 0.01
/im, exhibits a minimum somewhere in tlie range of 0.01 to 1 jum, and is strongly
influenced by tiie type of surface roughness. The wind-tunnel data of Sehmel et al.
(1973) for grass surfaces indicate that the deposition velocity is approximately
proportioned to both air velocity and particle size in the range of 2 to 12 m/sec and 1 to
100 iJim. Tliese grass data appear to correspond closely to field conditions provided that a
proper value is assigned to surface roughness.

Tamura (1976) has reported that more than 65% of the plutonium in soil samples
from Area 13 is associated with soil particles in the range of 20 to 53 /jm. Using the grass
data of Sehmel, Sutter, and Dana (1973) at 2.2 m/sec, the corresponding range of
deposition velocities is from 3 to 20 cm/sec. Particles on the order of 20 to 50/^m could
play an important role with respect to external contamination of vegetation, but particles
tills large are of little concern with respect to inhalation. Since respirable particles are
generally <10 [dm. the corresponding deposition velocities suggested by the grass data
would be <1 cm/sec.

Deposition Models. Both Healy (1974) and Sehmel (1975) present results of models
used to predict deposition velocities. Healy's results indicate that deposition velocity is
proportional to air velocity and is strongly dependent on atmospheric stability. Sehmel's
results indicate that deposition velocity increases as a nonlinear function of air velocity,
exhibits a minimum value as a function of particle size, and is not strongly dependent on
atmospheric stability. Both sets of results indicate a strong dependence on surface
rougliness. To apply either model to field conditions, we must estimate or measure the
surface roughness and velocity profile, both of which are variable.

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 46 7

For most applications to NTS, the grass data of Selunel et al. (1973) appear to be the
best analog. The trend of these data, in the range of 2 to 12 m/sec air velocity and 1- to
100-jum particle diameter, is approximately

^=3x10-^ dp (3)

where Vj is the deposition velocity (cm/sec), U is the wind velocity (cm/sec), and dp is
the particle diameter (jum). For lO-jLtm particles, Eq. 3 yields values similar to Healy's
results for neutral atmospheric stability.

Resuspension Factor. The resuspension of plutonium from soil is often expressed as the
ratio of air concentration (microcuries per cubic meter) to surface soil concentration
(microcuries per square meter). Many such measurements have been made at NTS (Mork,
1970; Anspaugh and Phelps, 1974) and in the vicinity of Rocky Flats, Colo. (Volchok,
1971). The measured magnitudes of this ratio range generally from 10~^ to 10~^ ' m~^ .
To estimate "acceptable soil concentrations," Anspaugh (1974) used a value of 10~^
m~^ for NTS. These ratios are, to say the least, extremely variable with respect to time
and environmental factors, such as wind speed and direction, rainfall, and disturbances
affecting aerodynamic properties of soil surfaces. Other factors affecting this ratio are the
aerodynamic properties of plutonium-bearing particles and their susceptibility to saltation
and resuspension. There is evidence that the ratio tends to decrease with time after fallout
contamination of soil (Anspaugli et al., 1973; Anspaugh, 1974; Kathren, 1968).
Anspaugh et al. (1975) have proposed a model in which the air/soil ratio decreases as a
function of time from a maximum of lO"'* to a minimum of 10~^ m~^ , i.e.,

^=10-Vxp[-k(f)'^] +10-^ (4)

where Ca = air concentration (juCi/m'*)

Css = soil surface concentration (juCi/m^)
k = 0.15day-'^
t = time from deposition (days)

This model is consistent with data collected over the years at NTS.

Resuspension Models. Many attempts have been made to develop mathematical models
to simulate resuspension (Amato, 1971; Mills and Olson, 1973; Killougli and McKay,
1976). Most of these are based on models of wind erosion developed by Bagnold (1960)
and, as a function of wind speed, take the form

Ca = K (U - Ut)' ^ (5)

where Uy is a tliieshold wind speed (m/sec) and K is a constant (sec/m^).

Others (Sehmel and Orgill, 1973; Shinn and Anspaugh, 1975) have used a power-law
expression of the form

Ca = K U" (6)

--SS

468 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

where K and n are empirical constants derived from the data. Sehmel and Orgill (1973)
found n = 2.1 when they fit Volchok's (1971) data for plutonium resuspension at Rocky
Flats to Eq. 6. Shinn and Anspaugli (1975) also found n = 2.1 for dust flux at NTS. We
approximately fit Sehmel's (1975) data for calcium molybdate resuspension at Hanford
and also arrived at a value of about 2.1 . The only contrary data are Shinn and Anspaugli 's
results for a plowed field in Texas, wliich yielded n = 6.4.

The empirical value of about 2 for n when derived for different tracers, different soils,
and different climates (provided that the soil is undisturbed) tends to provide indirect
confirmation for the theoretically derived form of Eq. 5. However, both K and Uj in
Eq. 5 are functions of particle size, soil moisture content, surface roughness, relative
humidity, and the time period over which the wind speed is averaged. Some attempts
have been made to theoretically include many of these factors (especially particle size),
but the theory does not seem to describe adequately the variations in the data. Thus K
and Ux must be treated as empirical constants for the present. Consequently there is no
practical benefit in using Eq. 5 in preference to the simpler Eq. 6. However, at least one
experimental measurement of resuspension and wind speed must be made to set the value
of K in Eq. 6 for the particular area.

Mass Loading. In the absence of data to implement Eq. 6 for a given area, Anspaugh
(1974) suggests that a mass-loading factor (L^) of 100 jug(soil)/m^(air) be used for
predictive purposes. If we assume that the radioactivity of one square meter is associated
with 50 kg of soil (5-cm depth x 10'* cm^/m^ x 10~^ kg/cm^), a mass-loading factor of
100 Mg/m'^ is equivalent to a resuspension factor of 2 X 10"^ m~' . The theoretical basis
for the mass-loading approach is described by Anspaugh (1974). Anspaugh et al. (1975)
provide comparisons showing that predicted air concentrations based on L^ = 100 jUg/m'^
are in good agreement with measured air concentrations.

We shall use the suggested mass-loading factor to represent average conditions at NTS,
but it must be noted that higlier than average wind velocities (Shinn and Anspaugh, 1975)
or mechanical disturbances, such as plowing (Milliam et al., 1976), could cause the
mass-loading factor to be temporarily much higher than 100 jug/m^. It should also be
noted that some recent work by Sehmel (1977) suggests little if any experimental
justification for this approach.

Vegetation

As shown in Fig. 1, vegetation can be contaminated externally by deposition of
resuspended material or internally by uptake from soil or by both processes simulta-
neously. Other mechanisms of external and internal contamination have been identified
or postulated, but direct deposition from air and root uptake appear to be the processes
most important to consider when attempting to develop a general model.

In the following paragraphs we discuss the mechanisms involved in contaminating
vegetation and present mathematical expressions to simulate the dynamics of the
contaminating mechanisms. We also discuss the parameters in these expressions and their
variations under the influence of different environmental factors. However, we conclude
that there are too few data to develop an adequate dynamic model, and we are forced to
use a simple steady-state model with a constant vegetation-to-soil contamination factor in
the overall transport model.

General Hypothesis. Externally deposited material can be removed from plant surfaces
by weathering, i.e., the mechanical action of wind and rain, and it can be diluted by plant

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 469

growth. Internally deposited material can also be diluted by growth but not by
weathering. Processes that remove biomass from vegetation (e.g., grazing, cropping, root
decay, and dehiscence of above-ground parts) also remove plutonium. If they exceed
growth rates, these processes can reduce the total amount of plutonium in the vegetation
compartment of an ecosystem. Different plant species can vary widely with respect to
their ability to retain externally deposited plutonium or to assimilate plutonium from
foliar deposits or soil, and translocation witliin the plant can result in large differences
regarding plutonium concentrations in different plant organs. In this discussion we do not
attempt to distinguish one plant species from another. We assume that plutonium is
uniformly distributed in edible plant materials and that processes which remove biomass
from the vegetation compartment have no effect on the concentration of plutonium in
the remaining biomass.

Differential equations expressing the principal processes described above can be
written:

''^ = kavCa - (>^\v + K + >^a) Yve (7)

dyvi
dt

ksvCs-(X, + AA)yvi (8)

where yve = concentration in vegetation of externally deposited plutonium (pCi/g)
kgv = air-to-vegetation deposition rate coefficient (m^/g • day)
Ca = concentration of plutonium in air (pCi/m^ )
X\v = weathering rate coefficient (day~^ )
Xg = vegetation growth rate coefficient (day~^ )
Xa = radioactive decay rate coefficient for ^^^Pu(day~^)
yvi = concentration in vegetation of internally deposited plutonium (pCi/g)
ksv = soil-to-vegetation uptake rate coefficient (day"')
Cs = concentration of plutonium in soil (pCi/g)

Equations 7 and 8 represent the external and internal components of plutonium in
vegetation. The former is due to foliar deposition; the latter, to root uptake. It is assumed
that plutonium taken up via roots can be translocated to stems and leaves, but this rate is
difficult to estimate. Consumers of vegetation are connected to both compartments
simultaneously, and this is the same as summing the two components. Assimilation of
externally deposited materials and their translocation to other parts of the plant have
been demonstrated experimentally for various kinds of substances applied externally to
foUage, but, in the case of plutonium (which is most probably deposited on foUage in the
form of insoluble particles), foliar assimilation is assumed to be zero. A recent study
(Cataldo, Klepper, and Craig, 1976) has demonstrated that translocation of foliarly
deposited plutonium to roots and seeds can occur. However, the accumulation ratios
observed in the absence of a solution vector (simulated rainfall) were on the order of
10~^ for both fresh and aged PuOa ; i.e., tlie observed concentrations in leaf tissue were
about 200,000 to 500,000 times higher than the observed concentrations in seed and root
tissue, respectively.

Althougli foliar deposition and root uptake of plutonium have been studied
separately in a variety of experiments, there is no reliable method for distinguishing

4 70 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

between the two components in a plant sample if both are present. If plants are grown in
contaminated soil or culture media and only the aerial parts are assayed, we can assume
ihat the activity detected was internally deposited. If the aerial parts of plants are
collected from a recently contaminated area that had no plutonium in the soil before the
contaminating event, we can assume that all or nearly all the activity detected was
externally deposited. If, on the other hand, plant samples are taken from an area that was
contaminated years ago, it is likely that most of the plutonium contained therein is due
to external contamination by resuspended soil particles and that only a small fraction is
due to internal contamination by root uptake.

Recent evidence (Romney et al., 1975; Wildung and Garland, 1974) suggests the
possibility that (1) the biological availability of plutonium in contaminated soil may
increase with time after environmental release and that successive crops of annuals thus
would take up successively greater amounts of plutonium; (2) perennial plants may
accumulate plutonium to a much greater extent than previously indicated by short-term
uptake experiments; and (3) plant uptake may increase with decreasing plutonium
concentration in the soil. Either or all of these factors could have the effect of making
biological transport of plutonium to man progressively more important relative to
physical transport and inhalation. These considerations are interesting because most
assessments of the potential hazards of environmental plutonium attribute little
importance to biological transport and ingestion compared with physical transport and
inhalation.

Foliar Deposition. The estimation of deposition velocity, V^, i.e., the ratio of surface
deposition rate to air concentration, was discussed earlier as applied to soil surfaces.

Deposition velocities have also been determined experimentally for different kinds of
plants and several kinds of vegetation with respect to various aerosols and particulates. In
this study, however, it was more convenient to base estimates of air-to-plant deposition
rates on the product of air concentration, soil/air deposition velocity, and a plant (or
vegetation) interception factor. The plant interception factor is defined as the amount
initially deposited per gram (dry weiglit) of plant material divided by the amount initially
deposited per unit area of soil surface.

Much of the available information concerning die interception of airborne radionu-
clides by plants has come from studies made in the fallout fields produced by nuclear test
detonations at NTS. Romney etal. (1963) summarized early studies in the vicinity of
NTS. They found that levels of fallout deposition on plants varied with respect to (1)
downwind distance from the detonation point, (2) lateral distance away from the midline
of the fallout field, (3) a variety of morphological features associated with different plant
species, and (4) the level of fallout deposition on soil surfaces near the contaminated
plants. Although there was no constant relationship between fallout concentrations on
plants (disintegrations per minute per gram) and on soils (disintegrations per minute per
square meter), there was a good correlation between radioactivity in plant samples and in
the fraction of fallout samples with particle sizes less than 44 jum. Apparently, the larger
particles were deposited close to the detonation point and were not as readily intercepted
by plants as were the smaller particles, which were deposited farther downwind; or, if
they were intercepted, they were apparently more easily removed by weathering
processes. In the laboratory, Romney et al. (1963) found that 50 to 90% of the gross
radioactivity on fallout-contaminated plants could be removed by washing with water or
a wetting agent, such as versene. These and similar studies demonstrated the possibility of

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 471

predicting fallout interception by plants, but they did not provide quantitative methods
for doing so.

Miller and Lee (1966) carried out extensive studies of fallout interception by plants.
The plants were cultivated in gardens near San Jose, Costa Rica, and the fallout was
provided by continuing eruptions of Irazu, a nearby volcano. Miller and Lee also
developed a comprehensive theoretical model of fallout interception by plants. The
model assumes different sets of constants for different fallout particle size classes,
different morphological characteristics of foliage, and different meteorological condi-
tions. Unfortunately, their model is practically unworkable, in spite of its elegance,
because it requires the use of constants and other parameter values that are rarely, if ever,
available for predictive purposes.

The values of the interception factor determined experimentally by Miller and Lee
(1966) varied only sliglitly with respect to the different species of cultivated plants they
studied, and the only meteorological condition consistently correlated with large
differences in measured values of the interception factor was relative humidity. The
particles intercepted by plants were essentially the same sizes as those deposited on
adjacent soil surfaces. In both cases the mass median diameters for the volcanic dust
deposited as fallout were generally between 50 and 100 jLtm. Tlte weiglited averages of
interception factors for all the plant types tested (mostly garden vegetables) were
95.7 ± 66.9 cm^/g for damp exposure conditions (relative humidity greater than 90%)
and 47.4 ± 29.7 cm^/g for dry exposure conditions.

Interception factors based on nuclear testing experience are about one or two orders
of magnitude lower. For detonations involving tlie incorporation of large quantities of
soil material in the initial cloud, estimates range from 1.9 to 11 .1 and have a mean of 3.7
cm^/g. For detonations involving the incorporation of little or no soil material in the
initial cloud, the estimated values are about an order of magnitude lower.

Miller and Lee (1966) noted that the "foliar samples obtained at the weapons test
experiments were apparently subjected to an unknown degree of weathering before they
were taken, while the primary samples [in our studies] were collected at the end of a 12-
to 24-hr period of exposure to more or less continuous fallout from Irazu, and the weight
of dust deposited on leaves was often greater than the dry weight of the leaves." Heavy
dust deposits such as these are easily dislodged by the sliglitest mechanical disturbance,
and moderate rains were observed to remove more than 90% of the material deposited.

Martin (1965) studied the interception and retention of *^Sr and ^^M by desert
shrubs (primarily Atriplex confertifolia and Artemisia tridentata). His estimates of the
plant interception factor were based on concentrations of ^^Sr and ' '^M in plant samples
collected 5 days after fallout deposition and estimates of the theoretical deposition rates
for these two radionuclides on unobstructed soil surfaces in the same locations ranging
from about 10 to about 100 miles downwind from the detonation point. Estimates for
different study areas range from 1.49 to 11.05 cm^/g, with the higher values occurring in
the more distant areas. The overall mean for ^^Sr was 4.09 cm^/g, and the overall mean
for ^^M was 4.00 cm^/g, approximately an order of magnitude lower than Miller and
Lee's average value for dry deposition conditions. Although the discrepancy may appear
to be large, it may be due to the effects of weathering during the 5 days between fallout
deposition and the collection of plant samples.

Weathering Rate. To estimate the effective rates of ^^Sr and '^M loss from
fallout-contaminated plants, Martin (1964) coUected additional sets of plant samples at

472 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

intervals of 10, 15, 30, and 60 days after fallout. Factoring out the loss of ' ^ ' I due to
vaporization and the losses of both radionuclides due to radioactive decay, the data
indicated that the weathering (environmental) half-time for these two radionuclides
increased with respect to time after fallout. During the interval from 5 to 15 days after
fallout, the average weathering half-time was about 20 days. The value obtained for the
interval from 15 to 30 days after fallout was about 30 days, and it increased for the
interval from 30 to 60 days postdetonation to about 130 days. Tlie D+5-day
concentrations can be compared with assumed D+0 concentrations, which would
reconcile the difference between Martin's estimate of the plant interception factor and
Miller and Lee's estimate of the foliage contamination factor. This procedure suggests an
average weathering half-time of approximately 1.4 days during the interval from 0 to 5
days after fallout.

These observations lead us to the hypothesis that tlie decay-corrected concentration
of a radionuclide in fallout-contaminated plant material is a very rapidly declining
exponential function of time at times soon after the contaminating event but approaches
a lower asymptote. Since this hypothesis appears to be correct, the effective rate at which
a radionuclide is removed from surfaces following external deposition cannot be
expressed precisely by a single coefficient because the weathering half-time increases as a
function of time after contamination. If the initial deposition is a heavy one, a significant
fraction of it (perhaps as much as 90%) can be removed by weathering in a matter of
hours, or a few days at most. A portion of what remains after this initial period of fast
weathering (something in the range of 10 to 60%) is so tightly trapped that it cannot be
removed even by vigorous washing (Romney et al., 1963). Presumably, this nonremovable
fraction is composed predominantly of particles that are small and mechanically trapped
on plant surfaces.

The situation in the plutonium-contaminated area at NTS is one in which foliar
deposition of resuspended particles and the loss of these particles from foliage is a more
or less continuous process. If the turnover rate is rapid, a foliage/soil steady state would
be quickly established.

Plant Growth Rates. As indicated earlier, the growth of new plant tissue may dilute both
the external and the internal concentrations of plutonium or other transuranium elements
in plant materials. Since different plant parts may grow at different rates, it is obvious
that the growth rate of interest with respect to external contamination is the growth rate
of leaves (and other edible parts formed above ground). If we assume that internal
plutonium due to root uptake is uniformly distributed to all parts of the plant, the
growth rate of interest with respect to dilution of root uptake is the overall growth rate,
i.e., the growth rate of leaves plus the growth rate of all other plant parts.

Plant growth is not a continuous process, nor is it the same for all species in a given
area or for all the parts of a given plant. In the temperate zone, at least, plant growth is
confined to the warm season, and the rate of growth is not uniform througliout the
growing season because different plant organs develop at different times. Ignoring the
morphogenic aspects of plant growth (i.e., the differentiation and development of
structure), growth is most simply conceived as an increase in biomass (i.e., dry weight of
tissue per unit area). For armuals, the biomass at the beginning of the growing season
consists of seeds; for herbaceous perennials, for which the aboveground parts die back
during the winter, it consists mostly of roots and other belowground parts; for woody
perennials, it consists of roots and stems (mostly dead tissue) plus, in the case of

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 473

evergreens, leaves and vestiges of fruits produced the previous growing seasons. In
addition to these seasonal and species variations, it is reasonable to suppose that the
roots, stems, leaves, fruits, and other organs of a given species grow at different times and
at different rates and that additional variations can be expected in response to
environmental factors, such as temperature, soil moisture, and availability of nutrients.

To attempt a mathematical description of vegetation growth that includes all factors
mentioned above (and others not mentioned) would be a monumental undertaking. What
is needed at present is a simple expression of growth rate, as a continuous function, which
will provide a reasonable, but conservative, estimate of the potential overall concentration
of plutonium in plant materials that have been contaminated externally by airborne
deposits and/or internally by root uptake from soil.

To obtain a rougli estimate of growth rate, we can define Xg as follows:

_ln[l+(Pn/Bo)]
^g 365 ^^^

where Xg is the growth rate coefficient averaged over the year (day~^ ), ?„ is the net gain
in biomass during a growing season (g/m^ ), and Bq is the biomass at the beginning of the
growing season (g/m^ ).

Odum (1971) has estimated that the average gross primary productivity (GPP) of
deserts and tundras is about 200 kcal m~^ yr""' . Since the fraction of GPP (0.2) used up
in respiration does not appear as new tissue, the dilution growth rate is proportional to
0.8 GPP = 160 kcal m"^ yr"^ At 4.5 kcal/g (dry weight) (Odum, 1971), this amounts to
a net gain of Pn = 36 g/m^ (approximately). The biomass of desert vegetation varies from
place to place. The mean biomass for Area 13 is Bq = 289 g/m^ (Wallace and Romney,
1972). Substituting these values of Pn and Bq in Eq. 9, Xg = 3 x 10""* day~^ . If we
assume that internally deposited plutonium is uniformly distributed above and below
ground, this would be the value to use in Eq. 8. If we assume that two-thirds of Pp is
above ground and one-third is below ground, the dilution growth rate for the external
(aboveground) component (Eq. 7) would be Xg = 2 x 10""* day~^ .

Root Uptake and Plant/Soil Concentration Ratio. For plutonium to enter plants via
root uptake, it must first reach the roots. Plowing, of course, accomplishes this
"transport" quite rapidly by mixing the soil, but the downward movement of plutonium
in an undisturbed soil profile is such a slow process that much of the plutonium deposited
on the surface may stay near the surface for many years. To circumvent the variability
inherent in these and other soil processes affecting the behavior of plutonium in soils
(factors reviewed by Price, 1973a; Francis, 1973), we have made the simplifying and
conservative assumption that plutonium deposited on soil is diluted by only the top 5 cm
of soil and that root uptake is related to the resulting concentration in surface soil; i.e.,
the probable concentration of plutonium in the root-zone soil is deliberately overesti-
mated.

Most of the available data (Price, 1973a; Francis, 1973) on plutonium uptake by
plants has been derived from short-term greenhouse experiments. Typical values thus
derived for the plant/soil concentration ratio range t>om 10"'^ to 10~^. Uptake has
been shown to be enhanced by the reduction of pH or the addition of chelating agents.
There is some evidence that plutonium uptake by plants may increase with time
(Romney, Mork, and Larson, 1970) and that the mobihty of plutonium, i.e., its ability to

474 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

move in the soil and its availability for root uptake, may increase with time as a result of
chelation due to bacterial decay of soil organic matter.

The evidence that the soil-uptake concentration ratio may increase exponentially as
a function of time is scanty but deserves consideration because of its implications.
Romney, Mork, and Larson (1970) grew ladino clover (Trifolium repens L.) on a
plutonium-contaminated soil under greenhouse conditions and cropped it repeatedly over
a period of 5 yr. The resulting estimates of Crv were as follows:

Crv.
(d/min)g"'' plants (dry weight)

Year (d/min) g"' soil (dry weight)

1.91 X 10"'

4.14 X 10-^

4.38 X 10-'

7.10X 10-'

13.95 X 10-'

A least-squares fit of these data to an exponential function yields

Crv = 1.31 X 10"^ exp (0.452 t)

where t is time (years) and 0.452 is the apparent growth rate coefficient; i.e., the data
indicate that Crv would be expected to double in about 1.5 yr. By extrapolation to 20
yr, the concentration ratio would be 0.1 1, a value within the range of field observations
for perennials (see below) but misleading nonetheless. Romney, Mork, and Larson (1970)
attribute the ''apparent" increase of Crv to root growth during the time the experiment
was being conducted and not to any change in the biological availability of the plutonium
contained in the contaminated soil.

In fact, the plant/soil ratios observed under field conditions are generally too high to
be explained by root uptake. Romney et al. (1975) collected soil and plant samples from
contaminated areas at NTS. The mean plant/soil ratios for different groups of paired
samples ranged from 0.004 to 0.44 and were inversely related to the mean soil
contamination of each group, which ranged from 5.9 x 10~^ to 0.12 luCi/g. The
weighted mean ratio for 506 paired samples was 0.096 ± 0.004. The plants in this study
were desert shrubs growing in areas that were contaminated with plutonium as a result of
nuclear safety tests conducted from 1953 to 1964. Ratios obtained by growing plants in
these same soils were on the order of 10""^ to 10""*, which indicates that no more than
1% of the plutonium in plant samples from contaminated areas at NTS is likely to be due
to root uptake.

Environmental monitoring data from Hanford (Bramson and Corley, 1973; Nees and
Corley, 1974; Fix, 1975) indicate ratios in the range of 0.05 to 1.0 for soil concentrations
in the range of 2 x 10~^ to 4 x 10~^ fJ^Ci/g. Similar data from Savannah River
(McLendon et al., 1976) indicate ratios from 0.009 to 0.97 for soil concentrations in the
range of 1.3 X 10"^ to 1.6 x 10"^ juCi "^Pu/g and 2x lO"''^ to 4.6 x 10"^ [JiCi
^^*Pu/g. The higlier plant/soil ratios are usually found at lower soil concentrations.

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 475

In general, plant/soil ratios for ^"'^Pu, which are based on plant and soil samples
collected under field conditions, range from 1 to 10~^, whereas ratios based on
laboratory studies, which preclude external contamination, range from 10"^ to 10~^.
Considering the situation at NTS, we believe it is reasonable to assume that
approximately 99% of the plutonium associated with the vegetation compartment is due
to external contamination and that no more than 1% is due to root uptake.

Data on the plant/soil ratio for other transuranium elements are very limited.
Romney et al. (1975) measured ^'*' Am concentrations of the vegetation from the areas
at and near NTS. Grouped according to species and location, the mean plutonium/
americium ratio in vegetation ranged from 2.0 to 28.3, with typical values being about
10. Similar groups of soil (Gilbert et al., 1975) ranged from 5.3 to 26, with typical values
also being about 10. These analyses indicate that the long-term plant/soil ratio for
americium is not significantly different from that for plutonium.* The data on the
short-term plant/soil ratio indicate significant differences that may be related to the
solubility of the element. Price (1973b) measured the uptake of ^^^Np. ^^^Pu, ^"^^ Am,
and ^'*'*Cm by tumbleweed and cheatgrass from various solutions applied to the soil. The
americium uptake was about 2 to 30 times as great as that of plutonium, curium uptake
was about 2 to 40 times as great as that of plutonium, and neptunium uptake was about
100 to 1000 times as great as that of plutonium. Bennett (1976) summarized much of the
short-term data and concluded that americium and curium uptakes were about 10 to 30
times as great as that of plutonium.

Variation of Plant I Soil Ratio. Data presented by Romney et al. (1975) also demonstrate
that the mean concentrations of plutonium in soils and plants decrease with increasing
distance from ground zero locations, whereas the vegetation/soil ratios within sampling
strata show a tendency to increase. Tamura (1976) provides a graph of soil activity vs.
distance from ground zero and fits the data to a power curve of the form y = ax"^, where
y is soil activity and x is distance from ground zero. This is an interesting notion to
pursue because we expect both soil concentration and particle size to decrease with
increasing distance from ground zero. If foliar retention is greater for small particles, we
would expect vegetation/soil ratios to increase as particle size and soil contamination
decrease with increasing distance from ground zero. We have no comparable curve for
vegetation, but we assume it would be of the same form. On the basis of this assumption,
the vegetation/soil ratio could be expressed as a function of distance from ground zero as
follows:

yv = avx ^^

(10)

*Gilbert et al. (1975, p. 407), using Double Tracks data, show that the average vegetation/soil
ratio tor americium is about 1 .5 times the average vegetation/soil ratio for plutonium; but, considering
the range of ratios contributmg to the averages, we are not persuaded that 1 .5 is significantly different
from 1.0. At other sites the ratios were not different.

476 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

where yy and C^ are plutonium concentrations (pCi/g) of vegetation and soil, respectively,
X is distance from ground zero, and a and b are constants derived from least-squares
analysis. If bj > by, the vegetation/soil ratio will increase with increasing distance from
ground zero and decreasing soil concentration.

To obtain an expression for plutonium concentration in vegetation as a function of
soil concentration, R. 0. Gilbert (personal communication) performed a regression
analysis to fit the available data (636 paired samples of vegetation and soil) to the
following equation:

in(yv) = ln(a)+bln(Cs)

V= A + bS

where V= In (yy)

S = ln(Cs)

A = In (a)

^ ^ rSV^^(l/n)(SV)^'.
ISS^ -(l/n)(SS)^J

A = V - bS

V and S = the mean of V and S, respectively.

[This method of calculating b is indicated because measurements of both yy and Cs are
subject to error. See Ricker (1973) for a discussion of this method of calculation.]
The results of tliis analysis are summarized as follows:

Standard

Parameter

deviation

V= -3.4961

2.3096

S= -0.9322

3.0382

b = 0.7602

0.0221

A= -2.7871

0.0458

r = 0.8084 (correlation coefficient)

n = 636

With antilogs, the mean vegetation/soil ratio is

yy(nCi/g)

Crv

Cs(nCi/g)

0.0303
0.3937

0.0770

which is somewhat lower than the value (0.096) obtained from the grouped data
presented by Romney et al. (1975).

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 471

The equation for vegetation concentration (nanocuries per gram) as a function of soil
concentration (nanocuries per gram) is

yv = 0.0620 C,«-'^ (11)

The corresponding equation tor the vegetation/soil ratio is

Crv =^= 0.0620 rr°-''' (12)

Thus, for areas in which soil concentrations are 10, 1.0, 0.1, and 0.01 nCi/g, the predicted
vegetation/soil ratios would be 0.036, 0.062, 0.107, and 0.187, respectively.

Equations 1 1 and 12 demonstrate the dependence of vegetation concentration on soil
concentration and the fact that the vegetation/soil ratio tends to increase as soil
concentration decreases. The use of either equation for predictive purposes may be
limited by the extreme range of soil and vegetation values and/or by various site specific
factors that are not considered in the regression analysis. It miglit be better to apply the
regression analysis to sampling strata means as shown in Table 3.

Except for strata 3 and 4 in Area 13, the measured and predicted values given in
Table 3 seem to agree quite well, but the equations for Area 13 and GMX-5 (footnote to
Table 3) predict higher values (especially at liigher soil concentrations) than would be
obtained from Eqs. 1 1 and 12, which are based on samples from all study areas.

Discussion. Equations 11 and 12 shed some Ught on how vegetation/soil ratios may be
expected to vary with respect to soil concentration, but they do not explain why. To
approach this and related questions, we refer back to Eqs. 7 and 8, the proposed
differential equations tor the external and internal components of plant contamination.
For the time being at least, we can dismiss Eq. 8 from further consideration because the
greenhouse studies have shown that root uptake cannot account for more than a small
fraction of the vegetation/soil ratios observed at NTS.

Equation 7, for external contamination, has the following solution for a constant air
concentration (Ca): .

yve = ^ h'^' (1 -exp HXw + h + XA)t] } (13)

Aw + Au + Aa

where the parameters are defined following Eq. 7, and

kav = VdFv (14)

Ca = L,Cs (15)

where Vj is the deposition velocity on soil (cm/day), Fy is a vegetation interception
factor (cm'^/g vegetation), and L^ is a mass-loading factor (g soil/cm^ air).

If we assume a steady state (large t) between vegetation and soil, the vegetation/soil
ratio can be expressed in the parameters of Eqs. 13, 14, and 15 as follows:

Yve VjFvLa

478 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Mean Concentrations of ^ ^ ^ '^ ^ ^ Pu in Soil and

Vegetation and Vegetation/Soil Ratios by Strata in Area

GMX-5 and Area 13

2 3 9,240pu „Q/g

Vegetatioi

Soil

Vegetation (means)

i/soil raiio

Strata

(means)*

Measured*

Predictedf

Measured*

Predictedf

Area GMX-5

0.084

0.0083

0.0099

0.13

0.12

0.059

0.0092

0.0075

0.16

0.13

0.73

0.064

0.055

0.075

4.5

0.26

0.23

0.052

0.051

7.3

0.31

0.34

0.050

0.046

Area 1 3

0.036

0.0052

0.0067

0.15

0.19

0.1

0.013

0.017

0.14

0.17

0.4

0.17

0.058

0.44

0.14

1.1

0.077

0.14

0.069

0.13

2.4

0.28

0.10

0.12

14.0

1.2

1.36

0.078

0.10

*Based on data from Romney et al. (1975).
tyv = 0.13 q-''' for Area GMX-5. Yv = 0.07 C°-*
equations are based on the strata means in columns 2 and 3

for Area 13. Both

where Xv = Xw + Xu + Xa is the effective decay rate coefficient for plutonium-bearing soil
particles externally deposited on vegetation.

As noted in the section on air, the deposition velocity is a function of particle size
(Eq. 4). In the soils of Area 13 (Tamura, 1976), most of the plutonium is associated with
coarse silt (20 to 53 /im), and the estimated deposition velocity (Vj) for particles of
50-jum diameter could be as higli as 20 cm/sec, or 1.73 x 10^ cm/day.

The plant interception factor (Fy) determined by Miller and Lee (1966) for freshly
deposited volcanic dust (50 to lOOjum) was 47.4 cm^/g for dry exposure conditions.

For predictive purposes Anspaugli (1974) has suggested that a mass-loading factor
(La) of 100 iJg/m^ (10"'° g/cm^) be used. This is the amount of dust we would expect
to find in the GMX area (Shinn and Anspaugli, 1975) when the wind velocity averages
about 1 .4 m/sec (3 mph).

Substituting these values in Eq. 16, assuming a vegetation/soil ratio of 0.1, and solving
for Xv indicates an effective half-life of about 8.5 days. So

yve_(1.73 X 10^ cm/day) (47.4 cmVg)(10''" g/cm^)

In (2)/8.5

0.10

This exercise proves nothing. It merely demonstrates that Eq. 16 might explain the
higli vegetation/soil ratios observed at NTS. Shinn and Anspaugh (1975) have
demonstrated that mass loading (L^) increases with wind velocity. The effective half-life
may decrease with wind velocity. Sehmel (1975) has shown that deposition velocity (V^)
decreases as the particle size decreases for dp > 1 lum. If small particles are more readily

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 479

retained by vegetation than larger particles, the plant interception factor and the effective
half-life may increase as the particle size decreases. Variations of the factors of Eq. 16 as
functions of particle size and wind velocity may account for the range of vegetation/soil
ratios implied by Eq. 11 , but there are still too many unknowns to develop a descriptive,
dynamic model for the vegetation compartment.

For predictive purposes we shall assume that the soil and vegetation compartments
are in steady state and that the mean vegetation/soil ratio is 0.1. This ratio is somewhat
conservative since it tends to overestimate the plutonium content of vegetation in areas
where soil concentrations are greater than 100 pCi/g.

The steady-state assumption is justified to some extent by the fact that movement of
contaminated particles from soil to foliage and back to soil is a more or less continuous
process. Since the turnover time is apparently short (between 1 and 2 days), a steady
state should be established quickly and characterized by a constant vegetation/soil
concentration ratio. The choice of 0.1 is in the range of statistical mean (0.096 ± 0.0004)
obtained from actual soil and vegetation samples.

Cattle

Transport Pathways. For present purposes the only herbivore assumed to contribute to
man's plutonium intake is the cow. Both dairy cattle and beef cattle are considered. The
principal plutonium inputs to these herbivores (Fig. 1) are by inhalation and by ingestion
of contaminated soil and vegetation. Figure 2 illustrates the assumed pathways of
plutonium transport to man via beef cattle and dairy cattle and provides estimates of
some of the parameters required to estimate potential concentrations of plutonium in the
muscle, liver, and milk of beef and dairy cattle maintained in a contaminated area at NTS.

Formulation of Cow Model. A general equation for the concentration of plutonium in
the muscle, liver, or milk compartment of Fig. 2 can be derived as follows:

f = ;^ (rb fbi - Ai y.) (17)

where i = muscle, liver, and milk

yj = concentration of plutonium in compartment i at time t (pCi/g)
mj = weight of compartment i (g)
r^ = rate at which plutonium enters blood (pCi/day)
fbi = fraction (dimensionless) transferred from blood to compartment i (Fig. 2: 0.07,

0.12, and 0.007)
Xj = effective elimination rate coefficient (day^' ) for plutonium in compartment i
(based on effective half-lives, T, in Fig. 2)

Estimated values for some of the parameters of Eqs. 17 and 18 are given, as indicated
above, in Fig. 2. The transfer fractions from the gastrointestinal tract to blood and from
blood to muscle, liver, and milk, the weight of muscle and liver, and the effective half-life
in milk are based on experimental results reported by Stanley, Bretthauer, and Sutton
(1974). The other values were assumed (Martin and Bloom, 1976) for purposes of
estimation. Equation 18 ignores the retention of plutonium in the lungs of cattle and

480 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

INGESTION
RATE

INHALATION
RATE

GASTROINTESTINAL

TRACT

r < 24 hr

LUNGS

10-5

3 X

L_[

BLOOD
r < 24 hr

0.07

0.007

0.12^

MUSCLE

m = 125 kg
T = 2,000 days

LIVER

m = 3.95 kg
T = 30,000 days

MILK

m = 25 kg/day
T = 16 hr

MA^'

■"

Fig. 2 Pathways of plutonium transport to man via beef cattle and dairy cattle.

assumes that all inhaled plutonium is transferred to the gastrointestinal tract or the blood
within 24 hr.

If we assume that the parameter values given in Fig. 2 are reasonably close to the true
values, the only parameters remaining to be determined are the plutonium ingestion and
inhalation rates and the rate at which plutonium enters the blood (rb).

Plutonium Ingestion Rates. Kleiber (1961) shows that the basal metabolic rate of
mammals (heat production by a fasting animal) is proportional to the three-fourths power
of body weight and that the feeding capacity (maximum energy intake) of domestic
animals, such as the cow, is about five times the food intake required for basal
metabolism. Data given in The Merck Veterinary Manual (Siegmund, 1967) for the
digestible energy (DE) requirements for maintenance of mature cows are based on

DE= 163.5 W°''3

(19)

where DE is the digestible energy required for maintenance (kcal/day) and W is the body
weight (kg).

The additional DE requirement for milk production in the range from 20 to 35 kg
milk/day at 5% butterfat is 1850 kcal (DE)/kg milk. The additional DE requirement for
growth ranges from about 8600 to about 19,800 kcal/kg gained, depending also on body
weight. According to McKell (1975), the average digestibility of desert vegetation is about
36% compared with 52% for good alfalfa hay and up to 80% for some concentrates
(Siegmund, 1967). The average energy content of most plant materials is about 4.5 kcal/g
(dry weight) (Golley, 1961), and the digestible energy content of desert vegetation is

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 481

about 1.6 kcal/g (dry weiglit). On the basis of these considerations, the vegetation
ingestion rate for a cow grazing desert vegetation can be estimated as follows:

^ 163.5 W"-^^ kcal/day .

'"^ 0.36 X 4.5 kcal/g ^^

where V,n^ is the vegetation ingestion rate (g/day) for maintenance of a mature cow
grazing desert vegetation. For cows that are gaining weight, producing milk, or pregnant,
the energy requirement, and thus the vegetation ingestion rate, would be higlier than
estimated by Eq. 20.

Grazing cattle also ingest soil. In earlier papers (Martin, Bloom, and Yorde, 1974;
Martin and Bloom, 1976) we assumed that the soil ingestion rate might be as high as 2000
g/day. Data recently reported by Smith (1977) indicate that this value is probably too
high. The amounts of sediment (soil) recovered from the reticulum and rumen of three
cows that had been grazing in Area 13 before sacrifice were 8.5, 57.3, and 278 g,
respectively. As Smith points out, "These data suggest that the total amount of soil
ingested is much less than 2 kg per day, and that a reasonable estimate would be between
0.25 and 0.5 kg."

Smith, Barth, and Patzer ( 1976) estimate that a 409-kg cow that grazed for 177 days
in the inner compound of Area 13 ingested a total of 100 /jCi of ^^^'^"^^Pu, or 0.565
pCi/day. This estimate was based on plutonium concentrations in the rumen contents of
fistulated steers allowed to graze in the same compound. Gilbert, Eberhardt, and Smith
(1976) made an independent estimate of 0.620 /iCi/day based on plutonium concentra-
tions in Eurotia lanata {347o of the cow's diet) and Atriplex canescens {(A% of the cow's
diet), as reported for Area 13 by Romney et al. (1975). The average wet weight of
vegetation ingested by the fistulated steers was 30 kg/day, and the average dry/wet ratio
was about 0.2 (R. 0. Gilbert, personal communication). In other words, the cow's
vegetation ingestion rate was estimated to be about 6 kg/day. On the basis of Eq. 20, a
409-kg cow would have to ingest about 8 kg/day to meet its energy requirements for
maintenance. Neither Smith, Barth, and Patzer (1976) nor Gilbert, Eberhardt. and Smith
(1976) include soil ingestion in their estimates of the plutonium ingestion rate for the
Area 13 cow.

On the basis of the methods outlined above, we would estimate this cow's plutonium
ingestion rate as follows:

Iv = 8139 g vegetation/day x 0.1 x 5.5 x lO""* /jCi/g soil
= 0.448 juCi/day

Is = 250 g soil/day x 5.5 x 1 0""^ A/Ci/g soil
= 0.138juCi/day

Iv + Is = 0.585 AJCi/day

where ly and 1^ are the plutonium ingestion rates through vegetation and soil,
respectively. In this calculation, 8139 g/day is the vegetation ingestion rate, which is
based on Eq. 20 for a 409-kg cow: 0.1 is the assumed average vegetation/soil ratio;
5.5 X 1 0 "* A^Ci/g is the average soil concentration o^ the Area 13 inner compound
(Gilbert, Eberhardt, and Smith, 1976); and 250 g/day is the assumed soil ingestion rate.

482 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

The resulting estimate of tiie cow's total plutonium ingestion rate, 0.585 /jCi/day, is
comparable to the independent, site-specific estimates of Smith, Barth, and Patzer (1976)
(0.565 AtCi/day) and of Gilbert, Eberhardt, and Smith (1976) (0.620 AtCi/day).
Considering the probable variability of the measurements and parameters involved, the
agreement of results is remarkably good.

The principal differences between our method and the method of Gilbert et al. are:
(1) our estimate of the vegetation ingestion rate, which is based on energy requirements (8
kg/day), is higher than their estimate, which is based on the rumen contents of a
tlstulated steer (6 kg/day); (2) our estimate of the average concentration of plutonium in
vegetation is lower than theirs, 55 pCi/g vs. 103 pCi/g; and (3) we included an input for
soil ingestion.

If we assume that the digestibility of Area 13 vegetation is 49% instead of 36%, a
ration of 6 kg/day would be adequate to meet the cow's maintenance energy
requirements. The weighted mean vegetation/soil ratio for Area 13 is about 0.15, or 50%
higlier than average. If we use this ratio instead of 0.1 and 6 kg/day instead of 8 kg/day,
Iv = 0.495 juCi/day and Is = 0.565 - 0.495 = 0.07 jL/Ci/day, or 127 g soil/day.

Inhalation Rate. Standard Man's respiration rate is 20 m^/day (International Commis-
sion on Radiological Protection, 1959), and his digestible energy requirement for
maintenance metabolism (no weight gain or loss) is 2600 kcal/day (National Research
Council, 1968). We assume that for man and cattle respiration rates are proportional to
digestible energy maintenance requirements. Tlie DE requirement (Eq. 19) for a 409-kg
cow is 13,185 kcal/day. The cow's estimated respiration rate is therefore
13,185x20/2600=101 m^/day. The cow's plutonium inhalation rate (I^) can be
estimated by

20 DE
2600

Ia=^Z7^XLaXCs (21)

where la = plutonium inhalation rate (pCi/day)

DE = digestible energy required (Eq. 19) for maintenance metabolism
(kcal/day)
Respiration rate = 20 m^/day divided by 2600 kcal/day = 7.69 X 10"^ m^/kcal for the
Standard Man
La = mass4oading factor (lOOjLig soil/m^ air) as recommended by Anspaugh

(1974)
Cs = average plutonium concentration (pCi/g) in the soil of the area grazed
by the cow
For the 409-kg cow of the Area 13 inner enclosure, where C^ = 550 pCi/g, la = 5.56
pCi/day.

Ingestion vs. Inhalation. It is obvious from the preceding discussion that ly > la, but the
accumulation of plutonium in organs or tissues other than the gastrointestinal tract or
lungs and the excretion of plutonium in milk depends (Fig. 2 and Eq. 17) on the rate (r^,)
at which plutonium reaches the blood. The rate at which ingested plutonium reaches the
blood is simply the plutonium ingestion rate (ly) multiplied by the fraction transferred
from the gastrointestinal tract to the blood. For the Area 13 cow discussed above,

= 565,000 pCi/day x (3 x 10~^)
= 16.95 pCi/ day

MODEL FOR ESTIMATING Pit TRANSPORT AND DOSE 483

where r^v is the rate at which ingested plutonium reaches the blood (pCi/day) and fgb is
the fraction transferred from the gastrointestinal tract to blood. Since this cow was left in
the inner compound of Area 13 for 177 days before sacrifice (Smith, 1977), the total
^ ^^Pu expected to have entered the blood via ingestion would be 3000 pCi.

The transter of inhaled plutonium to blood is considerably more complicated than
indicated by Fig. 2. Assuming resuspended particles to have an activity median
aerodynamic diameter (AMAD) of 0.5 jum and applying the Task Group on Lung
Dynamics model (Fig. 4 and Table 1 1) to the cow, we obtain the following expression for
the rate at which inhaled plutonium would be expected to enter the blood:

rba =0.0021 la + 0.0833 Xiyi +X2y2 +3 X 10"^ yj (22)

dy,

dy2

- = 0.18I, -Xiyi (23)

dya
dt

0.225 Xiyi -X2y2 (24)

0.75 lio.2079 +-y^U 0.667 Xiyi (25)

where r^g = rate (pCi/day) at which inhaled plutonium enters the blood

0.0021 = fraction of inhaled plutonium transferred directly to blood from

the upper respiratory tract
0.0833 = fraction of plutonium deposited in the lungs and then transferred
to blood
yi = amount (pCi) present in the lung at time t
Xi = In (2)/ 500 days = the lung clearance rate

y2 = amount (pCi) present in the lymph at time t
X2 - In (2)/ 1000 days = the lymph clearance rate

ya = amount (pCi) present in the gastrointestinal tract at time t
0.18 = fraction deposited in the lungs and cleared with a 500-day
half-life
0.225 = fraction transferred from lung to lymph and then cleared to
blood with a 1000-day half-hfe
0.75 (day ) = average residence time of plutonium in the gastrointestinal tract
0.2079 = fraction transferred directly from the upper respiratory tract to
the gastrointestinal tract
0.12 = fraction cleared from lungs to the gastrointestinal tract with a
1-day half-hfe
Xo =ln(2)/l day
0.667 = fraction cleared from lungs to the gastrointestinal tract with a
500-day half-life

Note that Eq. 22 consists of four terms. These terms represent transfers of inhaled
plutonium to blood from the upper respiratory tract, the lungs, the pulmonary lymph,
and the gastrointestinal tract, respectively. Equations 23, 24, and 25 are the differential

484 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

equations for the long-term components of the lungs and lymph nodes and for the
gastrointestinal tract, respectively. The solution to Eq. 22 is

rba(t) = Ia (0.0021 +0.015(1 - e^^> ') + 5.61 X 10"^ (e-^^* -e"^-*)

+ 3 X 10"' [0.381 +0.12(1 -e"-^'^)]} (26)

Integration of Eq. 26 and evaluation of the integral from t = 0 to t = 1 77 days give the
amount reaching the blood in 1 77 days as 0.676 la . Applying this result to the Area 1 3 cow,
the total ^^^Pu expected to have entered the blood via inhalation would be 3.76 pCi.
Compared with 3000 pCi for ingestion (see above), this is a negligible quantity.

On the basis of this comparison, we shall assume that plutonium concentrations in
cow milk, muscle, and liver are due to ingestion only, or, conversely, that the
contribution from inhalation is negligible. Even after 10 yr of exposure in Area 13, the
contribution from inhalation would be no more than 0.56% of the total.

Comparison of Model Predictions and Field Data. Plutonium concentrations in lungs,
liver, and muscle are given in Table 4 for four cows included in the Area 13 grazing study.

TABLE 4 Concentrations of ^ ^ ^ '^ "* ^ Pu in Tissues of
Cattle Grazed in Area 13 (NTS)

2 3 9 ,2

^<'Pu,pCi/kg

Outer compound*

Average

Inner compoundf

Tissue

1 4

2 M

Lungs

Liver

Muscle

74.5 51.4
14.5 15.8
0.05 0.195

18.2
10.9
lost

48.0
13.7
0.12

NRt NR
38.9 6.13 kg
0.17 189kg

*Data extracted from Smith, Barth, and Patzer (1976).
t Data extracted from Smith (1977).
:j:Not reported.

Cow 2 (Smith, Barth, and Patzer, 1976) was placed in the inner compound for 177 days.
Cows 1, 4, and 6 grazed the outer compound for 433 days (Smith, 1977). Cow 1 weighed
252 kg, and cow 4 weighed 300 kg; the weiglit of cow 6 is not known. We shall assume
that the average weight of the cows from the outer compound was 275 kg. Tissue weights
for these cattle were not reported; but we estimate for a 275 -kg cow that the average lung
weight is about 2.1 kg and the average liver weiglit is 4.8 kg, based on a study made by
Smith and Baldwin (1974). At 45% of body weight (Smith, 1977), the muscle weight for a
275-kg cow would be about 125 kg.

Lungs. On the basis of these and other considerations, the concentration of ^^^Pu in
the lungs of the cows from the outer compound of Area 13 can be estimated as follows:

P _(76mVday)(10~^ g/m^)(215 pCi/g)(Q.18) 1 -e''^^^^
^""^ 2AYg X

= 55 pCi/kg (vs. 48 pCi/kg, Table 4)

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 485

where 76 m^/day = estimated respiration rate for a 275-kg cow (Eq. 21)
10~^ g(soil)/m^(air) = estimated mass-loading factor (Eq. 21 )

215 pCi/g = average concentration of "^^^Pu in the soil of Area 13*

0.18 = assumed fraction of inhaled plutonium deposited in the lungs and
cleared with a 500-day half-life
2.1 kg = estimated weight of the lungs
X = In (2)/ 5 00 days

Liver and muscle. The plutonium ingestion rate for a 275-kg cow can be estimated as
follows:

Iv = 6158 g vegetation/ dayt X 0.1 X 70 pCi/g soil + 250 g soU/day x 70 pCi/g soil
= 60,606 pCi/day

The concentration in liver for cows in the inner compound can now be estimated as
follows:

^ _ (60,606 g/ day )(3 x 10'^)(0.12) 1 -e~^"^
^'•^^■- 4Jki X

= 19.6 pCi/kg (vs. 13.7 pCi/kg, Table 4)

where 3 X 10"^ = fraction transferred from the gastrointestinal tract to blood (Fig. 2)
0.12 = fraction transferred from the blood to the liver (Fig. 2)
4.8 kg = estimated weight of the liver
X = ln(2)/30,000days(Fig.2)

Similar calculations were made for the other cases given in Table 4. The observed
values from Table 4 and the estimated values [pCi/kg (wet weight)] are compared below;

pCi/kg (wet weight)

Outer compound

Inner compound

Observed

Estimated

Observed

Estimated

Lungs

Liver

Muscle

48.0

13.7
0.12

55.0
19.6

0.4

NR
38.9

0.17

61.0
60.7
1.12

These comparisons suggest that the model for beef cattle may be somewhat conservative,
but the order-of-magnitude agreement between observed and estimated values appears to
be good, better than might be expected, as a matter of fact. However, partial data for

*Tlie average soil concentration m the outer compound is 70 pCi/g, but some of the resuspended
material in the air of the outer compound is assumed to come from the soil of the inner compound,
t Based on i-.q. 20 for a 275-kg cow.

486 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

four cows, two areas, and two grazing times are hardly an adequate basis for model
validation. The best we can conclude from these comparisons is that they provide no basis
for rejecting the model. The discrepancy between the experimental values and values
predicted by the model is less than an order of magnitude, and we have little reason to
expect better than an order of magnitude accuracy.

The milk cow. The model milk cow is assumed to weigh 650 kg and to produce milk
at a rate of 25 kg/day. Such an animal would require a digestible energy intake of 64,750
kcal/day, i.e., 18,500 kcal/day for maintenance plus 25 kg/day x 1850 kcal/kg for milk
production (Siegmund, 1967). To meet this high energy requirement and, at the same
time, to provide a conservatively higli estimate of plutonium transport to man via milk,
we shall assume that the model milk cow consumes 10 kg of desert vegetation per day
and 15 kg of alfalfa hay grown in the same contaminated area per day. The remainder of
the diet consists of commercial concentrates containing no plutonium. For the model
milk cow, the plutonium ingestion rate is estimated as follows:

Iv = Cs (250 g soil/day + 0.1 x 10,000 g vegetation/day + 0.017 x 15,000 g alfal fa/ day)
= 1505Cs(pCi/day)

where C^ is the soil concentration (pCi/g) and 0.017 is the alfalfa/soil ratio, which is
assumed to be one-sixth the desert vegetation/soil ratio due to plowing and mixing of the
soil to a depth of 30 cm. The equation for estimating the concentration in milk is

(1505X3 X lCr^)(0.007)
"^m ilk

Cmilk - 25 X. - ^ ^^^^

= 1.37 X 10-^Cs(pCi/kg)
where Xmilk = In (2)/0.75 (Fig. 2).

Man

In the preceding discussion we have considered the dynamics of the plutonium transport
system (Fig. I) and have attempted to establish mathematical relationsliips between
compartments. Our present knowledge of the food-chain kinetics of plutonium in
contaminated areas at NTS is not adequate for modeling the dynamic aspects of all parts
of the transport system. To simplify estimation of the plutonium inhalation and ingestion
rates for herbivores (cattle), we assumed a steady-state system and constant intake rates.
We now apply the same simplifying assumptions to estimate potential plutonium
inhalation and ingestion rates for the hypothetical Standard Man.

Inhalation Rate. The plutonium inhalation rate (Am ) is defined as the product of the
respiration rate (Bm) and the concentration of plutonium in air. The concentration of
plutonium in air is, of course, quite variable, but, since it is due to resuspension of
contaminated soil, it can be related to the average concentration in surface soil (Cg).

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 487

For predictive purposes, Anspaugh (1974) has suggested the use of a mean
mass-loading factor of 100 jug soil/m^ air. We combine this factor with the further
assumption that the specific activity of plutonium in resuspended materials is the same as
that in the associated soil and estimate Am as follows:

Am =BmLaCs= 0.002 Cs (28)

where Am = plutonium inlialation rate for man (pCi/day)
Bm = respiration rate (20 m^/day)
La = mass-loading factor (100 ;Ug/m^)

Cs = average concentration of plutonium (pCi/g) in the soil of the contaminated
area

The observed mass-loading factor during cascade impactor runs at NTS was 70 Mg/m^
(Anspaugh, 1974), and the specific activity of particles recovered from the impactors was
about one-third as high as that of surface-soil samples from the same locations (Phelps
and Anspaugh, 1974). Compared with these observations, the estimate of Am provided
by Eq. 28 may be conservatively high by a factor of about 4 under average conditions.
High winds or mechanical disturbances, such as vehicular traffic, plowing, etc., could
cause the mass-loading factor to increase temporarily to very higli levels. However, a
comparison of observed and predicted air concentrations based on La = 100 jug/m^
showed very good agreement (Anspaugh et al., 1975).

Ingestion Rate. The plutonium ingestion rate is defined as the sum of products of the
rates at which different kinds of contaminated materials are ingested and the
concentration of plutonium in each kind of material. The formula used for estimating a
probable ingestion rate for use in this study was

n=6

Hm = Cs L liDi (29)

i=l

where Hm is the plutonium ingestion rate for man (pCi/day), Ij is the ingestion rate for
substance i (g/day), and D, is the discrimination ratio (dimensionless) for substance i.

The kinds of materials considered, their assumed ingestion rates (Ij), and associated
discrimination factors (Di) are Usted, together with their products and sum, in Table 5.
The methods, experimental data, and assumptions used to estimate the discrimination
factors (Dj) are explained in the following text.

Soil. The assumption that the Standard Man of the model accidentally ingests soil at
an average rate of 0.01 g/day is purely speculative but not unreasonable considering the
amount of dust that can be raised in desert environments by activities that disturb the soil
surface.

Vegetation. To estimate the plutonium concentration in native vegetation, we
assume an average vegetation/soil ratio of 0.1 . As explained earlier, this ratio should tend
to overestimate the concentration of plutonium growing in areas of relatively high soil
concentration at NTS. To distinguish between native vegetation and cultivated plants

488 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

(alfalfa hay), we assumed a sixfold dilution of soil concentration due to plowing to a
depth of 30 cm; i.e., the plant/soil ratio for cultivated plants is 0.017 instead of 0.1. In
preparing Table 1 , we assumed that 90% of the external contamination of "leafy
vegetables" and that 99% of the contamination associated with "other food plants"
would be removed by washing, peeling, etc., during preparation for consumption. In spite

TABLE 5 Estimation of Standard Man's Plutonium
Ingestion Rate

Substance suggested

I.*

Dj*

liDi

Percent

Soil

o.oit

1.0

1.0 X 10'"

5.151

Leafy vegetables

81t

1.7x 10"^

1.4 X 10"'

70.935

Other food plants

2221:

1.7 X 10-'

3.8 X 10"=

19.441

Beef muscle

273$

9.4 X 10"*

2.6 X 10"'

1.322

Beef liver

13§

4.7 X 10"'

6.1 X 10"'

3.148

Cow milk

436$

1.4x 10-^

6.1 X 10""

0.003

SliDi

= 1.9 X 10"'

100.000

*See Eq. 29 and explanation in text.

fAssumed accidental ingestion rate.

$From U. S. Department of Agriculture (1973).

§From Organization for Economic Cooperation and Development (1970).

of this assumed reduction, leafy vegetables and other food plants account for 90% of
Standard Man's estimated plutonium ingestion rate (Table 5).

Muscle, liver, and milk. The model beef cow weighs about 275 kg. Its plutonium
ingestion rate (ly), owing to ingestion of 6.2 kg native vegetation/day and 0.25 kg
soil/day, is about 870 Cs (pCi/day). Given this ingestion rate and the parameters noted in
Fig. 2, the discrimination ratios for muscle and liver (Table 5) were estimated as follows:

Cmuscle_(870)(3x 1 0" ' )(0.07) / 1 - exp {-730 [In (2)/2000] }

(\ -exp {-730 [In (2)/2000] }\
\ ln(2)/2000 /

Cs 125,000 \ ln(2)/2000

= 9.4 X 10"^

Ciiver_(870)(3 X 10-')(

Cs 4800

= 4.7 X 10"

0.12) /i - exp {-730 [In (2)/30,000| }\
\ ln(2)/30,000 /

In these examples t was set equal to 730 days (2 yr), and this is the assumed average age
of beef cattle at the time of slaughter.

The method of estimating the discrimination factor for milk was described earlier in
the description of the model milk cow (see Eq. 27).

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 489

Discussion. Models for calculating organ burdens and cumulative organ doses due to
ingestion and inhalation of ^^^Pu are discussed in considerable detail later in this chapter.
On the basis of the ICRP-II model for ingestion and the Task Group on Lung Dynamics
model for inhalation, radiation doses to the respiratory system would be entirely due to
inhalation, and doses to the gastrointestinal tract would be primarily due to ingestion.
Doses to organs receiving the radionuclide from blood (bone, liver, kidney, etc.) would be
due to both ingestion and inhalation. The relative importance of inhalation vs. ingestion
can be compared by comparing the two components of organ burden after a period of
chronic exposure. Such a comparison is provided in Table 6.

TABLE 6 Fractions of " ^ ^Pu in Bone, Liver,

or Kidney Due to Chronic Ingestion and Inhalation

for a Period of 50 yr*

Fraction due

Ingestion/inhalation

to ingestion

to inhalation

0.0005

0.9995

0.0053

0.9947

100

0.0506

0.9494

200

0.0964

0.9036

400

0.1758

0.8242

1000

0.3478

0.65 22

*Estimated burdens based on ICRP Publications 2 and
19 (International Commission on Radiological Protection,
1959; 1972).

On the basis of our estimates for the hypothetical Standard Man at NTS
[inhalation = 0.002 C^ (pCi/day) and ingestion = 0.2 C^ (pCi/day)] , the ingestion/
inhalation ratio would' be 100, and ingestion would contribute about 5% of the 50-yr
bone burden. As indicated by Table 6, the relative importance of ingestion vs. inhalation
increases as the ingestion/inhalation ratio increases. Any factor tending to increase the
transfer from the gastrointestinal tract to blood would have the same effect as an increase
in the ingestion/inhalation ratio. A factor tending to decrease the inhalation rate would
also increase the ingestion/inhalation ratio. The point of Table 6 is that, to have a
significant effect on internal organ burden, dose, or dose commitment, the ingestion rate
must exceed the inhalation rate by a factor of 100 or more.

Dose-Estimation Models

Plutonium reaches man by ingestion of contaminated food and water or by inhalation of
contaminated air. Part of this plutonium is distributed througliout the body where it may
remain for some time. While it remains within the body, organs that retain the plutonium
will receive a radiation dose that depends on the weight of the organ, the amount of
plutonium retained, and the time that the plutonium is retained. The several models we
have used to estimate plutonium distribution in man are discussed in this section.

490 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Dose and Dose Commitment

In all the dose-estimation models, the formula for estimating the radiation dose to a
critical organ of man, i.e., one of those organs which tends to receive the highest radiation
dose , is

dD E

^ = my ^30)

where t = time (days)

D = dose to the reference organ (rem)

E = 51.2159e, a dose-rate factor (g • rem idCi^^ day"^^ )

e = effective energy absorbed in the reference organ per disintegration of

radionuclide (MeV/dis)
y = plutonium burden in the organ (juCi)
m = either the mass of the organ if the organ is not part of the gastrointestinal tract

or twice the mass of the contents if the organ is part of the gastrointestinal

tract (g)

The values of the parameters in Eq. 30 for ^^^Pu and other transuranium elements are
given in Table 7. Most of these values were reported by the International Commission on
Radiological Protection (1959; 1964). The masses of deep lung and other portions of the
respiratory tract are the values used by Snyder (1967) and Kotrappa (1968; 1969). The
mass of thoracic lymph nodes was assumed to be the value (15 g) reported by Pochin
(1966). The mass of abdominal lymph nodes was assumed to be less than the mass of
thoracic lymph nodes and was arbitrarily set at 10 g. The dose accumulated in the organ
from the beginning of the exposure period (t = 0) to some later time (t = To) is given by

E r'D

D = — I ydt (31)

m Jo

If ingestion and inhalation of plutonium were halted at time Tq and the individual were
to live to some later time Tl, each organ would accumulate an additional dose from the
plutonium already within the body at time Tq. The dose commitment is the sum of the
dose accumulated to Tq plus the additional dose, or

Dc = D + Da (32)

where D^ is the additional dose (rem) and Dq is the dose commitment (rem).

ICRP Committee II Model

The report of the ICRP Committee II (International Commission on Radiological
Protection, 1959) contains a model and data that were used to estimate maximum
permissible concentrations (MPC's) of radionuclides in air and water. The model, as

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 49 i

TABLE 7 Parameters for Calculating Radiation Doses from
Transuranium Radionuclides

Effective energy,*

MeV/disintegration

Organ: GIT

Lung§

Bone

Liver

Kidney

Radionuclide

TAt

Mass,g: 150^

500

7,000

1,700

300

70,000

^^^Np

8x 10"

0.62

250

"'Np

2.33

0.14

0.16

0.63

0.16

0.15

0.22

238p^j

33000

0.55

284

239p^j

8.9 X 10*

0.5 2

266

2 4 0 p^,

2.4 X 10*

0.5 2

266

24 1 p^

4800

0.010

0.013

0.048

0.013

0.012

0.014

242p^

1.4 X 10»

0.49

253

243py

0.208

0.18

0.88

0.18

0.19

244py

2.8 X 10'°

1.14

292

'*' Am

1.7x 10=

0.56

283

242m^n^

5.6 X 10"

0.745

53.2

266

53.2

53.1

53.2

^^^Am

0.667

0.734

53.1

266

53.1

^^^Am

2.9 X 10*

0.79

54.2

273

54.2

^^^Am

0.0181

0.52

2.6

0.52

^'^Qm

162.5

0.61

315

^*^Cm

13000

0.61

299

^^^Cm

6700

0.59

299

'^'Cm

7.3 X 10*

0.55

277

^''Cm

2.4 X 10*

0.54

278

'^'Qm

3.3 X 10'°

0.54

278

^^«Cm

1.7 X 10'

11.5

453

2244

453

45 3

^*'Cm

0.044

0.31

5.2

^*«Bk

290

0.026

0.13

0.026

2 5 0g^

0.134

0.41

0.52

1.5

0.83

24 9Q-

1.7 X 10=

0.63

301

2 5 0cf

3700

0.61

311

25 1Cf

2.9 X 10=-

0.59

295

252cf

803

2.1

210

1100

210

253cf

18.0

0.78

343

2 5 4 pi-

56.0

120

3800

18900

3800

*Includes energy from daughter products with half-times less than 1 yr.

fr^ is radioactive half-hfe of radionuclide, days.

JGIT is gastrointestinal tract (principally, lower large intestine). Mass of contents of lower large
intestine is 150 g.

§ICRP-II used 1000 g for mass of lungs. All other models use 500 g. Effective energy to lymph
nodes and all portions of the respiratory tract are assumed equal to lung. Masses for these organs are:
nasopharyngeal region, 1.35 g; tracheobronchial region, 400 g; abdominal lymph nodes, 10 g.

applied to plutonium or other transuranic elements, is shown in Fig. 3. This model
distinguishes between insoluble and readily soluble compounds that are inhaled. In the
absence of other data, it is assumed that 25% of a soluble compound is exlialed, 50% is
swallowed and reaches the gastrointestinal tract (GIT) almost immediately, and 25% is
immediately transferred directly to the blood.

492 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

INHALATIOixi

EXHALED 25% -^

INGESTION

GASTROINTESTINAL
TRACT (18-hr TRANSIT)

LIVER
(30.000 days)

KIDNEY
(32,000 days)

BONE
(73,000 days)

12.5% (INSOLUBLE
COMPOUNDS)

12.5%

(INSOLUBLE

COMPOUNDS)

1 00%

FAST

(1-day
transit)

SLOW

(365
days)

DEEP LUNG

Fig. 3 International Commission on Radiological Protection Committee II model for
plutonium. Fraction from blood to total body is 100%. Biological half-time in total body
is 65 ,000 days.

None of the soluble compound resides in the lung for any significant length of time
and therefore does not contribute a radiation dose to the lung. For insoluble compounds,
in the absence of other data, it is assumed that 25% is exhaled, 50% is swallowed and
reaches the GIT immediately, and 25% is immediately transferred directly to tissues deep
in the lung. Of the amount transferred to the deep lung, half ( 12.5% of that inhaled) is
coughed up within 24 hr and is swallowed (reaches the GIT). The remainder (12.5% of
that inhaled) resides in the deep lung with a half-time of 365 days and is then transferred
to the blood. To avoid underestimating radiation doses, we use the insoluble parameters
for calculating the transfer to the lung and GIT but use the soluble parameters for the
transfer to all other organs.

On the basis of Fig. 3 and 'he above discussion, the equations for the ICRP II model
are

•"GIT - fuRG Am + fuRDLI A,n + H

111

{33}

•"B - *URB Am + fGlTB 'GIT

YGIT - 'GIT TgIT

(34)
(35)

dypLS
dt

- 'liRDLS Am - (Aa + Adls) VDLS

(36)

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 493

YDL = yDLS + fURDLF Am TdlF (37)

dyiiver

dVK
dt

dybone
dt

dyTB
dt

= fBL TB -(>^A + >^L)yiiver (38)

= fBK re - C^A + Xk) yK (39)

= fBBN TB -(^A + XB)ybone (40)

= ^BTB TB -(^A + XTB)yTB (41)

where rg, tgit - rates that plutonium reaches the blood and gastrointestinal tract (GtT),
respectively (juCi/day)
Am, Hm = plutonium inhalation and ingestion rates, respectively (juCi/day)

f = fraction of plutonium transferred from one location to another within
the body with the subscript notation as follows: URG is upper
respiratory tract (URT) to GIT; URB is URT to blood; GITB is GIT to
blood; URDLF and URDLS are URT to the fast and slow portions,
respectively, of the deep lung; BL, BK, BBN, and BTB are blood to
liver, kidney, bone, and total body, respectively
Tdlp,Tgit ~ transit times for the fast portion of the deep lung and the GIT (lower
large intestine), respectively
X = biological elimination rate constant with the subscript notation as
follows; DLS is deep lung (slow portion), TB is total body, L is liver, K
is kidney, and B is bone (day ~ ^ )
X = In (2)/r, where r is the biological half-time for the organ (days)
y = plutonium burden in the organ with the subscript notation either
self-evident or identical to that for X except DLF is the fast portion of
the deep lung and DL is the total deep lung (/iCi)

Values of the parameters in Eqs. 33 through 41 can be obtained from Fig. 3 for
plutonium. The values for other transuranic elements, if different from plutonium, are
given in Table 8. The radioactive half-times for each transuranic nuclide are given in
Table 7.

Task Group on Lung Dynamics Model

The ICRP Task Group on Lung Dynamics (Morrow et al., 1966) provided a more detailed
description of the inhalation pathway and arrived at the model indicated in Fig. 4. A
subsequent report [ICRP Publication 19 (International Commission on Radiological
Protection, 1972)] provided specific data for applying this model to plutonium and other
transuranic elements. The Task Group model treats the respiratory tract as a series of
compartments in which the amount initially deposited depends on the particle size and
the clearance rate (biological half-time) depends on the type of compound.

494 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 8 Parameter Values Applicable to the ICRP 11 Model
for Transuranic Elements Other than Plutonium*

Element
Neptunium Americium Curium Berkelium Californium

Biological Half -Time, days

Deep lung

(slow portion)

120

Liver

54,000

3.480

3,000

Kidney

64,000

27,000

24,000

Total body

39,000

20,000

24,000

Fraction from Blood to Orga

Liver

0.05

0.35

0.40

Kidney

0.03

0.02

Bone

0.45

0.25

0.30

120

65,000

120

65,000

0.80

*A11 other parameter values are identical to those of plutonium.

On the basis of Fig. 4, the equations for the compartments of the respiratory tract
and lymph nodes plus the transfer to the GIT and blood are

rciT - ^bYNPb "•■ ^dVTBd "^ VVPf + ^YPg + ^

(42)

fB - ^aYNPa + ^^cYTBe + ^eYPg + ^iYLM; + tjrCIT

dyNp,

—^ = faD3 Am - (Xa + Xa) YNP^

^ = fbD3A,^-(XA+Xb)YNPb

dyiBc

YNP - YNPa + YNPb

fc D4A„i -(Xa +Xc)ytBc

^^=fdD4An,-(XA + Xd)YTBd

YTBf,g = (XfYPj. + Xgypg) TjBf,^

YTB = YTBc ■'' YTBj + YTBf^g
dYP,

= feD5Am -(Xa + Xe)yp,

(43)

(44)
(45)
(46)
(47)

(48)

(49)
(50)
(51)

MODEL FOR ESTIMATING Pn TRANSPORT AND DOSE 495

Q
O
O

lU)

I
L

(a)

(c)

o . <>

NP3

faDsAm

NP.

ffaDsAm

(b)

NASOPHARYNGEAL REGION

D4"

fcD4Am

<ZL

TBd
^dDaAm,""

TRACHEOBRONCHIAL

REGION

(e)

3z:

feDsAr

^DsAm
(h)

3ZL

(f)

ffDgA^

3z:

(d)

(9)

fgDsAr

(i)

f ''^ PULMONARY REGfON

LMi [ LMf

LYMPH

o
<
cc

Fig. 4 Schematic diagram of the Task Group on Lung Dynamics model. [From
Houston, Strenge, and Watson (1975).]

^ = ffD5Ani-(XA + Xf)ypf

(52)

dyp

i = f„D5An,-(XA + X„)yp

dt ^

g/J'i'g

(53)

^ = fhD5Am-(XA + Xh)yPh

yp = ype + ypf + ypg + yph

dyLM-

— ^= fiXhypj^ - (Xa + Xi) yLMi

(54)
(55)
(56)

496 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

dyLMf _

dl ^

t'i) ^hYPh ~ ''^AYLMf

(57)

YLM - YLMi + YLMf

(58)

where r, X, y, Hm, f, and A^ are as defined in llie ICRP II model and the subscripts,
where different, refer to the compartments and the pathways listed in Fig. 4 and Table 9;

TBf,^

the residence time of material following pathways f and g in the

tracheobroncliial region (days); and D3, D4, and D5 are the fractions of inhaled material
that are deposited in the nasopharyngeal, tracheobronchial, and pulmonary regions,
respectively, of the respiratory tract (see Fig. 4 and Table 9).

Values of most of the parameters in the above equations are given in Table 9 (U. S.
Nuclear Regulatory Commission, 1976). TjBf is assumed to be 1 hr, or %4 day
(Snyder, 1967; Kotrappa, 1968; 1969), and f j ts identical to fciTB (0.003%) in the

TABLE 9 Task Group Lung Model Parameter Values

Fraction of Inhaled Particles Deposited in the Respiratory
System vs. Particle Diameter*

Fraction of inhaled quantity retained

Particle :

size

Nasopharyngeal

Tracheobronchial

Pulmonary

(AM AD),

region (D,)

region (D J

region (D^ )

0.05

0.001

0.08

0.59

0.1

0.008

0.08

0.50

0.3

0.063

0.08

0.36

0.5

0.13

0.08

0.31

1.0

0.29

0.08

0.23

2.0

0.50

0.08

0.17

5.0

0.77

0.08

O.Il

Clearance Parameter Valuesf

Translocation class

Days

Weeks

Years

Compartment

\'*

t'kt

t"kt

fkt

0.01

0.50

0.01

0.10

0.01

0.50

0.40

0.90

0.40

0.99

0.01

0.95

0.01

0.50

0.01

0.20

0.05

0.20

0.50

0.20

0.99

0.50

0.80

5 0

0.15

500

0.05

0.40

500

0.40

0.50

0.20

0.05

500

0.15

0.50

1.00

1000

0.90

*Estimaled trom data of Morrow et al. ( 1 966).

tAs amended by ICRP Publication 19 |lntcrnational Com-
mission on Radiolopical Protection (1972)].

^k is clearance pathway (see l"ig. 4); T^ is biological half-time
(days) for pathway k; f;^ is fraction cleared by pathway k.

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 497

ICRP II model. The equations for GIT, liver, kidney, bone, and total body are identical to
the ICRP II model and have identical parameter values except for the following values,
which apply to all transuranic elements: fsL = 0.45, feBN ~ 0.45, tl = 40 yr, and tb =
100 yr.

Stuart, Dionnc, and Bair (SDB) Model

Stuart, Dionne, and Bair (1968) developed models to describe the distribution and
retention of plutonium in the body following a single inhalation. These models were
based on the results of several studies with dogs, and these results were extrapolated to
where they miglit apply to man. Tlie short-term model is shown in Fig. 5, and the
long-term form is shown in Fig. 6. Stuart, Dionne, and Bair (1971) revised the long-term
model, and these revisions are incorporated in Fig. 6. Stuart et al. (1971) combined the
nasopharyngeal and tracheobronchial regions of the Task Group model into one
compartment but expanded the pulmonary region into two compartments, one with a
constant biological half-time of 3 yr and another with a variable half-time.

They also added compartments for abdominal lymph nodes and treated the transfers
from the pulmonary region to lymph nodes in a slightly different manner than the Task

70% OF TOTAL DEPOSITED

99.7%

NASOPHARYNX-
TRACHEOBRONCHIAL
(8 min)

0.3%

BLOOD
(8 hr)

1 1

30% OF TOTAL
DEPOSITED (INITI

STOMACH
(1-hr TRANSIT)

ALVEOLAR DEPOSITION

)

PULMONARY
LUNG

1 00%

SMALL INTESTINE
(4-hr TRANSIT)

1 00%

SKELETON -^

25%

UPPER LARGE

INTESTINE
(13-hr TRANSIT)

KIDNEYS

1 00%

-i_

URINE

LOWER LARGE

INTESTINE
(24-hr TRANSIT)

LIVER

71%

Fig. 5 Block diagram of the short-term form of the Stuart, Dionne, and Bair (SDB)
model.

498 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

85%

CONSTANT
(3 vr)

PULMONARY*
LUNG (100%)

15% VARIABLE
(50 days INITIALLY
DOUBLING EVERY
YEAR)

1 00%

SUMMATION
OF FECES

10% '

THORACIC
LYMPH NODES

COMPARTMENT

(INFINITE)

80%

70%

COMPARTMENT

(1 yr)

ABDOMINAL
LYMPH NODES

COMPARTMENT

(INFINITE)

70%

COMPARTMENT

(1 yr)

30%

10%

BLOOD

30%

1 00%

SKELETON
(100 yr)

25%

KIDNEYS
(40 yr)

99%

SUMMATION
OF URINE

90°

LIVER
(40 yr)

71%

*BASED ON 100% OF THE INITIAL ALVEOLAR DEPOSITION.

Fig. 6 Block diagram of the long-term form of the Stuart, Dionne, and Bair (SDB)
model.

Group model. In addition, they included feedback pathways for Hver to the
gastrointestinal tract (GIT) and for liver, kidneys, and skeleton to blood. Since they
applied this model to inhalation only, they did not consider the transfer from GIT to
blood since the fraction transferred is so small (less than 0.0 17f).

On the basis of Figs. 5 and 6, the equations for the compartments of the respiratory
tract and lymph nodes plus the transfer to GIT and blood are

^GIT = fuRG^NPTBYNPTB + ^LCG^LCYLC + ^LVYLV + fLG^LYliver (59)

fB - ^URB^NPTBYNPTB + fLCB-^LCYLC + ^TLB^TLYTL

+ ^ALB^ALYALNM + fLB^'^LYHver + ^KB^KYK + ^BYbone (60)

dYNPTB
dt

(0.7) (D3 + D4 + Ds ) Am - (Xa + Xnptb) YNPTB

(61)

dYLC

(0.3) (0.85) (D3 + D4 + Ds ) Am - (Xa + Xlc) Ylc

(62)

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 499

^^ = (0.3) (0.15) (D3 + D4 + D5) Ani - (Xa + Xlv) Ylv (63)

Xlv = [In (2)/50] exp [-In (2)t/365] (64)

yp = yLc + yLv (65)

dy-pLNM

5^ = fLCTLXLCyLC -(Xa +XTL)yTLNM (66)

dyiLNR

dyALNM

- flLTLXlLyTLNM -XaYTLNR (67)

^^ - fLCALXLCyLC -(Xa +XAL)yALNM (68)

dyALNR

- fALAL^ALYALNM - XaYaLNR (69)

where r, X, y, f, D, and Am are as defined in the ICRP II and Task Group models. The
subscripts, where different, are as follows: NPTB refers to the combined nasopharynx and
tracheobronchial region, IX refers to that portion of the pulmonary lung with a constant
biological half-time, LV refers to the variable half-time portion, and P refers to the
combined portions; TLNM and ALNM refer to those portions of the thoracic and
abdominal lymph nodes, respectively, in which plutonium is mobile, and TLNR and
ALNR refer to the corresponding portions where plutonium is retained indefinitely; LCG
refers to transfer from LC to GIT, LG is from liver to GIT, LCB is from LC to blood,
TLB is from TLNM to blood, ALB is from ALNM to blood, LB is from liver to blood, KB
is from kidney to blood, LCTL is from LC to TLNM, and LCAL is from LC to ALNM.
Values of the parameters in the above equations can be taken from Figs. 5 and 6. The
equations for GIT, liver, kidney, and bone are identical to the ICRP II and Task Group
models, but the parameter values are those indicated in Fig. 6. Stuart et al. (1971) did not
give estimates for total body.

Modifications to the SDB Model

We (Bloom and Martin, 1976) modified the SDB long-term model to include ingestion
and the possibility of transfer from the blood to lungs and to lymph nodes. We also
incorporated many of the long-term transfers of the Task Group model. However, we also
simplified the model somewhat by removing the variable half-life lung compartment from
the SDB model. The resulting model is shown in Fig. 7. In this figure the upper
respiratory tract (URT) refers to the nasopharyngeal and tracheobronchial regions, and
deep lung (DL) refers to the long-term component of the pulmonary compartment in the
Task Group model.

The equations for our modifications to the SDB model are

rciT = (D3 + D4 + D5) fuRcAm + foLcXoLyOL + ^LcXLyiiver + H^ (70)

500 TRAN SURA NIC ELEMENTS IN THE ENVIRONMENT

NHALATION

INGESTION

69.79%

30%

GASTROINTESTINAL

TRACT

(18-hr TRANSIT]

0.21%

22.65%

FECES

DEEP
LUNG
(3 yr)

LIVER
(40 yr)

8.5%

30%

RETAINED
INFINITE)

THORACIC
LYMPH NODES I

MOBILE
(1 yr)

70%

RETAINED
(INFINITE)

ABDOMINAL
LYMPH NODES

Fig. 7 Block diagram for modifications to the Stuart, Dionne, and Bail (SDB) model.
Fraction from blood to total body is 100%. Biological half-time in total body is 65,000
days.

TR - (D3 + D4 + D5) fuRBAm + fGITBTGIT + foLB^^DLYDL + fTLB^TLYTLNM

+ fALB'^ALYALNM + fLB'^LYliver + fKB^KYK + -^BYbone (71)

= (D3 + D4 + D5) fuRDLAm + fBDLFB - (^A + >^Dl) YDL

dYTLNM

dYTLNR
dt

= fOLTL^DLYDL + ^BJUB - i^A + ^TL) YTLNM

- flLTL^TLYTLNM + foLTR^DLYDL - '^AYTLNR

dYALNM
dt

- foLAL^DLYDL + ffiALfB " (^A + '^AL)yALNM

(72)
(73)
(74)
(75)

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 501
dVALNR

^alal^alYalnm - /^aYalnr (76)

— 3J^=fBLrB -(^A + ^L)yiiver (77)

dVK

dybone
dt

dyTB

= fBKrB-(^A + XK)yK (78)

= fBBNrB -(?^A + ^B)ybone (79)

- fBTB^B - (^A + ^TBJYTB (80)

where r, X, y, f, D, Am, and H^ are as defined in the ICRP II, Task Group, and SDB
models. The subscript URDL refers to transfer from URT to DL, BDL is from blood to
DL, DLTL is from DL to TLNM, BTL is from blood to TLNM, DLTR is from DL to
TLNR, DLAL is from DL to ALNM, and BAL is from blood to ALNM.

Parameters for the above equations can be obtained from Fig. 7. Most of these values
are identical to those of the SDB model (Stuart, Dionne, and Bair, 1971). Our
modifications are designed to make the model more applicable to chronic inhalation and
to include ingestion and transfer of plutonium from blood to DL and from blood to the
lymph nodes. The value for the transfer from GIT to blood is identical to the ICRP II
value. The values for the other parameters are based on data reported by Ballou, Park,
and Morrow (1972) for the translocation of a soluble form of plutonium (plutonium
citrate) in dogs. Stuart, Dionne, and Bair (1971) recommended a variable half-time for
part of the transfer rate from the lungs to blood, but it was felt that a constant 3-yr
half-time for the entire lung was an adequate representation of their dog data. It was also
much easier to use the constant half-time for the mathematical description of chronic
inhalation.

Like the SDB model, we assumed that 30% of the material initially deposited in URT
is rapidly transferred to DL. Since the short-term form of the SDB model transfers 99.7%
of the material in URT to GIT in a short time (8 min), the effective transfer from URT to
GIT is 69.79% (0.997 x 0.70). The corresponding transfer from URT to blood is 0.21%
(0.003 X 0.70).

Model Comparisons

The preceding models were compared by computing organ burdens and radiation doses
resulting from unit intakes by ingestion or inlialation. The units were either 1 /jCi as a
single intake or 1 juCi/day continuous intake, and the calculations were carried to 50 yr.
The resulting organ burdens and doses after 50 yr are given in Tables 10 and 1 1.

Stuart, Dionne, and Bair (1971) (SDB) applied their model to single inhalations, and
the variable half-time in the pulmonary lung did not present any difficulty. However,
there is some doubt as to the interpretation of the variable half-time for the chronic case.
A strict interpretation would imply that a significant fraction (15%) of material
continuously deposited in the pulmonary lung is eliminated with a half-time that exceeds

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504 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

140 yr after 10 yr of chronic inhalation. The lung burden could thus reach very high
levels. A fraction may indeed be retained with a long half-time, but we find it difficult to
believe that this fraction could be so liigh (15%). We therefore examined the long-term
behavior of the burden in this portion of the lung resulting from a single inhalation.
Equations 63 and 64 can be solved to yield

yLv = 0.15 yp(0)exp(-XAt)exp {-7.3[1 - exp (-0.001 899t)] } (81)

where yp(0) is the initial amount deposited in the pulmonary lung.

If we neglect radioactive decay (X^ = 0)> the initial fraction (15%) is reduced to 0.4%
after 1 yr, 0.013% after 5 yr, and asymptotically approaches a constant value of about
0.01%. On the basis of these results, we assume that this fraction (0.01%) is retained
indefinitely and the remaining 14.99% is removed with a 3-yr biological half-time.

As shown in Tables 10 and 1 1 , inhalation is the critical pathway for plutonium to all
organs except GIT. The organ burdens and radiation doses from inhalation are generally
1 ,000 to 10,000 times as great as the corresponding burdens and doses from ingesting the
same amount of plutonium. Tliis is due to the relatively large fraction (0.2 to 25%) that
reaches the blood directly from inhalation vs. the relatively small fraction (0.003%-) from
ingestion. For ingestion bone is the critical organ for the ICRP II and Task Group models,
whereas liver is the critical organ for the SDB and modified models. This difference is
explained by the fraction transferred from blood to the organ, which is 71% to the liver
and 25% to bone for the SDB and modified models, whereas the corresponding values for
the ICRP II model are 15 and 80%, respectively, and those for the Task Group model are
45 and 45%, respectively. Where the fractions are equal, the bone has the larger burden
and dose because it has the larger biological half-time.

For inhalation the lung is the critical organ for all models except ICRP II. The dose to
lymph nodes is actually liigher, but ICRP (International Commission for Radiological
Protection, 1959) does not recognize lymph nodes as critical organs. For the ICRP II
model the bone is the critical organ because this model has the highest fraction of inhaled
material that reaches the blood immediately (25%) and the shortest biological half-time in
lung (365 days).

In spite of differences in translocation pathways and biological half-times, the
radiation doses to critical organs are surprisingly similar for a given intake situation. This
leads us to use the Task Group model because it is recognized by ICRP [ICRP Publication
19 (International Commission on Radiological Protection, 1972)] , and the results using
this model are not too different from the more elaborate SDB and modified models.
Althougli still the official model of ICRP, the ICRP II model is generally considered to be
outdated. The Task Group model was used to calculate the accumulated doses and dose
commitments (to 70 yr) due to constant intake rates [Am - 0.002 Cs (pCi/day) and
Hjn = 0.19 Cs (pCi/day)] , and the results are shown in Figs. 8 and 9.

Practical Applications

Our purpose in this discussion is to show how the results of a transport- and
dose-estimation model can be applied to the practical problem of deciding whether and to
what extent environmental decontamination might be required to Umit or reduce
potential health hazards. The procedure suggested for this purpose and outlined below is
analogous to the procedure followed by ICRP in calculating maximum permissible

MODEL FOR ESTIMATING Pii TRANSPORT AND DOSE 505

10,000

20,000

DAYS

Fig. 8 Predicted cumulative doses due to ^ Pu in different organs of Standard Man

concentrations (MFC's) of radionuclides in air and water. The principal steps involved are
(1) identification of the critical exposure pathway, (2) identification of the critical organ
or organs, (3) selection of maximum-permissible-dose criteria, (4) calculation of the
corresponding MFC of plutonium in soil (MPC)s, and (5) comparison of the (MPC)s with
estimated inventories of plutonium in the surface soils of contaminated areas at NTS.

Critical Pathway

The estimated plutonium ingestion rate for a hypothetical Standard Man living in a
contaminated area at NTS is about 100 times the estimated inhalation rate, but, owing to
the very small fraction of plutonium transferred from the GIT to blood (3 x 10~^), the
GIT is the only organ that receives a significant dose from ingested plutonium. The
preferred dose-estimation model, based on ICRP recommendations (the Task Group
model. Fig. 4), shows that inhalation accounts for 100% of the plutonium deposited in
the lungs and thoracic lymph nodes, and, for an ingestion/inhalation ratio of 100,
inhalation accounts for about 95% of the plutonium in bone, liver, and kidney after 50 yr
of chronic exposure (Table 6). Clearly, inhalation is the critical pathway.

Critical Organ

According to th*^ Task Group model (Fig. 4), thoracic lymph nodes receive the highest
dose (Figs. 8 and 9), but the critical organs recognized by ICRP (International

506 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

10,000

20,000

DAYS

Fig. 9 Predicted dose commitments due to ^ ^ ' Pu in different organs of Standard Man

Commission on Radiological Protection, 1959) are bone, if the plutonium is "soluble,"
and lung if the plutonium is "insoluble." For relatively short exposure times, the model,
which makes no distinction between soluble and insoluble, predicts tliat the cumulative
dose to lungs would be considerably higher than that to bone (Fig. 8), but the cumulative
doses to 70 yr are about the same. For exposure times longer than 70 yr, the dose to
bone would be higher than the dose to lungs because of the relatively short biological
half-life of plutonium in lungs (500 days) compared with tliat of bone (36,500 days).
Since the estimated dose to lungs is higher tlian that to bone and the exposure periods
(<70 yr) and the permissible dose to lungs are lower (see below), the lung (i.e., the
pulmonary region of the respiratory tract) is the critical organ.

Permissible Dose Criteria

Current ICRP recommendations (International Commission on Radiological Protection,
1966) concerning "dose hmits for individual members of the public" indicate that the
dose to lungs should not exceed "1.5 rems in a year." Annual dose rates to a given organ
can be estimated on the basis of predicted organ burdens as a function of exposure time.
For present purposes we shall consider only the equilibrium lung burden, which, for
practical purposes, is constant for chronic exposure times in excess of about 10 yr.

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 507

Estimation of Acceptable Soil Concentration

The acceptable soil concentration (ASC) is defined as the average concentration of ^^^Pu
in the soil of an area occupied by the hypothetical Standard Man which would result in a
dose to the lungs equal to or less than the permissible dose. This value is estimated as
follows:

^3^.(1.5/365)

(E Yeq/m)

= 2817pCi2^^Pu/gsoU

where (1,5/365) = "permissible dose rate" (rem/day)

E for 23^Pu = 51.2159 X 53 = 2714 (rem/day )/MCi, or 0.0027144 (rem/day )/pCi
m = 500 g for lung
Yeq = 0.002 Csy

y = lung burden after 50 yr = 134.34 y£'\ per jLtCi inhaled per day (see
Table 1 1 )
0.002 Cs = plutonium inhalation rate (pCi/day), i.e., 20 m^ air/day X 10~^ g
soil/m^ air, and Cs is the average soil concentration (pCi/g)

Comparison of ASC and Soil Inventory Data

Tables 1 and 2 summarize the mean soil concentrations and the estimated inventories of
2 3 9,2 4 0pjj -j^ jj^g surface soils (0- to 5-cm depth) at NTS. In each contaminated study
area, soil sampling was stratified according to contour intervals (strata) previously
established by field instrument tor the determination of low-energy radiation (FIDLER)
surveys. The pertinent results for Area 13 are given in Table 12.

Only stratum 6 exceeds ASC = 2800 pCi/g, but stratum 5 is close enough to be
included in the contaminated region. Complete decontamination of strata 5 and 6(1.1%

TABLE 1 2 Estimated Inventory of ^ ^ ^ -^ "* ^ Pu in
Surface Soil (0- to 5-cm Depth) in Area 13

Area,*

2 3 9 ,2 4 Opy *.).

Soil concentration,!

Strata*

MCi/m =

pCi/g

1,245,000

1.9 ± 0.34

36 ± 7.8

2,547,000

5.8 ± 1.4

100 ±25

108,000

23 ±4.3

400 ± 75

74.000

54 ± 8.8

1100 ± 150

19.000

110± 19

2400 ± 430

24,000

4,017.000

220 ± 340

14000 ± 6400

Area weighted

mean 201

*l'rom Gilbert (1977, Table 1).

fl-rom Gilbert (1977, Table 1). Mean ± standard error.

jFrom Gilbert et al (1975, p. 393). Mean ± standard error.

508 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

of Area 13) would remove about 48% of the total plutonium in Area 13 and reduce the
average soil concentration from 201 to 106 pCi/g. If it were decided to decontaminate all
areas at NTS in wliich the average soil concentration exceeds 2 nCi/g, decontamination
would be required for about 0.2 km^ (about 50 acres) of the 11.5 km^ (about 2850
acres) included in the soil inventory study (Table 1). If the decontamination criteria were
further reduced to 1 nCi/g, the area requiring decontamination would be about 0.5 km^
(111 acres). In other words, the plutonium contamination at NTS is so concentrated in
areas near ground zero sites tliat decontamination of from 2 to 4% of the total soil
inventory area would reduce average soil concentrations by 40 to 50%.

It should be noted that possible decontamination activities of these sites are
complicated by potential damage to the desert ecosystem. Further information on this
subject can be found in reports by Wallace and Romney (1975) and Rlioads (1976).

Discussion

On the basis of the preceding results (ASC = 2800 pCi/g) and the mass-loading factor of
100 jug soil/m^ air, the expected air concentration would be 2.8 x lO"^^'' juCi/cm^ . The
maximum permissible concentration in air (MPC)a indicated by ICRP Publication 2
(10~^^ juCi/cm^) is higlier than this by a factor of about 3.6. Using (MPC)a = 10"'^
jLtCi/cm^ and a mass-loading factor of 100 jug/m^, we would find the acceptable soil
concentration to be 10 nCi/g instead of 2.8 nCi/g, which would be equivalent to assuming
a mass-loading factor of 355 instead of 100 jug soil/m^ air.

Another conservative factor in our estimate of ASC is that the lung deposition factor
(D5 = 0.31) is based on the assumption that the mean size of resuspended soil particles is
0.5 /am (AMAD). The value obtained from cascade impactor studies in the GMX area was
3 /jm (AMAD), wliich would indicate D5 < 0.2. Changing only this parameter would
increase the estimate of ASC by a factor of 1 .55 to 4266 pCi/g.

The least conservative factor involved in arriving at ASC = 2.8 nCi/g is the assumed
mass-loading factor of 100 iig soil/m^ air. As demonstrated by Shinn and Anspaugh
(1975) and Anspaugh et al. (1975), this estimate appears to be adequate for undisturbed
areas and normal winds, but liigh winds or mechanical disturbances, such as vehicular
traffic, plowing, excavation, etc., miglit increase the mass-loading factor to several
milligrams per cubic meter. If we assume, for example, that the hypothetical Standard
Man at NTS were exposed, for one reason or another, to mass-loading factors of 5000
jUg/m^ during 30 days each year, the average mass-loading factor would increase to about
500 A'g/m'^ , and our estimate of ASC would decrease to about 560 pCi/g.

The point of this discussion is that the notion of an "acceptable soil concentration" is
not fixed but is very much dependent on how man plans to use a contaminated area.
Under present conditions the ASC for contaminated areas at NTS is 2.8 nCi/g. If these
same areas were to be used for agricultural purposes or for any other purpose that would
tend to increase the average mass-loading factor, a lower ASC would be indicated. In
other words, the notion of an "acceptable soil concentration" is an attractive one. It
implies the existence of a numerical criterion that can be used to make important
determinations concerning the need for or effectiveness of countermeasures to ensure
safety. How to determine an ASC value is a different matter, and how to determine
whether a particular ASC value is entirely appropriate under a given set of physical,
social, and political circumstances is a far different matter.

MODEL FOR ESTIMATING Pu TRANSPORT AND DOSE 509

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VoL II, Physical Sciences, Pt. 1, Atmospheric Sciences, USAEC Report BNWL-1751(Pt.l),

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White and P. B. Dunaway (Eds.), ERDA Report NVO-153, pp. 207-215, Nevada Operations

Office, NTIS.
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Nevada Test Site, in Transuranium Nuclides in the Environment, Symposium Proceedings, San

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Organs and Tissues in Dairy Cattle,/. Dairy ScL, 57(9): 1055-1060.
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512 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Stanley, R. E., E. W. Bretthauer, and W. W. Sutton, 1974, Absorption, Distribution, and Excretion of
Plutonium by Dairy Cattle, in Tlie Dynamics of Plutonium in Desert Environments, P. B. Dunaway
and M. G.White (Eds.), USAEC Report NVO-142, pp. 163-185, Nevada Operations Office, NTIS.

Stuart, B. W., P. J. Dionne, and W. J. Bair, 1968, A Dynamic Simulation of the Retention and
Translocation of Inhaled Plutonium Oxide in Beagle Dogs, in Proceedings of the Eleventh AEC Air
Cleaning Conference, Richland, Wash., Aug. 31 -Sept. 3, 1970, M.W. First and J. M. Morgan, Jr.
(Eds.), USAEC Report CONF-700816, NTIS.

, P. J. Dionne, and W. J. Bair, 1971, Computer Simulation of the Retention and Translocation of

Inhaled 239-Pu02 in Beagle Dogs, in Pacific Northwest Laboratory Annual Report for 1970 to the
USAEC Division of Biology and Medicine, Vol. 1, Life Sciences, Pt. 1, Biological Sciences, USAEC
Report BNWL-1550(Pt. 1), Battelle, Pacific Northwest Laboratories, NTIS.

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Dunaway (Eds.), USAEC Report NVO-15 9, pp. 5-16, Nevada Operations Office, NTIS.

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Volchok, H. L., 1971, Resuspension of Plutonium 239 in the Vicinity of Rocky Flats, in Proceedings
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Henderson, and M. F. Milligan (Eds.), USAEC Report LA-475 6, Los Alamos Scientific Laboratory,
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USAEC Report NVO-15 3, pp. 251-337, Nevada Operations Office, NTIS.

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Wildung, R. E., and T. R. Garland, 1974, Influence of Soil Plutonium Concentration on Plutonium
Uptake and Distribution in Shoots and Roots of Barley,/. Agric. Food Chem. 22: 836-838.

A Model of Plutonium Dynamics
in a Deciduous Forest Ecosystem

CHARLES T. GARTEN, JR., ROBERT H. GARDNER, and ROGER C. DAHLMAN*

A linear compartment model with donor-controlled flows between compartments was
designed to describe and simulate the behavior of plutonium (^^^'^^^Pu) in a
contaminated forest ecosystem at Oak Ridge, Tenn. At steady states predicted by the
model, less than 0.25% of the plutonium in the ecosystem resides in biota. Soil is the
major repository of plutonium in the forest, and exclianges of plutonium between soil
and litter or soil and tree roots were dominant transfers affecting the ecosystem
distribution of plutonium. Variation in predicted steady-state amounts of plutonium in
the forest, given variability in the model parameters, indicates that our ability to develop
models of plutonium transport in ecosystems should improve with greater precision in
data from natural environments and a better understanding of sources of variation in
plutonium data.

Systems analysis techniques have been useful in simulating the fate and dynamics of a
variety of substances in ecosystems, including radionuclides (Olson, 1965; Wheeler,
Smith, and Gallegos, 1977), pesticides (Webb, Schroeder, and Norris, 1975), and stable
elements (Shugart et al., 1976). Both descriptive and predictive purposes are considered
in the building of these models. Past applications of ecosystem modeling of radionuclide
behavior in the environment have included (1) projection of the time-dependent
distribution of material within the system and (2) manipulation of the model system to
determine the sensitivity of various components to variation in transfer coefficients. The
latter exercise allows identification of critical pathways affecting radionuclide distribu-
tion in the system.

This chapter describes an ecosystem model of plutonium (^^^'^^°Pu) behavior in a
Tennessee forest. In ecosystem models the complexities of community structure and
ecological processes are often simplified to an abstract compartmental system linked by
linear differential equations (Hudetz, 1973). The model described was formulated on the
basis of data from International Biological Program (IBP) studies on deciduous forests
and investigations at a plutonium-contaminated forest in eastern Tennessee. A primary
objective of the model was to holistically describe plutonium behavior in the forest. To
this end Monte Carlo simulations of the transfer of plutonium from soil to biota and a
sensitivity analysis of the model forest were performed.

*Present address: U. S. Department of Energy, Washington, D. C.

513

314 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Design of the Model

The model was conceived to describe plutonium cycUng in a deciduous forest. Studies
from a plutonium-contaminated forest adjacent to White Oak Creek on the U. S.
Department of Energy's Oak Ridge reservation were used as much as possible to design
the model and estabhsh its parameter values. In 1944, during the Manhattan Project, the
White Oak Creek floodplain was contaminated with 2 3 9,2 4 0p|j ^^^ mixed fission
products. After 33 yr a deciduous forest dominated by white ash {Fraxinus americana,
eight trees per 100 m^) and sycamore {Plantanus occidentalis, three trees per 100 m^) has
developed on the site (Van Voris and Dahlman, 1976).

r^r\ MOi

l^/! (T

365

0.92

LEAVES

n i

0.95

LITTER

0.0092

§•

>\\

GROUND
VEGETATION

WOOD

♦

SOIL
FAUNA

1
1

SOIL

-* ' 1.7x10"''

ROOTS

0.03

Fig. 1 Diagrammatic model of plutonium transfers in a deciduous forest ecosystem
showing abstracted compartments and annual transfer coefficients. The basic model
included compartments linked by solid lines. Dashed lines indicate transfers and
compartments coupled after calibration of the basic model.

Initially, a six-compartment model with 10 transfers was set up to represent
plutonium dynamics in soil and vegetative components of the ecosystem (Fig. 1). Average
annual biomass values (grams of dry weight per square meter) and plutonium
concentrations (picocuries per gram of dry weight) were multiplied to arrive at the
amount of plutonium (picocuries per square meter) in each compartment of the forest
(Table 1). A majority of the transfers in the model were calculated on the basis of
biomass flux (grams per square meter per year) from the donor compartment. In lieu of
site-specific data, fluxes were derived from data collected for eastern deciduous forests in
the Oak Ridge area during the IBP (Harris, Goldstein, and Henderson, 1973; SoUins,
Reichle, and Olson, 1973; Harris et al., 1975). Values for annual transfer coefficients and
their derivation are given in Table 2. Parameter values for the six-compartment model
were arrived at independent of model performance. Later, some parameter adjustment
(Table 2) was necessary to caUbrate the predicted amount of plutonium (picocuries per
square meter) in the forest after a 30-yr computer simulation; the calculated inventory
was based on field data.

PLUTONIUM DYNAMICS IN A DECIDUOUS FOREST ECOSYSTEM 515

TABLE 1 Standing-Crop Biomass, Plutonium Concentrations, and

Areal Amounts of Plutonium in a 30-yr-old Contaminated

Deciduous Forest at Oak Ridge, Tennessee

Plutonium

Biomass,

concentration,

^3' '^"opu content.

Component

g/m^

pCi/g

pCi/m^

Soil*

2.6 X 10 =

1.7 X 10^

Tree rootsf

3,000

12,000

Litter!

500

3,000

Tree woodt

10,500

0.003

Ground vegetation^

110

0.15

Tree leavesf

400

0.003

1.2

*Mass is based on a 20-cm soil depth and a soil density of 1.3 g/cm^ ;
concentration is the floodplain average.

t Biomass is estimated from mensuration data (Van Voris and Dahlman,
1976) and regression equations (Harris, Goldstein, and Henderson, 1973);
concentrations are based on field measurements.

:[: Biomass and concentration data are based on field measurements.

After the six-compartment model had been cahbrated, plutonium transfers to animal
components of the ecosystem were simulated. The complexities associated with transfers
of plutonium to consumers and soil fauna in the model necessitated our simpHfying
assumptions to arrive at parameter estimates. For example, pathways for resuspension of
plutonium-contaminated soil to atmosphere and subsequent inhalation by animals were
not represented in the model. Resuspension factors (Anspaugh et al., 1975) range from
10~^° to 10~^ ^ m~^ for the forest, based on plutonium concentrations measured in soil
and air (Dahlman and McLeod, 1977). Since this resuspension factor is in the lower range
of values measured in natural environments [i.e., 10~^ to 10~^^ m~* (Hanson, 1975)],
food-chain transfer to animals is modeled as the chief transport pathway.

The forest is modeled as a closed system since inputs of plutonium (e.g., fallout
resulting in a cumulative plutonium concentration in soil of 0.02 pCi/g) are negligible
relative to the existing soil contamination (plutonium in soil ranges from ~25 to ~150
pCi/g). Forest outputs include surface mnoff, erosion, and groundwater seepage, but
these processes are beyond the scope of this model. We believe that solution-phase
transport is negligible as an output from the ecosystem because plutonium is strongly
sorbed to soil (Bondietti, Reynolds, and Shanks, 1976). In addition, the downward
movement of plutonium in soil is not considered important over the time frame of the
simulations. Although there is evidence of a downward movement of plutonium in soil
over time (Bennett, 1976; Jakubick, 1976), soil fauna could promote redistribution of
plutonium from the subsoil to the soil surface. Reichle et al. (1973) calculated a 44-yr
turnover time for the top 25 cm of forest soil due solely to earthworm activity.

Since the purpose of the model is descriptive, we have not made millennium
predictions of plutonium dynamics in the forest. Loucks (1970) points out that there is a
tendency in forests toward perturbation at time intervals ranging from decades to
centuries. Therefore simulations v^th the model were limited to time spans of less than
500 yr. Over this period of time, decreasing plutonium concentrations from radioactive
decay have a negligible effect on model predictions.

316 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Modeling Methods

The model was formulated as a linear compartment model with donor-controlled flows
between compartments. The behavior of the state variables (Xj) was described by simple
differential equations of the form

dXi
"dF

= (inputs to Xj) - (losses from X[)

TABLE 2 Values and Derivation of Annual Transfer Coefficients for Plutonium
Dynamics in a Contaminated Mixed Deciduous Forest*

Transfer

Derivation

Value, yr"

9.7 X 10-'

0.95
1.7 X 10"*

(1 10 g (ground vegetation produced) m ^ yr ' ] x (0.15 pCi

Pu/g)/[1.7 X 10' pCiPu/m' (soil)]
95% lost to litter via plant mortality (105 g m"^ yr"' )/(110

g/m^)
The initial value, ( 750 g (root production) m"^ yr"' ] x (4 pCi

Pu/g)/[ 1.7 X 10'' pCi Pu/m^ (soil)] , was adjusted downward to

make predictions of plutonium in roots at 30 yr match the cal-
culated inventory in Table 1
The initial value, [460 g (wood production) m~^ yr~' ] x (0.003 1.7 x 10~*

pCi Pu/g)/[3000 g (roots)/m^ x 4 pCi Pu/g] , was adjusted up-
ward to make predictions of plutonium in wood at 30 yr match

the inventory in Table 1
[400 g (leaf production) m"' yr"' ]/[ 10,500 g (standing 0.038

wood)/m^ ]
92% of the forest canopy is returned in leaf fall to the litter 0.92

each year (368 g m"^ yr"' )/(400 g/m')
Based on average values of branch-bole mortality in deciduous 0.0092

forests (97 g m"' yr"' )/[ 10,500 g (wood)/m^ ]
Based on measurements of root mortaUty in deciduous forests 0.23

(700 g m"' yr-' )/[3000 g (roots)/m' ]
Average decomposition rate for Fraximis and Plantanus litter 0.48

(240 g m-^ yr"' )/[500 g (litter)/m' J
Parameter fitting; represents physical resuspension of soil to 8.5 x 10"'

litter and not biological uptake
Small mammals are assumed to consume 33% of their biomass 0.03

each day [3.5 g (consumed) m"^ yr"' ]/[110 g (ground

vegetation)/m^ ]
Assumes most of the plutonium in small mammals is present in the 365

gut and has a turnover time of 1 day ( 1 day/365 days)"'
Assumes 14% of the annual leaf fall is processed by soil fauna 0.10

(50 g m-' yr-' )/[500 g (litter)/m' ]
Assumes a 30-day turnover time for soil fauna (30 days/ 12

365 days)-'
Assumes ~1% of the amount ingested is soil [0.035 g (soil)/yr] x 1.4 x 10"''

(65 pCi Pu/g)/[1.7 X 10' pCi Pu/m' (soil)J

•The diagrammatic model is shown in Fig. 1. Biomass data for deciduous forests were obtained
from Harris, (^Idstein, and Henderson, 1973; Harris et al., 1975; and SoUins, Reichle, and Olson,
1973. If plutonium concentrations are not shown, they cancel out of the calculation.

Soil to ground

vegetation
Ground vegetation

to litter
Soil to roots

Roots to wood

Wood to leaves

Leaves to litter

Wood to litter

Roots to soil

Litter to soil

Soil to litter

Vegetation to
consumers

Consumer to litter

Litter to soil fauna

Soil fauna to soil

Soil to consumer

PLUTONIUM DYNAMICS IN A DECIDUOUS FOREST ECOSYSTEM 517

The model uses time-invariant transfer coefficients. Since this is an annual model seasonal
variations in transfers are not represented. The system was modeled interactively on a
computer with a differential-equation modeling program (Rust and Mankin, 1976). The
average steady-state value of all compartments and their variabilities was determined with
COMEX (Gardner, Mankin, and Shugart, 1976). COMEX is a computer program that
uses Monte Carlo methods for analyses of donor-controlled linear compartment models.
Its features are: (1) selection of transfer coefficients for each simulation from a
multivariate normal random-number generator (the distribution is determined by
specifying means and variances for all transfers); (2) solution of the system by matrix
calculus; and (3) output of results to SAS, a statistical analysis package (Barr et al.,
1976), for analyses of state variables and relationships between state variables and model
transfers. From COMEX it is possible to evaluate the sensitivities of model compartments
to changes in transfer coefficients and provide estimates of variation in model predictions,
given variation in the model parameters.

Results and Analyses

First, the model was used to simulate plutonium dynamics in vegetative components of a
forest on the basis of the data set for the forest at Oak Ridge (basic model). Second,
simulations including the animal components of the forest were performed (expanded
model). Third, the variabihty in state variables under equilibrium conditions was
calculated for both cases, given variability in the model transfer coefficients. Last, a
correlation analysis was performed on the expanded model to identify important
plutonium transfers.

The accumulation of plutonium in vegetative components of the forest (basic model)
was simulated, starting from an initial condition of 1.7 x 10^ pCi Pu/m^ (soil). The
results are shown in Fig. 2. The amount of plutonium in litter, ground vegetation, roots,
wood, and leaves reached steady state in approximately 120yr. After this time, less than
0.1% of the total soil plutonium had transferred to aboveground components. Tree roots
and litter were the principal biological reservoirs of plutonium in the forest. Expansion of
the model by couphng it to animal compartments resulted in only a sHght alteration in
model performance. At steady state the consumer and soil-fauna components contained
<0.01 and 21 pCi Pu/m^, respectively. The addition of soil fauna lowered the
steady-state amount of plutonium in litter from =^3000 to ^2500 pCi/m^. All other
components were negligibly affected.

The selection of values of transfer-coefficient variability for the COMEX simulations
proved difficult. An accurate assignment of variation to each parameter would require
information on the form of the probabiHty density function of each transfer coefficient
(e.g., a frequency distribution of litter decay rates measured throughout the forest).
Presently, such statistical information is rarely available for a single parameter much less
an entire ecosystem. It is not uncommon for coefficients of variation (CV = standard
deviation/mean) to range from 0.5 to 3.0 for field measurements of biomass and
plutonium concentrations. Nevertheless, we have assigned CV values of 0.2 to transfer
coefficients on the assumption that variation is inversely related to sample size, and, with
a sufficient number of measurements from the forest, transfers could be quantified with
standard deviations much smaller than those normally encountered. We also assume that
in large samples these parameters will be normally distributed. Therefore the question
posed was, "Given that transfer coefficients can be measured with CV values equal to

518 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

I r

Tree wood

Ground vegetation

Tree leaves

IO,UUU

1 1 1 1 1

Tree roots

10,000

—

5,000

Litter

—

1 1 1 1 1 1

60 80

TIME, yr

100

120

140

Fig. 2 Simulated uptake of plutonium from contaminated soil (65 pCi/g) by major
biotic components of a deciduous forest.

0.2 and that the normally distributed model parameters are permitted random excursions
about their mean values, what is the variation in predicted amounts of plutonium in the
forest at steady state?"

Coefficients of variation in predicted values for plutonium in the biotic components
of the basic and expanded forest model ranged from 0.28 to 0.46 when CV values were
set at 0.2 for all transfer coefficients. Considering the basic model (i.e., vegetation only),
the predicted amount of plutonium Jn the forest exhibited considerable variation (e.g.,
the range for Utter was 1300 to 8900 pCi/m^). Similar variation was observed in the
expanded model (Table 3). Consequently, even if parameters for plutonium transport in
the forest could be measured with CV values approaching 0.2, which is unUkely, the
variation in predicted plutonium would be greater than the variation in model transfer
coefficients.

PLUTONIUM DYNAMICS IN A DECIDUOUS FOREST ECOSYSTEM 519

TABLE 3 Mean, Minimum, and Maximum Predicted Amounts of

Plutonium and Coefficients of Variation (CV) for Ecosystem

Components of the Model Deciduous Forest*

Mean,

Minimum,

Maximum,

Compartment

pCi/m'

cvt

SoUt

1.698 X 10^

1.696x10^

1.699 X 10'

0.025

Tree roots

13,300

4,920

34,100

0.325

Litter §

2,600

1,230

8,480

0.278

Tree wood

145

0.438

Soil fauna

7.8

109

0.460

Ground vegetation

7.5

0.301

Tree leaves

2.0

0.6

5.5

0.425

Consumers

0.0084

0.0039

0.022

0.309

*Statistics are based on 300 deterministic simulations to steady state with
varying transfer coefficients having CV values of 0.2.
fCV = standard deviation/ mean.
t A 20-cm soil depth.
§ Litter contaminated with some soil plutonium.

Correlations between transfers and the predicted steady-state values of plutonium in
each compartment of the model are given in Table 4. The magnitude of the correlations
between transfers and state variables indicates that the amount of plutonium in the
modeled forest was most sensitive to changes in transfers of plutonium from soil to other
components, especially roots and litter. The influence of these transfers is related to the
large pool of soil plutonium and its central location within the complex of model
pathways (Fig. 1).

Discussion

Three questions emerge from the experience of building this model:

• What does the model reveal about the behavior of plutonium in forest ecosystems?

• What are the possible sources of variation in model predictions?

• What does the uncertainty in model predictions tell us about our present ability to
develop ecosystem-scale models of plutonium behavior in the environment?

Plutonium is expected to accumulate in forest components that are characterized by
large biomass and long turnover times. For example, among the biotic components,
wood, roots, and litter made up 96% of the biomass and contained 99% of the plutonium
inventory. Nevertheless, the amount of plutonium in biota at steady state was always
<0.25% of the inventory in soil; most of the ecosystem's mass is in the soil, which has a
longer turnover time than any biotic component. Therefore the fractional transfer of
plutonium from soil to forest biomass is extremely small and is expected to remain so
indefinitely in the absence of major changes in forest structure or existing environmental
conditions.

The model transfers having the greatest effect on the amount of plutonium in soil,
tree roots, wood, and litter include (1) reciprocal exchanges between soil and tree roots,
(2) reciprocal exchanges between soil and forest litter, and (3) transfers from roots to
wood and wood to leaves. Because of the potential importance of these transfers to the
distribution of plutonium among forest components, more research is needed to

320 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

accurately quantify them before other model parameters are considered. We know the
least about critical transfers to and from soil because they are the most difficult to
quantify. For example, in this model three of the preceding transfers required some
parameter fitting to calibrate the performance of the model against field measurements.
Although potentially important to animal components because of the radiotoxicity of
plutonium, the addition of plutonium transfers to animals had a minor influence on the
major plutonium reservoiis in the forest.

Sources of variation in model predictions include spatial variation, temporal
variation, and measurement error in transfer coefficients. These sources are independent
of any bias due solely to model structure or coupling. COMEX simulations of plutonium
behavior in the model forest exclude temporal (annual and seasonal) variation in transfers
because the model transfers are considered time invariant. In this sense our simulations
differ from the time variant or stochastic modeling of plutonium behavior in
pinon-juniper forests in New Mexico (Wlieeler, Smith, and Gallegos, 1977), where
ecosystem behavior in response to random variations in climate was considered. Variation
in model predictions from the COMEX simulations can be interpreted as a consequence
of spatial variation in transfers, measurement error, or both, depending on perspective.
Intraforest variation in transfer coefficients resulting from different edaphic or
microclimatic conditions within a single forest will produce local differences in the
amount of plutonium (picocuries per square meter) in biota. Intraforest variation,
however, could be negligible relative to differences in transfer coefficients between
distinct forest stands (e.g., in different counties). Therefore the average predicted amount
of plutonium in each forest, given a uniform soil contamination, could vary, depending
on site conditions and forest species composition.

An example of how such variation between forests could bear on the assessment of
the environmental impact of plutonium is provided by considerations of fire. Assume that

TABLE 4 Correlations Between Varying Transfer Coefficients and Predicted
Steady-State Values for Each Forest Model Compartment*

Compartment

Ground Tree Tree Tree Forest Soil

Transfer Soil vegetation Consumers roots wood leaves litter fauna

Soil to tree roots -0.67 0.68 0.54 0.51

Soil to litter ' 0.68 0.45

Tree roots to soil 0.69 -0.69 -0.49 -0.52

Soil to ground vegetation 0.71

Litter to soil -0.65 -0.39

Soil fauna to soil -0.57

Ground vegetation to litter -0.74

Tree roots to wood 0.40 0.35

Consumers to litter -0.77

Tree wood to leaves -0.50

Litter to soil fauna 0.34

Tree leaves to litter -0.48

Soil to consumers 0.52

*Data are based on 300 independent simulations. Transfers were varied simultaneously before each
simulation, using a Monte Carlo random-n.umber generator. Only correlations greater than 0.30 are
reported.

PLUTONIUM DYNAMICS IN A DECIDUOUS FOREST ECOSYSTEM 521

a regional assessment involves the impact of a fire that burns litter, ground vegetation,
tree wood, and leaves. From field experiments transfer coefficients for plutonium have
been measured in forest stands over the region of interest with a precision such that CV
values are nearly 0.2 for all model parameters. Given that the soil is contaminated with
1.7 X 10^ pCi Pu/m^, the predicted amount of plutonium in the forest, at steady state, at
risk of release by fire from an average forest is 26.7 /iCi/ha. This amount could range
from 12 to 87 /.tCi/ha, however, depending on which forest was contaminated.

Attempts to model plutonium dynamics in ecosystems are hampered by uncertainties
in predictions arising from problems with system identification, quality of data, and lack
of validation. System identification, which involves determining transfers and model
structure in a way that model performance fits real-world data, is a problematic area in
ecosystem modeling because of the variability in ecological data and our lack of control
over the natural environment (Halfon, 1975). The recommendation of O'Neill (1973) and*
of Shelley (1976) to "build the simplest model appropriate to achieving the objective"
was followed in designing the present plutonium forest model. In simple models the
effects of measurement error on predictions are reduced, but inaccuracies arising from
systematic bias are increased (O'Neill, 1973).

Even when an optimal model structure can be found (i.e., one that simultaneously
minimizes inaccuracies due to systematic bias and measurement error), problems remain
because of natural variation in plutonium data from ecosystems. By simultaneously
varying all transfers, COMEX is a statistically based simulation technique that permits an
assessment of the variation in predicted amounts of plutonium in the forest, recognizing
that variance in transfers exists. Even with small amounts of variation about the
model parameters (CV = 0.2), there are typically eightfold differences in model
predictions. As Shelley (1976) points out, we cannot expect the variability in predicted
values from ecosystem models to be less than the variation in the data used to calibrate
the model. The COMEX simulations of plutonium in forest biota support this argument.
Given that CV values on plutonium data from field studies range from 0.5 to 3.0, we
conclude that the ability to adequately model plutonium transport in ecosystems is
strongly dependent on better data from natural environments and on an understanding of
the causes of variation in the data.

Data on plutonium in the White Oak Creek floodplain forest were used to calibrate
this model. Calibration data cannot be used for model validation (Shelley, 1976), and
criteria for validity (Mankin et al., 1977) are difficult to define. The model cannot be
judged valid, but it has been useful for the identification of areas where research is needed
to better our understanding of plutonium dynamics in forests and thereby develop more
precise models. Future field studies should provide the data necessary for systems analysis
and comparison of plutonium dynamics in forests and other ecosystems. It is unlikely,
however, that these models will be validated over the full time frame of some simulations
(e.g., >100yr) before advances in ecosystem analysis make the models obsolete.
Nevertheless, the long radioactive half-life of ^^^Pu (24,400 yr) and its potential for
accumulation in the biosphere, necessitate some predictions in lieu of none at all. The
model reported here represents an hypothesis that presents testable predictions about the
dynamics and distribution of plutonium in deciduous forests.

Acknowledgments

We thank S. I. Auerbach, W. F. Harris, C. A. Little, D. E. Reichle, H. H. Shugart, Jr., and
anonymous technical reviewers for their helpful comments and suggestions. The research

522 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

was sponsored by the U. S. Department of Energy (formerly Energy Research and
Development Administration) under contract with Union Carbide Corporation.
Publication No. 1334, Environmental Sciences Division, Oak Ridge National Laboratory.

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of Plutonium in Soils, Health Phys., 29; 571.
Barr, A. J., J. H. Goodnight, J. P. Sail, and J. T. Helwig, 1916, A User's Guide to SAS 76, Sparks Press,

Raleigh, N. C.
Bennett, B. G., 1976, Transuranic Element Pathways to Man, in Transuranium Nuclides in the

Environment, Symposium Proceedings, San Francisco, 1975, pp. 367-383, STI/PUB/410,

International Atomic Energy Agency, Vienna.
Bondietti, E. A., S. A. Reynolds, and M. H. Shanks, 1976, Interaction of Plutonium with Complexing

Substances in Soils and Natural Waters, in Transuranium Nuclides in the Environment. Symposium

Proceedings, San Francisco, 1975, pp. 273-287, STI/PUB/410, International Atomic Energy

Agency, Vienna.
Dahlman, R. C, and K. W. McLeod, 1977, FoUar and Root Pathways of Plutonium Contamination of

Vegetation, in Transuranics in Natural Environments, Symposium Proceedings, Gatlinburg, Tenn.,

Oct. 5-7, 1976, M. G. White and P. B. Dunaway (Eds.), USDOE Report NfVO-178, Nevada

Operations Office, NTIS.
Gardner, R. H., J. B. Mankin, and H. H. Shugart, Jr., 1976, The COMEX Computer Code

Documentation and Description, ERDA Report EDFB-IBP-76-4, Oak Ridge National Laboratory,

NTIS.
Halfon, E., 1975, The Systems Identification Problem and the Development of Ecosystem Models,

Simulation, 25: center section.
Hanson, W. C, 1975, Ecological Considerations of the Behavior of Plutonium in the Environment,

Health Phys., 28: 529.
Harris, W. F., R. A. Goldstein, and G. S. Henderson, 1973, Analysis of Forest Biomass Pools, Annual

Primary Production and Turnover of Biomass for a Mixed Deciduous Forest Watershed, in

Proceedings of the Working Party on Forest Biomass of lUFRO, H. Young (Ed.), pp. 41-64,

University of Maine Press, Orono, Maine.
, P. SoUins, N. T. Edwards, B. E. Dinger, and H. H. Shugart, 1975, Analysis of Carbon Flow and

Productivity in a Temperate Deciduous Forest Ecosystem, in Productivity of World Ecosystems,

D. E. Reichle, J. F. FrankUn, and D. W. Goodall (Eds.), pp. 116-122, National Academy of

Sciences, Washington, D. C.
Hudetz, W. J., 1973, Ecosystem Models and Simulation, Simulation, 21: 81.
Jakubick, A. T., 1976, Migration of Plutonium in Natural Soils, in Transuranium Nuclides in the

Environment, Symposium Proceedings, San Francisco, 1975, pp. 47-62, STI/PUB/410, Interna-
tional Atomic Energy Agency, Vienna.
Loucks, O. L., 1970, Evolution of Diversity, Efficiency, and Community Stability, /4m. Zooi, 10: 17.
Mankin, J. B., R. V. O'Neill, H. H. Shugart, and B. W. Rust, 1977, The Importance of VaUdation in

Ecosystem Analysis, in New Directions in the Analysis of Ecological Systems, Part 1, pp. 63-72,

G. S. Innis (Ed.), Simulation Councils Proceedings Series, Simulation Councils, Inc.
Olson, J. S., 1965, Equations for Cesium Transfer in a Liriodendron Forest, Health Phys., 11: 1385.
O'Neill, R. v., 1973, Error Analysis of Ecological Models, in Radionuclides in Ecosystems,

Proceedings of the Third National Symposium on Radioecology, Oak Ridge, Tenn., May 10-12,

1971, D.J. Nelson (Ed.), pp. 898-908, USAEC Report CONF-710501, Oak Ridge National

Laboratory, NTIS.
Reichle, D. E., D. A. Crossley, Jr., C. A. Edwards, J. F. McBrayer, and P. SoUins, 1973, Organic Matter

and ' ' ■'Cs Turnover in Forest Soil by Earthworm Populations: Application of Bioenergetic Models

to Radionuclide Transport, in Radionuclides in Ecosystems, Proceedings of the Third National

Symposium on Radioecology, Oak Ridge, Tenn., May 10-12, 1971, D.J. Nelson (Ed), pp.

240-246, USAEC Report CONF-710501, Oak Ridge National Laboratory, NTIS.
Rust, B. W., and J. B. Mankin, 1976, Interactive Differential Equations Modeling Program, ERDA

Report EDFB-IBP-74-2, Oak Ridge National Laboratory, NTIS.

PLUTONIUM DYNAMICS IN A DECIDUOUS FOREST ECOSYSTEM 523

Shelley, P. E., 1976, Field Data for Environmental Modeling — Adjacent or Integral? m Environmental

Modeling and Simulation, W. R. Ott (Ed.), Report EPA-600/9-76-016, pp. 583-585, U.S.

Environmental Protection Agency, Washington, D. C.
Shugart, H. H., D. E. Reichle, N. T. Edwards, and J. R. Kercher, 1976, A Model of Calcium Cycling in

an East Tennessee Liriodendron Forest: Model Structure, Parameters, and Frequency Response

Analysis, f'co/o^', 57: 99.
Sollins, P., D. E. Reichle, and J. S. Olson, 1973, Organic Matter Budget and Model for a Southern

Appalachian Liriodendron Forest, USAEC Report EDFB-lBP-73-2, Oak Ridge National Labora-
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Van Voris, P., and R. C. Dahlman, 1976, Floodplain Data: Ecosystem Characteristics and '^''Cs

Concentrations in Biota and Soil, ERDA Report ORNL/TM-5526, Oak Ridge National

Laboratory, NTIS.
Webb, W. L., H. J. Schroeder, Jr., and L. A. Norris, 1975, Pesticide Residue Dynamics in a Forest

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Wheeler, M. L., W. J. Smith, and A. F. Gallegos, 1977, A Preliminary Evaluation of the Potential for

Plutonium Release from Burial Grounds at Los Alamos Scientific Laboratory, ERDA Report

LA-6694-MS, Los Alamos Scientific Laboratory, NTIS.

A Review of Biokinetic and Biological
Transport of Transuranic Radionuclides
in the Marine Environment

T. M. BEASLEY and F. A. CROSS

Present understanding of the uptake, retention, and loss of transuranic radionuclides by
marine biota is limited. Laboratory experiments have demonstrated that for certain
species assimilation of plutonium and americium from labeled food is an efficient process
and that direct uptake from seawater is important in the bioaccumulation of all
transuranic radionuclides studied to date. Organisms appear to play an important role in
the vertical transport of these radioelements from the surface layers of the ocean to
greater depths.

A discussion of the biokinetic behavior of transuranic radionuclides in marine organisms
should address the rates at which these radioelements are ingested, assimilated, and
egested as well as the rates at which they are lost from the organism's tissues over time
(e.g., turnover time and biological half-life). In terms of oceanic processes, such
information is of little value, however, unless it can be used in predicting the ultimate fate
of transuranics released into the marine environment. We have chosen, therefore, to
address both these aspects of transuranic behavior and believe that this chapter, along
with the chapters by Noshkin (this volume) and by Eyman and Trabalka (this volume),
will provide a comprehensive summary of the behavior of transuranics in aquatic
ecosystems.

Background

By far the greatest amount of information to date dealing with transuranic radionuclides
in marine organisms has been confined to the determination of absolute amounts of these
radioelements in both whole animals and selected tissues with the subsequent
computation of a concentration factor to indicate the degree of biomagnification
between the organism and its environment. The use of this approach will be discussed
later.

The review articles by Noshkin (1972), Cherry and Shannon (1974), and Eyman and
Trabalka (this volume) describe the general features observed to date in aquatic
ecosystems relative to the accumulation of plutonium in aquatic organisms. Although
these data do not give information as to the rates of uptake and loss of the transuranics in
aquatic organisms, they are useful in clarifying one unfortunate error in terminology that
requires rectification. In the truest sense of the word, aquatic organisms do not
discriminate against transuranic radionuclides; if they did, the radioelements could not be
measured in the organisms deriving these entities from their labeled environments. The

524

TRANS URANIC RADIONUCLIDES IN MARINE ENVIRONMENT 525

data in hand simply indicate that uptake by organisms either by assimilation or by surface
adsorption is greater for certain species than for others.

Biokinetics of Transuranics in Marine Organisms

The pubhshed information deaUng with the uptake, assimilation, and loss of the elements
plutonium, americium, curium, and neptunium in marine organisms is limited. Apart
from the general lack of concern regarding transuranics as potential marine pollutants
until the late 1960s, there are other reasons why progress in this field has been retarded.
First, the number of laboratories having access to fresh flowing seawater and the culture
facilities necessary to undertake such research are limited. Second, the extensive radiation
protection measures required to conduct even modest tank experiments with these
alpha-emitting radionuclides coupled with the analytical task of making large numbers of
low-level alpha measurements have discouraged most investigators. Finally, thos5
laboratories which have been involved in marine transuranic measurements are reluctant
to house even small amounts of these radioelements so as to preclude the possibiHty of
sample contamination, which would compromise their low-level determinations.

It is surprising that as early as 1966 Todd (1968) and Todd and Logan (1966) had
demonstrated the feasibiUty of using ^^'^Pu (T^ = 45,6 days), which decays by electron
capture, as a tracer for metaboHc studies. However, it was not until 1974 that Bair et al.
(1974) used this isotope in a dual-labeling experiment with 239,240p^ j^^ ^ comparative
study of the distribution and excretion of the element in beagle dogs, and only recently
Fowler, Heyraud, and Beasley (1975) used the isotope to perform metabolic studies with
marine organisms. Because of its high specific activity (curies per gram), it is possible to
approach lov/ atom concentrations in labeling solutions more comparable to those found
in environments contaminated by the lower specific-activity isotopes ^^^■^'^^Pu and
^^^Pu. In addition, the 100-keV X ray emitted in the deexcitation of ^^''Np permits easy
detection by Nal(Tl) scintillation techniques and therefore permits whole-body counting
techniques to be used with small marine organisms. Therefore, for plutonium many of the
obstacles for laboratory tank experiments can be minimized, even though current
production costs of ^.^^Pu are high ($500 per microcurie) and small amounts of ^^^Pu
and ^^^Pu are present in the purified ^^''Pu.

Although laboratory experiments can be performed with care using ^^^ Am as a tracer
and counting its 60-keV X ray by scintillation techniques, relatively high activity levels
must be used and small experimental animals must be used to preclude serious geometry
problems associated with the absorption of the weak X ray in the organisms. For curium
and neptunium, there are no isotopes of long enough half-Hfe and decay characteristics to
permit in vivo measurements; thus one must use the more demanding techniques of total
alpha counting by thick source measurements (Cherry, 1964; Guary and Fowler, 1977) or
chemical isolation of the isotopes and alpha spectrometry. It is not an exaggeration,
therefore, to say that much time and effort will be expended before sufficient data are in
hand to present a reasonable picture of the biokinetic behavior of transuranics in aquatic
organisms.

Plutonium

Perhaps the first open-literature publication dealing with the uptake and tissue
distribution of plutonium in a marine organism was the work of Ward (1966). Using the
lobster Homarus vulgaris, she demonstrated that direct uptake of ^^^'^'^'^Pu from labeled

526 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

seawater did occur. Near equilibrium was reached in the exoskeleton and gills after 50
days of exposure. At 220 days the gut and hepatopancreas concentration factors
approached that of the shell and gills (^100) but gave evidence of still being far from
equilibrium. Flesh at day 220 showed concentration factors of ==3, but the shape of the
uptake curve suggests that higher values would have been reached had the experiment
been continued. Approximately 90% of the total plutonium taken up by the lobster was
found in the exoskeleton; some 4.6 and 1.2% was retained by the hepatopancreas and
flesh, respectively. As would be expected from these results, the major portion of the
plutonium accumulated by the lobster is lost during molting.

Between 1966 and 1975 very little information was pubhshed in the open literature
deahng with the actual rates of uptake and loss of plutonium in marine organisms. Zlobin
(1966; 1971), Zlobin and Mokanu (1970), and Zlobin and Perlyuk (1971) presented data
on the uptake of ^^^'^''Opu in marine algae, principally the brown a\gae Ascophyllum
nodosum, which suggested that the element was actually assimilated by the plant rather
than simply adsorbed to it. However, the subsequent work of Wong, Hodge, and Folsom
(1972), Hodge, Hoffman, and Folsom (1974), Folsom, Hodge, and Gurney (1975), and
Folsom and Hodge (1975) using other macroalgae suggests that adsorption is the more
likely mechanism for accumulation and that plutonium might be attached to large
macromolecules or micelles, which have slow diffusivities but great affinity for a variety
of surfaces. In any case evidence now exists that marine plants (phytoplankton and
free-floating and rooted algae) do accumulate transuranics to a relatively high level and
that the rate of accumulation is rapid. This process applies to both fallout-derived
plutonium (Noshkin, 1972) and that introduced from fuel reprocessing plants (Hethering-
ton, Jefferies, and Lovett, 1975; Hetherington et al., 1976; Fraizier and Guary, 1976;
1977). For phytoplankton equilibrium between the algae and water can be established in
as Httle as 5 to 10 days (Gromov, 1976).

The first laboratory experiments using ^^"^Pu to determine plutonium biokinetics in
marine organisms appear to be those of Fowler, Heyraud, and Beasley (1975). Using
mussels (Mytilus gallop) ovincialis), shrimp (Lysmata seticaudata), and marine worms
(Nereis diversicolor), they followed both uptake and loss of plutonium in the organisms
after direct uptake from seawater and, in the case of mussels and shrimp, from labeled
food as well. The valence state of ^^'^Pu tracer was chemically adjusted to either the
quadrivalent (-1-4) or hexavalent (+6) state before the isotope was introduced into the
experimental aquaria. No subsequent attempt was made to determine the valence state of
the isotope during the course of the experiment. In all three organisms direct uptake from
seawater occurred quite readily. For mussels exposed to filtered seawater containing
Pu(+6), concentration factors ranged from 20 to 60 after 26 days of exposure, and a large
percentage of the plutonium taken up resided in the shell and byssus threads. In those
cases where the byssus was removed from the mussel, greater than 80% of the activity was
associated with the shell. The activity was firmly bound to the shell material, and the
shell showed only minor losses even when rinsed for as long as 8 hr in 0.17V HCl solution.
Mussels that accumulated Pu(+6) directly from seawater showed a two-component loss
when placed in unlabeled seawater; the biological half-life (T^i^) for the fast pool
containing 35% of the total plutonium was 7 days; that for the slower turnover pool,
which contained 65% of the total plutonium, was 776 days (>2yr). Mussels that had
accumulated Pu(+4) from both food and water showed more rapid turnover owing both
to a shorter labeling time and presumably to a more rapid clearance of labeled material

TRANSURANIC RADIONUCLIDES IN MARINE ENVIRONMENT 527

voided as feces. Direct uptake of Pu(+4) by mussels from filtered seawater was not
investigated.

For shrimp direct uptake from Pu(+6)-labeled seawater was slow and was strongly
influenced by molting. A single individual that did not molt during a 25-day exposure
period reached a concentration factor of only 19. Three individuals that molted during
the first 18 days of exposure lost between 92 and 100% of their total body content of
plutonium, a value that may be artificially high owing to the capacity of such material to
further adsorb plutonium from labeled solutions once they have been cast. Animals that
molted twice during the loss period showed virtually no plutonium in these second
exuviae, which indicated that their whole-body content was indeed the result of
systemically deposited plutonium. Excretion in L. seticaudata following a single feeding
of Pu(+6)-labeled brine shrimp (Artemiaj was rapid during the first 3 days but then
decreased sharply to an exponential rate for 1 month until only 1% of the initial burden
remained. Shrimp fed daily rations of labeled ^rrem/a for 15 days did not accumulate
higher levels of plutonium than those fed a single ration of labeled food. Although shrimp
that were starved after a single feeding of labeled Artemia retained a significant fraction
of the initial dose up to day 8 (40%), they quickly eliminated this material once feeding
was resumed. It is therefore likely that accumulation in tissues other than the exoskeleton
in the shrimp would be a slow process.

Interestingly, the marked decrease in the rate of excretion after a single feeding of
labeled food and the gut clearance of this material suggested that the assimilation
efficiency in the shrimp greatly exceeded the tenths to hundredths of 1% assimilation
efficiencies reported for terrestrial mammals (Thompson, 1967). This appears to be the
case whether the plutonium is derived from food or directly from water. This was perhaps
one of the first indications that marine invertebrates were capable of retaining a
substantial portion of the plutonium they derived from these two major routes.

For worms direct uptake of plutonium from water for either the +4 or +6 valence
state was both rapid and efficient since, after 15 days of exposure, concentration factors
approached 200 for both valences. Eight days following uptake [Pu(+6)] , worms placed
in unlabeled seawater rapidly lost some 30% of their plutonium in 4 days; thereafter the
rate of loss slowed dramatically, giving a T^yj of 79 days (computed between days 4 and
35). Once again a surprisingly high percentage of the initial plutonium body burden
appeared to have been retained by the organism, but it was not determined whether the
plutonium had been systematically incorporated into tissue or sequestered by the
external mucus. Moreover, it was clearly shown that the exometabolites excreted by
worms into seawater can render the plutonium less available to fresh worms introduced
into this conditioned water.

Not only do these experiments give interesting insight into the rates of accumulation
and loss of the plutonium in the animals used but they also confirm the general tissue
distributions found in similar species that accumulate plutonium from fallout and in
those at Thule, Greenland (Aarkrog, 1971; 1977). Crustacea contain large amounts of
plutonium in their exoskeletons, molluscs retain the majority of their plutonium in the
shell and byssus threads, and polychaetes efficiently accumulate plutonium and are
expected to evidence higli levels of the element when exposed to contaminated water.
Finally, if one were to assess the relative importance of the pathways by which the
element is accumulated by marine organisms, direct uptake from water may be significant
and in some cases more significant than uptake from labeled food. By contrast, for

528 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

molluscs, which derive their plutonium from reprocessing wastes, tissue concentrations
often exceed those found in the shell, and concentration factors for benthic fishes appear
higher as well (Hetherington, Jefferies, and Lovett, 1975; Hetherington et al., 1976; Pillai
and Mathew, 1976; Guary and Fraizier, 1977; Guary, Masson, and Fraizier, 1976). This
discrepancy may result from differences in (1) bioavailability of physicochemical form,
(2) absolute levels in the environment, (3) duration of exposure, and (4) environmental
processes in various ecosystems.

Data on the biokinetic behavior of plutonium in marine fishes are scarce. Pentreath
(personal communication, 1977) has studied the direct uptake of ^^''Pu(+6) from
seawater by the plaice (Pleuronectes platessaj, as well as assimilation from Nereis sp.
injected with the isotope. Sixty-three days following direct uptake from water, the
whole-body concentration factors were <1. Complete dissections of the fish showed that
^^''Pu had concentrated in all fish in the stomach (concentration factor, 1 to 2), upper
gut (1 1 to 26), and lower gut (2 to 10). Concentration factors for skin and gill were 1 and
2 to 3, respectively. Plutonium was detectable in the livers of six of the seven fish used in
the experiment (0.8 to 1.7), and only traces were measurable in blood cells, plasma, and
bone of two fish.

The retention of ^^'''Pu by plaice fed injected Nereis sp. [both Pu(+4) and Pu(+6)]
and subsequently fed unlabeled A^ere/s sp. as maintenance rations ranged between 0.4 and
3.0% 5 days after exposure. The measured Tb^ values of ^^"^Pu for 10 fish in the
experiment ranged from 9 to 49 days. By contrast, the retention of parenterally
administered ^^^Pu(+4) in five fish that had been injected in the right dorsal muscle was
relatively high; T^i^ values ranged between 642 and 877 days. Similar results were
obtained from fish that had been injected directly in the body cavity (Tbvi, ^^~ ^° WOO
days). Redistribution of the isotope within the fish 158 days postinjection was marked;
the highest accumulations occurred in the liver, kidney, and spleen. Similar distributions
were observed for both injection sites.

Unlike the plaice the thornback ray (Raja clavata) appears to assimilate more ^^"^Pu
from Nereis sp. when the isotope is injected into the worm. Measurable amounts of the
isotope are detected in the liver at dissection (Pentreath, personal communication, 1977).
Crab digestive gland, which was incubated with ^^"^Pu and subsequently fed to both
plaice and rays, produced results similar to those from experiments in which A^ems sp.
was used. Clearly, the digestive physiologies of the two fish are sufficiently different that
enhanced plutonium uptake occurs in the ray.

The most recent evidence suggesting efficient assimilation of plutonium from labeled
food by invertebrates is the work of Fowler and Guary (1977a) in which crabs {Cancer
maenas and Cancer pagurus) were fed ^^''Pu-labeled TV. diversicolor. Remarkably high
assimilation efficiencies ranging from 20 to 60% were observed. Of the ^^''Pu absorbed
across the gut wall, 43 to 85% was found in the hepatopancreas, 8 to 43% in the shell,
and 5 to 10% in the gill. It appears to make no difference whether the initial plutonium
labeling solution contains Pu(+6) or Pu(+4). Initial results from experiments in which
mussel (M. galloprovincialisj tissue labeled by ^^''Pu uptake from water and phyto-
plankton was fed to starfish indicate similarly high assimilation efficiencies (Fowler and
Guary, 1977b). These latter experiments would tend to support the earlier contention of
Noshkin et al. (1971) that food-chain magnification in the simple food chain mussel-
starfish can occur. That crabs efficiently take up plutonium from labeled environments
has equally been demonstrated by the work of Guary, Masson, and Fraizier (1976) and
Guary and Fraizier (1977).

TRANSURANIC RADIONUCLIDES IN MARINE ENVIRONMENT 529

It would thus appear that assimilation of plutonium in invertebrates does not
constitute an inefficient process nor does it parallel the low gastrointestinal adsorption of
plutonium by terrestrial vertebrates where the plutonium has been administered by
gavage (Thompson, 1967). What role digestive physiology or the manner in which the
label is administered plays in these strikingly dissimilar results remains to be elucidated. If
biochemically -bound plutonium is more readily assimilated across gut walls of experi-
mental animals than plutonium in either ionic or chelated forms (citrate or tartrate), then
it remains to be demonstrated whether substantial differences exist in either the
distribution or the elimination of the plutonium retained by the animal. Should such
differences be observed, the implications for revising current radiation protection
standards are obvious.

Americium

As far as we are able to determine, the only published data dealing with the biokinetic
behavior of americium in marine organisms are those of Fowler and Heyraud (1974).
Using the brine shrimp (Artemia) and the euphausiid (Meganyctiphanes norvegica),
Fowler and Heyraud studied ^'*^Am uptake directly from seawater (brine shrimp and
euphausiid) and a combined food— seawater pathway (brine shrimp) in a short-term
experiment. Artemia accumulated ^'*^Am efficiently from water; concentration factors
of 1700 were attained in as little as 48 hr. By contrast, concentration factors from these
animals in a labeled phytoplankton suspension reached 400 in the same period of time.
Filtration of the seawater showed that some 81% of the ^"^ ^ Am was associated with the
algal cells; 2.0% was retained by 0.45-jum filters, and some 17% remained in the filtrate.
Recalculation of the concentration factor based on the activity in water and the 0.45-/im
filterable fraction alone gave concentration factors virtually identical to those found
when accumulation was from water only. Clearly, little ^'*' Am (if any) was assimilated
by the food pathway in this experiment. When placed in fresh seawater, both groups of
Artemia retained less than 1% of their accumulated ^"^^ Am after 3 hr, which suggested
that very little of the ^^^ Am was incorporated metabolically.

Euphausiids appeared to accumulate ^'^^ Am less efficiently than the brine shrimp and
at a somewhat slower rate. After the euphausiids had been exposed for 64 hr to labeled
seawater, concentration factors had reached a value of 125, and there was an indication
that equilibrium was being approached. When placed in unlabeled seawater, a single
euphausiid lost only 40% of its ^'*^Am burden during the first 8 days. At molting,
however, virtually all the ^'^^ Am was lost v^th the cast molt.

Aside from these observations, we have little data concerning the uptake rate,
assimilation, and loss of ^"^^Am in marine organisms. However, Fowler and Guary
(1977b) have observed relatively high assimilation efficiencies for starfish fed ^"^^Am-
labeled mussel tissue, as observed for ^^"^Pu, and retention times appear to be long.
Starfish fed a single ration of ^"^^ Am-labeled mussel tissue retained 85 to 95% of the
ingested dose 5 weeks postexposure. At dissection, ^90% of the retained ^'^^Am was
found in the pyloric caeca.

The relationships between the behavior of americium and plutonium in the marine
environment have recently been summarized by Bowen (1975) and by Livingston and
Bowen (1976a). For marine organisms there appears to be no consistent trend regarding
their bioavailability, i.e., that one is preferentially taken up by biota in favor of the other.
Virtually all the comparisons to date rest with measurements of these isotopes in biota

530 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

deriving the elements either from fallout or from fuel reprocessing wastes. However,
Pentreath and Lovett (1976) and Beasley and Fowler (1976b) have presented additional
data contrasting the relative amounts of the two isotopes in plaice collected from the
Irish Sea and the relative uptake of the two elements by polychaetes exposed to labeled
sediments. In plaice (Pleuronectes platessa) there was an indication that ^^'Am was
taken up in preference to 239,240p|^j^ although the variable discharge rates of the two
isotopes from the Windscale effluent before sampling and the mobility of the fish
themselves in the discharge area make a definitive statement regarding differential uptake
problematical (Pentreath and Lovett, 1976). For these reasons, and others raised by
Fowler and Beasley (1977), accurate estimations of the concentration factors for the fish
could not be made. For the polychaetes (Nereis diversicolor), the case is somewhat
clearer. Whether exposed to labeled Windscale sediment or to that collected from the
Bikini Atoll, uptake of plutonium was greater than that of americium (Beasley and
Fowler, 1976b).

Curium

At this time (early 1978), we are unable to find any reference to biokinetic experiments,
either field or laboratory, which deal specifically with the element curium. Moreover,
until proven otherwise, it would not be prudent to extrapolate experimental findings
derived from ^'^^Am studies to unequivocally predict curium behavior in marine
organisms. We base this statement on the previously published data of Sugihara and
Bowen (1962) and Bowen and Sugihara (1965), in which a distinct difference was noted
in the behavior of fallout '^''Ce and '^^Pm with respect to their uptake on particulate
matter in the oceans. In addition, there continue to be unsettling reports in the literature
which suggest a differential behavior between isotopes of the same element due
principally to differences in their specific activities or to potential differences due to their
existence in different physicochemical forms as a result of "hot-atom chemistry" when
formed by decay of a parent radionuclide (Volchok et al., 1975; Hakonson and Johnson,
1973; Emery, Klopfer, and Weimer, 1974; Emery and Garland, 1974; Bowen and
Livingston, 1975). In the case of plutonium isotopes (Beasley and Fowler, 1976a),
polychaetes exposed to labeled sediments from Windscale and Bikini Atoll and to spiked
Mediterranean sediments showed no preferential uptake of the isotopes measured (^^^Pu
and ^^^'^^"^Pu). There have been no comparable experiments, however, dealing with
^^^Am or ^'*^'^'*'^Cm. The rather large differences in specific activities between ^'*' Am
(3.45 Ci/g), ^^^Cm (6.83 x 10^ Ci/g), and to a lesser degree ^'^'^Cm (82.8 Ci/g) still leave
open the possibility of differential behavior between the isotopes, particularly if
concentration effects become operative at high activity levels, as have been noted
between mammahan experiments using ^■^^Pu and 2 3 9,240p^ (Bair et al., 1974).

Neptunium

The recent paper by Guary and Fowler (1977) on the biokinetic behavior of ^^^Np in
mussels and shrimp appears to be the only published paper on this subject. Direct uptake
of ^^"^Np from water by both the mussel (M. galloprovincialis) and the shrimp
(L. seticaudata) appears to be much less than that observed from plutonium. Whole-body
concentration factors of 15 to 20 for both species were observed after exposure for 3
months. Tissue distributions for the mussel followed patterns previously seen for

TRANSURANIC RADIONUCLIDES IN MARINE ENVIRONMENT 53 1

plutonium, i.e., high concentrations in the shell and lesser amounts in flesh. The
exoskeleton of the shrimp accumulates the major portion of neptunium, and loss rates are
strongly influenced by molting. When the shrimp were placed in unlabeled seawater, the
loss was biphasic; the largest part of the ^'^'Np initially present was lost with a TbVj of 4
days, and about 3% of the original activity exhibited a Tbi^ of 252 days. Mussels held in
the laboratory had Tbij values of between 180 and 226 days; faster turnover was observed
in animals held at temperatures of 25°C than in those held at 13°C. Mussels placed in the
sea showed whole-body Tt^ values of some 81 days; the faster turnover rates were
attributed to active growth of the organism.

As yet there is no information regarding the assimilation of ^^"^Np by marine
organisms fed labeled food and subsequent turnover rates of assimilated material.

Extrapolation of Laboratory-Derived Information to Natural Conditions

The entire subject of biokinetics of radioactive and stable isotopes of elements and the
extrapolation of these laboratory-derived data to the real world has been a point of
discussion among aquatic scientists for a number of years. It is extremely difficult to
design laboratory experiments that are short term and simplistic relative to oceanic
processes which will provide predictive information on the accumulation and redistribu-
tion of radionuclides by marine organisms. In addition, field verification of laboratory-
derived information often is not possible. Inherent problems in experimental designs for
studying the transfer of radionuclides in marine food chains are discussed in a recent
review by Cross, Renfro, and Gilat (1975).

Laboratory experiments must be designed which will present the radionuclide to the
test organism in a manner similar to that which occurs in nature; i.e., it must be allowed
to distribute between particulate and dissolved fractions realistically and must be of a
similar specific activity. In addition, feeding rates of the test organisms must reflect
natural conditions, as should population densities in the experimental aquaria. These
conditions require a combination of laboratory and field observations and basic
information on the general ecology of the test organism, which often is not available.
Realistic estimates of uptake rates, assimilation efficiencies, and turnover rates can only
be obtained by carefully designed experiments (Cross, WiUis, and Baptist, 1971; Cross
et al., 1975; Cross, Renfro, and Gilat, 1975; Willis and Jones, 1977). Perhaps one of the
major problems in radiotracer experiments is that of incomplete labeling (Willis and
Jones, 1977).

Another important aspect of biokinetics of radionuclides which warrants discussion
here is the use (or misuse) of concentration factors. At this time we do not know what
fraction of an element in nature is bioavailable relative to concentration factors, and we
obviously are not in agreement (Lowman, Rice, and Richards, 1971; Cross, Renfro, and
Gilat, 1975; Fowler and Beasley, 1977). Examples can be found in the literature which
base concentration factors on (1) dissolved concentrations in the water, (2) total
concentrations in the water, (3) concentrations in food organisms, and (4) concentrations
in sediment. The use of any of these four fractions will depend on the author's prejudice
relative to the source of the element to the organisms under study. Presently, we do not
know what physicochemical forms of elements are most bioavailable to marine organisms.
In fact, we have not even developed adequate experimental designs to determine the
relative importance of food and water in conveying elements to aquatic animals (Cross
and Sunda, 1979). In reality, bioavailable fractions of elements probably consist of a

532 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

complex interaction of food, water, and sediment, depending on trophic level and feeding
type.

Another problem in using concentration factors in field studies is that concentrations
of elements in water or particulate materials are highly variable and organisms under
study usually are constantly moving both horizontally and vertically in the ocean.
Concentrations of elements in organisms represent an integration of environmental levels
to which the organisms have been exposed, and environmental levels at the point of
capture may not actually reflect the history of exposure. Therefore there is no known
bioavailable fraction or constant environmental level that can serve as a realistic value on
which we can base concentration factors. For these reasons we believe that the use of
concentration factors in aquatic studies should be approached more cautiously in the
future than in the past.

Biological Transport of Transuranics in the Ocean

Lowman, Rice, and Richards (1971) have discussed the relative importance of biological
vs. physical processes in the distribution of radionuchdes and trace metals in the ocean
and have included an excellent bibhography on the subject. It has become increasingly
apparent that the biological processes operative in the mixed layer of the ocean greatly
influence the vertical transport of materials from it and that biological activity near the
sediment— water interface may be a key mechanism for transport of recently deposited
materials to depth in the sediment. Although we have chosen to address only the data
relating to transuranics in this regard, the bulk of the evidence for vertical biological
transport of materials from the mixed layer to depth has come from measurements of
trace metals (Kuenzler, 1969; Fowler et al., 1973; Small and Fowler, 1973; Small, Fowler,
and Keckes, 1973), natural radionuchdes (Broecker, Kaufman, and Trier, 1973; Cherry
et al., 1975; Beasley et al., 1977), artificial radionuchdes (Osterberg, Carey, and Curl,
1963), and synthetic organic compounds (Elder and Fowler, 1977). Vertical mixing of
sediments by biological activity has been well established (Davidson, 1891; Dapples,
1942; Emery, 1953; Gordon, 1966; Glass, 1969; Rhoads, 1974), and this process
obviously will affect the vertical distribution of radionuclides associated with sediments.

Bowen, Wong, and Noshkin (1971) first demonstrated that plutonium subsurface
maxima occur in the upper 1000 m of the ocean and attributed its removal from the
mixed layer to biogenic particle fluxes. Subsequently Noshkin and Bowen (1973)
proposed a model to explain both the vertical distributions observed in the water column
and the small but measurable amounts of plutonium found in deep-sea sediments as a
function of fallout delivery. This heuristic model invokes a mixed-particle population,
30% sinking at 392 m/yr, 40% sinking at 140m/yr, and 30% sinking at 70 m/yr; no
assumptions concerning the exact nature of the particulate matter involved are required.

Direct evidence for the association of plutonium and americium with particulate
matter in the upper mixed layer of the oceans has been confirmed by several
investigators. Livingston and Bowen (1976b) have found that as much as 70% of fallout
plutonium can be removed by Milhpore filtration of open North Atlantic surface seawater
(presumably 0.45 )um); in coastal waters near Woods Hole, Mass., something over 90% of
the plutonium is associated with the particulate phase. Evidence is accruing which would
suggest that americium is also associated with particulate matter, although the number of
analyses are fewer than those for plutonium. Silker (1974) has measured both the soluble
and particulate plutonium in Pacific surface waters and has found that some 55 ± 7% of

TRANSURANIC RADIONUCLIDES IN MARINE ENVIRONMENT 533

the plutonium is being retained by alumina; the remainder is fixed to glass-fiber prefilters
with an effective pore size of 0.3 ^im. More recently, Holm et al. (1977) measured the
particulate plutonium and americium retained by 0.45-/um MilUpore filters from open
ocean surface Mediterranean waters (Tables 1 and 2). Substantially lower fractions of
plutonium were found on the filters than was reported by Livingston and Bowen or by
Silker. The average percent of plutonium retained for 14 samples (1700 to 7680 liters)
was only 3.8 ± 0.2%. By contrast, the average percent of the ^^'Am retained by the
filters was 10 ± 1%.

That fractionation of ^'*'Am and plutonium isotopes is occurring in the water
column is inferred from tlie profile data of Livingston and Bowen (1976a) by the
frequency with which the ^^ ' Am/*^^"^''^^°Pu ratio in deep waters from the Atlantic
exceeds those ratios observed either on land (Krey et al., 1976) or in coastal sediments
(Livingston and Bowen, 1976a). This ratio at present is calculated to be 0.22, which is in
good agreement with undisturbed soil measurements (0.22 to 0.25) and shallow coastal
sediments (0.20) (Krey et al., 1976; Livingston and Bowen, 1976a). Deep-water samples
from the North Atlantic, however, appear to have '^^ Kml'^'^'^^^Vxx ratios significantly
different (higher) from this range of values. Profile data from the North Pacific are fewer
in number but do not appear to evidence the same trend. Although the actual mechanism
for this apparent fractionation is yet to be proven, it is conceivable, as suggested by
Livingston, Bowen, and Burke (1976), that ^^' Am is preferentially more associated with
inorganic particles than are plutonium isotopes. Certainly the data from the Mediter-
ranean make the hypothesis an attractive one. Fukai, Ballestra, and Holm (1976) have
shown that ^'^'Am is depleted in Mediterranean surface waters relative to ^^^'■^^^Pu.
The mean ratio was 0.055 ± 0.007 from nine stations throughout the Mediterranean (July
to September 1975), a value significantly different from 0.22. From core samples taken
from the Mediterranean by Livingston, Bowen, and Burke (1976), the surface-sediment
ratios averaged 1.2 ±0.4 and are equaled only by values from the northeast South
American slope sediments off the Guiana coast (0.7 to 1.2). The Millipore filter data of
Holm et al. (1977) shown in Tables 1 and 2 evidence a distinct enrichment of
^'^^ Am/^^^"^'*'^Pu; this ratio in unfiltered seawater is 0.055 ±0.007, whereas the same
ratio on 0.45-/jm filters is 0.13 ±0.05 (la level).

It is well known that the Mediterranean has a high proportion of terrigenous detritus
in its waters and sediments (Emelyanov and Shimkus, 1972), and biological productivity
is known to be low (Brouardel and Rink, 1956). These facts, coupled to the data already
in hand from the Mediterranean basin, appear to us to be increasingly compelling
evidence that, in fact, fractionation of ^^^Am and plutonium isotopes can occur,
depending on the nature of the particulate matter in the water column. Slightly damaging
to the argument, however, is the fact that ^"^ ' Am/'^'^^''^'**^Pu ratios in open-ocean
plankton appear to be very nearly those observed for the mixed-layer waters (Livingston
and Bowen, 1976b). The same appears to be true iox plankton samples taken from Lake
Michigan (Wahlgren et al., 1976). One piece of critical information that is still missing,
however, is the measurement of '■*' Am/^^^'^'*°Pu ratios in zooplankton metabolic
particulate products (molts and fecal pellets), which would be invaluable in helping to
clarify this question.

That molts and fecal pellets from zooplankton can transport plutonium to depth has
been clearly shown by the recent laboratory data of Higgo et al. (1977), in which
239,240p|j nieasurements on the molts and fecal pellets of the euphausiid Me^a/n'cr/-

534 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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TRANS URANIC RADIONUCLIDES IN MARINE ENVIRONMENT 535

TABLE 2 23 8p^/23 9 + 240p^3nji24,^j^/23 9 + 240p^^^jj^jjy

Ratios in Mediterranean Surface Waters

Station

(2 3 8py/2 3

'+=^°Pu)x 100

e^'Amr

'^'+^^''Pu)x 100

No.

In water

In particulate

In water

particulate

11 ± 1

3.5 ±0.8

3± 1

10 ±2

4.1 ±0.8

11± 2

4.8 ± 0.8

18±4

4.3 ± 0.6

9± 3

5 ± 1

3.6 ±0.8

15 ± 2

5 ± 1

4.8 ±0.8

4± 2

21 ± 3

5 ± 1

4.1 ±0.8

6±2

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6±2

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3± 1

11± 3

9±2

3.8 ± 1.4

4± 1

Average*

7 ± 3

4.2 ± 0.5

4± 1

13 ±5

^'Iirrors are given in standard deviations.

phanes norvegica were combined with soluble plutonium excretion experiments to arrive
at the flux rates of plutonium through the organism. Essentially 99% of the plutonium
taken up by M. norvegica is excreted by fecal pellets. A crude estimate of the removal
time of plutonium from the mixed layer by fecal-pellet production alone is 3.6 yr.
Although the uncertainties of such an estimate were willingly admitted by the authors, it
is interesting to note the approximate order-of-magnitude agreement between this
estimate and that of Hodge, Folsom, and Young (1973) and Folsom (1975), which
ranged from half removal times of 3.5 yr to complete removal times of 1 yr from the
surface layers of the ocean.

Although evidence is accumulating on vertical transport mechanisms for the
transuranics as a result of planktonic biological processes (at a painfully slow rate!), we
are unaware of any published data on the redistribution of these elements either
horizontally or vertically by other marine organisms except reports that attribute the
redistribution of plutonium and americium from surface to deeper sediments to
bioturbation (Livingston and Bowen, 1976a; Livingston, Bowen, and Burke, 1976).

Conclusions

The amount of information now in hand Concerning the biokinetic behavior of
transuranic radionuchdes in marine organisms is astonishingly small and, as discussed
earlier, is plagued with difficulties inherent with experiments of this type. Even so there is
now evidence that certain marine organisms exhibit relatively high assimilation
efficiencies for the transuranics which are quite unlike those seen for terrestrial
vertebrates. There is increasing evidence to suggest that, despite similar aqueous
chemistries, fractionations among transuranics in the ocean can occur and are mediated
by the nature of the particulate matter to which they are adsorbed. Removal rates to
deep water and sediments would thus be quite different through thd world oceans. A
single attempt at numerically estimating the importance of zooplankton metabolic
products as a transport mechanism of plutonium to depth has shown that such
particulates are Ukely a key mode of removal. Yet this evidence comes from investigations

536 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

using only one species of a rather large macroplankton whose particulate products (molts,
fecal pellets, and carcasses) would be expected to reach depths at fairly rapid rates. The
microzooplankton, which constitutes by far the larger portion of the biomass at trophic
level II (herbivores), most certainly contributes to this process, yet the basic biological
data concerning fecal-pellet and molt production rates under varying food conditions
comparable to data obtained for selected macrozooplankton species have yet to be
formulated in a way that could be used to refine vertical transport estimates for the
transuranics. Finally, the importance of the food vs. water pathway for uptake of
transuranics by marine biota has been established for only a very few species, and even
fewer definitive experiments have investigated direct uptake from labeled sediment. Until
substantial progress is made in each of these areas, our understanding of the behavior of
transuranics in the biotic component of the marine environment and its attendant
influence on the movement of these elements in the ocean, including transport back to
man, will continue to be inadequate.

Acknowledgments

We have attempted to select information that would give the reader a general overview of
the field and the directions that are being followed by researchers in the discipline. There
are Ukely omissions of some of the work of our colleagues, and for this we apologize. The
interested reader is urged to consult the articles cited to gain further insight into the
complexities of the subject as well as a full appreciation of the arduous tasks ahead.
Support for this work has been received from the U. S. Department of Energy under
contract Ey-76-5-06-2227, Task Agreement No. 30, and from a cooperative agreement
between the U. S. Department of Energy and the National Marine Fisheries Service.

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TRANS URANIC RADIONUCLIDES IN MARINE ENVIRONMENT 539

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Pentreath, R. J., 1977, Fisheries Radiobiological Laboratory, Lowestoft, United Kingdom, private

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Radioactive Contamination of the Marine Environment, Symposium Proceedings, Seattle, 1972,

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Agency, Vienna.
Thompson, R. C, 1967, Biological Factors, in Plutonium Handbook, Vol. II, pp. 785-829, Gordon

and Breach Science PubUshers, New York.
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pp. 27-63, National Academy of Sciences, Washington, D. C.
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Transuranics and Possible Chemical Homologues in Lake Michigan Water and Biota, in

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Morsk. Rybn. Khoz. Okeanogr., 29: 169-175.

340 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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159-168.

Geochemistry of Transuranic Elements
at Bikini Atoll

W. R. SCHELL, F. G. LOWMAN, and R. P. MARSHALL

The distribution of transuranic and other radionuclides in the marine environment at
Bikini Atoll was studied to better understand the geochemical cycling of radionuclides
produced by nuclear testing between 1946 and 1958. The reef areas, which are washed
continually by clean ocean water, have low levels of radionuclide concentrations.
Radionuclides are contained in fallout particles of pulverized coral. In the water these
particles may dissolve, be transported by currents within the Atoll, or enter the North
Equatorial Current by tidal exchange of water in the lagoon. The transuranic elements are
distributed widely in sediments over the northwest quadrant of the Atoll, which suggests
that this area serves as a settling basin for particles. The distribution of plutonium in the
water column indicates that plutonium in th( sediments is released to the bottom waters
and then is transported and diluted by the prevailing currents. Upon interaction with the
lagoon environment, plutonium occurs in several physicochemical states. Laboratory tests
and field studies at Bikini show that approximately 15% of the plutonium is associated
with the colloidal fraction. Different ^^^Pu/^^^ '^'*^Pu ratios found in sediments,
suspended particulates, and soluble fractions suggest that ^^^Pu may be more ^'soluble"
than ^^^'^^^Pu. Different isotope ratios for the physicochemical states of plutonium
radionuclides may be due to differences in decay rates and/or the mode of formation.

Bikini Atoll was one of the sites used for nuclear weapons testing between 1946 and
1958. In the 19 yr since cessation of testing, physical decay and environmental processes
have removed or reduced significantly many of the radionuchdes that resulted. However,
several fission and neutron-induced radionuclides, such as ^°Sr, ^^^Cs, ^°Co, ^^Fe,
'^^Eu, and ^°^Bi, which have half-lives of 2 to 30 yr, can still be measured easily in
sediments, soils, and some biota. In addition, unburned fissile and device materials of
uranium and plutonium, as well as many of the neutron-induced transuranium
radionuclides, such as americium and neptunium, which have half-lives of 10^ to 10^ yr,
still remain in the Atoll ecosystem. These transuranic elements generally decay by
alpha-particle emission, and their measurement requires detailed chemical analysis of
samples.

In 1946 the Marshallese living on Bikini were evacuated from the Atoll during the
U. S. nuclear testing program. Today at Bikini Atoll a potential health hazard from these
long-hved radionuclides may still exist to the returning Marshallese people. The potential
release of transuranic elements to coastal marine environments other than Bikini is
indicated by the projected increase in the global use of plutonium in power reactors by
more than 10^ times between 1971 and 2000 (Shapley, 1971). This exposure to the

541

542 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

population from plutonium will come from power-reactor waste products and/or
accidents.

In the marine environment the distribution of transuranium elements occurs through
biogeochemical processes, which effect a transfer of materials between the sediments,
waters, and biota of the ecosystem. Surprisingly few data have been published on the
redistribution of radioactivity between the sediments and waters of the lagoon in the
contaminated environment of Bikini Atoll. This has been partly due to the secondary
importance placed on sediment studies compared with studies of the uptake of
radioactivity by flora and fauna and with studies of the radiation exposure to inhabitants
of the Atolls. The sediment environment at Bikini has been disturbed significantly at the
sites of the 23 nuclear detonations, and a significant time period may have been required
to achieve quasi-steady-state concentrations of sediments and radionuclides. Recent
studies on the concentrations of long-lived radionuclides remaining in the lagoon
environment have indicated that nearly steady-state processes may now exist (Noshkin
et al., 1974). Although the problems of data interpretation presented by the complex
sources of the introduction of radioactivity into the lagoon still remain, the present
situation at Bikini offers unique opportunities and advantages for the study of the
physical and biogeochemical processes, which have governed and will continue to govern
the fate of radionuchdes in this marine ecosystem. The data and interpretations at Bikini
will help in assessing the general hazards of plutonium in the marine environment. The
objective of this chapter, then, is to review the current status of studies on the processes
and mechanisms that control the distribution of transuranic elements in the Bikini lagoon
ecosystem.

Sources in the Bikini Ecosystem

The formation of transuranic elements at Bikini resulted from the detonation of fission
and fusion devices of different sizes using different fissile materials (■^"'^U, ^"^^Pu, and
^^^U). The largest test was the Bravo event of the Castle series (1954) — 15 Mt
equivalent TNT. This device consisted of the fission-fusion— fission process [^^^U
(^^^Pu)— LiD(T)— ■^^^U] . The transuranic elements now present in the lagoon environ-
ment are from unburned fissile material, energetic particle-induced activation products,
and decay products which were incorporated in or on coral material. Recently,
information has been obtained that •^'*^Cm was used as a fallout tracer of the transuranic
elements in several nuclear devices. Significant amounts of ■^^^Pu would now be present
in debris from the alpha-particle decay of ■^'*'^Cm (ti^, 162 days). The formation of the
coral fallout particles resulted from interaction of vaporized device and soil materials in
the fireball with the environmental materials that were swept into the expanding fireball
and cloud at later times (Adams, Farlow, and Schell, 1960).

The locations of the 23 detonations reported at Bikini are shown in Fig. 1 , and the
detonation parameters are given in Table 1 . The yields of the largest detonations reported
were: Bravo, 15 Mt in 1954 at location B;Zuni, 3.53 Mt in 1956 at location C;and Tewa,
5.01 Mt in 1956 at location G. There was also a "several megaton" airburst detonation in
1956 wliich probably resulted in relatively minor contamination of lagoon sediments.
Typically, two types of sites were used for testing nuclear devices at Bikini, and each
probably gave rise to fallout particles of distinctly different compositions and structures.

The first was for devices exploded over water deep enough to prevent the
incorporation of large quantities of soil in the ensuing fireball and cloud (sites A, F, D,

TRANSURANIC ELEMENTS AT BIKINI ATOLL 543

Fig. 1 Approximate locations of nuclear tests at Bikini Atoll.

and E in Fig. 1). Devices detonated on barges at Bikini under these conditions contained
large quantities of iron and coral which were used as barge ballast (Adams, Farlow, and
Schell, 1960). Spherical particles (< 1 jum) of "dicalcium ferrite" (2 CaO X Fe203)
formed from vaporization of the barge and ballast contained about 85% of the
radioactivity in the fallout droplets (Schell, 1959); the saturated sodium chloride (sea
salt) droplets, in which these insoluble solids were suspended, contained the remaining
15% of the measured radioactivity (Farlow and Schell, 1957).

The second common site was the shallow water or island environments where the
largest tests were conducted (sites B, J, G, I, C, and H in Fig. 1). From explosions of this
type, Adams, Farlow, and Schell (1960) found that condensation of the vaporized
materials typically occurred as impurities into and on the surfaces of the coral soils swept
into the fireball, which produced two distinct types of fallout particles, spherical and
angular. The spherical particles consisted of CaO, which was partially hydrated to
Ca(0H)2 . A surface coating of Ca(0H)2 and/or CaCOa was present owing to the reaction
of the particles with water vapor and atmospheric COo during the fallout. These particles
were formed by high-temperature (>2570°C) vaporization of coral with subsequent
condensation of the oxide as spherical particles, which lost their normal porosity. The
radioactivity was almost uniformly distributed throughout the particles. The angular
particles consisted of Ca(0H)2 with a thin outer coating of CaCOa. Some of these
particles contained unmelted coralline sand fragments as the central core; the bulk of the
radioactivity was in the outer carbonate shell. The angular shape of these particles, the
lack of incorporated radioactivity, and the presence of occasional unmodified sand grains
suggested that these particles were formed from nonvolatilized coral that had been heated
enough to melt and decarbonate (800 to 900°C) while incorporating only an outer
surface of condensing radionuclides. Occasionally 10-jL/m and smaller oxide spherical

344 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Announced Nuclear Detonations at Bikini Atoll

Operation

Height,

Map ref.

and event

Date

Location

Yield

(Fig. 1)

Crossroads

Able

6/30/1946

+520

Air

Nominal

Baker

7/24/1946

-90

Water

Nominal

Castle

Bravo

2/28/1954

Surface

15 Mt

Romeo

3/26/1954

Barge

Koon

4/6/1954

Surface

llOkt

Union

4/25/1954

Barge

Yankee

5/4/1954

Barge

Redwing

Cherokee

5/20/1956

4320

Air

Several
megatons

Zuni

5/27/1956

Surface

3.53 Mt

Flathead

6/11/1956

Barge

Dakota

6/25/1956

Barge

Navaj o

7/10/1956

Barge

Tewa

7/20/1956

Barge

5.01 Mt

Hardtack Phase I

Fir

5/11/1958

Barge

Nutmeg

5/21/1958

Barge

Sycamore

5/31/1958

Barge

Maple

6/10/1958

Barge

Aspen

6/11/1958

Barge

Redwood

6/27/1958

Barge

Hickory

6/29/1958

Barge

Cedar

7/2/1958

Barge

Poplar

7/12/1958

Barge

Juniper

7/22/1958

Barge

particles were observed adhering to these particles. Because of the dense property of the
CaO/Ca(OH)2 particles, their atmospheric hydration was dependent on the aqueous
environment encountered during fallout and sedimentation. Complete hydration, which
was observed in laboratory tests over several weeks' time, was found to be accompanied
by a 100% increase in particle volume and in the development of a crumbly, fluffy
structure. The CaO/Ca(OH)2 particles began to dissolve slowly when wet with seawater.
The freed calcium ions reacted with sulfate ions in the seawater to form calcium
sulfate-dihydrate (gypsum), whereas tlie hydroxyl ions reacted to form insoluble
Mg(0H)2 . A hard shell of Mg(0H)2 formed around the particle, which, during the period
of observation, apparently stopped any further reaction with seawater; a region of
Ca(0H)2 remained on the inner surfaces of the spherical particles. The remaining
radioactivity was associated with the Ca(0H)2 in the center of the sphere. Some of the
freed calcium ions in the spheres also formed CaCOs by reaction with bicarbonate ions in
seawater. The time history of the distribution and redistribution of transuranic elements
has been intimately associated with these particles and with their redistribution by the
physical circulation system of the lagoon.

TRANSURANIC ELEMENTS AT BIKINI ATOLL 545

Distribution in Surface Sediments

Tlie size of the detonation craters and the extent of the impact on the reef ecosystem are
shown in Figs. 2 and 3. The Bravo crater, a dish in the reef, is approximately 550 m in
diameter and about 40 m deep. The several large craters in the reef are evident from the
photographs. Finely divided coral particles, which resulted from the detonations, appear
on the reef flat near Bravo crater and on the lagoon terrace extending south toward the
islands of Bokdrolulu and Bokaetoktok. The sampling stations for the biogeochemical
survey trip* in 1972 are shown in Fig. 4.

A thin-source survey method for alpha radioactivity was developed to initially scan
the surface sediments collected in 1972 (Marshall, 1975). The results of this rapid total
alpha analysis are shown in Fig. 5. The highest total alpha radioactivity is shown not to be
in the bomb craters but to be distributed widely over the northwestern quadrant of the.
lagoon. Thus the principal source of transuranic elements to the water is a large area in
the lagoon; the maximum concentrations are near the Namu Island— Bravo crater area.
The plutonium and americium concentrations were determined in the surface sediments,
and the results of these analyses are shown in Figs. 6 and 7, respectively, and in Table 2
(Marshall, 1975; Nevissi and Schell, 1975; Schell and Walters, 1975). The general
distribution pattern of plutonium and americium in the isopleths of Figs. 6 and 7 is the
same as that shown previously for the results obtained by the total alpha method of
analysis of sediments, which indicates that most of the alpha radioactivity is derived from
plutonium and americium.

Many of the sediment samples collected for analysis in the study by Marshall (1975)
consisted predominantly of coralline particles, which were much smaller in size than
natural Marshall Island Atoll sediments, as described by Emery, Tracey, and Ladd (1954)
and Anikouchine (1961). The sediments were probably pulverized by the detonations and
were distributed in the lagoon; the finely divided particles contained the highest
concentrations of radioactivity. The proportion of the finely divided material (<16^(m)
in each sample was estimated visually.

Surface sediments collected from stations C-1, C-3, C-4 (Bravo crater), B-2, and B-20
(lagoon) consisted entirely of fine-grain material. Surface sediments collected from
stations C-7, C-8, B-21, and B-30 contained 45 to 95% fine-grain material. Sediments
collected from stations B-18 and B-19 contained approximately 20 to 40% pulverized
material. All other sediments contained widely varying portions of fine material but
generally less than approximately 10 to 15% by volume (Marshall, 1975).

Two observations were made regarding the distribution of pulverized sediments and
the distribution of ^^^'^"^^Pu, for example. Sediments collected at stations C-5, C-10,
and C-11 (S-16), which had much lower concentrations of radionuclides than did
sediments collected at the nearby stations, C-1, C-2, C-3, C4, C-6, C-8,and C-11 (S-19),
respectively, also contained lower proportions of fine-grain material. Although a similar
relationship held for most of the sediments collected, there were three obvious
exceptions. These exceptions occurred for sediments collected at stations B-21, B-22, and

*This sampling trip was initiated by the Energy Research and Development Administration. The
Puerto Rico Nuclear Center vessel R. V. Palumbo was used for the trip, and the chief scientists were
Frank Lowman, Puerto Rico Nuclear Center, Victor Noshkin, Lawrence Livermore Laboratory, and
William Schell, University of Washington.

(Text continues on page 551.)

546 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

•■"^^-i^^

(a)

(b)

Fig. 2 Aerial photograph of northwestern Bikini Atoll, (a) Tewa crater area, including
Namu Island, (b) Bravo crater area and the western reef.

TRANS URANIC ELEMENTS AT BIKINI ATOLL 547

(a)

^^^^^^k

1 II I !■ ■■ 1^

ic-Mimwitia-ffim

(b)

Fig. 3 Aerial photograph of Bikini Atoll, (a) Southwestern reef toward the deep passes
at Bokdrolulu and Bokaetoktok Islands, (b) Zuni crater area.

548 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Wave direction

Wind direction
during sannpling

D-2

BOKDROLULU

, ftiBOKAETOKTOK

C:;;)B-7 B-8

^^, C-11,12

Fig. 4 Sampling stations and wind and wave directions measured during the joint
sampling program in 1972 (Adapted from Noshkin, 1974).

Kilometers
J — f — f — I — I — I — ( — I — I — I — I — I — I — I I

.\ 0 14

\ 10-fathom contour

114 144 106 124
STATIONS

Fig. 5 Distribution of total alpha concentrations at Bikini Atoll lagoon. Q)ncentrations
in picocuries per gram of sediment.

TRANSURANIC ELEMENTS AT BIKINI ATOLL 549

40 45

STATIONS

Fig. 6 Distribution of ^ '' '^^"Pu concentrations at Bikini Atoll lagoon. Concentrations
in picocuries per gram of sediment.

38 24 24

STATIONS

Fig. 7 Distribution of ^ "" Am concentrations at Bikini Atoll lagoon. Concentrations in
picocuries pet gram of sediment.

330 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Distribution of 2 ^^'240pu, 2 ^«Pu. and 2^ 'Am and

the '^^Pu/"'''^'*°Pu Ratio in Surface Sediments

Station

Sample
location

Concentration,*

pCi/g (dry weight)

location

2 39 ,2 4 0 p„

238pu

2 38 py/2 3 9 ,2 4 0py

^^'Am

B-2

S-20

107.0 ± 11

1.11 ± 0.41

0.010 ± 0.004

69.5 ± 2.6

B-3

S-23

19.7 ± 1.3

0.80 ± 0.18

0.041 ± 0.009

9.35 ± 0.52

B-4

S-2I

3.17 ±0.65

0.27 ± 0.15

0.087 ± 0.018

1.74 ±0.16

B-6

S-14

3.44 ± 0.78

1.52 ± 0.06

B-7

S-18

5.16 ±0.51

0.54 ± 0.12

0.104 ±0.025

2.41 ±0.18

B-8

S-12

4.11 ±0.47

0.234 ± 0.078

0.057 ± 0.02

1.09 ± 0.14

B-10

S-5

0.381 ±0.083

0.015 ± 0.022

0.272 ± 0.087

B-15

S-1

3.71 ±0.38

0.085 ± 0.046

0.022 ± 0.012

1.88 ± 0.17

B-16

S-8

3.13 ±0.27

0.050 ±0.019

0.016 ±0.006

1.80 ±0.15

B-16

S-7

5.73 ±0.43

0.067 ± 0.045

0.012 ± 0.008

2.83 ±0.27

B-18

S-9

64.2 ± 1.7

2.11 ± 0.22

0.033 ± 0.004

46.5 ± 0.8

B-19

S-24

87.0 ±6.2

0.70 ±0.18

0.008 ± 0.002

62.6 ± 1.6

B-20

S-22

121.0 ± 7.0

0.53 ± 0.13

0.004 ± 0.001

77.3 ± 1.4

B-21

S-15

27.4 ± 2.4

0.42 ± 0.31

0.015 ± 0.012

15.4 ±0.4

B-22

S-11

52.1 ± 2.2

0.27 ± 0.07

0.005 ± 0.001

33.0 ± 1.1

B-23

S-17

6.39 ±0.36

0.199 ±0.043

0.031 ± 0.008

3.06 ± 0.17

B-24

S-10

43.0 ± 3.4

0.21 ±0.07

0.005 ± 0.002

28.9 ± 0.5

B-25

S-13

9.27 ± 0.45

0.134 ±0.048

0.014 ± 0.004

5.87 ± 0.28

B-26

S-6

10.6 ± 0.5

0.218 ± 0.052

0.020 ± 0.006

4.34 ± 0.42

B-27

S-4

8.52 ± 0.41

0.073 ± 0.021

0.009 ± 0.003

3.28 ± 0.18

B-30

S-3

42.5 ±2.1

1.59 ±0.24

0.037 ± 0.006

6.77 ±0.26

B-30

S-2

38.4 ± 2.1

1.21 ± 0.20

0.032 ± 0.006

5.65 ±0.24

C-1

S-30(K)t

71.6 ±3.8

4.16 ± 0.54

0.058 ± 0.008

37.6 ± 1.9

C-4

S-29

40.3 ± 1.6

1.72 ± 0.17

0.043 ± 0.004

24.1 ±0.9

C-5

S-28t

45.7 X 2.2

0.67 ± 0.23

0.015 ± 0.004

24.7 ± 0.8

C-6

S-33

46.6 ± 1.4

0.33 ± 0.14

0.007 ± 0.002

33.5 ± 1.0

C-8

S-32

43.3 ± 2.2

0.69 ± 0.27

0.016 ± 0.006

31.7 ± 2.0

C-8

S-31

41.5 ± 1.9

0.550 ± 0.097

0.013 ± 0.002

28.8 ± 1.5

C-10

S-34

12.5 ±0.4

C-11

S-19

28.9 ± 1.1

15.0 ± 0.9

0.518 ± 0.002

3.46 ± 0.41

C-11

S-16

5.36 ± 0.36

1.32 ± 0.19

0.246 ± 0.030

0.71 ± 0.11

S-26

2.20 ± 0.19

0.206 ± 0.053

0.093 ± 0.022

0.72 ± 0.11

D-8

S-27

30.6 ± 4.5

0.51 ±0.23

0.017 ±0.008

*Mean ± 2 standard deviations.

t Surface sediment C-1 S-30 (n) was not analyzed for plutonium. Sediment C-1 S-30 (K) is a
portion of the remaining grab sample.

:j:Contained sediment from upper several centimeters (see Table 4).

TRANSURANIC ELEMENTS AT BIKINI ATOLL 551

B-24, which were located to the south and east of the area of the highest radionucHde
concentrations measured at stations B-2 and B-20. Station B-21 was located at the
extreme southern end of the region of high radionuclide concentration, and the fmely
divided sediments found there were similar in appearance to those collected at stations
B-2 and B-20. Even thougli both stations B-21 and B-20 had similar proportions of fine
sediments, only about 23% of the 2 3 9,2 4 op^ measured at station B-20 was found at
station B-21. In contrast, stations B-22 and B-24, which were located some distance
downstream and to the east of the area of high 2 3 9,24 0p^ concentrations (i.e., stations
B-2 and B-20), contained low proportions of finely divided material (less than ~15%) but
contained 43% and 39%, respectively, of the total 2 3 9,2 4 Op^^ Pleasured at station B-20.
These observations can be explained by two processes: the first is by dilu-
tion of the concentration of radioactive particles deposited at station B-21 by
material of a lower concentration (resulting from biological activity or erosion of the
reef); the second is by physical or chemical fractionation of the radioactivity in, or from,
debris particles that are transported in suspension. (Physical fractionation could arise
from differences in the concentrations and rates of dissolution of different-size particles.)
The plutonium concentrations of relatively larger size particles deposited at station B-21,
for instance, may have been lower than those of smaller size particles deposited farther
downstream at stations B-22 and B-24. Second, since chemical fractionation of the
radionuclides may be a function of the length of time the particles remained in
suspension, particles deposited at station B-21 may also have lost a higher proportion of
their surface-associated radioactivity than those deposited at stations B-2 and B-20. The
relatively high concentrations of the sediments collected at stations B-22 and B-24 would
be consistent with the deposition of finely divided material of a high specific activity.

Distribution in Sediment Cores

Measurements of the concentration distribution of elements in the sediment column are
fundamental to the study of the exchange of materials across the sediment- water
interface. In the Bikini lagoon measurements of both the transuranic and fission-product
radionuclides with depth were considered to be particularly informative since debris from
several detonations have been added to the lagoon at different times.

Nine sediment cores were collected from various locations in the lagoon. Three types
of profiles of the radionuclide concentration with depth were observed. These occurred in
(1) crater sediments (stations C-3 and C-12), which had either relatively homogeneous or
constant distributions of most radionuclides with depth; (2) northwest quadrant lagoon
sediments (stations B-2, B-20, and B-21), which had large proportions of finely pulverized
material and which had radionuclide concentrations that changed regularly with depth;
and (3) central and eastern lagoon sediments (stations B-15, B-16, B-27, and B-30), which
had variable radionuclide concentrations with depth (Marshall, 1975).

Crater Cores

The distribution of radionuclides measured in the sediment core collected from the center
of Zuni crater (station C-12) showed approximately constant transuranic and fission-
product concentrations with depth (Table 3). No appreciable portion of the sediments in
the Zuni crater core was fmely pulverized. A unique concentration sequence of the
following order was found: 2 39,240p^ ^ ^^s^^ ^ 2 38p^ >^°Co > ' ^'Cs > ^"^^ Am >
2^^Bi.

552 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Distribution of ^"^ ' Am, ^^^ '^'^^Pu, "^Pu, and ^ ' ^Po and the
2 3 8 p^,/2 3 9 ,2 4 0 py j^j j^^ ^jjj^ p^p^j^ j^^ Scdimcnt Cores

Core deptli

Concentration,* pCi/g (dry weight)

^^' Am

2 3 9 ,2 4 0 py

2 3 8pu

238py|239,240pj,

2.0po

Core C-I2

Oto 2

5.7 + 1.5

35.5 ± 1.9

19.0 ± 1.1

0.56 ± 0.04

0.860 ± 0.054

2 to 4

6.33 ± 0.70

4 to 6

6.04 ± 0.87

6 to 8

6.27 + 1.0

38.8 ± 1.4

18.8 ± 0.8

0.48 ± 0.03

8 to 10

6.04 ± 0.97

lOto 12

6.4 ± 1.1

12 to 15

6.12 ±0.61

36.4 ± 1.4

18.6 ± 0.8
Core C-3

0.51 ± 0.03

Oto 2

29.8 ± 1.0

49.4 ±4.0

2.98 ± 0.62

0.063 ±0.013

1.18 ±0.13

2 to 4

34.4 ± 2.1

4 to 6

43.2 ± 2.2

6 to 8

67.7 ± 2.4

98.7 ±5.6

8.27 ± 0.96

0.083 ± 0.011

2.99 + 0.16

8 to 10

26.1 ± 0.8

lOto 12

17.5 ±0.9

26 to 28

13.6 ±0.5

28 to 30

13.6 ±0.6

30 to 32

8.32 ±0.28

6.24 ± 0.46

0.163 ±0.064

0.026 ± 0.010

0.712 ± 0.045

32 to 34

1.44 ±0.06

48 to 50

19.0 ±0.6

50to52

23.1 ±0.5

35.4 ± 2.2

1.28 ±0.26

0.036 ± 0.007

1.08 ± 0.07

54 to 56

26.1 ± 0.9

Core B-2

Oto 2

103.0 ± 1.0

107 0 ±4.0

1.30 ± 0.22

0.012 ± 0.002

Lost

2 to 4

85.7 ± 1.4

4 to 6

80.2 ± 0.8

102.0 ±4.0

1.32 ±0.22

0.013 ± 0.002

0.677 ± 0.072

6 to 8

75.4 ± 1.1

97.2 ±3.6

1.28 ±0.22

0.013 ±0.002

0.572 + 0.072

6 to 10

56.2 ± 1.0

10 to 12

48.7 ± 1.6

16.4 ± 1.5

0.21 ±0.14

0.013 ± 0.008

0.505 ± 0.032

12 to 14

20.2 ± 0.4

14 to 16

14.4 ±0.5

16 to 18

8.27 ±0.21

18 to 20

5.17 ± 0.27

6.76 ±0.58

0.320 ± 0.088

0.047 ±0.013

0.194 ±0.018

20 to 22

2.66 ± 0.20

22 to 24

1.65 ± 0.16

24 to 26

0.73 ± 0.14

0.919 ± 0.028

0.01 ± 0.002

0.011 ± 0.002

Lost

26 to 28

0.178 ± 0.073

28 to 30

0.24 ±0.13

30 to 32

0.124 + 0.084

32 to 34

0.203 ± 0.058

34 to 36

0.165 ±0.008

0.002 ±0.001

0.010 ± 0.007

0.032 ± 0.018

36 to 38

0.070 ±0.050

38 to 40

0.92 ± 0.18

1.63 ± 0.08

0.011 ±0.007

0.007 ± 0.004

0.090 ± 0.010

TRANS URANIC ELEMENTS AT BIKINI ATOLL 553
TABLE 3 (Continued)

Core depth,

Concentration,* pCi/g (dry weight)

^"'Am

2 3 9,240py

238py

2 3 8py/2 3 9,240py

2.0po

Core B-20

Oto 2

81.7+ 1.6

101.0 ± 3.0

0.303 ± 0.088

0.003 ± 0.001

1.88 ± 0.10

2 to 4

57.4 ± 1.7

4 to 6

53.3 ± 1.2

6 to 8

61.3 ± 1.8

71.8 ±4.6

0.41 ± 0.22

0.006 ± 0.003

1.72 ± 0.09

8 to 11

40.1 ± 1.3

11 to 12

25.8 ± 1.0

34.7 ± 1.4

0.126 ±0.060
Cote B-21

0.004 ± 0.002

0.901 ± 0.054

Oto 2

16.7 ±0.5

23.6 ±2.1

0.58 ± 0.20

0.024 ± 0.009

0.473 ± 0.036

2 to 4

19.0 ±0.4

4 to 6

18.5 ±0.7

29.8 ± 1.0

0.604 ± 0.062

0.020 ± 0.002

6 to 8

13.2 ±0.8

8 to 10

12.8 ±0.7

19.1 ± 1.2

0.38 ± 0.10

0.020 ± 0.005

lOto 12

4.34 ± 0.20

12 to 14

2.47 ± 0.22

4.34 ± 0.38

0.075 ± 0.032

0.017 ± 0.008

0.225 ± 0.036

14 to 16

7.02 ± 0.35

Core B-15

Oto 2

2.86 ±0.19

3.73 ±0.48

0.067 ±0.011

<0.034

2.21 ± 0.11

2 to 4

1.72 ±0.15

4 to 6

1.80 ± 0.21

6 to 8

1.75 ±0.25

4.19 ± 0.20

0.085 ± 0.022

0.020 ± 0.005

8 to 10

2.56 ±0.27

lOto 12

3.75 ±0.31

3.32 ± 0.18

0.056 ± 0.012

0.017 ±0.004

12 to 14

2.89 ± 0.49

14 to 16

2.74 ± 0.29

6.99 ± 0.34

0.130 ±0.036

0.018 ± 0.005

*Mean ± 2 standard deviations.

A long core (56 cm) of entirely pulverized sediment was collected from the center of
the Bravo crater (station C-3). Three segments of this core (the 0- tal2-, 26- to 34-, and
48- to 56-cm regions) were cut into 2-cm sections for the radionuchde measurements. The
concentrations of radionuclides (Fig. 8 and Table 3) measured in the two lower regions of
the core were similar to the uniform concentrations measured in the Zuni crater core. In
the surface 12 cm, however, a well-defined layer of high-radionucUde concentrations was
centered at the 6- to 8-cm depth. Elevated concentrations of all radionuclides were
measured in this section, which contained the highest concentrations of ^^^Pu
(8.3 pCi/g), ^^^Bi (432 pCi/g), and ^^Co (306 pCi/g) measured in any Bikini sediments
except for the one higher ^^^Pu concentration (19.0 pCi/g) that was found in Zuni crater
(station C-12) sediments. The ordering sequence of radionuclide concentrations in
different regions in the core differed greatly. This ordering sequence can be compared
with that for surface sediments shown in Fig. 9. The sequence in the 0- to 2-cm section of
the core differed from that in lower sections and from that found in the three other
surface grab samples collected across the crater; these grab samples also differed from
each other. In the 2- to 12-cm region of the core, the order in the sections was Bi > Co >

554 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

X ^°Co X 0.2

O I37cs

A 2°"'Bi X 0.2

241

239,240

40 60 80

ACTIVITY, pCi/g

100

120

Fig. 8 Distribution of radionuclides in the sediment core collected at station C-3.

Eu > Am > Cs. In the 26- to 34-cm region of the core, three or four sections had the
order Co > Bi > Eu > Am > Cs, and in the 48- to 56-cm region of the core, the sequence
was Co > Am > Eu > Bi > Cs or Co > Am > Bi > Eu > Cs. Since plutonium was
measured in only one section from each region, its placement is not included as
characteristic of the larger regions. The higli 2 3 8p^^2 3 9,24 0p^ ratios found in the upper
12 cm of this core, however, showed that the origin of these radionuchdes was different
from those found in the lower regions. The radionuclides measured in the uppermost
sections may be remnants from one of the smaller post-Bravo tests conducted in this area.
The 2 3 8py|2 3 9,24 0p^ ratios of 0.036 and 0.026 fouud in the two deeper segments of the
core were similar to several other ratios found in surface sediments of the lagoon that
were collected away from the region south of tlie Tewa crater (station C-8).

Northwest Lagoon Cores

Three sediment cores were collected from this region of the lagoon (stations B-2, B-20,
and B-21). Pulverized sediments were found at all three stations, although the
radionuclide concentrations were significantly lower at station B-21 than at stations B-2
and B-20.

TRANSURANIC ELEMENTS AT BIKINI ATOLL 555

6i > Pu - Eu > Co > Am > Cs

A Pu Am ' Eu
• Eu ■ Pu ■ Am
■ Pu > Eu ;. Am

\0.039' VSl^~i'~

Co > Bi > Cs

Pu Eu - Co ■ Cs - Am > B

Co ■ Pu > Eu > Am > Bi > Cs

Fig. 9 Distribution of the ordering sequence of radionuclide concentrations in the
surface sediments of Bikini AtoU lagoon. The ^ ' * Pu/^ 3 « .2 4 o p^ ratios at each station for
the year 1972 are also shown.

Station B-2. The distribution of "^'^''^Pu, ^^ ^ Am, ^^''Bi/ ^^Eu, ' ^ v^.^^ ^^ eo^.^ ^
the sediment core collected at station B-2 is shown in Fig. 10, and that of ^"^^Am,
2 3 9,2 4 0p|j^ 2 3 8py^ ^^^ ^^ ^Po is shown in Table 3. Several features of this long core were
similar to features in other sediment cores collected from the northwest quadrant. The
sediments in this core consisted of mixtures of Halimeda and pulverized fine coral
material; Halimeda predominated at the 8- to 10-cm section. The distribution of ^^ ^ Am,
2 3 9,240p^^ 1 5 5£u^ ^^^ ^ '^ ''Cs Concentrations at station B-2 was similar to that at station
B-20 except that (1) the absolute concentrations of ' ^ ^Cs measured were lower than the
^"^^Am and '^^Eu concentrations by a factor of 10 and (2) the 2 3 9,240p^ concentra-
tions measured were slightly more irregular with depth than the concentrations measured
for ^^' Am or ^^^Eu. In the top 11 cm of the core, the concentrations of ^^^'■^^^Pu,
^^*Am, '^"^Eu, and ^^''Cs decreased regularly with depth to 50% of their respective
concentrations, which were measured in the surface layer. In the 12- to 26-cm region of
the core, the concentrations of ^^^'^"^^Pu, ^'^^ Am, ^ ^^Eu, and ^^''Cs decreased nearly
logarithmically with depth. In the 28- to 38-cm region of the core, the concentrations
again decreased slowly with depth.

The distribution of ^^Co and ^^^Bi concentrations in the core was unusual in that
decreasing concentrations (with increasing depth) were not found in the upper 10 cm of
the sediment core. Although the concentration of ^^Co was relatively constant in the
upper 12 cm of the core, the concentration of ^^^Bi increased 50% between the 2- to 4-
and 8- to 10-cm sections. Below the 8- to 10-cm section in the core, the decrease in
concentrafion of ^'^'^Bi was similar to that of ^"^^ Am, ^ ^^Eu, and ^^''Cs; however, the
concentration of ^°Co was almost constant with depth. The distribution of ^^^Pu/
2 3 9,2 4 Op^ ratios measured in different sections of the core was separated by the value of

556 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

0.05 0.1

0.3 0.5

1 3 5

ACTIVITY, pCi/g

30 50

Fig. 10 Distribution of radionuclides in the sediment core collected at station B-2 in
1972.

0.047, which was found in the 18- to 20-cm section. Below this 18- to 20-cm section, the
ratios decreased with depth from 0.0108 to 0.0069, whereas above the 18- to 20-cm
section, the ratios ranged from 0.0138 to 0.01 13.

The ordering sequence of the radionuclide concentrations measured in this core varied
with depth. Below the 0- to 2-cm section, the order changed from that shown in Fig. 9 to
Pu > Eu > Am > Co > Cs > Bi in the 2- to 6-cm region, to Pu > Eu > Am > Co > Bi >
Cs in the 6- to 26-cm region, to Co > Pu > Eu > Am > Bi > Cs in the 26- to 38-cm
region, and to Pu > Eu > Am > Co > Bi > Cs in the 38- to 40-cm region.

Station B-20. The distribution of ^^^•^'^^Pu, ^^^ Am, and ' ^^Eu concentrations in this
core (Fig. 1 1 and Table 3) was again quite similar and decreased with depth by 50% at
about the 9-cm section. At 1 1 cm in the core, a sharp break occurred between the finely
divided material in overlying sections to coarse sand. Considering the range and
distribution of 2 3 8p^j^2 3 9,2 4 0pu j-^i^jq^ measured in surface sediments across the lagoon,
the ratios found in the three sections of this core were uniquely low, which possibly
indicates a common sourcef"^) for the majority of the plutonium contamination in the
sediment column collected at this station. Only in the 6- to 8-cm section of this core does
the radionuclide ordering sequence differ from that found in surface sediments (Fig. 9).
In this section the ordering of ^"^ ' Am and ^ ^ ^ Eu concentrations was reversed from those
in other sections. This order, Pu > Am > Eu, was found in sediments only in the four
sections from the bottom of the Bravo crater core and from the eastern lagoon.

Bismuth-207 concentrations were below the limit of detection in most sections of the
core. However, the concentration of ■^°^Bi in the 0- to 2-cm section was at least four to
five times as high as that in any lower section.

The concentrations of ^ ^Co and ' ^ ''Cs decreased, respectively, to 50% of their largest
concentration at the 9- and 11 -cm levels in the core. However, neither of these

TRANSURANIC ELEMENTS AT BIKINI ATOLL 551

60co X 2

2°^Bi X 200

•

241 Am

□

155eu

239, 240 r

60 80

ACTIVITY, pCi/g

Fig. 1 1 Distribution of radionuclides in the sediment core collected at station B-20 in

1972.

radionuclides showed steadily decreasing concentrations in the upper layers. The
concentration of ^*^Co in the 0- to 2-cm section of the core was significantly lower than
that in lower sections. The concentration of '^^Cs at the 0- to 9-cm level of the core
showed no appreciable change with depth.

Station B-21. The concentration profiles of all the radionuclides measured in this
sediment core are rouglily similar to those in station B-20 in that the concentrations
increased to a maximum at between 5 and 7 cm (except at 3 to 5 cm for ^'*' Am) and
then decreased to 50% of their highest measured concentrations at depths of 10 cm, after
which increasing proportions of Halimeda began to appear. As in the B-2 core, increased
concentrations of radionuclide distributions were measured in the lowest section of this
core. As in both of the other cores from this region of the lagoon, the 2 3 8p^^23 9,2 40p^j
ratios measured with depth in the core (Table 3) showed only a slight decrease with
depth.

In the 8- to 16-cm region of the core, the ordering sequence of radionuclide
concentrations changed from those found in surface sediments (Fig. 9) to the order Pu >
Eu > Am > Co > Bi > Cs. This sequence was the same as that observed below the 6-cm
section at station B-2 and in surface sediments at the far western region of the Atoll.

358 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Central and Eastern Lagoon Cores

The four sediment cores collected from the central and eastern regions of the Atoll
(stations B-27, B-16, B-15, and B-30) were similar in three respects: First, there was no
significant net increase or decrease in the concentration of radionuclides measured
between the upper and lower sections in any of these four cores; second, the distribution
profiles of 2^9'2^0pu, ^'^'Am, '^^Eu, ^^^Cs, ^^'^Bi, and ^'^Co concentrations were
roughly similar with depth in the individual cores; and third, the 2 3 8p|j^2 3 9,2 4 0p^ ratios
measured in all but the lower section of the cores from station B-27 were similar to that
in the surface-sediment section. Because of the very short (6 cm) length of the cores at
stations B-16 and B-30, no further interpretation of the observed radionuclide profiles
was warranted. Except for the ordering of '^^'Am in one section and the 2 3 9,2 4 op^^
concent! ation in the 10- to 12-cm section immediately below, the ordering sequence of
radionuclide concentrations in the 16-cm core from station B-15 was the same as that in
the surface sediments (Fig. 9). In the 10-cm core collected from station B-27, the
sequence of ^^Co, ^°^Bi, and ^^^Cs concentrations measured did not change with depth
from that shown in Fig. 9. However, in the 4- to 8-cm region, '^^ ' Am concentration was
higher than ^^^'^^*^Pu, and in the 8- to 10-cm section, the ordering sequence was the
same as that at lagoon stations B-16, B-26, and B-22 to the west.

The constancy of the concentrations of all radionuclides measured to depths of 10 cm
(core B-27) and 16 cm (core B-15, Table 3) showed that a considerable penetration of
radionuclides occurred in these sediments, which appeared physically to be normal lagoon
deposits. Assuming a negligible natural sedimentation rate, the penetration of radio-
nuchdes into these sediments is significantly greater than that observed by Held (E. Held,
University of Washington, unpubUshed results) in Rongelap Atoll sediments. However,
these two sediment cores were the longest obtained from any station in the Atoll having
unpulverized sediments, which suggests that these sediments may have been significantly
less consolidated than average. This could explain both the length of the core collected
and the radionuclide concentrations measured with depth.

Sedimentation Rates at Station B-2

Measurements of the concentrations of ^^°Pb and ^^^Ra with depth in core B-2 were
used to determine the effective sedimentation rates based on tlie ^"'Pb age dating
teclinique (Goldberg, 1963; Koide, Soutar, and Goldberg, 1971). The average ^^^Ra
concentration of 0.131 pCi/g, which was measured by gamma counting, was used to
determine the concentrations of unsupported ^'°Pb. Tlie unsupported ^^°Pb concentra--
tions measured in the 0- to 2-, 4- to 6-, 6- to 8-, and 10- to 12-cm sections decreased
logarithmically with depth, which indicates a constant sedimentation rate for the upper
layers. Below 12 cm the ^^°Pb concentrations were not significantly different from the
^^^Ra concentrations measured, which indicates no unsupported '^^^Pb. The effective
sedimentation rate was determined by calculating a Hnear regression of the unsupported
^^*^Pb concentrations in the upper 1 1 cm of sediment. A sedimentation rate of 0.58
cm/yr (correlation 0.98) was calculated for the upper 11 cm of sediment. Thus the
approximate date calculated for the deposition of the 11-cm section was 1953, a date
consistent with the period of nuclear testing at Bikini.

These data indicate that two different processes were responsible for the deposition
of the 40 cm of sediment sampled at this station (B-2): (1) slow accumulation of
sediment occurred in the upper layers (1 1 cm) and (2) at some point below 1 1 cm in the

TRANSURANIC ELEMENTS AT BIKINI ATOLL 559

core, rapid accumulation of the sediments containing no unsupported ^'°Pb pre-
dominated. Figure 10 shows that both ^°Co and ^""^Bi have concentration profiles that
are markedly different from those of other radionuclides in the core above and below
about 11 cm. This may indicate that not only the process of deposition but also the
source of contaminated debris may have differed for the two depth regions in the
sediment column at station B-2.

Given the dynamic hydrological environment at Bikini, the most significant
contamination of the sediment environment a priori would arise from the large surface
bursts (such as Bravo, Koon, and Zuni) whose fireballs strongly interacted with the soil or
sediment and from similar interactions of deep lagoon or barge bursts (such as the Baker
and Tewa tests). At Bikini the initial introduction of highly contaminated debris to the
lagoon from detonations o{ this type can be described as fallout deposition of a large
mass of chemically altered coralline soils reduced in size and containing the condensed
radionuclides. A large mass of crushed coralline material of a relatively low specific
activity must also have been ejected by the detonations. The areal distribution of two
different types of materials (altered and unaltered coral) would overlap at progressively
greater distances away from the detonation craters, and mixed particles would descend at
rates depending on their sizes and shapes. In the aqueous environment the particles would
be transported a distance that would be determined by their settling velocities, sizes,
densities, and the speed of the prevailing lagoon currents. The result of these physical
forces with time would be to yield a concentration of fine particles in the surface
deposits. The net result of the radionuclide concentrations would be to yield sediment
concentrations (picocuries per gram) that were progressively more dilute (by natural
sediments) at increasing distances downstream.

The sedimentation rate measured in the upper 1 1 cm of sediment collected at the
station near the Bravo crater (station B-2) showed that the material was deposited at a
constant rate between the 1950s and 1972. Although the initial source for the material
deposited at these locations was the detonation craters, the present location of the
source(s) supplying the material for redistribution at these lagoon stations is not known.
The importance of this point should not be underestimated because the location and
extent of the source of these fine sediments may determine the continued availability of
the radionuclides for redistribution and uptake by biota.

It is clear, from the large size of the Bravo, Tewa, and Zuni detonation craters, that a
huge quantity of pulverized sediment was removed from the reef immediately after the
detonations. However, as noted by Welander et al. (1966), lagoon currents were capable of
maintaining a large flow of the finely divided sediment out of certain craters at Enewetak
long (>1 yr) after the testing stopped. It is quite likely that much of this material at
Bikini was deposited outside the detonation craters and was the source for part of the
material redistributed in the lagoon. The ^^^Pu/'^'^^''^'*°Pu ratios measured in the craters
and at various stations in the northwest quadrant suggest three possibihties for the source
of the redistributed material deposited at station B-2: (1) from locations between station
B-2 and the Bravo crater; (2) from (1) above and from the area between station B-2 and
the northern reef (near station B-19); or (3) from (1) or (2) above and from within the
detonation craters. The reason for making these hypotheses is that the 2 3 8py^2 3 9,2 4 0p^
ratios in the top 1 1 cm of redistributed sediments at station B-2 are about 0.0125,
whereas the ratios measured in the fine surface sediments collected in Bravo crater are
about 0.05, and the ratios at station B-19 are about 0.008. Thus a mixture of sediments
from different sources may be deposited at station B-2.

560 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Concentrations and Physicochemical States in the Water Column

At each of the sediment sampUng stations, piutonium and americium were measured in
water collected at 2 m from the bottom and at 1 to 2 m from the surface. During the
collections in November 1972, the wind and wave conditions were typical of the winter
season — ESE and ENE as given in Fig. 4 (Noshkin et al., 1974). During the collections in
July 1976, the wind direction and velocity were essentially the same as those in
November 1972. The physical circulation of the water in the lagoon may be traced by
using the piutonium concentrations and distributions measured at lagoon stations away
from the northwestern quadrant.

Distribution of'^^^''^^^Pu Concentrations

With the use of all total piutonium concentration values obtained in 1972 by F. G.
Lowman, Puerto Rico Nuclear Center, V. Noshkin, Lawrence Livermore Laboratory, and
W. R. Schell, University of Washington (unpublished data) and the few additional values
obtained in 1976 for the lagoon water and by averaging the values when duplicate
collections were available for the same station, isopleths of the piutonium concentrations
for the surface and deep water of the lagoon have been constructed and are shown in
Figs. 12 and 13, respectively. These isopleths show clearly the distribution patterns of
piutonium, caused by the lagoon circulation, from its main source in the sediments of the
northwestern quadrant of the lagoon. With the use of 2 3 9,24 0pjj concentrations as the
tracer, the transport and circulation of the water in the lagoon have been estimated. The
surface water appears to be diluted by incoming ocean water through the wide pass near
Eneu and by oceanic water over the northeastern reef. The outlet of lagoon water is

Kilometers

Fig. 12 Distribution of 2 3 9,24opy concentrations in surface water (2 m) at Bikini
Atoll. Concentrations in picocuries per cubic meter.

TRANS URANIC ELEMENTS AT BIKINI ATOLL 561

103121 62 104
STATIONS

Kilometers

Fig. 13 Distribution of 2 ^ 9 ,2 4 o p^ concentrations in deep water (2 m above bottom) at
Bikini Atoll. Concentrations in picocuries per cubic meter.

through the deep passes in the southwestern part of the Atoll. It is here that lagoon water
exits into the North Equatorial Current. The pattern for the transport of deep water
appears to be that of oceanic water entering the lagoon through the Eneu passage and
moving as far as the deep passes at the southwestern part of the lagoon. Oceanic water
also either enters througli or over the reef at the northwestern part of the lagoon and
dilutes the deep water in this region. As shown in Fig. 6, the higliest 239,240pjj
concentrations in sediments are near the northwest reef. The highest 2 3 9,2 4 Op^^
concentrations in the deep water appear to be displaced southward slightly from the
higher concentrations that were measured in the sediments and to be transported in an
easterly direction. For such a complex water circulation pattern in the lagoon, water must
pile up and descend near the western reef and upwell near the eastern reef. This process
was identified by Von Arx (1954). Water must flow in opposite directions at the
southcentral part of the lagoon. The tongue of oceanic water at the bottom in the
southern part of the lagoon has a lower ^^^'^'^^Pu concentration than the surface water
that is leaving the lagoon through the deep passes. The same general circulation pattern
was inferred from ^ ^ Fe measurements in the same samples except that the source was
more concentrated near the northern reef (Schell, 1976).

The distribution of plutonium in the lagoon water indicates that the near-reef areas
contain low concentrations of plutonium; this is probably due to dilution by oceanic
water that enters the lagoon from over the reef. Thus the organisms that inhabit the
lagoon terrace or seaward reef areas are exposed to very low levels of plutonium even
though much higher levels exist inside the lagoon. The predominant source of fish for the
returning Marshallese will be those reef fish, which should contain low levels of
transuranic radionuclides. The radionuclide levels are probably not much higher than

562 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

those found at Atolls that have not been contaminated by fallout (Noshkin, Eagle, and
Wong, 1976).

Physicochemical States of ^^^^^"^^Pu and ^^^Pu

In a model of the marine environment that can be used to predict effects, the water is the
central component; the water receives the inputs of metals and provides the transport
medium for uptake into the biosphere or loss to the sediments. Because of this central
role, the behavior and physicochemical states of the transuranic elements in seawater
must be known if a useful model of the system is to be constructed. Once the speciation
and behavior of these elements are known, their relative hazards can be predicted. Higli
concentrations of salts in seawater interfere with the methods used to measure the
physicochemical states and must be removed before the analysis for the transuranic
elements. Consequently separation and preconcentration from the saltwater matrix are
required at some stage of the analysis. Ideally the best method would be direct in situ
measurement of the elements in the field. Unfortunately, with present technology direct
field measurement is not feasible. The next best thing is to extract and concentrate the
elements in the field. This would eliminate some of the problems resulting from
contamination, losses during transport and storage of samples, and changes in chemical
speciation on storage in containers.

The Battelle large volume water sampler (BLVWS) is a sampling system that can be
used to concentrate low levels of trace metals or radionuclides from natural waters in the
field (Silker, Perkins, and Rieck, 1971). This method effectively eliminates the need for
preservation and storage of water samples, extraction in the field, and evaporation or ion
exchange in the laboratory to concentrate the elements to a level sufficient for analysis.
The advantage of the BLVWS technique over the "conventional" techniques is that both
the total concentration and the physicochemical-state concentrations of the particulate
and soluble fractions can be measured. Collection efficiencies of the soluble fraction are
determined individually for each element during collection. In addition, much larger
volumes of water permit lower concentrations to be measured.

An evaluation of a new sampling and measurement technique for transuranic elements
requires detailed studies of the precision and accuracy of the technique in both controlled
and natural environments. It also requires simultaneous measurements of samples that
have been collected by the more conventional methods. For the past several years, we
have attempted to set up experiments that would test the validity of the BLVWS
technique for plutonium measurements in both laboratory and field studies (Huntamer,
1976;Nevissi and Schell, 1975;Schell, Nevissi, and Huntamer, 1978).

Description of the Sampler. The BLVWS is a field collector that can process as much as
4000 liters of water in 3 hr with the large sampler (28-cm diameter) and about 800 liters
of water with the small sampler (13-cm diameter), depending on the particulate loading.
The filtering section of the BLVWS normally consists of eight filters arranged in parallel.
The number of filters used can be expanded or reduced by removing or adding plates to
the BLVWS. The water, after passing through one of the filters, is then channeled througli
the sorption beds. The sorption beds generally consist of two to four 0.6-cm-thick
sections (Fig. 14). The use of individual sorption beds rather than one thick bed permits
the calculation of the collection efficiency for individual elements and permits easy
variation of the sorption-bed thickness. It also allows for the use of a mixture of different
sorption beds if desired.

TRANS URANIC ELEMENTS AT BIKINI ATOLL 563

INLET

7777777MM77777777777777777777777777M,

' " " ' ' ^ ' ■ ■ ■ ■ ■■■■.,...■.,,'

m/mmmmm/mm///

'/////////////////////////////////////////// ^

0.3-pm MILLIPORE FILTER

FIRST BED

SECOND BED

THIRD BED

FOURTH BED

OUTLET

Fig. 14 Schematic representation of the BLVWS showing the flow of water through the
filters and sorption beds.

Water is forced through the BLVWS with an electric pump. The smaller BLVWS
requires a lower flow rate than the larger samples; so the flow is controlled by the use of a
valve and a water "bypass;" The entire pumping system is shown schematically in Fig. 15.
The volume of the water sampled is measured with a recording water meter.

Collection Efficiencies. Most of the sampling procedures and techniques used in
seawater analysis assume 100% efficiency for the collection and measurement or use
radioactive tracers to determine the chemical yield of the samples. One problem with the
use of tracers is that they are usually not added to the samples in the same chemical form
as the element in the sample. If the chemical states of the sample and tracer element are
not identical, the isotopic equilibrium might not be reached for a long time, and the
chemical yield may be in error.

Because of the short column lengths used in the BLVWS, collection efficiencies
seldom are 100% except for a few elements that are totally retained on the first sorption
beds, such as ^"^^Am, ^^^Bi, ^^Fe, and ^^^Eu (Schell, Nevissi, and Huntamer, 1978).
Elements that are not collected quantitatively can be determined by the differences in the
amounts collected on successive sorption beds, as outlined by Held (1971), Schell, Jokela,
and Eagle (1973), and Schell, Nevissi. and Huntamer (1978).

This method of determining collection efficiencies, referred to as the "BLVWS
ichnique," is an empirical method which assumes that a constant fraction of the

564 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

POLYETHYLENE PIPE

PUMP

[jllllllH WATER BYPASS

PLASTIC SCREEN

Fig. 15 Schematic representation of the BLVWS.

available solute is removed by each sorption bed. When this approach is used to measure
the amount of solute, N, on the individual sorption beds, N^ and N^+i , the collection
efficiency, E, between beds m and m+ 1 can be determined from Eq. 1 (Schell, Nevissi,
and Huntamer, 1978).

Nm -Nm+i

^(m, m + i )

this efficiency can then be used to obtain the concentration of the solute in the soluble
phase, Cs-

E(m, m + i)

m-1
m

m- 1

(2)

The total concentration, Ct, is found by adding the contribution of the particulates,
Cp, to the soluble.

Ct = Cs + Cp (3)

where E = collection efficiency between m and m + 1 beds

Njy, = concentration of solute retained on the mth sorption bed
Nm+i = concentration of solute retained on tire m + 1 sorption bed
Cs = concentration of solute in the soluble fraction of the water
Cp = concentration of solute in the particulate fraction retained on the filters
Ct = total concentration of solute in the water volume sampled

Tank Experiments with the BLVWS. The BLVWS has been evaluated at different
salinities in the laboratory by a series of tank experiments. The experimental procedures
are discussed in detail in an M.S. thesis by Huntamer (1976) and in papers by Schell,
Nevissi, and Huntamer (1978) and Nevissi and Schell (1976). Six elements, americium.

TRANS URANIC ELEMENTS AT BIKINI ATOLL 565

bismuth, cobalt, europium, iron, and plutonium, have been evaluated in the tank
experiments by using two different sorption beds, AI2O3 and Chelex-100. These elements
are of interest in that they are either products of nuclear weapons testing or nuclear
reactor operations.

The results show that the BLVWS can be used to measure the concentration of the
above elements in marine waters over a wide range of salinities. The precision of the
BLVWS method has been evaluated by comparing duphcate samples taken with the
BLVWS. For example, the mean percent variation between duplicate samples (collected
simultaneously from the same tank) in experiments using AI2O3 was ±10% for
plutonium.

The results of the tank experiments indicate that, with the use of either AI2O3 or
Chelex-100 sorption beds, the BLVWS is a suitable sampling method for some elements.
In addition, the behavior of individual elements on the sorption beds provides qualitative
information on the physicochemical speciation of the elements. Table 4 gives a summary
of the physicochemical states observed at high salinity (>31 %©) with the BLVWS in
both the tank and field studies at Bikini lagoon. The results are compared with the
chemical species predicted by equilibrium calculations or determined by the other
measurement methods. Americium was found to be 80 to 100% particulate (i.e.,
>0.3 jum) and consequently was collected efficiently on the Millipore filters and/or first
AI2O3 bed at all salinities. Plutonium was collected by the BLVWS technique, using

TABLE 4 Physicochemical States of the Trace Metals in Marine Waters Estimated

by the BLVWS Method Compared with the Physicochemical States

Measured and Predicted with Other Methods

Elements

Samples

Particulate, %

CoUoidal,* %

Soluble, %

Predicted and observed species

Americium

Tank

76 to 87

13 to 24

Particulates

Field

30 to 100

Oto70

Bismuth

Tank

69 to 90

10 to 31

Bi(OH)= + , Bi(OH)f + ,t insoluble

Field

0to22

78 to 100

alkaline solution J

Cobalt

Tank

0to2

98 to 100

CoCl"^,Co^ + ,(CoSO°),§^

Field

organics**

Europium

Tank

29 to 71

Particulates (freshwater)tt

Field

11 to 100

0to89

Iron

Tank

72 to 95

2 to 28

Fe(0H)+,Fe(0H)7,§

Field

31 to 60

40 to 69

particulates f

Polonium

Tank

100

Particulates:!:

Field

95 to 100

0to5

Plutonium

Tank

69 to 93

4 to 28

CoUoidal.tJPuOjCCOj)^-,

Field

2 to 60

40 to 98

Pu3 + ,PujOH+t$

*Colloidal species based on the complete retention of the fraction passing through a 0.3-Mm
Millipore filter on the first Alj O3 bed.
tStumm(1967).
$Nozaki and Tsunogai (1973).
§Stumm and Morgan (1970).
H Sibley and Morgan (1976).
**Lowman and Ting (1973).
tfRobertsonetal. (1973).
$tAndelman and Ruzzell (1970).

366 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

AI2O3 beds, at all the salinities tested. The efficiencies derived from the concentrations
on each sorption bed can be used to obtain the total concentration in the water.

The tank experiments, in addition to testing the BLVWS technique, also indicated
that chemical speciation beyond particulate and soluble fractions can be made. There
appeared to be evidence of a "colloidal" form of plutonium that passed through a 03-jJim
Millipore filter but was efficiently sorbed on the first AI2O3 bed. Tlie colloidal
plutonium, first discussed by Nevissi and Schell (1975), appeared at the higlier salinities,
31.6 %o, and was identified by a much greater collection efficiency on the first AI2O3
bed compared with the second and third beds, as shown in Fig. 16. An abrupt change in

100 FT

LU
>

1= 100
<

IX.

31.576 %c

1 —

0 %«

31.576 %o

11.756 °'oo

31.608 %o

1 2. 1 79 %o

31.608 %o

20.923 %o

31.608 %o

19.950 %o

31.608 %o

I I I I

4 0

AI2O3 BED NUMBER

Fig. 16 Relative concentrations of plutonium sorbed on AljOj beds at different
salinities using prefiltered water (0.3 ^m) that had been aged for 2 to 3 weeks before
sampling with the BLVWS. o, extrapolated values for soluble activity. •, activity
measurements.

TRANSURANIC ELEMENTS AT BIKINI ATOLL 567

100 pr

Mill

- D-6 (2 m)

c^ 10

Q-

0.1

100

B-3 (A)(2 mi

B-3 (16 m)

B-3 (29 m)

B-3 (B)
(16 m)

II I ll I I II

B-32 (B)

(2 m)

B-32 (32 m)-

4 0

AI2O3 BED NUMBER

3
Q.

0.1

I II I

B-32 (A)
(2 m)

1 —

I I I I

B-32 (2 m)

S-1

I I II

B-32 (2 m)

S-2

B-32 (2 m)
S-3

B-32 (2 m)
S-4

B-32 (17 m)

40 2 40 2 40 2 40

AI2O3 BED NUMBER

4 0

Fig. 17 Concentrations of plutonium sorbed on Ai2 03 beds from BLVWS samples
collected at Bikini Atoll in 1976. ., extrapolated values for soluble activity. •, activity
measurements.

slope occurred in bed 2 for the high-salinity samples. This effect was not observed at the
lower salinities. By extrapolation back to the first bed, the colloidal fraction was
estimated. For plutonium the colloidal fraction of the total averaged 15 ±3% with a
range of 1 1 to 1 7%.

The BLVWS collections of 2 3 9,2 4 op^ -^^ water samples collected in July 1976 at
Bikini Atoll have been evaluated to determine the fraction of the total concentration
present in the colloidal state. The colloidal fraction has been determined, as before, by
extrapolating the amounts of 2 3 9,2 4 Op^ ^^gllected on the second, third, and fourth bed
to the first bed, as shown in Fig. 17. The least-squares regression line with its error
through these data extrapolated to bed 1 gives the soluble fraction in bed 1. The

568 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

difference between this extrapolated value (soluble) and the total concentration measured
in bed 1 is defined as the amount present in the colloidal state. Table 5 gives the amount
of 2 3 9,240py present in the total sample, i.e., soluble, colloidal, and percent colloidal.
The average of the colloidal concentration of 2 3 9,24 0pjj ^ ^.j^g 22 lagoon samples
measured from Bikini Atoll was 12 ± 8% (SD). This average can be compared with the
tank studies at 31.6 %q salinity in which 15 ± 3% (SD) was found in the colloidal state.
Since only two AI2O3 beds in the BLVWS were used for the collections at Bikini
Atoll in 1972, no colloidal fraction could be determined, and a systematic error may have
been present in the total plutonium concentration data. The 1976 results have been
evaluated with only two beds to observe if significant systematic errors were present in
the 1972 data. Of the total plutonium present in 14 samples tested, 91% (range, 78 to
100%) would have been measured if only the first two beds had been used in the
calculations of total concentrations. Thus the 1972 data for individual samples measured
by the BLVWS technique should be reliable to within about 20%.

Speciation of"^^ ^Pu and ^ ^ ^ '^^^Pu into Paniculate and Soluble Fractions. The isotope
ratios of 2 3 8pjj^2 3 9,24 0p^ ^ surface sediments and in the overlying batch water samples
are shown in Figs. 18 and 19, respectively. The range of ^^^Pu/^^''^'*°Pu ratios in the
surface sediments is 0.004 to 0.52; the average value determined from 28 measurements is
0.029 ± 0.027 (SD). The 2 3 8py^2 3 9,24 0py ^^^^^^ -j^ ^^ ^^^gj. samples collected at 2 m
above the bottom show that at several stations the 2 3 8p^^23 9,240p^ ratios in water are
similar to those of underlying sediments. However, the ^^^Pu/^^^'^'^^Pu ratios in water,
which range from <0.02 to 1.11, are considerably more variable than those in the
sediments. It appears that there may be enrichment in ^^^Pu/^^^'^'^^Pu ratios in the
overlying water compared with those in sediment.

The 238p^^239,240py ^^^j^^ f^^ ^.j^g BLVWS collection of plutonium in surface- and

deep-water samples with 0.3-ium Millipore filters and two AI2O3 sorption beds are shown
in Figs. 20 and 21, respectively. The 2 3 8p|j^23 9,240py j^^jq^ ^gre not usually constant
for the three fractions of the same surface- or deep-water samples. This unexpected
finding was at first questioned, and the analysis and counting of the samples were
repeated. Possible contamination by ^'^^ Am and ^^^Th, which decay by alpha particles
of nearly the same energy as ^^^Pu, was rechecked. The results showed that the original
values were real and that the particulate fraction had 2 3 8p|j^2 3 9,24 0py ratios that were
significantly different from the two soluble (<0.3 /im) fractions. In fact, the two soluble
fractions (first AI2O3 bed and second AI2O3 bed) also had different 2 3 8py/2 3 9,2 4 0p^
ratios. The source of plutonium isotopes to the water column is the contaminated lagoon
sediments, but only a few water samples have 2 3 8py^2 3 9,24 0p^ ratios equal to those of
the sediment. In the surface waters of the lagoon, the ranges of ^^^Pu/^^^'^^°Pu ratios
were: MiUipore filters, 0.014 to 0.57; first AI2O3 bed, 0.06 to 0.64; second AI2O3 bed,
0.09 to 0.46. The 238p^j^239,240pjj j2Ltios in surface-water samples from the craters and
outside the lagoon were even more variable. In fact, the 2 3 8p^^2 3 9,24 0py ratios in a few
collections made outside the lagoon were greater than 1; i.e., more ^^^Pu than

2 3 9,24 0p^^ was present in the samples where the total concentration was less than

3 pCi/m^ .

An external source of ^^®Pu, such as SNAP-9A, which burned up in the atmosphere
over the Indian Ocean in 1964 (Volchok, 1969), may contribute to the higher
^^^Pu/^^^'^'*°Pu ratios in the water column at Bikini. However, the amount of this
material at any location would be small since 17 kCi of ^^®Pu was dispersed throughout

TRANSURANIC ELEMENTS A T BIKINI A TOLL 569

the earth's atmosphere and hydrosphere. Thus, on the basis of the above findings, the
conclusions must be that different properties of the plutonium isotopes exist when the
isotopes interact with various components of the marine environment at Bikini Atoll.

The ^^^Pu appears to be more soluble than the 2 3 9,240py ^ lagoon samples, as
evidenced by the higher ratios found in the soluble fractions than in the particulate
fractions from the BLVWS collections; this preferential solubility is also illustrated by the
fact that the 2 3 8py^2 3 9,2 4 0py ^^^^^ j^ ^ig^gj- [^^ many "batch" samples from the water

column than in samples from the surface sediments. A source of ^^^Pu that is different
from bomb plutonium is indicated; most bomb debris would have much greater
2 3 9,240py than ^^^Pu concentrations except for those devices which used ^"^^Cm as a
tracer. Since both ^^^Pu and ^'*'^Pu were measured together by alpha spectroscopy, some
of the differences in the 2 3 8py^2 3 9,24 0p^ ratios possibly could be ascribed to variability
in the ^^°Pu isotope in samples. However, an evaluation of this radionuchde would
require a more detailed study using mass spectrometry to measure the ^"^^Pu
concentrations.

Plutonium in seawater at a pH 8.0 to 8.2 forms oxy— hydroxy— carbonatoplutonyl
complexes. The size of the aggregates of the plutonyl complexes would depend on the
number of plutonium atoms available and on the charge field surrounding the aggregates
or clusters (the cluster hypothesis). At Bikini Atoll the coralline particles that
experienced the effects of the fireball contain the plutonium isotopes. The release of
plutonium into the water column from these particles may depend on recoil from the
alpha decay of the plutonium isotopes; this decay, would break the bonds between
plutonium clusters and the coral matrix. The ^^^Pu clusters would have a higher
probability of being released from the coral particles than ^^^Pu because of the
differences in alpha-decay half-lives (86 yr for ^^*Pu and 24,400 yr for ^^^Pu) and
possibly by ^^^Pu formation from the decay of ^'^^Cm (t^ of 162.5 days); thus it is
reasonable to assume that ^^*Pu could be more soluble than ^^^Pu. However, the
magnitude of this preferential solubihty has not yet been determined.

If the clusters containing ^^*Pu are smaller (i.e., in effect, more soluble) than those
containing ^^^'^^*^Pu, then the results of the measurements made at Bikini lagoon and
deep ocean areas could be explained. The larger clusters of ^^^ '^"^^Pu could attach to the
riatural particles and could be removed from the water column at a more rapid rate than
the more-soluble ^^^Pu clusters. The availabihty of these different physicochemical states
of plutonium may help decide the potential hazards of transuranic elements in the
aquatic food chain to man. A concentrated effort is needed to collect additional data and
to interpret further these prehminary findings.

Conclusions

The measurements of the radionucUdes in Bikini lagoon sediments show that bomb
craters are only one of the sources for the transuranic elements in the ecosystem.
Sediments in the northwest quadrant of the lagoon contribute significantly to the
concentrations of the radionuclides found in the water and biota. Coral particles that
have been altered by the bomb and the environment contain the radionuclides. These
particles must be transported and subsequently deposited at different locations; this is
indicated by the high sediment rate found at station B-2 (0.58 cm/yr) and by the changes
observed in the radionucUde concentrations found in the sediment-core profile.

(Text continues on page 576.)

570 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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TRANSURANIC ELEMENTS AT BIKINI ATOLL 573

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574 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

0.007(0.002) C-6

0.02(0.008)

Kilometers

T 1 1 r—

BR^^OCRATER/^'J'N^ .0.10(0.02) .0.06(0.02)

15
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(0.008) (0.0041
STATIONS

0.52(0.002) t
0.25(0.03)

0.09(0.02)

Fig. 18 Ratios of ^ssp^^asg ,2 4 0py concentration in surface sediments collected in
1972.

Kilometers

STATIONS

Fig. 19 Ratios of ^ ^*Pu/^^' '^''"Pu concentration in 90-liter batch samples collected in
July 1976 at Bikini Atoll and coprecipitated with MnOj-

TRANSURANIC ELEMENTS AT BIKINI ATOLL 575

Kilometers

1 — I — I I

0.10(0.02
0.55(0.09)»

0 18(0.04)
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AREA

(\>\ • 0.56(0.04)
'Svfc' 0.35(0.08)

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0 0810 011
STATIONS 0451013)

Fig. 20 Ratios of ^ ^ * Pu/^ a 9 ,2 4 o p^ concentration in surface-water fractions from
BLVWS collections in 1972: Top number, particulates (error); middle number, first
Alj O3 bed (error); bottom number, second Alj O3 bed (error). — (-), no data available.

Kilometers

STATIONS

Fig. 21 Ratios of 2 3 8pu/2 3 9,24 0py concentration in deep-water fractions from
BLVWS collections in 1972: Top number, particulates (error); middle number, first
Alj O3 bed (error); bottom number, second Alj O3 bed (error). — (-), no data available.

516 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

The distribution of plutonium concentrations in the water column follows the general
lagoon circulation pattern with the tiighest concentration in the northwest quadrant and
with decreasing concentration gradually toward the east and south. The outlet of the
radioactive lagoon water is through the deep passes in the southwestern part of the Atoll
where the exit into the North Equatorial Current occurs. The physicochemical state
studies of plutonium show that approximately 15% of the total concentration is present
in the colloidal fraction; varying amounts are found in the soluble and particulate
fractions, depending on location. The americium is found primarily associated with the
particulate fraction (>0.3/nm). Measurements of the isotope ratios of ^^^Pu/^^^'^'^^Pu
in the particulate, first AI2O3, and second AI2O3 fractions of a water sample are not
constant. The ratios in these water fractions also vary with location in the lagoon. The
^^^Pu appears to be more soluble than the ^^^'^^*^Pu in lagoon samples since the ratios
are higher in the soluble fractions (first AI2O3 and second AI2O3) than in the particulate
fraction. In an attempt to explain these observations, it is speculated that the ^^^Pu
forms smaller clusters than the 2 3 9,24 0pjj Q^jj^g jq possible differences in specific
ionization. This could be a result of the differences in decay half-lives (^^^Pu, 86 yr;
2 3^Pu,24,400yr).

References

Adams, C. E., N. H. Farlow, and W. R. Schell, 1960, The Compositions, Structures and Origins of

Radioactive Fallout Particles, Geochim, Cosmochim. Acta, 18: 42-56.
Andelman, J. B., and T. B. Ruzzell, 1970, Plutonium in the Water Environment, \n Radionuclides in

the Environment, R. F. Gould (Ed.), Advances in Chemistry Series, American Chemical Society,

Washington, D. C.
Anikouchine, W. A., 1961, Bottom Sediments of Rongeiap Lagoon, Marshall Islands, M.S. Thesis,

University of Washington, Seattle, Wash.
Emery, K. O., J. I. Tracey, Jr., and H. S. Ladd, 1954, Bikini and Nearby Atolls: Part 1, Geology,

Professional Paper No. 260-A, U. S. Geological Survey.
Farlow, N. H., and W. R. Schell, 1957, Physical, Chemical and Radiological Properties of Slurry

Particulate Fallout Collected During Operation Redwing, Report USNRDL-TR-170, U. S. Naval

Radiological Defense Laboratory, NTIS.
Goldberg, E. D., 1963, Geochronology with Pb-210, in Radioactive Dating, Symposium Proceedings,

Athens, November 1962, pp. 121-131, STI/PUB/68, International Atomic Energy Agency, Vienna.
Held, E., 1971, Amchitka Radiobiological Program Progress Report. July 1970-April 1971, USAEC

Report NVO-269-11, University of Washington, College of Fisheries, NTIS.
Huntamer, D. D., 1976, An Evaluation of the Battelle Large Volume Water Sampler for Measuring

Concentrations and Physico-Chemical States of Some Trace Elements in Marine Waters, M.S.

Thesis, University of Washington, Seattle, Wash.
Koide, M., A. Soutar, and E. D. Goldberg, 1971, Marine Geochronology with Pb-210, Earth and

Planet, 5cz. Letters, 14: 442446.
Lowman, F. G., and R, Y. Ting, 1973, The State of Cobalt in Sea Water and Its Uptake by Marine

Organisms and Sediments, in Radioactive Contamination of the Marine Environment, Symposium

Proceedings, Seattle, July 10-14, 1972, STI/PUB/313, International Atomic Energy Agency,

Vienna.
Marshall, R. P., 1975, Concentrations and Redistributions of Plutonium, Americium and Other

Radionuclides on Sediments at Bikini Atoll Lagoon, M.S. Thesis, University of Washington,

Seattle, Wash.
Nevissi, A., and W. R. Schell, 1975, Distribution of Plutonium and Americium in Bikini Atoll Lagoon,

Health Phys., 28: 539-547.
, and W. R. Schell, 1976, Efficiency of a Large Volume Water Sampler for Some Radionuclides in

Salt and Fresh Water, in Radioecology and Energy Resources, Proceedings of the Fourth National

Symposium on Radioecology, Oregon State University, May 12-14, 1975, pp. 277-282, C. E.

TRANS URANIC ELEMENTS AT BIKINI ATOLL 577

Gushing, Jr. (Ed.), Ecological Society of America Special Publication Series, No. 1, Dowden,

Hutchinson and Ross, Inc., Stioudsburg, Pa.
Noshkin, V. E., R. J. Eagle, and K, M. Wong, 1976, Plutonium Levels in Kwajalein Lagoon, Nature,

262: 745-748.
, K. M. Wong, R. J. Eagle, and C. Gatrousis, 1974, Transuranics at Pacific Atolls. I. Concentrations

in the Waters at Enewetak and Bikini, USAEC Report UCRL-51612, Lawrence Livermore

Laboratory, NTIS.
Nozaki, Y., and S. Tsunogai, 1973, A Simultaneous Determination of Lead-210 and Polonium-210 in

Sea Water, y4/2fl/. Chem. Acta., 64: 209-216.
Robertson, D. E., W. B. Silker, J. C. Langford, M. R. Peterson, and R. W. Perkins, 1973, Transport and

Depletion of Radionuclides in the Columbia River, in Radioactive Contamination of the Marine

Environment, Symposium Proceedings, Seattle, July 10-14, 1972, STI/PUB/313, International

Atomic Energy Agency, Vierma.
Schell, W. R., 1959, Identification of Micron-Sized, Insoluble-Solids Fallout Particles Collected During

Operation Redwing, Report USNRDL-TR-364, U. S. Naval Radiological Defense Laboratory.
, 1976, Concentrations and Physical-Chemical States of ^^Fe in Bikini AtoU Lagoon, in

Radioecology and Energy Resources, Proceedings of the Fourth National Symposium on

Radioecology, Oregon State University, May 12-14, 1975, pp. 271-276, C. E. Cushing, Jr. (Ed.),

Ecological Society of America Special Publication Series, No. 1, Dowden, Hutchinson and Ross,

Inc., Stioudsburg, Pa.
, 1977, Concentrations, Physico-Chemical States and Mean Residence Times of * '"Pb and ** "Po

in Marine and Estuarine Waters, Geochim. Cosmochim. Acta, 41: 1019-1031.
, T. Jokela, and R. Eagle, 1973, Natural ^'°Pb and ^"'Po in a Marine Environment, in

Radioactive Contamination of the Marine Environment, Symposium Proceedings, Seattle,

July 10-14, 1972, pp. 701-724, STI/PUB/313, International Atomic Energy Agency, Vienna.
, A. Nevissi, and D. D. Huntamer, 1978, Sampling and Analysis of Am and Pu in Natural Waters,

Mar. Chem., 6: 143-153.

, and R. L. Watters, 1975, Plutonium in Aqueous Systems, Health Phys., 29: 589-597.

Shapley, D., 1971, Plutonium: Reactor Proliferation Threatens a Nuclear Black Maiket, Science, 172:

143-146.
Sibley, T. H., and J. J. Morgan, 1976, Equilibrium Speciation of Trace Metals in Freshwater. Sea Water

Mixtures, California Institute of Technology, unpublished.
Silker, W. R., R. W. Perkins, and H. C. Rieck, 1971, A Sampler for Concentrating Radionuclides from

Natural Waters, Ocean Eng., 2: 49-55.
Stumm, W., 1967, Metal Ions in Aqueous Solutions, in Principles and Applications of Water

Chemistry, S. D. Faust and J. V. Hunter (Eds.), John Wiley & Sons, Inc., New York.
, and J. J. Morgan, 1970, Aquatic Chemistry: An Introduction Emphasizing Chemical Equilibria in

Natural Waters, John Wiley & Sons, Inc., New York.
Volchok, H. L., 1969, Fallout ofPu-238from the SNAP-9A Burnup-IV, in USAEC Report HASL-207,

pp. 1-5 to 1-13, Health and Safety Laboratory, NTIS.
Von Arx, U. S., 1954, Circulation Systems of Bikini and Rongelap Lagoons, in Professional Paper No.

260-B, pp. 254-273, U. S. Geological Survey.
Welander, A. D., et al., 1966, Bikini-Eniwetok Studies, 1964. Part I. Ecological Observations, USAEC

Report UWFL-93(Pt. 1), University of Washington, College of Fisheries, Laboratory of Radiation

Ecology, NTIS.

Transuranium Radionuclides

in Components of the Benthic Environment

of Enewetak Atoll

V. E. NOSHKIN

Data on the concentrations and distributions of transuranium radionuclides in the marine
environment of Enewetak Atoll are reviewed. The distributions of the transuranics in the
lagoon are very heterogeneous. Tlie quantities of transuranics generated during the
nuclear-test years at the Atoll and now associated with various sediment components are
discussed. Whenever possible, concentrations of "^^^ Am and '^^^'^'^^^Pu are compared.
The lagoon is the largest reservoir of transuranics at the Atoll, and radionuclides are
remobilized continuously to the hydrosphere from the solid source terms and are cycled
with components of the biosphere. Although ^^^ ^^^Pu is associated with filterable
material in the water column, the amount that is relocated and redeposited to different
areas in the lagoon is small. Barring catastrophic events, little alteration in the present
distribution of transuranics in the sediment is anticipated during the next few decades.
The Atoll seems to fiave reached a chemical steady state in the partitioning of^^^'^^'^^Pu
between soluble and insoluble pliases of the environment. Tlie amount of dissolved
radionuclides predicted, with an experimentally determined K(j for '^^^'^^^^Pu, to be in
equilibrium with concentrations in the sediment agrees well with recently measured
average concentrations in the water at both Enewetak and Bikini atolls. Tlie remobilized
2 39+2 40p^ /ws solute-like characteristics. It passes readily and rapidly through dialysis
membranes and can be traced as a solute for considerable distances in the water. It is
estimated tfiat 50% of the present inventory of ^^^'^^^^Pu in sediment will be
remobilized in solution and discliarged to the North Equatorial Pacific over the next
250 yr.

Large inventories of several transuranium radionuclides (U. S. Atomic Energy Commis-
sion, 1973) persist in the marine environment of Enewetak Atoll. Forty -three nuclear
weapons tests were conducted by the United States at Enewetak between 1948 and 1958.
The testing produced close-in fallout debris which was contaminated with transuranics
and which entered the aquatic environment of the Atoll. More transuranics were
transported westward to Enewetak in airborne debris and water contaminated from
nuclear testing at Bikini Atoll. Global fallout deposited a small additional amount of
transuranics on the Atoll. Presently, the largest inventory of transuranics introduced from
these source terms is associated with components of the benthic environment.

Because of the high level of deposition, the Atoll is now its own transuranic source
term. Plutonium, for example, is not permanently fixed with the carbonates and other
material with which it was originally deposited in the lagoon and on the reef during
nuclear testing. Small amounts of plutonium are now remobilized, resuspended,
assimilated, and transferred continuously within the Atoll environment by physical,
chemical, and biological processes.

578

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 579

More than half the U. S. nuclear tests in the Pacific were conducted at Enewetak
Atoll. Surface and tower shots left craters and contaminated scrap on land and generated
radioactive debris that was redistributed to the adjacent reef and lagoon. Megaton tests
that left underwater craters and barge shots in the lagoon contributed significantly to the
present transuranic inventory.

The impact of nuclear testing and the fate of the residual radioactive materials
introduced to the aquatic environment at both Enewetak and Bikini atolls are the
subjects of reports too numerous to list herein. Not until late 1972, however, when a
radiological resurvey of Enewetak Atoll was conducted to gather data for the
development of cleanup and rehabilitation procedures for the resettlement of the
Enewetak people to their homeland, did extensive measurements of transuranics in the
Atoll environment begin. The information was published in a survey report (U. S. Atomic
Energy Commission, 1973), which contains data on most long-lived residual radio-
nuclides, including plutonium and americium, in components of the marine environment.
The survey was followed by other more-extensive investigations, which concentrated on
the measurement of transuranics to better assess the impact of these radionuclides on the
environment and inhabitants of the Atoll and to increase our understanding of the
mobilization, reconcentration, and redistribution processes from sources within the
environment.

This chapter contains a summary of data related to the concentrations of the
transuranium elements in components of the benthic and pelagic environment of the
Atoll lagoon. Data from the survey report (U. S. Atomic Energy Commission, 1973),
more-recent publications, and unpublished results from this laboratory are discussed.
Some published and unpublished data from Lawrence Livermore Laboratory (LLL)
studies at Bikini Atoll are presented when necessary for comparison with Enewetak data
and, in the absence of Enewetak data, for the clarification of characteristics of
transuranic radionuclide concentrations at the Atolls. Whenever possible, the Atoll data
are compared with those from other marine ecosystems.

Geography and Atoll Test History

Enewetak Atoll, with U. S. -assigned and native names and several landmarks, including
the locations of craters formed by nuclear tests, is shown in Fig. 1 . The U. S. -assigned
island names are used throughout this chapter.

The Atoll originally consisted of a ring of 42 low islands arranged on a roughly
elliptical reef, 40.2 by 32.2 km (Emery, Tracy, and Ladd, 1954), with the elongated axis
in the northwesterly direction. Nuclear testing completely destroyed the islands of Gene
and Flora, and only a sandbar now remains to distinguish the island of Helen. Only 39 of
the original 42 islands of the Atoll remain; these islands make up a total land area of
approximately 6.9 km^ , which is situated on the reef which has an area of 84 km^ . The
average depth of the lagoon is 47.4 m; the maximum depth is 60 m. The lagoon area is
933 km^ . The sedimentary components in Enewetak lagoon were studied extensively
during the late 1940s (Emery, Tracy, and Ladd, 1954). The main components in the
lagoon sediments included foraminifera, coral, Halimeda remains, shells of moUusks, and
tine material. Material finer than 0.5 mm in diameter was too fine to identify and was
classified as fine debris. Distributions and average abundance of the sedimentary
components were described (Emery, Tracy, and Ladd, 1954). Fine debris made up 57%
of the lagoon sediments and was abundant throughout the lagoon to witliin a few
hundred feet from the shore.

580 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Teiteiripucchi (Gene)
Elugelab (Flora
Bokinwotme (Edna)
Louj (Daisy)
Kirunu (Clara)
Bokomboko (Belle)
Bokoluo (Alice

Bokaidrikrik (Helen)
■ Boken (Irene)

Enjebi (Janet)
r- Mijikadrek (Kate)
■ Kidnnen (Lucy)
Bokenelab (Mary)
Elle (Nancy)

Aomon (Sally)
Bijire (Tilda)

Lojwa (Ursula)

Alembel (Vera)
Billae (Wilma)

LACROSSE
CACTUS

Biker
(Leroy)

Southwest
passage

Kidrenen (Keith
Ribewon (James

Boken (Irvin

Mut (Henry

Ikuren (Glenn* -'

Nautical miles

Runit (Yvonne)

"M" Zona
Boko (Sam)
Munjor (Tom)
nedral (Uriah)

Jinedrol (Alvin)
Ananij (Bruce)

Jinimi (Clyde)
Japtan (David)

Medren (Elmer)

Enewetak (Fred)
IT' 20' North

11/20 12 3 4 5

Fig. 1 Map of Enewetak AtoU with names and locations of the islands and the six
nuclear craters.

A detailed description of the forms, living habits, populations, and specific
relationships of the aquatic biological components at the Atoll is beyond the scope of this
report. A significant number of articles published between 1955 and 1974 describing the
research conducted at the Enewetak Marine Biological Laboratory were compiled
recently in a three-volume report (Mid-Pacific Marine Laboratory, 1976). The individual
reports dealing with specific ecological studies at the Atoll are too numerous to list. The
reader is referred to the compilation for descriptions of the biology and ecology of the
Atoll.

The most severe radiological impact on the aquatic environment of Enewetak
occurred during the nuclear-test years between 1948 and 1958. The types of nuclear
events, shot frequencies, geographical locations, yields, generated particles, conditions
after the tests, and other factors determined the resulting distributions of transuranics
and influenced the physical and chemical forms of the elements deposited in the benthic
environment. A brief liistorical review of testing at Enewetak, abstracted from several

TRANS URANIC RADIONUCLIDES IN ENEWETAK LAGOON 581

unclassified documents (U. S. Atomic Energy Commission, 1973; Circeo and Nordyke,
1964; Hines, 1962), explams a few conditions responsible for the transuranic distribu-
tions and inventories at the Atoll.

The test series at Enewetak began in 1948 (Operation Sandstone) when 37-, 49-, and
18-kt devices were detonated from 200-ft towers on the islands of Janet, Sally, and
Yvonne between April 14 and May 14. In 1951, testing was resumed (Operation
Greenhouse), and four tower shots were conducted during a 47-day interval. The island of
Janet was again the location of two ground zeros. In 1952, the first thermonuclear device
(Mike) destroyed the island of Flora on the northwest reef. The Mike event, a 10.4-Mt
surface detonation, occurred on October 31. Water surging from the point of the
explosion sent a wave over adjacent islands, including Janet, the site of three previous
ground zeros. The original crater where Flora had once been had an irregular outline and
was more than 1 mile in diameter. Before the crater was partially refilled by the returning
rush of coral sediment, it was almost 200 ft deep; it is presently 90 ft deep. The 1952
series of tests concluded with the King event, a high-yield airdrop over Yvonne Island. In
1954, a single device. Nectar, was detonated on a barge located over Mike crater. Not
only did this test greatly disturb the radionuclides already deposited in the crater
sediments but it also again sent a surge of contaminated water over adjacent islands,
including Janet. In 1956, the Redwing series began with a tower detonation on Yvonne
and included two additional cratering events, LaCross and Seminole. LaCross was a
39.5-kt device detonated on an earth-filled causeway built on the reef off the north end
of Yvonne. Seminole, detonated on the island of Irene, was first placed in a
15-ft -diameter tank that was itself then placed in a 50-ft -diameter tank filled with water
before it was fired. During 1958, the final year of testing at Enewetak, 22 tests of various
types were held at different Atoll locations during an 82-day period. The series opened
with an 86,000-ft balloon shot over the Atoll on April 28. On May 5, an 18-kt device
produced Cactus crater on the northwest end of Yvonne and west of LaCross crater.
During May 1 1 and 12, one of three tests conducted was the Koa event, a 1 .37-Mt nuclear
device housed in a tank of water and detonated on the east end of the Gene— Helen island
complex. A sizable crater was produced, which connected with Mike crater. On June 8,
the Umbrella device was detonated on the floor of the lagoon. Twenty days later, the
8.9-Mt Oak device was fired on a barge 4 miles southwest of Alice off the edge of the
reef. The test left a crater that breached to the lagoon. The Quince event on Yvonne
Island failed to produce a fission yield; so the plutonium within the device was dispersed
by a high explosive. Subsequently another nuclear device was successfully detonated over
the same area and undoubtedly further dispersed the nonnuclear -generated plutonium. In
addition to the nuclear tests, radionuclides were dispersed by plowing on many of the
islands during the test years. Unfortunately, none of the radiological safety reports during
these operations provided details to determine the eventual fate of the radioactive debris,
e.g., location and quantity of the disposal (U. S. Atomic Energy Commission, 1973).

From this brief summary, we can assume safely that the transuranic elements were
introduced to the aquatic environment not only as complicated carbonate particles fused
or condensed with other material from the environment or with devices and associated
structures but also as soluble and particulate species of transuranium oxide.

Despite the complexities in the formation processes, much of the behavior of the
transuranics is similar to that determined from investigations of fallout and other aquatic
pathways. The results from the Atoll studies therefore have great value in predicting
transuranic behavior and fate on a global aquatic scale.

582 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Transuranic Elements Identified at the Pacific Test Site Atolls Since 1972

Neptunium

Concentrations of ^^''Np in several 1972 samples of unfiltered lagoon and crater water
from Enewetak were determined by mass spectrometry (Noshkin et al., 1974). The
average concentration in six samples from the lagoon was 0.058 ± 0.013 fCi/liter. Water
samples from Mike and Koa craters averaged 0.45 ± 0.22 fCi/liter. Outside the lagoon and
to the east of the Atoll, concentrations in water samples from the open ocean surface
averaged 0.013 ±0.003 fCi/liter. Tliis comparison shows, as do results for all other
transuranics, that Atoll sources contribute the major fraction of the transuranic inventory
in the water column of the lagoon. The ^^'^Np concentrations in the lagoon and crater
water samples were less than 0.2% of the measured ^^^ ^''^Pu concentrations in those
samples.

Plutonium

Many types of samples from the Atoll contain ^^^Pu, ^^^Pu, ^'^"Pu, and ^"^'Pu. Most
reported values are the sum of ^^^Pu and ^"^^Pu activities determined by alpha
spectrometry. These radionuclides are distributed widely throughout the Atoll and have
been detected in nearly every type of marine and terrestrial sample analyzed to date.
Atoll water samples, sedimentary components (including fine unidentifiable carbonate
sands, coral fragments, Halimeda debris, foraminifera, and moUusk shells), living algae,
benthic invertebrate tissues, planktonic species, and marine vertebrate tissue all contain

239+240p^

The distribution of ^ ^^Pu is as wide among components in the marine environment as
is 239+240p^ ^^^ ^^ \o\Nex Concentrations. The ^^^Pu/^^^'^^'^^Pu ratio determined in a
variety of aquatic samples from different regions of the lagoon ranges from less than 0.04
to greater than 0.50.

A few activity ratios of ^'*°Pu/^'^^Pu were determined by mass spectrometry. The
ratios in two water samples collected from the lagoon during 1972 were 0.432 and 0.289
(Noshkin et al., 1974). Samples of mackeral bone and of viscera collected in 1972 near
the island of Glenn had ^'^^Pu/'^'^Pu activity ratios of 1.15 ±0.25 and 1.27 ±0.26,
respectively; goatfish viscera and tridacna tissue from nearby David had ratios of
0.68 + 0.07 and 0.66 ± 0.19 (Gatrousis, 1975), respectively. The activity ratios in 56 soil
samples from seven islands ranged from 0.066 to 1.42 and averaged 0.84 ± 0.37
(Gatrousis, 1975), and the average ratio in seven marine water and biota samples was
0.66 ± 0.40. Neither average value determined in the environmental samples differed
greatly from the average of 0.65 ± 0.05 for global fallout debris (Krey et al., 1976). The
similar isotopic ratio in mackeral tissue shows no obvious discrimination in uptake of
isotopes by tissues of organisms in the Atoll if feeding and living are restricted to specific
regions of the Atoll.

The average ^'^^Pu/'^^^Pu ratio in the yearly growth sections of a live sample of
Favites virens coral collected from the western basin in Bikini lagoon was 0.77 + 0.07
(Noshkin et al., 1975). This value is similar to the isotopic ratio in Enewetak samples.
However, the mean isotopic concentration ratio in soil and vegetation of Bikini and Eneu
islands is 1.15 (Mount et al., 1976), which is somewhat higher than the average in the
Bikini coral sample.

Since the ^'*°Pu/'^^^Pu activity ratio in some environmental samples exceeds 1, it
seems inappropriate to use the shorthand notation, ■^^^Pu, when referring to the sum of

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 583

^^^Pu and ^"^^Pu activities as has so often been done in the literature. Throughout this
report, '^^ ^"^^Pu will refer to the sum of the activities of the two radionuclides, and
^^^Pu will refer to only that isotope.

In two Enewetak lagoon water samples collected during 1972 (Noshkin, 1974),
^'*' Pu was measured by mass spectrometry. The ^^'Pu/'^^^ ^"^^Pu activity ratios as of
December 1972 were 1.14 and 2.56. hi the 1972 growth section of the previously
mentioned live coral from Bikini, the ^'^'Pu/'^'^^ ^''^Pu activity ratio was 11.7 and the
24ip^^239p^ ratio was 21.0 ±1.1 (Noslikin et al., 1975). As of Jan.l, 1975, the
^"^^Pu/^^^Pu ratio in soil samples from Bikini and Eneu islands and in Bikini Island
vegetation averaged 22.0 ±3.3 (Mount et al., 1976). Correcting the ^'*^Pu in the
November 1972 coral growth section for decay to Jan. 1, 1975, yields a ■^ "* ' Pu/^ ^ ^ Pu
ratio of 18.9 ±1.1. Bikini and Eneu islands and the sedimentary environment from which
the coral was obtained were contaminated principally with radioactive debris from the
1954 Bravo event. The good agreement between the ratios determined in the terrestrial
and marine samples indicates a lack of discrimination between ■^'^^ Pu and ^ ^^Pu isotopes
in processes in these environments. Bikini and Enewetak have very different isotopic
ratios and therefore different inventories of plutonium isotopes. The amount of ^ "* ^ Pu in
the environment regulates the projected inventory of ^^' Am through growth and beta
decay of the parent radionuclide. The amount of ^^' Am that will be generated at Bikini
from ^^' Pu decay will exceed the amount produced by this source at Enewetak.

Americium

The distribution of ^"^^ Am is also wide spread in the aquatic environment of the Atoll.
Although the highest concentrations of plutonium and americium are in the same areas of
the lagoon at Enewetak, the two transuranics are distributed differently. The ^"^^ Am/
239 240p^ j.^^-Q jj^ sediments collected from the lagoon during 1972 ranged from 0.06 to
0.93. Signitlcant errors, therefore, can be introduced if one transuranic is used to predict
the levels of others at any given location in the lagoon. No other americium isotopes were
detected in the aquatic environment of either Enewetak or Bikini.

Curium

No strenuous effort has been made to obtain an inventory of ^'*'^Cm or ■^'''^Cm in
Enewetak by alpha spectrometry. Curium activities were separated and measured in
several lagoon water samples: ^"^"^Cm activities were less than 0.2 fCi/liter. No ^'^'^Cm or
^"^"^Cm was detected in sediment samples from the Bravo crater at Bikini Atoll (Beasley,
1976).

Higher Transuranic Elements

No information is available to my knowledge on either berkelium or californium in
marine samples from Enewetak or Bikini.

Transuranic Elements in the Benthic Environment of Enewetak Atoll

Surface-Sediment Distributions and Inventories

The distributions of ^^^"^^"^^Pu and ^'^^ Am activities measured in December 1972 in the
2.5-cm-thick surface layer of sediment from the lagoon floor are shown in Figs. 2 and 3,

584 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

-3335± 518

-1336 ± 158

74.5 ±11.8

1110± 125
695 ± 80

Fig. 2 Activities of 239+2*0 Pu (millicuries per square kilometer) associated with the
sediment components in the top 2.5-cm layer of Enewetak lagoon.

respectively. Isolines were constructed to distinguish regions of the lagoon having similar
concentrations. The mean transuranic inventory in the surface layer and the range of
concentrations within the defined areas are shown in the two figures. Figures 4 and 5
show regions of the surface layer of sediments mXh. similar 2 38py^2 3 9+240pjj ^^^
^'*^ Am/^^^'''^'*°Pu ratios, respectively.

The transuranic concentrations in the surface layer of sediments were determined in
over 150 ball-milled surface samples of known thickness and in 20 core samples obtained
throughout the lagoon. The lagoon was divided into a grid consisting of a series of 6-km^
regions; at least one sediment sample was obtained from each region to provide
radiological data for areal distributions. All sediments are composed of different
quantities of fine- and coarse-grained carbonate material, shells, coral fragments, and
Halimeda debris. No attempt was made in assessing the sediment inventory to distinguish
concentration levels in specific sedimentary components. Figures 2 through 5 illustrate
the main features of the transuranic distributions in the surface layer of the lagoon
sediment. Isolated regions of relatively high concentrations of 2 3 9+2 4 op^ ^^^ evident in
some lesser contaminated areas of the lagoon; other small regions of high surface

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 585

82 + 6

Fig. 3 Activities of ^^'Am (millicuries per square kilometer) associated with the
sediment components in the top 2.5-cm layer of Enewetak lagoon.

radioactivity might have escaped detection. The areal distributions are based on available
data from the samples that were collected and analyzed.

The transuranics are distributed nonuniformly over the lagoon floor. Highest surface
concentrations are associated with the sediments near, but not necessarily adjacent to, the
locations of larger or more numerous nuclear tests. Highest plutonium concentrations are
associated with the sediments from the northwest quadrant in a north- and south-oriented
elliptical area that is roughly 2 to 3 km east of the islands of Alice and Belle and several
kilometers southwest of Mike and Koa craters. A second region of relatively high
concentration is in sediments off the shore of Yvonne Island. The activity in this region is
lower than that in sediments in the northwest. Most of the transuranic inventory in the
surface sediments can be separated roughly from the lesser contaminated deposits by a
line extending from the Southwest Passage to the island of Tom (Munjor), which is south
of Yvonne on the eastern reef. The surface 2 39+2 40p^ concentrations north of this line
range between 2 and 170 pCi/g (dry weight); those south of this line are less than 2 pCi/g.
All surface-sediment samples obtained during and since 1972 contained ^^^ ^'*°Pu. The
inventory (mCi/km^) in only the top 2.5-cm layer of sediment exceeds the activity
deposited to the earth's surface as worldwide fallout in any latitude band in the northern
or southern hemisphere (Hardy, Krey, and Volchok, 1973).

586 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Range, 0.10-0.15

:^'fr^1

Mean, 0.06 ± 0.02
Range, 0.04-0.10

Range,
0.06-0.38

^>--

I 'I

Mean, 0.11 ^ 0.02
Range, 0.08-0.14

/ /

Fig. 4 Activity ratios of =3 8py^239 +2 4opy j^, jj^^ surface sediments of Enewetak
lagoon.

Although the surface distribution of ^'^^ Am in the sediments appears similar to that
of "'^■'^^"Pu, the ratios of ^"^ ^ Am/^^^^^'^^Pu activities (Fig. 5) show that the
radionuclides are not well mixed throughout the surface deposits. The ratios in the
sediments range from 0.06 to 0.93. The mean ratio, however, determined by averaging
^^^ Am/^^^"''^^^Pu ratios from all surface-sediment samples, is 0.29 ± 0.17, and the ratio
determined from the mean surface concentrations (Table 1) is 0.30 ± 0.06. The average
ratio is similar to that found in central Pacific and northeast Atlantic sediments
(Livingston and Bo wen, 1976), which receive only worldwide fallout deposition but have,
in contrast, one-half the average concentration ratio of surface sediments at Bikini
(Nevissi and Schell, 1975).

The ratios of 2 38p^y2 39+24 0py activities in the surface sediments (Fig. 4) demon-
strate the nonuniformity among plutonium isotopes in components of the sediment in
the Atoll environment. There are, however, large geographical regions of the lagoon with
similar isotopic ratios in the sediment. On the other hand, small areas of the lagoon, such
as a 600-m strip on the lagoon side of Yvonne Island, contain plutonium with isotopic
ratios ranging from 0.05 to 0.38 (U. S. Atomic Energy Commission, 1973). In Cactus
crater, at the northern end of Yvonne, the isotopic ratio of 0.55 in the sediments is one
of the highest at the Atoll. The average concentration ratio in the lagoon sediments,

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 587

Mean, 0.24 + 0.04 /
Range, 0.21-0.29/^

Fig, 5 Activity ratios of ^"^ Am/^^'"'"^'"'Pu in the surface sediments of Enewetak
lagoon.

TABLE 1 Estimated Transuranic Sediment Inventory, Enewetak
and Bikini Atolls, Jan. 1, 1973

2 3 9+240

'Pu

2 4 0

Enewetak Atoll (area, 933 km^)
Areal activity to 2.5-cm depth, mCi/km^
Total radioactivity to 2.5-cm depth, Ci
Total radioactivity to 16-cm depth, Ci

Bikini Atoll (area, 629 km^)
Areal activity to 2.5-cm depth, mCi/km^
Total radioactivity to 2.5-cm depth, Ci
Total radioactivity to 16-cm depth, Ci

Total radioactivity due to global fallout
from weapons testing, January 1971,
kCi

267

145

122

493

249

135

114

460

1,185

167

642

543

2,190

475

492

229

263

4,809

289

309

144

165

3,025

182

1,470

686

786

14,405

1,140

319

22="

192

127

3,010

^Weapons, 8.6 kCi; faUout debris from SNAP 9A, 13.4 kCi.

588 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

determined from the mean surface concentrations, is 0.14. The average ratio determined
in the lagoon water samples during 1972, 1974, and 1976 is identical to the sediment
ratio. A steady-state condition is reached where plutonium isotopes are remobiUzed to
the aqueous phase in proportion to their concentrations in different regions of the
sediments and reef environments. One region of the lagoon sediments with lower or
higher isotopic ratios, for example, is not the dominant source term supplying plutonium
isotopes to the water column.

In 1977, several core samples were obtained from the lagoon basin near stations that
had been sampled in 1972. The ^^^ Am concentrations in surface-sediment layers sampled
in 1972 and 1977 were nearly identical, which showed that there was Httle change in the
surface concentrations of transuranics at many lagoon locations during those years. Only
small quantities of the transuranics were remobilized or reworked to greater depths in the
sediment column during these years. Little resuspended material from other areas of the
lagoon having different concentrations of transuranics was transported and deposited to
the areas that were resampled.

The largest inventory of transuranics at Enewetak Atoll is associated with the
components of the lagoon sediment. The estimated lagoon sediment inventories given in
Table 1 were determined from Figs. 2 and 3 by summing the products of the areas in the
lagoon by the average inventory of the transuranics present there. Approximately 250 Ci
Q^ 2 39+2 4 0py and 75 Ci of ^"^^ Am are unevenly distributed throughout the 2. 5 -cm-thick
surface-sediment layer of the lagoon. The total 2 39+2 40p^j inventory in island soils,
sampled to depths of 35 to 150cm, is estimated from available data (U.S. Atomic
Energy Commission, 1973; Noshkin et al., 1976) at <25 Ci. Transuranic distributions in
surface sediment at Bikini Atoll were constructed, and inventories were estimated from
published (Nevissi and Schell, 1975; Noshkin et al., 1975) and unpubhshed (Noshkin
et al., 1978a) data. Bikini sediment inventories were estimated from substantially fewer
data than were available from Enewetak. Future results from Bikini might change the
present estimates of transuranic inventories given in Table 1. Analysis of 25 cores (12 to
21 cm deep) from different locations in Bikini and Enewetak lagoons showed that only
21 ± 11% of the 2 39+2^0py ^^^ 16 ±6% of the ^^^^ Am (Noshkin et al., 1978a) in the
sediment column are associated with components in the top 2.5-cm layer. If the average
2 39+24 0p^ inventory in the surface sediment is only 21% of the total inventory to a
mean depth of 16 cm for the entire lagoon, then the estimated 2 39+24 op^ inventories in
the sediment column to a 16-cm depth at Enewetak and Bikini are 1.2 and 1.5 kCi,
respectively. However, in a few deeper cores, which are difficult to obtain from carbonate
deposits, 2 3 9+2 40py ^^^ 24 1 ^^ ^^^^ detected at depths below 20 cm. The inventories
computed to a depth of 16 cm then can be assumed only to represent lower limits. With
the average isotope ratios fron. samples from the Atoll environment (discussed earlier), an
estimate of the concentration for each plutonium isotope and ^'^^Am in the Atoll
sediments is made (see Table 1). Transuranic isotopes deposited from global fallout of
weapons debris are estimated from available data (Krey et al., 1976; Hardy, Krey, and
Volchok, 1973) and are also given in Table 1 .

The inventory at the Atolls of transuranics produced by weapons is only a small
fraction of the total quantity deposited to the earth's surface from global fallout debris.
Some specific marine environments were contaminated with substantial quantities of
transuranics from other source terms. These lagoon sediments, however, are the most
contaminated aquatic regions in the world that received transuranic inputs only from
nuclear weapons. The estimated ^^^Pu,^'*°Pu, ^'*^Pu, and ^"^^ Am inventories at Bikini

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 589

exceed the respective isotopic inventories at Enewetak, but ^^*Pu is higher in Enewetak
lagoon. Inventories of ^^^ Am at Bikini will increase by 25% from ^"^ ^ Pu decay, but only
a 10% increase over present ^^ ^ Am levels is expected at Enewetak from ^'^^ Pu decay.

Transuranic Elements Associated with Components in the Sediment Column

In line with the definition of Emery, Tracy, and Ladd (1954) for classifying fines as
material less than 0.5 mm in diameter, 23 surface samples and several cores were
separated into fine and coarse fractions. The dry weight of the fines ranged from 25 to
80% of the total dry weight of the surface volume (Noshkin et al., 1978a). A similar range
of fine material was found in Bikini Atoll sediments. At least 93% of the sediment weight
in Mike and Koa crater deposits was fine material. In over 98% of the sediment samples
from Enewetak lagoon, the ^"^'Am and ^^^ ^"^^Pu concentrations (pCi/g) associated
with the fine sediment components were greater than or equal to the concentrations
associated with the coarse fraction. The activity of ^^^ Am and 239+240p^ -^^ ^.j^^ ^^^^^
was 0.6 to more than 10 times that in the coarse fraction. These distributions between
size fractions are very unlike those encountered for fallout of ^ ''^''"^^^Pu in sediments in
Buzzards Bay, Mass., where the 2 3 9+2 4 op^ ^^^ ^^^ preferentially associated with the
fine fractions of sedimentary deposits (Bowen, Livingston, and Burke, 1976). This
difference is perhaps not unexpected because most of the transuranic inventory deposited
to the lagoon environment was probably associated with small particulate carbonates.
During the years since nuclear testing, some plutonium has exchanged slowly as a result
of chemical reactions with exposed surfaces of the larger sedimentary components.

The transuranic inventory at lagoon locations, however, is dependent on the local
abundance of the fine and coarse materials. Table 2 shows the ^"^^ Am concentrations in
the fine and coarse components of two core samples from midlagoon locations at
Enewetak. The fraction of the coarse components in the sediment column of core 6
decreases with depth and in core 1 increases with depth. The ^"^^Am concentration
associated with the fine fraction in the surface 2-cm section is 2 to 5 times the
concentration associated with the coarse fraction; but, because the fine material in the
surface layer of core. 6 accounts for only 25% of the total dry weight of the sediment
volume, 58% of the ^^^Am in the surface 2-cm layer is associated with the coarse
fraction. In core 1 , on the other hand, 95% of the total '^^^ Am in the surface 2-cm layer
is associated with the fine fraction. Although the ^'*^Am concentrations (pCi/g)
associated with the fine material at various depths in the sediment column exceed the
concentrations associated with the coarse components in both cores, the inventory of the
radionuclide (pCi/cm^) within any depth interval associated with the fine and coarse
components can be variable throughout the sediment column. Areal transuranic
distributions like those shown in Figs. 2 and 3 but associated with only the fine or only
the coarse component of sediment would differ.

The vertical distributions of the transuranics in the lagoon sediment are very complex.
No generalization about the shape of the concentration profile in any region can be made.
Table 2, for example, shows a ^'*^ Am peak associated wdth the fine components of core 6
at depths of 25 to 30 cm with Httle ^'^^ Am associated with the coarse components at
these depths. In core 1, the highest ^'^'Am concentrations are associated with the fine
components at depths of between 8 and 10 cm in the sediment column. The ^"^^Am
concentrations associated with the coarse component in both cores generally decrease
gradually with depth. Transuranic concentrations increase, decrease, or remain constant

590 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Americium-241 Associated with Components in
Core Samples of Sediment

Concentration,

pCi/g (dry weight)

Inventory, pCi/cm^

Radioactivity
associated with

Relative amount of

Fine

Coarse

Fine

Coarse

coarse component.

coarse fraction.

Depth, cm

fraction*

fraction

% (dry weight)

Core 6

0-2

5.97(23)

2.84(28)

0.57

0.78

57.8

74.6

2-4

4.99(7)

1.81(11)

1.54

0.99

39.1

64.0

4-6

6.44(5)

2.08(12)

2.68

0.95

26.1

52.4

6-8

5.51(6)

1.40(11)

2.88

0.63

17.9

46.4

8-10

3.10(6)

0.96(19)

1.74

0.39

18.3

41.6

10-15

0.72(18)

0.20(40)

0.38

0.09

19.1

45.9

15-20

<0.09

<0.05

48.6

20-25

5.83(7)

<0.09

3.75

<0.03

8.0

37.0

25-30

11.1(7)

0.16(42)

6.91

0.06

0.9

39.0

30-35

0.06

<0.08

<0.02

48.3

35-40

0.06

<0.04

<0.03

36.6

Core 1

0-2

42.5(3)

12.9(16)

43.4

2.2

4.8

14.6

2-4

30.4(6)

6.31(5)

24.7

3.7

13.0

41.8

4-6

34.8(5)

4.90(13)

26.1

2.3

8.1

38.0

6-8

41.1(3)

4.91(20)

33.7

2.4

6.6

37.2

8-10

51.0(3)

4.07(14)

37.6

2.4

6.0

44.4

10-15

25.2(3)

0.98(17)

10.5

0.7

6.3

62.8

15-19

Lost

0.36(17)

0.2

93.0

19-25

3.55(18)

0.23(34)

0.2

0.01

4.7

90.7

*The values in parentheses are the la counting errors expressed as the percent of the value listed.

with depth in sediment cores from other lagoon locations (Noshkin et al., 1978a). The
concentrations of ^^^ ^"^^Pu and ^^^ Am associated with the carbonate components in
four cores taken along a 1 .5-km transect across Mike and Koa craters are shown in Fig. 6.
The concentrations in the sediments from the Atolls' largest craters are surprisingly
nonhomogeneous. Turbulence and large-scale mixing of the sediments during and after
testing should have produced a much more uniform distribution than that found. The
2 39 24 0pjj concentration in the sediment column at station 17E is fairly uniform to a
depth of 50 cm. At station 16E, the concentration increases with depth to 35 cm. The
^'^^ Am concentration in the sediment column at station 16E decreases with depth. No
correlation is obvious between the ^"^^Am and 2 39+240p^ concentrations associated
with the components of these crater sediments. The craters should act as natural sediment
traps, but little sedimentation in the Mike and Koa craters has occurred since the bottom
depths were redetermined in 1964. In 1964 the maximum bottom depth of Mike crater
was 27.4 m below sea level (U. S. Atomic Energy Commission, 1973). We have found no
measurable change in the depth of the crater bottom during the period 1972 to 1977.
Only small quantities of resuspended or reef-generated particulate material are then
transported in the water masses to the western reef. Very little sedimentary material
therefore escapes from the lagoon, and any resuspended bottom material probably settles
out again on the lagoon floor close to its origin. The complex areal and vertical patterns

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 591

17E

239 + 240

241

Eastern
lip

16E 15E 14E

CRATER STATION IDENTIFICATION

Fig. 6 Vertical and areal distributions of ^ ^ '"""^ ''"Pu and ^ " ' Am activities in sediments
in Mike and Koa craters.

of transuranics detected in this relatively small region of the lagoon where the
distributions are expected to be more uniform are but examples of the complex patterns
in the lagoon.

Halimeda, shells, coral, and foraininifera fragments were sorted from the coarse
fraction of several sediment samples by hand. Table 3 shows the 239+240p|j concentra-
tions associated with each component in the surface layer from two locations in
Enewetak lagoon and at various depths in a core from Bikini lagoon. The ^^' ^"^^Pu
concentration associated with Halimeda fragments at station 40C only slightly exceeds
that in fragments from station 3D. The concentrations associated with the separated
foraminifera and coral fragments from station 40C are, however, at least 2.5 times higher
than those associated with their respective components at station 3D. The distribution of
2 39 2 40pjj j^ different among components in the sediment from different regions of the
lagoon. The fine fraction at these locations contained the highest concentration of
2 39 +2 40 p^^ To within our analytical precision, the ^^^Pu/^^^'^^'^^Pu concentration
ratio is identical in the components from both stations.

In the sediment column from Bikini station B3, 239+240p^ ^^^ associated with all
components that were separated. At all depths in the sediment column, the highest
concentration of 2 3 9+2 4 op^ -^^ ^.j^^ coarse components was associated with Halimeda
fragments. Sedimentation of labeled material to the lagoon occurs at a rate that is too
slow to account for the buried activities below a few millimeters in the sediment column.
Although the age of the Halimeda fragments, coral, and other components at depths
greater than a few centimeters must tlierefore predate the test years, 2 3 9+2 4 Op^ j^
associated with these components.

592 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Plutonium Concentrations Associated with
Sediment Components [pCi/g (dry weight)]

Plutonium concentration of ^^'^^""Pu [pCi/g(dry weight)] and
ratio of * Pu/^ ' ^ +^ " ° Pu at Enewetak Atoll (surface 2.5-cm layer)

MoUusk shells

Dead Halimeda fragments

Coral fragments

Fines «0.5 mm)

Foraminifera

Station 3D

Station 40C

2 3 9+240

0.64 ± 0.06
4.8 + 0.5
1.3 ±0.1

6.85 ± 0.05
1.1 ± 0.1

'Pll/^ 3 9+240

0.09 ± 0.03
0.08 ±0.01
0.08 ± 0.03
0.07 ± 0.01
0.07 ± 0.03

2 3 9+240

Pu/

2 3 9+240

Absent in sample

6.0 + 0.4 0.08 ± 0^01

10 ±5 0.13 + 0.07

23.5 ± 0.2 0.10 ±0.01

2.7 ±0.2 0.10 ±0.02

Plutonium concentrations at Bikini core Station B3,
pCi/g(dry weight)

Depth in sediment

Halimeda

Mollusk

Coral

column, cm

f ragmen ts

Foraminifera

shells

fragments

0-5

1.15 ±0.06

0.324 ±0.013

0.265 ± 0.013

0.101 ±0.004

5-10

0.95 ± 0.03

0.154 ±0.008

0.090 ± 0.004

0.130 ±0.004

10-15

0.62 ±0.01

0.093 ± 0.004

0.038 + 0.003

0.063 ± 0.004

15-20

0.33 ±0.01

0.024 ± 0.004

0.032 ± 0.002

0.073 ± 0.004

20-25

0.177 + 0.003

0.012 ±0.002

0.018 ±0.002

0.007 ± 0.002

25-30

0.013 + 0.001

0.002 ± 0.001

0.006 ± 0.001

0.001 ± 0.000

30-35

0.009 ± 0.001

0.001 ± 0.000

0.003 ± 0.001

0.008 ± 0.003

The possibility that subsurface remains labeled during testing were buried later in the
sediment column by large-scale turbulence can be discounted. Coral or Halimeda
fragments directly subject to a nuclear explosion probably would not retain their
identity. In recent yearly growth increments of a living sample of Favites virens coral
from station B3, the ^^^'''^'^^Pu concentrations averaged 104 ±12 pCi/kg (Noshkin
etal., 1975). Tliis value agrees well wdth the 239+240pjj concentrations in dead coral
remains in surface layers at station 33. hi no yearly growth increment from this coral
since 1954 was the 2 39+240py concentration below 104 ± 12 pCi/kg. Lower concentra-
tions are associated with coral remains deeper in the sediment column. From the
radiological record retained in the skeletal matrix of the Favites virens, coral labeled
during 1954 and 1958, for example, should have ^^^"""^"^^Pu concentrations of 39 x 10-^
and 4.5 x 10^ pCi/kg, respectively (Noshkin et al., 1975). These concentrations are
orders of magnitude larger than those in any subsurface coral remains. These data,
therefore, do not support the translocation of labeled coral material deeper into the
sediment column by physical processes during or after testing. Burrowing organisms could
redistribute some fraction of labeled sedimentary components to depths in the sediment
column. However, when the 2 3 9-t-240pjj activities associated with each component at
various depths are compared to the activity in the corresponding component at the
surface, the 239-i-240py activities differ. For example, between 5 and 10cm, the
2 39+240p^ concentrations associated with the coral, Halimeda, foraminifera, and shells
are, respectively, 1.29, 0.83, 0.43, and 0.34 times the concentrations associated with
those components in the surface layer. Burrowing and mixing processes by organisms are
not likely to move specific components selectively down through the sediment column.

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 593

The data indicate that all plutonium does not remain associated with the sedimentary
material with which it was originally deposited. Small quantities of plutonium are
remobilized continuously from the sediments to the lagoon water column by surface-
exchange mechanisms. Plutonium is also detected in the interstitial water extracted in situ
from sediments (Noslikin et al., 1978b) at concentrations higher than those in the
overlying bottom water at Enewetak Atoll. By equilibration, small quantities of
2 39+240p^ from the sediments are exchanged and released. Vertical diffusion moves the
radionuclides to the sediment water interface where the plutonium mixes with the lagoon
water mass. Remobilized plutonium can then be concentrated by members of the marine
food chain. Vertical diffusion can also move the exchanged plutonium in the interstitial
fluid deeper into the sediment colunin. Exchange of plutonium with exposed carbonate
surfaces might account for the concentrations associated with material deeper in the
sediment column.

Transuranic Elements Associated with the Calcareous Algae Halimeda

Debris from the calcareous algae Halimeda is the second most abundant component of
Enewetak lagoon sediments and covers an estimated 26% of the lagoon floor (Emery,
Tracy, and Ladd, 1954). Live species have recently been collected by divers and during
dredging operations from numerous locations at both Enewetak and Bikini. Because algae
were shown previously to concentrate plutonium (Noshkin, 1972), the role of this
benthic algae in recycling the transuranic elements at the Atoll should be assessed.

The mean concentration factor for ^^^ ^'^"Pu associated with algae species from
both atolls is 6 X 10"* and ranges from 1 x 10"^ to 32 X 10"^ (Noslikin et al.. 1978a). To
within the precision of our measurements, the concentration factors for plutonium at the
two atolls and of different Halimeda species from both atolls do not differ (Noshkin
etal., 1978a). Concentrations of ^^^ ^^°Pu associated with the hve algae ranged from
0.4 to 22 pCi/g (wet weight), and the concentrations in the water where the algae were
obtained ranged from 10 to 116 fCi/liter. Surface-sediment concentrations at the stations
(Noshkin etal., 1978a; Nevissi and Schell, 1975) were compared to the algae
concentrations at these sites. The average ratio of the 239-i-240py concentrations
associated with the Halimeda species [pCi/g (dry weight)] to that in the top 2.5-cm
sediment layer [pCi/g (dry weight)] was 0.24 ± 0.13, and the ^^^ Am concentration ratio
was 0.32 ± 0.24. Concentrations of 239-H240p^ ^^^ 241 ^^^ -^^ ^^^ sediment ranged from
9 to 82 pCi/g and from 1.1 to 67 pCi/g, respectively. On an equivalent weight basis, the
live benthic algae have lower 239-t-240p^ ^^^ ^"^^Am levels than sediments in the
immediate environment. The average plant/sediment concentration ratios of ^^^'''^'^'^Pu
and ^ ' Am are not statistically different. Thus there is no discrimination between
2 3 9-H2 40p^ and ^^^Am in processes beginning with remobilization of the transuranics
from the environment and ending with concentration by the algae.

Table 4 summarizes data on transuranic concentrations in algae, water, and sediment
from Cactus crater at Enewetak. The data show that the ^^^Pu/^^^'^^'*°Pu ratios in
plants, water, and sediment are identical. In this crater ecosystem marine algae do not
discriminate among the plutonium isotopes in the environment. The plant/sediment
concentration ratios of 2 3 9-^2 40p^ ^^^ ^^^ Am are nearly identical, wliich again shows
that the processes of environmental release and plant uptake of the two transuranics are
similar.

Halimeda monile*

Crater sediment

Crater water

2.53 ± 0.30

9.2 ± 2.0

18.68 ± 0.033

82 ±2

116±62fCi/Utert

0.54 ± 0.03

0.54 ± 0.02

0.53 ± 0.02

0.14 ± 0.02

0.11 ±0.02

594 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 4 Concentrations of Transuranium Radionuclides in
Halimeda Algae, Sediment, and Water from Cactus Crater

Radionuclides

^'" Am, pCi/g (dry weight)
2 3 9 +2 4 0 py^ pCi/g (dry weight)

2 3 8pii/2 3 9+240py

*"* Am (Halimeda /sediment)t 0.28

2 3 9+24 0py (Halimeda I sediment) t 0.23

Concentration factor for "'^^""Pu = 7.5 x lO'' [pCi/kg (wet weight) per pCi/kg H^O]

*Wet weight/dry weight - 2.13.

t Average concentration in eight water samples from the crater bottom collected between 1974
and 1977.

$ Values are expressed as pCi/g (dry weight) Halimeda/ pCi/g (dry weight) sediment.

The mean surface -sediment inventory of ^^^'''^'^"Pu at Enewetak is 249 Ci (Table 1),
The lagoon is 933 km^ in area, and the average specific gravity of the Halimeda and other
sediment components is 1.8g/cm^ (Emery, Tracy, and Ladd, 1954). Activities of
2 39 24 op^ associated with the algae are related to the activity in the surface sediment.
The mean wet/dry ratio of the Halimeda species is 2.3, and the average wet weight of the
plants, without holdfast, is 6.4 ± 3.8 g (Noshkin et al., 1978a). Therefore the average
2 39 2 40py concentration associated with tlie Uve Halimeda species at Enewetak is
0.62 pCi/g (wet weight). Approximately 4.0 pCi is associated with each plant. If the
number of Halimeda plants were known, the mean plutonium inventory associated with
the living Halimeda reservoir could be computed. Unfortunately no estimates of Halimeda
biomass at Enewetak are available. During the late 1940s, the mean sedimentation rate of
Halimeda at Bikini was estimated at 3.8 mm/yr (Emery, Tracy, and Ladd, 1954). If this
sedimentation rate is applicable to Enewetak Atoll, approximately 1 Ci of ^^^''"^'^^Pu is
deposited annually in the sediments in association with Halimeda detritus. This quantity
represents only 0.4% of the surface-sediment inventory and a smaller yet fraction of the
total inventory in the sediment column. If, however, the life-span of each plant is 1 yr,
for example, a quantity of ^^^^^^^Pu equivalent to half the present sediment inventory,
or 125 Ci, could be recycled with the algae in approximately 175 yr. Spies, Marsh, and
Colsher (1978) demonstrated that, when live Halimeda from Enewetak were cleaned and
treated with IN acetic acid, the acid-soluble fraction, or the carbonate material,
contained 58% of the total ^^^'''^'*°Pu, and 42% remained bound to the plant tissue. As
the plant decomposes after death, the organic material and associated radioelements are
released to the environment, leaving the skeletal carbonate matrix and its associated
transuranics in the sedimentary deposits. The transuranics associated with the organic
fraction released during decomposition are recycled to the benthic or pelagic environ-
ments. Over the long term the algae could play a key role in cycling the transuranics
between the sediments and the aqueous environment.

Plutonium Concentrations in the Lagoon Seawater

A considerable number of lagoon water samples have been collected for plutonium
analysis by this laboratory since 1972 (U. S. Atomic Energy Commission, 1973; Noshkin

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 595

et al., 1974; Noshkin et al,, 1976; Noshkin et al., 1978b). Several studies are in progress
at the Atoll which require data on concentrations in lagoon water; so the number of
samples and the locations sampled are predicated by the requirements of the current
program. Contours of ^^^"""^"^^Pu concentrations in the water show complex distribution
patterns (Noshkin et al., 1974) in various regions of the lagoon. The spatial patterns of
surface and bottom 2 3 9+24 Op^ concentrations in solution and in association with
filterable material are very different, as are the ^^^Pu/^^^'^^'*°Pu ratios in the water
mass. A detailed discussion of the plutonium levels in the pelagic environment of the
lagoon is in preparation (Noshkin et al., 1978b). Instead of relating all results from the
analysis of lagoon water samples collected since 1972 with hydrological, seasonal, or
spatial factors, I will summarize some of the data that are related to remobihzation and
redistribution of plutonium.

In 1972, 1974, and 1976, a sufficient number of water samples from the lagoon were
analyzed for 239+240p^ ^^ permit an estimate of mean concentrations in the lagoon. The
mean concentrations are summarized in Table 5. In 1972, the average 239+240p^
concentration in the lagoon was determined for 34 unfiltered surface and bottom
samples. A more-detailed water sampling program was conducted in 1974. In 1976, a
smaller number of water samples were collected around the perimeter of the lagoon 2 km
off the shore of the reef. Water samples collected during 1974 and 1976 were filtered
through l-jum filters. In the discussion to follow, the estimated average soluble
2 39+240 py concentrations shown in Table 5 refer to material passing through a 1-Mm
filter.

During July 1974, the soluble ^^^"""^"^^Pu in the lagoon water ranged in concentra-
tion from 2 to 75 fCi/liter. The percentage of the total activity associated with the
filterable material in the water samples during 1974 and 1976 ranged from 2% to 54%
and from 12% to 94%, respectively. The concentrations of plutonium radionucUdes in
solution above fallout background concentrations in the lagoon water are direct evidence
of the remobilization of trans uranics from the soUd phases of the environment. Dissolved
plutonium released from the sediments of Cactus crater was traced for considerable
distances along the reef by a plutonium radionuclide balance, which involved the change
in the 2 3 8py^239+240py j.^^-Q -^ ^j^g water, and dyes to trace the crater water (Noshkin
et al., 1978c). The dissolved plutonium moves in solution apparently without interacting
with the sediment deposits during transport. The dissolved plutonium passes readily
through dialysis membranes (Noshkin et al., 1978c). Equilibration between dissolved
plutonium in the crater seawater and low-activity seawater contained in dialysis bags is
achieved in 3 days (Noshkin et al., 1978c). These characteristics suggest that the
plutonium remobilized to the environmental waters has very solute-like characteristics. It
is tempting to suggest, considering the environment, that the remobilized chemical species
is some form of carbonate complex.

The average concentration of total 2 3 9+2 4 Op^^ ^ ^j^^ water was essentially the same
in 1972 and 1974, but a marked decrease was noted during 1976. In 1976, the average
concentration associated with the filterable material in the lagoon doubled over the mean
1974 level, and the mean soluble concentration was reduced to half. Forty percent fewer
samples were collected in 1976 than in 1974. During the 1974 program, samples were
taken at stations throughout the lagoon, whereas the 1976 samplings were restricted to
locations only 2 km from the reef. Similar 239+240py concentrations were found in
water samples from the few 1974 locations resampled in 1976, which suggests that any

596 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

computed mean concentration in the lagoon is contingent on the number and location of
samples. The mean ■^^^Pu/^^^"'"^'*°Pu ratios in the lagoon water samples in 1972, 1974,
and 1976 were virtually the same. Differences in sample ratios between soluble and
particulate phases were noted sometimes, but the average ratios associated with the two
phases from all stations were not significantly different.

During 1974 and 1976, 1.5 and 0.7 Ci of

239+240

Pu, respectively, were found in

solution, and 0.27 and 0.53 Ci, respectively, were associated with particulate material.
These latter quantities represent less than 0.2% of the plutonium inventory in the surface
sediment and less than 0.04% of the inventory estimated to a 16-cm depth in the
sediment column (Table 1). The average quantities of soluble plutonium in the water are
also small fractions of the sediment inventory. Therefore in recent years only small
fractions of the Atoll plutonium inventory are either remobilized to the solution phase or
resuspended to the water column.

During 1976, zooplankton samples contained less than l%ofthe'^^^ ^"^^Pu activity
in the total material filtered from an equivalent volume of water (Noshkin et al., 1978b).

TABLE 5 Plutonium Concentrations in the Water Column

at Enewetak Atoll

Sampling time

December 1972

July 1974

April 1976

Mean concentrations in lagoon water,
fCi /liter
Soluble (<1 jum)
Particulate Ol nm)

Toti

Mean

'Pu/

2 3 'i + 2 4 0

Water-column inventory, mCi/km^
Soluble
Particulate

Lagoon inventory, Ci
Soluble
Particulate

Water inventory, % of sediment inventory
Water inventory compared to top 2.5 cm
of sediment -surface inventory, %
Soluble
Particulate

Water inventory compared to top 16 cm
of sediment inventory, %
Soluble

Particulate

Total

35
6

0.12

0.13

1.65

0.29

0.76
0.57

1.94

1.33

1.54
0.27

0.71
0.53

1.70

0.68

1.81

0.62
0.11

0.73

0.13
0.023

1.24

0.29
0.21

0.50

0.056
0.045

Total

0.14

0.15

0.10

TRANSURANIC RADIONUCLIDES IN ENEWETAK LAGOON 597

The remaining 23 9 + 240p>^j -^^ ^^^^ particulate material is therefore associated with other
forms of suspended matter. Between 1960 and 1963, Johannes (1967) investigated the
composition of the suspended particles in the lagoon. Progressing from the eastern reef
toward the lagoon, suspended benthic algae and sediment particles became less abundant
with depth of the water as they settled to the bottom, and suspended macroscopic
organic aggregates, consisting largely of mucus released by coral, increased progressively
in size and number (Johannes, 1967). Often calcareous grains resuspended near the reef;
microorganisms, copepod fecal pellets, and other undifferentiated material were
incorporated with the aggregates. These materials and other particles produced in the
pelagic environment are the most important food components for lagoon zooplankton
and certain plankton-feeding fish (Gerber and Marshall, 1974). The small quantities of
plutonium ingested with tliis particulate debris are dispersed over the lagoon by these
organisms. Herbivorous fish play a role in the generation of particles in the water column
(Smith, 1973). These fish are not efficient assimilators; while satisfying their energetic
requirements, they disturb large quantities of material and release large amounts of
unassimilated material containing plutonium in their feces. Moriarty (1976) estimates
that a 200-g mullet, a species common to Enewetak, which feeds by scooping up bottom
material to sift and remove small algae, will pass 50 g of dry sediment through its gut per
day.

Bottom particles from the northwest quadrant of the lagoon, where higliest
plutonium concentrations in sediment are found, usually have high plutonium concentra-
tions, which indicates that a fraction of the plutonium in the particulate phase may
originate from turbulent resuspension of the sediment components in deep (60 m) water.
The resuspended material and associated plutonium are probably not transported for any
distance in the lagoon. Previous results indicate that the material is redeposited in the
same general area of its origin. Only a few of the variety of active processes capable of
generating and moving particulate plutonium in the water mass have been considered. It is
remarkable that these and other processes resuspend so little of the plutonium inventory.
Barring catastrophic events, the present distribution and inventory of plutonium in the
sediments will be only slightly altered during the years by relocation of labeled material
from other regions in the lagoon.

Laboratory studies with contaminated sediments and soils from Enewetak show that
plutonium is rapidly partitioned between the solid phase and solution and reaches
equilibrium after several days with an average distribution coefficient for plutonium of
1.8 X 10^. Table 6 shows this and some recent determinations of the distribution
coefficient for plutonium in laboratory and field experiments using a variety of
sediments. Considering the difference in the types of environmental samples represented
in Table 6, it is striking that the K^j for plutonium differs so little.

Table 1 shows that the mean plutonium inventory associated with the sediment
components in the top 2.5 cm at Enewetak is 249 Ci. The lagoon sediment has an average
density of 1 .8 g/cm^ (Emery, Tracy, and Ladd, 1954) and occupies an area of 933 km^.
The mean depth of the lagoon is 47.4 m. With these data and the K^ for plutonium of
1.8 X 10^, one can construct a simple model to predict the average concentration
expected in the lagoon water by assuming that the plutonium in solution is in equilibrium
with that in the sediments. At any time the amount of plutonium in solution is limited by
the saturation of the solution under equilibrium conditions. The rate at which water and
its dissolved plutonium is flushed from the lagoon is balanced by input of uncon-
taminated ocean water, which is rapidly saturated with remobilized plutonium from the

598 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 6 Recent Determinations of the Distribution
Coefficient for Plutonium

Sediment type

Range

Average*

Reference

Enewetak, coral soil

and sediment (laboratory

desorption study)
Enewetak, groundwater particulates
Trombay Harbour, suspended silt
Bikini Tewa crater sediment

(laboratory desorption study,

oxic-anoxic conditions)
Windscale area, 5% clay, 50% silt,

45% sand
Humboldt Bay, Calif., suspended

clay-silt particulates
Lake Michigan, suspended

particulates
Mediterranean sediment

(laboratory sorption study)

4 X lO^-S X 10'

1.4 X lO'-lO*
4.8 X 10''-1.3 X 10'
4 X 10''-4 X 10'

0.6 X 10^-22 X 10"

4.7 X lO^-ll.S X 10"

1.8x10' Unpublished

2.5x10' Noshkinetal., 1976
0.9 X 10' PiUai and Mathew, 1976
2.2x10' Moand Lowman, 1976

0.5 X 10' Hetherington, Jefferies,
and Lovett, 1975

0.8 X 10' UnpubUshed data, this
laboratory

3.0x10' Wahlgrenetal., 1976

1.3 X 10" -9.4 X 10" 0.5x10' Duursma and Parsi, 1974

Mean 1.4 x 10' cm^/g

*Quantity of 2 3 9+24opy ^Qund to the sediment per unit dry weight of sediment divided by the
amount of ^ ^ ' """^ " ° Pu in water per cubic centimeter.

Atoll source terms. If plutonium is cycled through an intermediate host, such as the
Halimeda, the rate at which it is released from decaying plants must be balanced by
uptake in the new growth to maintain a steady state condition. Given that steady state
conditions exist, the mean plutonium inventory in the lagoon water and the concentra-
tion expected in solution computed from the basic equation relating K^ to water and
sediment concentrations are 1 .4 Ci and 32 fCi/liter, respectively. There is general
agreement between the average quantity of ^^^''"^'^"Pu predicted and that measured in
solution (see Table 5). In 1976, the computed value differed from the measured mean
soluble concentration by a factor of 2. Although this is not a large discrepancy, the
average concentration, as was mentioned previously, probably does not represent the real
mean for the lagoon at the time sampled. From the appropriate dimensions for the Bikini
lagoon, the sediment data in Table 1 , and the Kj for 2 39+24 Op^^ ^j^^ average inventory in
the water column and the concentration at Bikini are computed to be 1 .7 Ci and
60 fCi/hter, respectively. During December 1972, the mean soluble 239+240p^ inventory
and the concentration in the lagoon water were 1.2 Ci and 42 ± 21 fCi/Uter, respectively
(Noshkin et al., 1974), and in January 1977 the respective values were 1 .4 Ci and 49 ± 21
fCi/liter (Noshkin et al., 1978b). These average values also are consistent with the
amounts predicted.

For many reasons it may be argued that some of this agreement is fortuitous.
Nevertheless, the general agreement found between computed and twice-measured
average concentrations in both lagoons between 1972 and 1977 demonstrates the general
usefulness of this simple equiUbrium model in predicting long-term average concentra-
tions in lagoon water. From radiological records retained in yearly growth of coral

TRANS URANIC RADIONUCLIDES IN ENEWETAK LAGOON 599

sections (Noshkin et al., 1975; Noshkin et al., 1978a), Bikini and Enewetak lagoon water
along with dissolved species are estimated to be exchanged approximately twice per year.
At this rate of exchange under steady-state conditions, shghtly more than 250 yr will be
required to reduce the plutonium inventory in the sediment by 50%. The rates of the
mobilization and migration processes of plutonium away from the Atoll to the equatorial
Pacific waters are much faster than the rate of radioactive decay. These figures and results
should be considered when the consequences of disposal methods for transuranic wastes
to the oceans are discussed.

Some massive corals collected from the atolls contain well-defined growth bands
dating from the collection time to the early 1950s. Each yearly growth concentrates
plutonium in proportion to the levels in the environment (Noshkin et al., 1975; Noshkin
et al., 1978a). Concentrations of 239+240p^ associated with grov^h increments dated
since 1965 in three Enewetak corals from different locations in the lagoon and one Bikini
lagoon sample are given in Table 7. The average amount of plutonium concentrated by
the coral from 1965 until the year of collection has been computed and is shown in
Table 7.

The average absolute concentrations in the corals are different, as expected, and
reflect the local environmental concentrations in the region. In only a few growth
sections are the ^^^ ^'^^Pu concentrations different from the mean by more than a
factor of 2, and only corals 1 and 2 show this magnitude of variation. Corals 1 and 2 were
obtained in the water on the lagoon side of the eastern reef. The patterns of current in

TABLE 7 Concentrations of ^ ^ ^ ^ ^ ^ ^ Pu in Yearly Growth Sections of

Enewetak and Bikini Coral

3 9 +2 -. opy concentrations, fCi/g (dry weight)

Enewetak

Coral 2

Bikini,

Coral 1

(Gonrastrea

Coral 3

Coral 4

(Fa via pallida)

retiformisj

(Favia pallida)

(Favites virens)

Year of growth

section

1974

7.4(13)*

1973

3.9(12)*

19.6(5)*

5.2(14)

1972

2.3(9)

35.5(4)

6.7(13)

130(7)*

1971

0.9(25)

13.5(6)

7.0(10)

130(7)

1970

1.3(14)

5.0(12)

6.3(12)

100(5)

1969

4.9(10)

41.7(4)

5.9(11)

100(4)

1968

7.9(8)

10.4(6)

6.9(10)

100(2)

1967

3.6(6)

12.9(7)

Lost

100(5)

1966

5.1(11)

11.2(7)

3.7(13)

90(5)

1965

3.2(9)

10.5(6)

6.0(8)

110(5)

Average concentration

3.7 ±2.1

17.8 ± 12.5

6.1 ± 1.1

108 ± 15

(1965 to year of

collection)

Date of collection

October 73

April 74

August 74

November 72

♦Values in parentheses are Ict counting errors expressed as the percent of the value listed.

600 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

this region of the lagoon are variable, and the windward reef community contributes a
significant detrital load with associated plutonium to the lagoon. Since growmg coral is a
point source in the environment, small changes in even the local circulation, to name one
of many factors, will greatly alter the plutonium concentration in the vicinity of the
coral. It is rather more surprising that, for the most part, the ^'^^ ^'*°Pu levels associated
with the last 9 yr of growth are nearly constant. This shows that the dissolved 2 3 9+2 4 Op^
levels available to the corals in a specific region have also been similar during the last 9 yr.
These results from coral and other studies demonstrate that Enewetak lagoon has
attained a chemical steady-state condition with respect to plutonium remobilization from
solid components to solution. Not only will the simple equiUbrium model explain average
concentrations in lagoon water but also it can be used to estimate local concentrations
expected in the waters from areas of the Atoll with different levels of contamination. By
using appropriate concentration factors for plutonium, one can estimate the quantities
accumulated by marine organisms anywhere in the lagoon. The data on biotic
concentration can be used to estimate the potential dose to man if part or all of the Atoll
were to supply his marine food requirements.

Acknowledgments

I wish to express my appreciation to several coworkers, K.Wong, R.Eagle, R. Spies,
K. Marsh, T. Jokela, J. Brunk, G. HoUaday, and L. Nelson, who provided much of the
previously unreported data and without whose efforts in the field our program in the
Marshall Islands could not be carried out. This work is supported by the Division of
Biology and Environmental Research of the U. S. Department of Energy, contract No.
W-7405-ENG-48.

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Algae, Halimeda. Linmol. Oceanogr. (in preparation).
U. S. Atomic Energy Commission, 1973, Enewetak. Radiological Survey. USAEC Report NVO-140,

Vols. I-III, Nevada Operations Office, NTIS.
Wahlgren, M. A., J. J. Alberts, D. M. Nelson, and K. A. Orlandini, 1976, Study of the Behaviour of

Transuranics and Possible Chemical Homologues in Lake Michigan Water and Biota, in

Transuranium Nuclides in the E)ivironnicnt. Symposium Proceedings. San IVancisco, 1975, p. 9,

ST1/PUB/41U, International Atomic Energy Agency, Vienna.

Plutonium and Americium Behavior

in the Savannah River Marine Environment

D. W. HAYES and J. H. HORTON

77?^ 23 9,240^ ^^^j '^^^Am concentrations in the Savannah River are about the same as
those measured in other U. S. rivers (0.25 fCi/liter and 0.05 ± 0.05 fCi /liter, respectively).
j^'23 9, p^^ 1^^ ^^^ Savannah River originated from the watershed ^3 9,2 4 0^^^ inventory,
then the net annual removal from the watershed represents 0.005% per year. This
indicates that thousands of years will likely pass before all the plutonium is displaced to
the river. Tfie plutonium and americium concentrations (highest, -^50 fCi/g (dry
weight)/ in sediment cores from the tidal freshwater and near the mouth of the
Savannah River estuary were comparable and are not greatly different from those of
other locations in which the only source is nuclear fallout from nuclear weapons testing.
The transuranic activity in these sediments represents less than 10% of the gross alpha
activity from the natural radionuclides that are present. Current concentrations of
plutonium and americium in seafoods make only a very minor contribution to the overall
dose commitment to humans.

Thirteen power reactors, two fuel fabrication plants, and a U. S. Department of Energy
nuclear production complex, the Savannali River Plant (SRP), are operating on rivers or
in coastal regions of tlie soutlieastern United States. Rivers and estuaries are a major
geographic feature of this region and can represent both transport paths and sinks for
transuranics. Studies are in progress to establish the distribution and transport properties
of transuranic elements in the rivers and estuaries of this region. Of particular interest is
tlie Savannali River and its estuary because located on the watershed are tlie Savannah
River Plant and tliree commercial power reactors. These facilities make tlie Savannali
River watershed one of the most intensively developed nuclear watersheds in the United
States. The SRP consists of three production reactors, two fuel separation plants, a fuel
fabrication facility, and a heavy water plant. The SRP has been in operation since 1952,
whereas die three power reactors located at the headwaters of the Savannali River have
operated for less than 10 yr. Included in this chapter are estimates of watershed loss rates
for plutonium, measurements of plutonium and americium concentrations in the water
and sediments of tlie Savannah River and its estuary, estimates of plutonium
concentrations in seafood, and dose-rate estimates for seafood ingestion.

Savannah River Basin and Estuary

The Savannali River basin has a surface area of 27,400 km^. It can be divided into three
physiographic provinces that transect the basin (Fig. 1). The Blue Ridge Mountains
include portions of North CaroUna, South Carohna, and Georgia and range in elevation

602

Pu AND Am IN SAVANNAH RIVER MARINE ENVIRONMENT 603

BLUE RIDGE
MOUNTAINS

THREE REACTORS
(DUKE POWER COMPANY)

BURTON
LAKE

PIEDMONT PLATEAU

ONE REACTOR

(GEORGIA POWER COMPANY)

UNDER CONSTRUCTION

SAVANNAH RIVER
PLANT

COASTAL PLAIN

RIVER SAMPLING
LOCATION,
HIGHWAY 301

SAVANNAH

ATLANTIC
OCEAN

Fig. 1 Map of the Savannah River basin in the southeastern United States showing
physiographic provinces, major reservoirs, nuclear power plants, and sampling points.

from 5500 ft at the headwaters to about 1000 ft at the Piedmont plateau. The hilly
plateau descends from 1000 ft to about 200 ft near Augusta, Ga. The gently rolling
(upper) to nearly level (lower) Atlantic Coastal Plains extend from Augusta, Ga., to the
ocean.

Of the 16 rivers in the southeastern United States with flows greater than 30 m^/sec
(1000 cfs), the Savannah River is fourth in volume flow with an average of about
340 m^ /sec (12,000 cfs). The flow is regulated by two large impoundments, the Clark

604 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Hill and Hartwell reservoirs, located in the Piedmont region (Fig. 1). Each reservoir has a
capacity exceeding 3 km^ and can contain the equivalent of one-half to one year's flow
for the region of ttie river where it is located.

The Savannah River estuary is relatively narrow (about 0.5 km) and is maintained at a
minimum depth of 11 m throughout its length of 35 km to accommodate shipping. To
maintain the depth of 1 1 m in the harbor requires practically continual dredging, and the
dredge materials are dumped on adjacent areas to the north of the harbor. The harbor
region has a tidal range of 2.1 to 2.4 m. The estuary classification is that of a moderately
stratified one. The estuary has been heavily polluted with raw sewage and industrial
waste, but these pollutants have been reduced considerably in the last few years.

Sampling

Water, sediment, and seafood samples from the Savannah River and its estuary were
collected and analyzed to permit transport, inventory, and dose-to-man calculations. The
location of the sampling station for plutonium transport in the Savannali River watershed
is shown in Fig. 1. This location was chosen because the sampler could be easily located
to take water samples near midstream. Montlily composite samples were collected.
Sample bias was avoided by taking four samples per day, about 300 cm^ each, from a
depth of 1 m with an automatic compositing sampler.

Estuary water and sediment sampling locations are shown in Fig. 2. Sediment was
collected in marshes where vegetation lended stability to the sediments. The sediments
were collected by inserting a 3.6-cm-diameter core barrel into the sediments. The cores
were then extruded, sectioned, and bottled. Estuary water samples were 50-liter grab
samples from a depth of 1 m.

Oysters and crab meat were obtained from a local wholesaler, who obtained the
oysters from Wassaw Island and the crabs from crab pots located in Wassaw Sound
(Fig. 2). Clams were collected from Port Royal Sound, which is about 32 km north of the
mouth of the Savannah River estuary. Shad were netted in the Savannah River, and
mullet and speckled trout were obtained from a local wholesaler whose boats work in the
Savannah River estuary and nearby waters.

Analytical Methods

The procedure developed by Wong, Brown, and Noshkin (1978) for concentrating
plutonium from large volumes of water was adapted for use on these water samples for
both plutonium and americium analyses. In the modified method the 50 liters of water in
the drum was adjusted to pH 2 with hydrochloric acid. Plutonium-236 and americium-
243 spikes were added, and the sample was equilibrated for 7 to 10 days. At the end of
the equilibrium period, 40 cn^^ of saturated potassium permanganate was added and the
pH was adjusted to 8 with sodium hydroxide. The potassium permanganate was reduced
by using a slight excess of sodium bisulfite. The hydrated manganese oxide was collected
on a \-iJim cotton filter by continually recirculating the sample through the filter at
12 liters/min for 25 min. Recirculation had the advantage of keeping the water vigorously
stirred as it was continually passed through the manganese oxide bed being collected on
the filter. The samples were ashed while wet to avoid rapid combustion. The plutonium
analyses were performed according to a procedure developed by Butler (1965) and
Sanders and Leidt (1961) or by the LFE Laboratories, Richmond, Calif. All americium
analyses were done by the LFE Laboratories.

Pu AND Am IN SAVANNAH RIVER MARINE ENVIRONMENT 605

STATE

GEORGIA

FRESHWATER

STATE

SOUTH CAROLINA

FORT PULASKI

BOTTOM HILTON HEAD

NEAR MOUTH ViSLAND
5

SAND

SAVANNAH BEACH

5 km

ATLANTIC OCEAN

Fig. 2 Map of the Savannah River estuary showing water and sediment sampling
locations.

Sources of Transuranics

The Savannah River receives transuranics by direct deposition of fallout from nuclear
weapons tests, watershed runoff, and discharges from nuclear facihties. In addition to
receiving the transuranics by deposition of fallout from nuclear weapons tests, watershed
runoff, and discharges from nuclear facilities, the estuary also receives transuranics from
the ocean via the movement of salt water some 35 km up the estuary.

Estimates of the total amount of transuranics deposited on the watershed from
nuclear weapons tests are based on analyses of soil cores in the southeastern United
States. These estimates range from 1.5 to about 2.2 mCi/km^ with 2 3 8p^|23 9,24 0p^
ratios of about 0.04 to 0.18 (McLendon, 1975). If fallout deposition is uniform over the
Savannah River watershed, then the inventory is approximately 55 Ci, of which about
1.5 Ci was deposited on the water impoundments. No americium data are available for
estimating its inventory in the southeastern United States.

Data on plutonium releases from the SRP, located 256 km from the mouth of the
Savannah River, are available (Ashley and Zeigler, 1975). Atmospheric releases have
totaled about 3.7 Ci since fuel reprocessing operations began in 1954. Of the 3.7 Ci,

606 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

about 2 Ci was released in 1955, which was before the installation of high-efficiency
filters on the air exhaust system, and about 0.8 Ci in 1969 when a sand filter failed.
Currently, atmospheric releases average about 10 mCi/yr. Most of the plutonium from the
SRP operations is probably on site because analyses of soil cores from the plant perimeter
and off-site soils have about the same concentration, 1.96 ±0.7 mCi/km^ at the plant
perimeter and 1.81 ±0.58 mCi/km^ at 160 km; other values at the same latitude are
about 2 mCi/km^ .

Savannah River Plant plutonium releases to surface water were estimated to be about
0.3 Ci for the 20-yr period from 1954 to 1974 and were fairly consistent over this
interval (Hayes, LeRoy, and Cross, 1976). Until 1971 plutonium releases were estimated
by measuring gross alpha only and by assuming that all alpha activity was from
plutonium. Since then plutonium has been analyzed for specifically. The water after
release .is subjected to cleanup by on-site streams (about 16 km in length), an on-site
swamp, and the Savannah River before it reaches the estuary.

The fate of ^ ^ ^Cs released to surface water by the SRP has been extensively studied
and can be used to estimate the fate of plutonium released to surface waters. Plutonium
and cesium have similar transport properties in most environmental systems owing to
their strong association with the very fine suspended solids in stream water and with
stream-bed sediments (Simpson et al., 1976). Some 500 Ci of ^^ ^Cs has been discharged
to effluent streams, and only 90 Ci (about 18%) has been measured at Highway 301
(Fig. 1). About 58% of the 500 Ci of ^ ^ '^Cs that has been discharged is estimated to have
been deposited in the SRP streams before reaching the on-site swamp, and the swamp is
estimated to contain about 120 Ci, or about 24%, of cesium that has been discharged
(Marter, 1974). If these ^^^Cs data are extrapolated to plutonium, about 0.054 Ci of
plutonium is estimated to have left the SRP site since start-up.

From the above data, the total amount of plutonium on the watershed is estimated to
be about 59 Ci.

Results and Discussion

Savannah River

Plutonium concentrations in tlie Savannali River water are lower than would be predicted
by considering the concentration in other fresliwaters. Concentrations measured for 3
months in the Savannali River at Higliway 301 (Fig. 1) varied from 0.13 to 0.32 fCi/liter.
In comparison. Lake Michigan contains 0.6 fCi/liter (Edgington et al., 1976), the Great
Miami River in Oliio contains about 1 fCi/liter, and the Neuse and Newport rivers in
North Carolina contain abou: 1.2 and 1.7 fCi/liter, respectively (Hayes, LeRoy, and
Cross, 1976). Concentrations in the Savannali River appear to be greatly influenced by
reservoirs. Sedimentation in the two large reservoirs on the Savannali River sliould remove
all except some of the very small plutonium-bearing particles. About 73% of the
Savannah River flow originates above Clark Hill Dam. Erosion is greater in the hilly
Piedmont region above the reservoirs than in the coastal plains; so removal of
plutonium-bearing particles from the waterslied below the reservoirs would be less rapid
than that above the reservoirs. In contrast, the Great Miami River of Ohio and the Neuse
and Newport rivers of North Carolina do not have large water impoundments on them.
Also, the percentage of the Savannali River watershed that is under cultivation is only
one-third as large as that of the Great Miami River.

Pu AND Am IN SAVANNAH RIVER MARINE ENVIRONMENT 607

The calculated rate of plutonium removal from the Savannali River watershed is
about 10% of that from the Great Miami River. On the basis of the 3-month average
plutonium concentration and measured flow rates (see Table 1 ), the estimated plutonium
transport in the Savannali River at Highway 301 is 0.22 mCi/month, or 2.6 mCi/yr. The
area of the Savannah River watershed above Higliway 301 is 81% of the total watershed.
So the amount of plutonium in the watershed above the sampling point is 0.81 x 55 Ci
from nuclear weapons fallout plus 4 Ci released by the SRP, or a total of 48.6 Ci. Thus
the annual removal rate is approximately 0.005%. The value reported for the 1400-km^
watershed of tlie Great Miami River is 0.05% (Sprugel and Bartelt, 1978).

TABLE 1 Plutonium Transport in the Savannah River

2 39,2 4 0py

Sampling

transport.

period

River flow,

i39,J40py *

mCi/sampling

(1976)

liters

/period

fCi/liter

period

6/8 to 7/7

1.25

X 10'^

0.13 ±

0.09

0.16

7/7 to 8/4

1.20

X 10'^

0.27 ±

0.08

0.32

8/5 to 9/7

7.31

X 10"

0.26 ±

0.11

0.19

Average

1.06

xlO'^

0.22

*Mean ± standard error.

Information concerning the transport and fate of americium in rivers and estuaries is
limited. The concentration of ^''^ Am in Savannah River water has not been accurately
determined; a few samples collected at Highway 301 indicate that it is about
0.05 fCi/liter, as compared with 2 3 9,2 4 op^ concentration of 0.25 fCi/liter. The ^"^^ Am
concentration is the same as that in the Mediterranean Sea (Fukai, Bullestra, and Holm,
1976) and Lake Michigan (Wahlgren et al., 1976) water, where ^"^^ Am concentration is
3 to 5% of the 2 3 9,240p^ concentration. If the same percentage existed in Savannah
River water, the concentration of ^"^ ' Am would be only about 0.01 fCi/liter.

Savannah River Estuary

The plutonium concentrations in tlie Savannah River estuary are not much different from
those in other estuaries in the southeastern United States; in fact, the concentrations in
this estuary are lower than those in some others. Water concentrations of ^^''^"^"Pu were
determined in the Neuse and Newport river estuaries of North Carolina for comparison
v^dth concentrations in the Savannali River estuary. The results (Fig. 3) show that the
concentrations in the Newport River estuary are about three times as great as those in the
Neuse River or Savannali River estuaries, which are about equal.

The three estuaries and the rivers supplying them are quite different. The Neuse and
Savannah rivers flow through both the Piedmont plateau and Atlantic Coastal Plains.
Only the Savannah River has its flow regulated by reservoirs. The volume of flow in the
Savannah River is about twice that of the Neuse River and 10 times that of the Newport
River. The Newport River estuary is extremely small and shallow witli depths of less than
1 m at mean low water as compared with at least 4 m in the other two. Suspended soHd
(5-/im fraction) concentrations in the Newport River estuary are about one and one-half

608 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

"I r

Newport River estuary

Savannah River estuary

10
SALINITY, %o

Fig. 3 Total plutonium concentrations in the waters of three southeastern U. S.
estuaries.

times as great as those in the Savannah River or Neuse River estuaries (Hayes, LeRoy, and
Cross, 1976); this may be due to shallow water in the Newport River estuary, which
could resuspend bottom sediments tlirougliout its depth. These sediments are likely to be
very fine since the Newport River flows entirely in the coastal plains where the slope is
small. The higher plutonium concentration in the Newport River estuary could be due to
the larger quantity of suspended solids.

Within the Savannali River estuary, the plutonium concentrations in the sediment
from the tidal freshwater region and near the mouth of the estuary were comparable
(Table 2). The values are not greatly different from those of otlier locations that received
transuranic input from nuclear weapons fallout only. Plutonium concentrations up to
200 fCi/g have been reported for the Great Lakes sediments (Edgington et al., 1976);
about 60 to 70 fCi/g for Atlantic coastal waters, e.g., Buzzards Bay (Livingston and
Bowen, 1975); and about 10 to 30 fCi/g for the Savannah River system (Hayes, LeRoy,
and Cross, 1976). Fallout 2 3 8p^/2 3 9,2 4 0p^ j.^^jq5 ^^^ generally less than 0.1. Ratios

Pii AND Am IN SAVANNAH RIVER MARINE ENVIRONMENT 609

TABLE 2 Plutonium, Americium, and Gross Alpha Activities (dry-weight basis)
in the Savannah River Estuary Sediments

Core

depth

2 3Spy

^"'Am

Gross

interval.

2 3 8 py *

239,240py «

2 3 9 ,2 4 0 Pj,

^^•Am.*

2 3 9 ,2 4 0 py

alpha,

L<ication

fCi/g

ratio*

fCi/g

ratio*

fCi/g

Tidal

0-5

4.3 + 1.3

27.2 ± 3.3

0.16 ± 0.05

11.5 ± 4.1

0.42 ± 0.16

24,000

freshwater

5-15

6.1 ± 1.2

35.5 ± 2.8

0.17 ± 0.04

4.0 ± 1.9

0.11 ± 0.05

20,000

15-30

2.8 ± 1.3

30.9 ± 2.8

0.09 ± 0.04

5.4 ± 1.9

0.17 ± 0.06

20,000

30-50

0.4 ± 0.4

10.6 ± 1.0

0.04 ± 0.04

2.0 ± 0.8

0.19 ± 0.08

18,000

50-70

0.05 ± 0.10

0.05 ± 0.05

0.23 ± 0.23

18,000

Mouth of

0-5

3.2 ± 1.1

50.6 ± 4.1

0.06 ± 0.02

11.1 ± 1.7

0.22 ± 0.04

10,000

estuary

5-15

1.7 ± 1.0

21.6 ± 2.2

0.08 ± 0.05

3.0 ± 2.3

0.14 ± 0.11

13,000

15-25

0.2 ± 0.2

2.9 ± 0.5

0.07 ± 0.07

1.8 ± 0.6

0.62 ± 0.23

12,000

45-65

0.5 ± 0.5

13,000

*Mean + standard error.

greater than this are usually indicative of other sources of plutonium in the system. The
ratios for the Savannali River estuary cores are reported in Table 2, and only in the
freshwater core in the upper 0 to 15 cm were ratios found to be different from those
from fallout. These ratios were about a factor of 2 greater than fallout ratios and
presumably resulted from the SRP releases to the river system.

If americium dynamics in estuaries are different from those in freshwater or
seawater systems, this difference would be evidenced by tlie ^^ ' Am/'^^^''^'*^Pu ratios.
The average value for such ratios in shallow near-shore sediments (Livingston and Bowen,
1975) and in Lake Michigan sediments (Edgington et al., 1976) varies from 0.14 to
0.34, with an average of 0.22, and no fractionation between americium and plutonium
has been found in these sediments, even when the radionuclides are being lost from tlie
sediment following upward migration (Livingston and Bowen, 1975). Except for one
value of 0.62, the ^^' Am/'^'^^^'^'^^Pu ratios for two sediment cores reported in Table 2
are not significantly different from those quoted in the literature. The indication is that
tlie chemistries of americium and plutonium are similar in diis estuarine sy stein and that
tire ^'^ ' Am has grown in from ^'^ ' Pu.

The transuranic alpha activity in these cores represents less than 1% of the gross alpha
activity from the natural radionuclides that are present. Indeed, modern civilization's
impact on the alpha activity of these cores is small compared with die natural
background.

Seafoods

At present plutonium levels, seafoods make a very minor contribution to the overall
radiation-dose comnntment to the populations in the southeast. Seafood samples that
represent all trophic levels consumed by people in the southeastern United States were
collected in and near the Savannali River estuary and analyzed for plutonium. The
plutonium concentrations decreased as expected from the molluscs to the fish, with die
oyster having the higliest concentration. 0.12pCi/kg, coinpared with 0.001 pCi/kg for
sliad (Table 3). The 50-yr bone-dose commitment from consuming 5.9 kg of oysters per

610 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 Southeastern Seafood Plutonium
Dose Commitments

pCi/kg

50-yr bone-dose

(wet weight)*

commitment,* mrem

Oysters

0.12

5.5 xlO""

Clams

0.05

2.3x10-*

Crabs

0.007

2.4 X 10-'

Mullet

0.005

5 X 10-=

Speckled trout

0.004

3.9 X 10-'

Shad

0.001

1.3x10-'

* Consumption assumed for the dose calculation,
5.9 kg/yr molluscs, 11.8 kg/yr fish.

year is less than 0.0004% of tlie annual background radiation dose of about 120 mrem
which is received by man in tlie southeastern United States.

Conclusions

Nuclear facihties operating on the Savannali River waterslied have contributed less than
10% as much plutonium to the waterslied as has nuclear weapons fallout. Transport of
plutonium from the watershed to the estuary is very slow and appears to be influenced by
two large reservoirs that serve as sinks for suspended plutonium-bearing particles.
Consequently plutonium concentrations in Savannah River water and estuary sediments
are no higher — and in some cases are much lower — than plutonium concentrations in
other rivers and estuaries on which there are no nuclear facilities.

At present plutonium levels, seafoods make a very minor contribution to the overall
dose commitment to the population in tlie soutlieastern United States.

Acknowledgment

The information contained in diis chapter was developed during the course of work under
contract No. AT(07-2>1 witli tlie U. S. Department of Energy.

References

Ashley, C. A., and C. Zeigler, 1975, Releases of Radioactivity at the Savannah River Plant, 1954
Through 1975, ERDA Report DPSPU-75-25-1, E. I. du Pont de Nemours & Company, NTIS.

Butler, F. E., 1965, Determination of Uranium and Americium-Curium in Urine by Liquid Ion
Exchange, ^«fl/. Chem., 37: 340.

Edgington, D. N., J. J. Alberts, M. A. Wahlgren, J. O. Karttunen, and C. A. Reeve, 1976, Plutonium
and Americium in Lake Michigan Sediments, in Transuranium Nuclides in the Environment,
Symposium Proceedings, San Francisco, 1975, p. 493, STI/PUB/410, International Atomic Energy
Agency, Vienna.

Fukai, R., S. Bullestra, and E. Holm, 1976, ^"'Am in Mediterranean Surface Waters, TVarM re, 264:
739.

Hayes, D. W., J. H. LeRoy, and F. A. Cross, 1976, Plutonium in Atlantic Coastal Estuaries in the
Southeastern United States, in Transuranium Nuclides in the Environment, Symposium Proceed-
ings, San Francisco, 1975, p. 79, STI/PUB/410, International Atomic Energy Agency, Vienna.

Livingston, H. D., and V. T. Bowen, 1975, Americium in the Marine Environment — Relationships to
Plutonium, in Environmental Toxicity of Aquatic Radionuclides: Models and Mechanisms.
Proceedings of the International Conference on Environmental Toxicology, Rochester, N. Y., June

Pu AND Am IN SAVANNAH RIVER MARINE ENVIRONMENT 611

2-4, 1975, M. W. Miller and J. W. Stannaid (Eds.), Ann Arbor Science Publishers, Ann Arbor,

Mich.
Marter, W. L., 1974, Radioactivity from SRP Operations in a Downstream Savannah River Swamp,

USAEC Report DP-1370, E. I. du Pont de Nemours & Company, NTIS.
McLendon, H. R., 1975, Soil Monitoring for Plutonium at the Savannah River Plsnit, Health Phys., 28:

347.
Sanders, S. M., and S. C. Leidt, 1961, A New Procedure for Plutonium Analysis, Health Phys., 6: 189.
Simpson, H. J., C. R. Olsen, R. M. Trier, and S. C. Williams, 1976, Man-Made Radionuclides and

Sedimentation in the Hudson River Estuary, Science, 194: 179.
Sprugel, D. G., and G. E. Bartelt, 1978, Erosional Removal of Fallout Plutonium from a Large

Midwestern Watershed,/. Environ. Qual., 7: 175.
Wahlgren, M. A., J. J. Alberts, D. M. Nelson, and K. A. Orlandini, 1976, Study of the Behavior of

Transuranics and Possible Chemical Homologues in Lake Michigan Water and Biota, in

Transuranium Nuclides in the Environment, Symposium Proceedings, San Francisco, 1975, p. 9,

STI/PUB/410, International Atomic Energy Agency, Vienna.
Wong, K. M., G. S. Brown, and V. E. Noshkin, 1978, A Rapid Procedure for Plutonium Separation in

Large Volumes of Fresh and Saline Water by Manganese Dioxide Coprecipitation, /. Radioanal.

Oiem., 42(1): 7-15.

Patterns of Transuranic Uptake by Aquatic
Organisms: Consequences and Implications

L. D. EYMAN and J. R. TRABALKA

Literature on the behavior of plutonium and transuranic elements in aquatic organisms is
reviewed. Tlie commonality of observed distribution coefficients over a wide array of
aquatic environments (both freshwater and marine) and the lack of biomagnification in
aquatic food chains from these environments are demonstrated. These findings lead to the
conclusion that physical processes dominate in the transfer of transuranic elements from
aquatic environments to man. The question of the nature of the association of plutonium
with aquatic biota (surface sorption vs. biological incorporation) is discussed as well as
the importance of short food chains in the transfer of plutonium to man.

For years plutonium and the transuranic elements were considered to be unimportant in
ecological transfers and food chains because of their low solubihty and uptake when
ingested by mammals. It is true that the mobility and availability of plutonium are
limited compared with cesium or strontium, the major concerns for years. However,
sufficient information has been developed in the past 5 yr to indicate that plutonium and
americium are available to a greater extent than was shown in earlier studies. Current
information demonstrates, even with the greater uptake by aquatic biota, that the
transuranic elements are not enriched in aquatic food chains but rather are discriminated
against. Data on plutonium and americium concentration in aquatic organisms are
diftlcult to interpret owing to differing degrees of surface contamination and/or gut
loading. Thus true biological accumulation is often masked by these contributions, which
frequendy indicate a higher degree of assimilation than is actually the case.

To assess die potential hazards of transuranic materials released to aquatic
environments, such as in low-level waste effluents from fuel-cycle processes and burial
grounds, some measure of their environmental behavior is needed, particularly as it relates
to the accumulation of these isotopes in man. The following review of the literature on
the behavior of plutonium and americium in aquatic environments is intended to provide
information on which such an assessment can be based.

Literature Review

Emery et al. (Emery, Klopfer, and Weimer, 1974; Emery and Fariand, 1974; Emery et al.,
1975; Emery et al., 1976) have described the behavior of plutonium and americium in a
pond at the Hanford plant. The pond is fed by a plutonium processing plant and laundry
wastes with a flow of about 10 m^ of water per minute. Percolation accounts for about
95% of the water loss from the pond. The pond is described as an ultraeutrophic system
with most of the plant nutrients supplied by laundry wastes. Analyses of sediment from

612

UPTAKE BY AQUATIC ORGANISMS 613

trenches leading to the pond indicate that most of the historic transuranic releases were
removed in the sediments of the trenches and thus did not reach the pond. This raises the
possibility that the plutonium and americium deposited years earlier may constitute a
supply of possibly solubilized radionuclides to the pond which could influence the
measured values.

Sediment contains more than 95% of the total plutonium pool in the pond. The
potential availability of ^^^Pu, ^^^'^'*°Pu, and ^'^^Am from sediments was estimated
from a series of extractions using sodium chloride, oxalate, and ethylene-
diaminetetraacetic acid (EDTA). Emery et al. (Emery, Klopfer, and Weimer, 1974;
Emery and Farland, 1974; Emery et al., 1975; Emery et al., 1976) corrected the
concentration ratios for pond biota to account for the estimated available fraction from
the sediments or water. However, for purposes of this presentation, we have chosen to
present concentration ratios (CR, ratio of sample plutonium concentration to source
plutonium concentration) for plutonium and americium related to the total
concentrations reported by Emery et al. in sediments and vrater. Table 1 clearly
demonstrates the difference in calculated CR values for aquatic biota when different
sources are assumed.

Tlie rarios of either ^"^^ Am or ^^^Pu to ^^^'^^^Pu in organisms divided by those
same ratios in various potential source compartments (sediment, interstitial water, and
overlying water) are given in Table 2. A value of 1.0 would indicate that the acceptor
compartment (biota) contains exactly the same isotopic ratios as the proposed source
compartment, i.e., that the ^"^ ^ Am/^^^'^'^^Pu ratios or ^^^Pu/^^^'^'^^Pu ratios are the
same in both biota and source. Therefore, if we assume that there are no significant
differences in metaboHsm of these isotopes by aquatic biota, the compartment that
exhibits ratios closest to 1.0 for both isotopes can be considered the prime source of
transuranic elements to the biota in this system. As shown in Table 2, interstitial water
from sediments most closely meets this requirement.

Marshall, Waller, and Yaguchi (1974) provided an earUer assessment of the role of
organisms in tlie removal of plutonium from Great Lakes waters and the potential for
plutonium to reach man via the food chain. Concentration ratios for 2 3 9,2 4 0p^j ^^^

TABLE 1 Concentration Ratios for Plutonium and Americium
in Aquatic Biota from U-Pond Calculated Using Different Sources*

Sediment

Interstitial water

Water

Pu Am

Algae floe
Snail

5
0.2

219 185
6 62

2 X 10"
5 X 10*

2x 10=
8 X 10"

Submergent cattail

0.2

0.6

7 34

5 X 10*

5 xlO"

Algae

Native goldfish, without gut

Emergent cattail

0.1

0.05

0.02

0.1
0.05

2 7

0.8 3

5 X 10*
2 X 10*
5 X 10=

8 X 10^
3 X 10^

Beetle (Coleoptera)

0.009

0.01

0.3 0.7

3 X 10=

8 X 10^

Submergent bulrush

0.003

0.1

1 X 10=

Emergent bulrush

0.003

0.001

0.1 0.2

1 X 10=

3 X 10'

Native goldfish muscle

0.002

0.005

0.06 0.3

5 X 10"

4 X 10'

*Data from Emery, Klopfer, and Weimer, 1974; Emery and Farland, 1974; Emery et al., 1975:
Emery et al., 1976.

614 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Quotients of Isotopic Ratios in Biota
and Potential Source Compartments

Ratios in biota/ratio

in source

Sources

Compared isotopic ratio

Sediment

Interstitial water

Overlying water

238py/239,240py
24.^^/239,240py

1.7-1.9
2.3-2.8

0.74-0.84
1.7-2.0

0.41-0.46
0.002-0.003

TABLE 3 Concentration Ratios for ^ ^ ^ Pu in Great Lakes
Biota, June 1972 to November 1973 (Wet Weight)

Sample

Number

Concentration ratios over water

Lake

Mean + standard error

Range

Superior

Mixed plankton

4000 ±

900

2680-5730

Zooplankton

630

Smelt

Michigan

Mixed plankton

5700 ±

800

620-15,300

Cladophora sp.

3800 ±

500

1060-6930

Zooplankton

350 ±

122-653

Mysis relicta

760 ±

587-989

Pontoporeia affinis

1600

1450-1830

Slimy sculpin

250 ±

128-560

Chub

37 ±

21-50

Ale wife

25 ±

17-30

Smelt

20 ±

6-33

Perch

4-29

White fish

5-23

Coho

Chinook

Lake trout

1-2

Huron

Mixed plankton

4460

3340-5680

Alewife

165

25-305

Smelt

Perch

Erie

Zooplankton

500 ±

150

316-788

Smelt

235

Perch

Ontario

Mixed plankton
Alewife

2420 ±
176

200

2030-2670

reported for biota of the five Great Lakes, with emphasis on samples from Lake Michigan
(Table 3). High CR's in mixed plankton compared with zooplankton are thouglit to be
due, at least partially, to plutonium associated with phytoplankton, which predominate in
mixed plankton samples. It is further suggested that much of the ^■^^Pu in zooplankton
samples is due to phytoplankton in their digestive tracts. Food-chain relationships
between most of the species analyzed are mentioned in this chapter and are listed in
Table 4.

UPTAKE BY AQUATIC ORGANISMS 615

TABLE 4 Primary Food Sources for Various Great Lakes Species

Species Primary food sources

Zooplankton (general) Phytoplankton

Mysis, Pontoporeia (crustaceans) Periphyton near surficial sediment layer

Sculpin Pontoporeia

Chub, alewife, smelt, and whitefish Zooplankton or mixture of zooplankton

and benthic invertebrates

Perch Mixture of invertebrates and fish

Coho salmon, chinook salmon, and lake trout Smaller fish

As summarized by Marshall, Waller, and Yaguchi (1974), "The results clearly indicate
that although the concentration of plutonium in phytoplankton is several thousand times
that in the water, it decreases by an order of magnitude in each successive link in the food
chain leading to man."

Wahlgren and Marshall (1974) studied the distribution of residual fallout plutonium
in Lake Michigan between the water and the various trophic levels of the food chain.
They found tliat the CR value for plutonium in phytoplankton compared with that in
water was about 5000. A reduction in concentration by a ratio of about 10 was
observed at each trophic level consisting of zooplankton, planktivorous fish, and
piscivorous fish. The top predators had concentrations only slightly greater than the lake
water. However, the concentration in benthic (bottom-feeding) fish was considerably
higher than tliat in the planktivorous fish. It should be noted that the fish were not
dissected but were analyzed in their entirety (including GI tract). From Fig. 1 we
estimate the planktivorous fish to have a CR of about 20 over water and the piscivorous
fish a CR of about 2, whereas benthic fish exhibit a CR of approximately 300.

The major role of phytoplankton in plutonium kinetics in aquatic systems has been
postulated to be one of the removal of a significant fraction of plutonium from tlie water
column (Walilgren et al., 1976; Hetherington, 1976). However, collection techniques are
such tliat phytoplankton cannot readily be separated from inorganic suspended
parficulate matter. The CR values for algae reported in Table 3, in fact, may be high
owing to the inclusion of suspended inorganic particulate matter, which would have a CR
value of approximately 10^ (see below). The correlafion between percent silicon content
and plutonium concentrations in phytoplankton samples (Yaguchi, Nelson, and Marshall,
1974) was attributed to the predominance of diatom frustules in the samples analyzed.
Wahlgren et al. (1976) reported a correlation between percent ash weight and plutonium
concentrafions. They also concluded that the plutonium was associated with diatom
frustules in the plankton samples. However, in neitlier case was the contribution of
associated suspended inorganic particulate matter quantified. Wahlgren et al. (1976)
reported a distribuUon coefficient

„ _ Concentrafion on solid phase (g/g)
f^d

Concentration in liquid phase (g/ml)

of plutonium for suspended sediment materials of about 3 X 10^. The inclusion of a
small amount of those materials in the ash residue of phytoplankton samples is a plausible
alternative interpretation of the observed correlafions.

616 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

103

102 —

101 —

10° —

10-1 1—

o
u

3 10-2
Q.
O)
CO
CM

10--^ —

10-^

- \

1 1

• Mean value

J Range of values

' <

" o '

1 \

1 1

1 1 1

1 ( )

1 1

'- 00 I,,

1-
<

■z.
o

D
O
CE

DC
CD

3
O

?^^

UJ 5

QC
LU

O
<

UJ
1-
DC

o
>

" 1

QC _j

X <

— -1

H
Z

UJ
>

o
o

CO 01

Q-

D I-
_l Q.

UJ >

I- -I

< D

Fig. 1 Plutonium concentrations in various Lake Michigan compartments. (Data from
Wahlgren and MarshaU, 1974.)

Edgington et al. (1976) measured the plutonium and americium concentrations in
Lake Michigan sediments and identified some of their characteristics. They calculated
that approximately 97% of the plutonium that has entered Lake Michigan now resides in
the sediments. Surface sediments in Lake Michigan now contain between 140 and 400
fCi/g of dry sediment, whereas the concentration in the water column is less than 1
fCi/liter. Earlier, Edgington et al. (1976) had indicated a negligible contribution to the
input of plutonium to Lake Michigan from runoff via tributary rivers and streams. Their
data suggest a significant redistribution of sedimentary material in the lake with a rapid
movement of the radioactivity from its site of deposition on the sediment surface to a
final site of deposition in the sediments. Apparently there are large areas of Lake
Michigan where no significant sedimentation occurs. These areas have a layer of tloc,
containing somewhat higher concentrations of plutonium, overlying the glacial till or
sand. Since no significant accumulations of sediment have occurred in these areas, it is
likely tliat tliese 1- to 2-cm-thick deposits are transitory and that the material is readily
resuspended.

The 2 3 9,2 4 0pjj ^j^ values for shoreline plants in Lake Ontario found by Bowen
(1974) were similar to those reported by Marshall, Waller, and Yaguchi (1974). However,
the plankton CR values ranged tVom 10 to 300 (vs. 600 to 15,000 as reported by
Marshall, Waller, and Yaguchi, 1974). Further, Marshall, Waller, and Yaguchi reported CR

UPTAKE BY AQUATIC ORGANISMS 617

values for Lake Ontario which ranged from 2000 to 2700. Observed differences in CR
values for plankton reported in these two studies have not been resolved. Bowen (1974)
and Marshall, Waller, and Yaguchi (1974) reported that benthic feeders accumulate
plutonium to higher levels than limnetic feeders. Some predator species (i.e., largemouth
bass, rock bass, white perch, and coho salmon) deposit most of the plutonium in the
bone, whereas other forms (i.e., northern pike, yellow perch, and freshwater drum)
deposit plutonium primarily in the liver. These apparent differences in tissue distributions
between species have not been satisfactorily explained.

Dahlman, Bondietti, and Eyman (1976) reported on the behavior of plutonium in the
biotic components of White Oak Lake. Results from analyses of various components of
the lake system support the finding by Marshall, Waller, and Yaguchi (1974) in Lake
Michigan, i.e., decreased concentrations of plutonium at higher trophic levels (Table 5).

TABLE 5 Concentrations of ^ 3 9 ,2 4 op^ ^^^ Concentration
Ratios for Fishes from White Oak Lake

Species

Plutonium content

Carcass*

Gl tract

Concentration ratiot

pCi/g

Standard error

pCi/g

Standard error CR Standard error

Largemouth bass 2x10 "

BluegiU 1 X 10"^

Goldfish 1 X 10-^

Shad 2x10"'

6x 10" =
8 xlO-"
1 xlO"^
3 xlO"'

6x10-^
4 xlO-"
8 xlO-^
4 X 10-'

4x 10-"
5 X 10-5
3 xlO-'
7 X 10-^

0.04
3
3
4

0.2
2
3
0.1

*Total fish minus Gl tract.

fConcentration ratio (CR) is defined as [plutonium] in organisms (wet weight)/ [plutonium] in
water. Water concentration of plutonium used in the calculation of CR values is 4 x lO"" pCi/g. The
CR values are for carcass.

Organisms living in or on the bottom of sediments in the White Oak Lake (filamentous
algae and benthic invertebrates) have plutonium concentrations that are two or three
orders of magnitude higlier than those in predatory fish, such as largemouth bass and
bluegill. Filamentous algae associated with sediments in shallow areas of White Oak Lake
had the highest concentration of plutonium of any biotic component measured.
Gastrointestinal contents of goldfish, gizzard shad, and bluegill had plutonium concentra-
tions intermediate between those of sediment and food organisms, which indicates the
importance of sediment as a dietary source of plutonium. Further evidence of sediment
ingestion by these fish was the fact that the ^^^Pu/^^^Pu ratio of gut contents was
similar to that measured in sediments.

Long-term, chemically and/or biologically mediated transformations of plutonium
compounds may be expected to occur in aquatic systems. These transformations may
result in plutonium being complexed by naturally occurring chelating agents, such as
carboxyhc acids (citrate), fulvic acids, or proteins. Several laboratory experiments were
carried out at Oak Ridge (Eyman, Trabalka, and Case, 1976; Eyman and Trabalka, 1977;
Trabalka and Eyman, 1976) to determine the uptake and distribution of monomeric
plutonium (IV) in chelated forms both in an aquafic vertebrate, the channel catfish
Ictalums punctatus, and in a littoral aquafic microecosystem. A primary finding was that
the gastrointestinal intake by catfish was significantly higher than that reported for

618 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

mammals. The highest observed retention (whole body) at 63 days was 3.8% of ingested
dose for ^^^Pu citrate, whereas retention of the fulvate was 0.6%. Reduced uptake of the
fulvate complex is thought to be due to either its high molecular weight (> 10,000) or its
stability in metabolic systems. Increased uptake of ^'^^Pu citrate is attributable to both
high gut permeability and instability of the complex in metaboHc systems. Chelation can
either enhance or reduce the uptake of ingested plutonium relative to plutonium
hydroxide (monomer) in channel catfish. Approximately two-thirds of the initial body
burden of plutonium (administered as citrate) was lost within 63 days after gut clearance.
These results suggest that the observed retention pattem of plutonium in channel catfish
was due to plutonium labehng of the gut followed by subsequent turnover by cell-renewal
processes. This suggestion is based on the observed slow gut-cell renewal times in fish (20
to 30 days) (Hyodo-Taguchi, 1968) compared with those in mammals (48 hr in mouse)
(Lesher-, Wahlburg, and Sacher, 1964).

Tissue-distribution studies in channel catfish revealed that relatively little (<10%) of
the intracardially injected plutonium citrate was excreted. Blood clearance rates were
similar to those found in small mammals, the plutonium being associated primarily with
the plasma protein transferrin. The fractional body burdens in bone, Uver, and kidney 17
days after injection were 31, 24, and 9% of the injected dose, respectively. High kidney
burdens relative to mammals are expected since the kidney functions as the major site of
hemopoiesis in teleosts. The absence of significant excretion of plutonium reinforces tlie
previous suggestion that a short half-life component of elimination following gut
clearance in gavage studies is due to plutonium uptake by, and subsequent turnover of,
cells in the gut wall.

A distribution coefficient of 9 x 10^ was observed for sediment in a year-old aquatic
microcosm spiked with ^^^Pu nitrate. A materials balance at 90 days postspike provided
the following estimates of plutonium distribution: 0.001% in water, 0.04% in biota, and
over 99.9% in sediments. Concentrations of ^^ ^Pu in whole animals, including fish, were
surprisingly uniform (within a factor of 10, 1.2 to 9.9% of mean sediment concentrafion).
This was related to gut loading of sediments and/or surface contamination. The uptake by
emergent macrophytes not exposed to surface contamination was quite small: <0.03 to
0.1% of the sediment concentration.

On the basis of this set of laboratory experiments, sorption to plant surfaces, on gut
walls, and on exoskeletons appears to provide the dominant sites for plutonium
deposition in or on submerged components of aquatic systems. Interestingly, the
sediment distribution coefficient observed in the laboratory microcosm study was well
within the range of values reported from a wide variety of laboratory and field studies
(Table 6).

In a study of crayfish from the Great Miami River, Wayman, Bartelt, and Groves
(1976) reported that most of the plutonium was concentrated in soft tissues rather than
in the sclerotized shell. Similar results were reported earlier by Nelson and Noshkin
(1973) for the Tridacna clam and lobster in marine studies at Enewetak Atoll. Noshkin
(1972) reported higher concentrations of 2 3 9,240p^ j^ ^g gj^^jj ^^ -^^ ^^^ body of

scallops, whelks, and moonshells collected off Cape Cod. Ward (1976) also observed that
the calcified shell appeared to accumulate 2 3 9,2 4 Op^ ^^ ^ rapid rate. A very high
proportion (89.5%) of the total plutonium was in the skeleton, which accounted for
about 43% of the total weight of the lobster. Concentration factors for shells were on the
order of 200, and gills were about 100. The flesh, which comprised about 28.7% of the
total body weight, contained only 1.2% of the plutonium present in the entire body. The

UPTAKE BY AQUATIC ORGANISMS 619

TABLE 6 Distribution Coefficients (R^) for
Plutonium Isotopes in Aquatic Systems

9.0xl0'»

237

1.3 - 9.4 X 10"

237

1.2- 7.9 X 10"

239

4.8 xlO" -

1.3 xlO'

239

5 xlO"

239

3 xlO'

239

5 xlO*

239

Kj* Isotope Environmentf Reference

L,F Trabalka and Eyman (1976)

L,M Duursma and Parsi (1974)

L,F Trabalka and Frank (1976)

M Pillai and Mathew (1976)

F Bowen (1974)

F Wahlgrenetal. (1975)

M Hetherington et al. ( 1976)

^„ Concentration on solid phase (grams per gram)

Concentration in liquid phase (grams per milliliter)"
tF, freshwater; M, marine; L, laboratory.

content of the cast shell was approximately twice that of the shell of the intermoult
lobster. Whether the differences reported are due to the chemical species of plutonium
the organisms were exposed to at the various sites or to interspecific physiological
differences is open to question at this time.

Some authors (Emery et al., 1975; Bair et al., 1974; Hakonson and Johnson, 1973;
Morin, Nenot, and LaFuma, 1972) have suggested that plutonium isotopes can behave
differently in biological systems. In environmental studies the discrepancies in behavior
can be explained by different physic ochemical forms of the isotopes in the original source
or by different sources of uptake. In laboratory experiments differences in metabolic
behavior can be attributed to differences in concentrations of the isotopes tested.
However, Fowler, Heyraud, and Beasley (1975) found that, if the different plutonium
isotopes (^^^Pu, ^^"^Pu, ^^^Pu, and ^^^'^"^^Pu) were present in the same physico-
chemical form, aquatic organisms were unable to discriminate between tliem in either
accumulation or excretion studies. They also concluded that the suggestion of Moghissi
and Carter (1975), reinforced by the work of Eyman, Trabalka, and Case (1976), is
correct, "that in environmental studies, ^^ ''Pu is most likely the best tracer for measuring
plutonium kinetics in biological systems."

Discussion

This chapter is not intended to be definitive, but, rather, we have attempted to evaluate
the most pertinent data sets. Although there appears to be a commonality (i.e., no
biomagnification, highest concentrations in benthic organisms and phytoplankton) among
all environmental data sets, whether marine or freshwater, there are also significant
discrepancies. The most interesting common factor we were able to extract from a wide
range of studies, both laboratory and field, is the similarity in the observed distribution
coefficient for plutonium in sediment (Table 6). Inconsistencies involve differences in CR
values for freshwater algae and some fissue distributions observed in both freshwater and
marine studies. These discrepancies are probably explainable on the basis of differences in
physicochemical form to which organisms were exposed at different sites.

From a review of the data available for both freshwater and marine environments,
there appear to be relatively few significant differences in the patterns of accumulafion or
in the observed CR's.

620 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

There is a lack of any definitive infomiation on the other three elements of interest,
americium, neptunium, and curium. One study of neptunium in the Columbia River
(Davis et al.. 1958) involved an isotope with a half-life of only 2.3 days (vs. 2x10^ days
for ^•'^Np); so the radioactive decay limited the quantities of this isotope in the
organism. Data from Lake Michigan indicate increased CR's for ^'^^Am over plutonium in
the lower trophic levels by factors of 1.5 to 5 (Wahlgren et al., 1976). However,
americium CR values for fish are less than or equal to 10 times those for plutonium: the
data were not adequate to calculate specific values. Data from the discharge of waste
from Windscale (Hetherington et al., 1976) indicate little difference between ^^^^Puand
^"^^ Am concentrations in tlsh within 10 km of the discharge for equal discharge rates of
the two isotopes.

Concentration ratios chosen for use in both freshwater and marine environments are
given in Table 7. For americium, neptunium, and curium (■^^' Am, ^■'^Np, ■^'*^Cm, and
^"^^Cm), we have applied a factor of 10 over the plutonium values. These values were

TABLE 7 Recommended Concentration Ratios

Americium,

curium, and

Material

Plutonium

neptunium

Sediment

100,000

Plankton

5,000

50,000

Benthic algae and macrophytes

5,000

50,000

Benthic invertebrates

1,000

10,000

I'ish

Bottom feeders

250

2,500

Plankton feeders

250

Piscivorous (fish eaters)

chosen because these isotopes have a greater availability than plutonium in terrestrial
systems (Dahlman, Bondietti, and Eyman, 1976), and it is expected that, when adequate
data are available, the same may be true in aquatic systems. Limited data from Lake
Michigan and Windscale indicate that the factor foi ^^ ' Am may be less than 10; however,
owing to the incomplete nature of the data, we have chosen a more conservative figure as
the factor for biota.

The very high affinity of plutonium for particulate matter in aquatic ecosystems
(distribution coefficient, ~10^) suggests tliat it may not be appropriate to use the
traditional expression of CR to estimate die concentration of this element in biota.
Rather, we feel that the observed concentrations of plutonium in aquatic biota should be
related to the primary abiotic source in the system, sediment (both suspended and
bottom). To express this relafionship, the temi Trophic Transfer Factor (TTF) has been
used by various researchers (Lipke, 1971; Trabalka and Eyman, 1976; Elwood,
Hildebrand, and Beauchamp, 1976). The concentration of an element in sediment or food
is substituted for the concentration in water, which is normally used in the calculation of
a CR. The undedying assumption is that, owing to the higli distribution coefficients
(Kd's) observed, element accumulation in tissues of higlier trophic levels will be
dominated by gut absorption rather than by direct uptake from water. Trophic transfer
of plutonium by aquatic animals is comprised of diree fractions: exterior surface

UPTAKE BY AQUATIC ORGANISMS 621

sorption, gut labeling, and absorption from the gut. Again we stress that external
contamination with sedimentary particulate matter and gut loading are not considered to
represent true uptake and should be considered separately. The TTF, then, serves as a
realistic measure of plutonium discrimination in food chains. It should be noted that the
CR values for the Hanford waste pond fall into line with other data sets when considered
on this basis (Table 1).

Some of the variation in TTF values observed can be explained by the relative trophic
position of the organisms analyzed. The number of intervening food-chain transfers
between the organism analyzed and the abiotic source of plutonium should be inversely
related to the observed TTF value.

Conclusions and Recommendations

To assess the potential transfer of plutonium to man from aquatic ecosystems, we
must concentrate on those food sources most closely linked to sediment as a measure of
maximum plutonium in human food. One could postulate, as an extreme, although
improbable, case, the direct consumption of sediment by man. Owing to the high Kd of
plutonium in sediments, direct ingestion could expose humans to plutonium con-
centrations up to several hundred thousand times tliose found in water. A more probable
projection, however, would be exposure to plutonium via a food chain involving a single
trophic transfer from sediment to an organism that is consumed by man. Such short,
single trophic transfer food chains should result in the highest plutoniumi concentrations
in human food derived from aquatic ecosystems (Critical Exposure Pathway). Some
examples would include bottom-feeding fishes, shellfish, and rooted macrophytes, such as
rice. Although we could find no data on the accumulation of plutonium in rice, this
information seems critical since it is representative of a single trophic transfer from
sediment to man and is a major dietary component of a large segment of the world
population. Further, both marine (Pillai and Mathew, 1976) and freshwater (Wahlgren
and Marshall, 1974) organisms associated with the sediment— water interface (i.e.,
benthos) contain plutonium burdens that are one hundred times as high as those of
free-swimming forms.

Present National Committee on Radiation Protection and Measurement (NCRP)
guidelines (Title 10, Part 20) for plutonium in food are derived by inference from
standards based on drinking water consumption at a fixed rate (International Commission
on Radiological Protection, 1959; National Bureau of Standards, 1959). The total
radioactivity ingested as food and/ or water cannot exceed the product of the Maximum
Permissible Concentration (MPC) times the consumption rate for water. Therefore, if an
individual is ingesting water contaminated at the MPC level, he or she cannot be exposed
to plutonium from any other dietary source. Weights of water and food ingested are
approximately equal for the Standard Man. It is apparent, therefore, that knowledge of
the expected dietary plutonium contribution from food is as important as that of the
contribution from water.

The MPC for plutonium in drinking water was derived by using a fractional gut
transfer factor of 3 X 10~^ for ingested plutonium (International Commission on
Radiological Protection, 1959). The value of the human gut transfer factor is based on
studies whereby laboratory mammals were fed plutonium in a variety of chemical forms.
The value of 3 X 10~^ is based on plutonium administered in the nitrate form. Actual

622 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

gut transfer factors reported in these studies ranged from 1 X 10~^ to 3 x 10~^,
depending on a number of factors (chemical form, species of test animal, age of test
animal, etc.) (Hodge, Stannard, and Hursh, 1973). To our knowledge no studies have
been published which determine the fractional gut transfer factor for plutonium
incorporated in actual food materials. On the basis of the variability in data reported
above, one cannot rule out the possibility that this factor may be significantly greater
than the values used by the NCRP. In addition, plutonium releases in low-level effluent
streams from fuel-cycle processes or from burial grounds may be in a more available
chelated form, which is either due to release in a chelated form or to long-term
environmental transformation to the chelated form. These potential routes need to be
quantified, and the dominant mobile forms need to be idenfified.

Even greater uncertainty arises when trying to predict probable plutonium con-
centrations in an aquatic food source for purposes of dosimetric calculations. This
predicted concentration is the product of three values, plutonium concentration in water,
Kd for plutonium in sediment, and the TTF for a given single transfer food chain (Critical
Exposure Pathway). Since the Kj is known to be very high (=^10^), if the TTF value is
greater than 10~ ^ , the potenfial dose from food intake could be higher than the value for
which the standards (based on water intake) were designed, notwithstanding the
uncertainties cited above. Since TTF values of 5 x 10"^ to 5 x 10~^ appear reasonable,
based on data presented in this chapter [laboratory results (Eyman and Trabalka, 1977;
Trabalka and Eyman, 1976; Trabalka and Frank, 1978; Beasley and Fowler, 1976a;
1976b) and apparent CR values in benthic biota of 500 to 5000], our hypothetical
individual consuming such aquatic food organisms exclusively (obtained from a water
source contaminated at MPC levels) could receive radiation doses significantly higher than
intended by the standards. Although the radiation doses projected above obviously
represent a purely hypothetical case, the conditions sufficient to produce such doses
cannot be excluded because of the present uncertainty associated with parameters used to
derive the dose estimates. Further, it should be recognized that certain cases involving the
assessment of the plutcnium contribution to human diet require that the TTF be
separated into its components (i.e., for animals that are not to be consumed whole). It is
not our purpose to "single out" plutonium or any other actinide as an unusual hazard.
Many isotopes released to the environment from various nuclear-fuel-cycle processes have
been subjected to similar scrutiny in attempting to assess transport to man. Several
(radionuclides of cesium, strontium, and cobalt) will undoubtedly contribute significantly
higher doses to man than expected for plutonium (Blaylock and Witherspoon, 1978). To
demonstrate that tlie conservative assumptions stated above are unwarranted, we must
develop a data base on the environmental behavior of actinides reasonably comparable to
existing information on cesium, strontium, and cobalt.

Future research on the transport of plutonium to man from aquatic ecosystems
should concentrate on those food chains which have the lowest number of trophic
transfers between abiotic sources in the system and man. Data generated from such
research will provide critical information necessary for the evaluation of present standards
by determining plutonium concentrations in critical aquatic organisms that serve as food
sources to man. Additional research on the relationship between the chemical
characteristics of plutonium in abiotic components of the system and observed
concentrations in edible aquatic foods will strengthen our ability to predict potential
transfer of plutonium to man from aquatic ecosystems.

UPTAKE BY AQUATIC ORGANISMS 623

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of ^^'Pu and "»Pu Nitrates in Beagle Dogs, Health Phys., 27: 392.
Beasley, T. M., and S. W. Fowler, 1976a, Plutonium and Americium: Uptake from Contaminated

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1958, Vol. 18, pp. 423-428, United Nations, New York.
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, D. C. Klopfer, T. R. Farland, and W. C. Weimer, 1975, Ecological Behavior of Plutonium and

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, D. C. Klopfer, T. R. Farland, and W. C. Weimer, 1976, The Ecological Behavior of Plutonium and

Americium in a Freshwater Pond, in Radioecology and Energy Resources, C. E. Gushing, Jr. (Ed.),
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Eyman, L. D., and J. R. Trabalka, 1977, Plutonium-237: Comparative Uptake in Chelated and
Non-Chelated Forms by Channel Catfish (Ictalurus punctatus). Health Phys., 32: 475.

, J. R. Trabalka, and F. N. Case, 1976, Plutonium-237 and 246: Their Production and Use as

Gamma Tracers in Research on Plutonium Kinetics in an Aquatic Consumer, in Environmental
Toxicity of Radionuclides: Models and Mechanisms, pp. 193-206, M. W. Miller and J. N. Stannard
(Eds.), Ann Arbor Science Publishers, Inc., Ann Arbor, Mich.

Fowler, S., M. Heyraud, and T. M. Beasley, 1975, Experimental Studies on Plutonium Kinetics in
Marine Biota, in Impacts of Nuclear Releases into the Aquatic Environment, Symposium
Proceedings, Otaniemi, Finland, STI/PUB/406, International Atomic Energy Agency, Vienna.

Hakonson, T. E., and L. J. Johnson, 191 3 , Distribution of Environmental Plutonium in the Trinity

624 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Site Ecosystem After 27 Years. USAEC Report LA-UR-73-1291, Los Alamos Scientific

Laboratory, NTIS.
Hetherington, J. A., 1976, The Behavior of Plutonium Nuclides in the Irish Sea, in Environmental

Toxicity of Aquatic Radionuclides: Models and Mechanisms, M. W. Miller and J. N. Stannard

(Eds.), Ann Arbor Science PubHshers, Inc., Ann Arbor, Mich.
, D. F. Jefferies, N. T. Mitchell, R. J. Pentreath, and D. S. Woodhead, 1976, Environmental and

Public Health Consequences of the Controlled Disposal of Transuranic Elements to the Marine

Environment, in Transuranium Nuclides in the Environment, Symposium Proceedings, San

Francisco, 1975, pp. 139-154, STl/PUB/410, International Atomic Energy Agency, Vienna.
Hodge, H. C, J. N. Stannard, and J. B. Hursh (Eds.), 1973, Uranium Plutonium Transplutonic

Elements, Springer-Verlag, New York.
Hyodo-Taguchi, Y., 1968, Rate of Development of Intestinal Damage in the Goldfish After

X-Irradiation and Mucosal Cell Kinetics at Different Temperatures, in Gastrointestinal Radiation

Injury, pp. 120-126, M. F. Sullivan (Ed.), Excerpta Medica Foundation, Amsterdam.
International Commission on Radiological Protection, 1959, Recommendations of the International

Commission on Radiological Protection, Publication 2, Pergamon Press, New York.
Lesher, S., H. W. Wahlburg, and G. A. Sacher, 1964, Generation Cycle in the Duodenal Crypt Cells of

Germ-Free and Conventional Mice, Nature, 202(4935): 884-886.
Lipke, E. J., Jr., 1971, Effects of Environmental Parameters on the Uptake of Radioisotopes in

Freshwater Fish, Ph.D. Dissertation, University of Michigan, Ann Arbor.
Marshall, J. S., B. J. Waller, and E. M. Yaguchi, 1974, Plutonium in the Laurentian Great Lakes: Food

Chain Relationships, in Proceedings of the XIX Congress of Association of Limnology, Winnipeg,

Manitoba, Canada, Aug. 23-28, 1974.
Moghissi, A. A., and M. W. Carter, 1975, Comments on "Comparative Distribution and Excretion of

^ ' 'Pu and ' =■ 'Pu Nitrates in Beagle Dogs," Health Phys., 28: 825.
Morin, M., J. C. Nenot, and J. LaFuma, 1972, Metabolic and Therapeutic Study Following

Admmistration to Rats of ^ ^ *Pu, Health Phys., 23: 475.
National Bureau of Standards, 1959, Maximum Permissible Body Burdens and Maximum Permissible

Concentrations of Radionuclides in Air and in Water for Occupational Exposure, Handbook 69,

GPO.
Nelson, V., and V. E. Noshkin, 1973, Eniwetok Radiological Survey, USAEC Report NVO-140,

Nevada Operations Office, NTIS.
Noshkin, V. E., 1972, Ecological Aspects of Plutonium Dissemination in Aquatic Environments,

Health Phys., 22: 537-549.
Pillai, K. C, and E. Mathew, 1976, Plutonium in the Aquatic Environment: Its Behaviour, Distribution

and Significance, in Transuranium Nuclides in the Environment , Symposium Proceedings, San

Francisco, 1975, pp. 25-45, STl/PUB/410, International Atomic Energy Agency, Vienna.
Trabalka, J. R., and L. D. Eyman, 1976, Distribution of Plutonium-237 in a Littoral Freshwater

Microcosm, Health Phys., 31: 390.
, and M. L. Frank, 1978, Trophic Transfer by Chironomids and Distribution of Plutonium-239 in

Simple Aquatic Microcosms, ^ecW; Phys., 35: 492.
Wahlgren, M. A., and J. S. Marshall, 1974, Distribution Studies of Plutonium in the Great Lakes, in

Proceedings of the 2nd International Conference on Nuclear Methods in Environmental Research,

University of Missouri, Columbia, Mo., July 29-31, 1974, CONF-740701, NTIS.
, J. J. Alberts, D. M. Nelson, and K. A. Orlandini, 1976, Study of the Behavior of Transuranic and

Possible Chemical Homologues in Lake Michigan Water and Biota, in Transuranium Nuclides in the

Environment, Symposium Proceedings, San Francisco, 1975, pp. 9-24, STl/PUB/410, Interna-
tional Atomic Energy Agency, Vienna.
Ward, E. E., 1976, Uptake of Plutonium by the Lobster Homarus vulgaris. Nature, 209(5023):

625-626.
Wayman, C. W., G. E. Bartelt, and S. E. Groves, 1976, Further Investigations of Plutonium in Aquatic

Biota of the Great Miami River Watershed, Including the Canal and Ponds in Miamisburg, Ohio, in

Radiological and Environmental Research Division Annual Report, January -December 1975,

USAEC Report ANL-75-60 (Pt. 3), Argonne National Laboratory, NTIS.
Yaguchi, E. M., D. M. Nelson, and J. S. Marshall, 1974, Plutonium in Lake Michigan Plankton and

Benthos, in Radiological and Environmental Research Division Annual Report, July 1972- June

1973, USAEC Report ANL-8060, Argonne National Laboratory, NTIS.

The Migration of Plutonium

from a Freshwater Ecosystem at Hanford

RICHARD M. EMERY, DONALD C. KLOPFER, and M. COLLEEN McSHANE

A reprocessing waste pond at Hanford has been inventoried to determine quantities of
Plutonium that have accumulated since its formation in 1944. Expressions of export were
developed from these inventory data and from informed assumptions about the vectors
that act to mobilize material containing plutonium. This 14-acre pond provides a realistic
illustration of the mobility of plutonium in a lentic ecosystem. The ecological behavior of
plutonium in this pond is similar to that in other contaminated aquatic systems with
widely differing limnological characteristics. Since its creation this pond has received
about 1 Ci of ^^^'^^^Pu and ^^^Pu, most of which has been retained by its sediments.
Submerged plants, mainly diatoms and Potamogeton, accumulate more than 95% of the
plutonium contained in biota. Emergent insects are the only direct biological route of
export, mobilizing about 5 X 10^ nCi of plutonium annually, which is also the estimated
maximum quantity of the plutonium exported by waterfowl, birds, and mammals
collectively. There is no apparent significant export by wind, and it is not likely that
plutonium has migrated to the groundwater below U-Pond via percolation. Although this
pond has a rapid flushing rate, a eu trophic nutrient supply with a diverse biotic profile,
and interacts with an active terrestrial environment, it appears to effectively bind
plutonium and prevent it from entering pathways to man and other life.

The dissemination of plutonium in our environment continues to be a major issue
centering around the development and appUcation of nuclear energy. In addressing this
problem, investigators have made efforts to inventory the worldwide plutonium burden in
terms of fallout and point-source deposition (Electric Power Research Institute, 1976). A
number of locations have been identified as having above-background plutonium
concentrations, such as weapons-testing sites, sites of accidental releases where plutonium
has escaped its container, and sites of controlled releases associated with waste
management areas. For some of these areas, such as the Eniwetak and Bikini Atolls and
the Mortandad Canyon leading to the Rio Grande, plutonium inventories are being
investigated in attempts to estimate quantities that migrate away from these sites over
time (Schell and Watters, 1975; Hakonson, Nyhan, and Purtymun, 1976). The export of
plutonium away from any contaminated aquatic site has not been sufficiently studied to
provide a quantitative example of the environmental mobility of this element. In this
regard waste ponds can serve as useful study sites since they often have diverse ecological
profiles, receive additions of plutonium over extended periods of time, and have frequent
contact with terrestrial forces that can mobilize plutonium.

Waste ponds likely constitute the most probable and greatest percentage of freshwater
environments that become contaminated with plutonium from local sources, whether

625

626 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

from controlled low-level releases or by accident. The few that exist in this country are
located at Rocky Flats, Colo. (Johnson, Svalberg, and Paine, 1974); Oak Ridge, Term.
(Dahlman, Bondietti, and Eastwood, 1975); the Savannah River plant near Aiken, S. C.
(D. Paine, 1977, Battelle, Pacific Northwest Laboratories, private communication); Idaho
Falls, Idaho (D. Markham, 1976, Rockwell Hanford Operations, Richland, Wash., private
communication), and Hanford near Richland, Wash. (Emery, Klopfer, and Weimer,
1976). These ponds are managed in association with fuel separation, reprocessing, and
reactor testing operations. The three unmanaged freshwater systems reported to have
received small amounts of plutonium include the Miami River near Miamisburg, Ohio
(Bartelt, Wayman, and Edgington, 1975), Sawmill Creek at Argonne, 111. (Singh and
Marshall, 1977), and streams leading to the Rio Grande River near Los Alamos, N. Mex.
(Hakonson, Nyhan, and Purtymun, 1976). Although amounts of plutonium released to
these aquatic systems are usually quite small, concentrations accumulated in waste ponds
are often significantly above background levels.

Information about the ecological transport of plutonium from any of these systems
would be of special interest since they represent the results of actual contamination
events as they exist today. The distribution and fate of plutonium in a waste pond at
Hanford, specifically called U-Pond (Fig. 1), have been studied since 1973 (Emery,
BClopfer, and Weimer, 1974; Emery et al., 1976). The results appear to provide a good
example of the behavior of plutonium in a freshwater environment. U-Pond has an
estabUshed ecosystem and has been exposed to plutonium longer than any other aquatic
environment. Since 1944 plutonium has reached this 14-acre pond via waste ditches,
which have received occasional pulses of transuranic elements from the clean-up of minor
accidental spills of low-level contaminants within the reprocessing laboratories.

One of the major goals of the study at U-Pond has been to obtain sufficient
information about the pond's ecosystem and the distribution of plutonium within it so
that plutonium export routes can be assessed quantitatively. Although it is often difficult
to measure with reasonable certainty the parameters necessary for describing these export
routes, the purpose of this wark is to formulate the best expressions of export given the
conditions that Umit this process. The objectives are to determine ranges of quantities of
plutonium in the pond's ecosystem and assess the amount of plutonium being exported in
relation to this inventory. To accompUsh this task, we estimated the pond's plutonium
inventory quantities on a basis of minimum, mean, and maximum values for each
ecosystem compartment to postulate the amount of these inventories that is exported
yearly.

Methods and Materials

To examine the pond's inventory and export conditions, we separated the ecosystem into
two categories, the aquatic system and the contacting terrestrial system. The entire
inventory and export scenario is shown in Fig. 2. The aquatic system is divided into 10
compartments:

1. Nonfilamentous algae (including sestonic diatoms).

2. Filamentous algae.

3. Submerged macrophytes.

4. Emergent macrophytes.

5. Lower invertebrates (excluding insects and gastropods).

6. Resident insects (those with aquatic adult stages).

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 627

7. Emergent insects (those with emerging adult stages).

8. Gastropods.

9. Goldfish (anthropogenic ally introduced).

10. Sediments (down to 10 cm and including organic floe generated from decom-
posing plant material).

Water itself is not considered as a compartment, but the particulate contents
contained by the water mass are accounted for in other compartments. Suspended
particles greater than 0.1 jum (seston) are fractioned by weight into inert particles and
algae on the basis of microscopic inspection and dry weight— ash weight comparisons.
Weiglits of the sestonic algae, which are mostly diatoms, are added to the nonfilamentous
algae compartment, and weights of inert particles are placed with the sediments. Particles

I LABORATORY

0 2 4 6 8 10
Kilometers

Fig. 1 Map of Hanford site showing location and detail of U-Pond.

628 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

I I
I I

<
<

t/5

►

5
o

Li-
ce
LU

NOIiV100H3d

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1/5

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MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 629

of plutonium less than 0.1 jum and soluble plutonium are in very dilute concentrations
{< 1 pCi/liter), which are lower than standards that regulate plutonium concentrations in
drinking water. It is assumed that tliis small fraction of plutonium remains in the pond
water until it enters any one of the ecosystem compartments that could provide a route
of export. Dissolved and suspended materials in the pond have a short duration since
nearly all the water leaves the pond by percolation after a mean residence time of 40 hr
(Emery, Klopfer, and Weimer, 1974). The pond has no surface outflow.
The contacting terrestrial system has four compartments:

1 . Waterfowl.

2. Birds (other than waterfowl).

3. Mammals.

4. Airborne particulates.

Since these compartments have a transient association with the pond, they also serve as
routes of plutonium export. In addition, the transient insect population in the pond,
along with the emergent macrophytes, provides means for plutonium to leave the pond.
The emergent macrophytes would require assistance from one of the other export routes
to release any of their plutonium content to adjacent areas. Tliis is also true for the
plutonium that resides in the shoreline sediments. Thus the only export vectors through
which plutonium can leave U-Pond are:

• Percolation.

• Emergent insects.

• Waterfowl.

• Birds.

• Mammals.

• Wind (containing airborne particles).

Methods of sample preparation and plutonium analysis of pond samples, which
involve drying, ashing, chemical separation, and electrodeposition, are described by
Emery, Klopfer, and Weimer (1974).

Several techniques were used in the measurements of annual production of biomass in
the aquatic system. For all compartments except sediments, the annual production is
expressed as the quantity of biomass that is generated in 1 yr. The material quantity of
the pond's sediments is expressed as the dry weight of sediments to a depth of 10 cm.
The inventory of plutonium in the sediments is then the amount of plutonium in the
upper 10 cm. The dry weight of the upper 10-cm layer of sediments is 3.4 x 10^ kg,
which extends over an area of 5.6 x lO'^ m^ .

Nonfilamentous algae are composed mainly of sestonic diatoms (not always a true
phytoplankton population) and, to a much lesser extent, Tetraspora, which rests on the
bottom in loose globular masses. The annual production of sestonic diatoms is estimated
by using the weight of the average fraction of seston that is diatoms. This was done by
microscopic examination of concentrated seston samples to determine the mean
percentage of the total number of the particles that is diatoms (28%). Seston
concentrations were then multiplied by 0.28 to obtain an estimate of the concentration
of diatoms (in milligrams per liter). This concentration was proportioned to the volume
of the pond to estimate an instantaneous standing crop of diatoms in the pond. Sestonic
diatom standing crops were sampled seasonally to establish an annual mean. Since the
mean residence time of this water mass is 40 hr, this mean standing-crop value is

630 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

multiplied by the number of 40-hr intervals in a year (219) to estimate the annual
production. The production of Tetraspora was estimated by direct observations of
appearance in sampling grids made periodically during the growing season. The mean dry
weight of a volumetric quantity was established and then proportioned to the volume of
Tetraspora observed to occur in U-Pond.

Inventories of filamentous algae and submerged macrophytes are based on periodic
measurements of primary productivity and standing crop [see Emery, Klopfer, and
Weimer (1974) for a detailed description] . A method described by Verduin (1964) was
used to measure pond-wide primary production, which accounts for the photosynthesis of
all plant life in the pond inclusive of phytoplankton, macroalgae, and submerged
macrophytes. Results of these measurements express the net accumulation of plant mass
per unit time. Hence a summation of the montlily rates of net productivity for the entire
year provides one estimate for the annual quantity of submerged plant biomass that may
accumulate plutonium. For verification of this estimate, submerged plant standing crop
was measured at the beginning and end of the growing season by the areal sampling
methods described by Emery, Klopfer, and Weimer (1974). Areal sampling of submerged
plant biomass was also done periodically to provide a direct accounting of changes in
standing crop. So that standing crop could be measured, in this way plant material was
harvested from known surface areas, dried, weighed, and projected on a weight basis to
the area of the pond observed to be covered by these algae and macrophytes. Results of
both methods of estimating the pond's production of submerged plant biomass were in
reasonable agreement with each other (less than an order of magnitude difference).

Emergent macrophytes represent 10 to 15% of the pond's biomass at the peak of the
growing season. The annual production of emergent plants was estimated by the same
method used for submerged plants.

Pond invertebrates live mainly in association with rooted macrophytes and to a lesser
extent in the organic floe covering much of the pond's bottom. A O.lS-ft'^ Ekman dredge
was used to estimate the area concentrations of invertebrates living in the organic floe.
The more densely populated surfaces of macrophytes were quantitatively sampled by
direct collection and enumeration of invertebrates from the plants (mainly Potamogeton),
for which dry weights were also determined. Dry weights of the various types of
invertebrates were assessed per unit weight of plant material and then projected to the
pond-wide weiglit of tlie specific plant types to estimate an invertebrate standing crop.
All invertebrate life was assumed to appear in the pond on an annual basis; hence the
maximum standing crop served to represent the weight of invertebrates produced
annually by U-Pond.

The goldfish population in U-Pond is largest during the summer months. During the
colder months this population appears to decrease by several orders of magnitude, and
few goldfish are observed during the winter. Since this goldfish population appears to
reproduce annually and undergo substantial depletion in the colder months, the only
valid expression of annual production can be based on the summer standing crop. For this
reason the standing crop observed in August 1974, the only time when goldfish were
counted, will serve as the expression of annual production.

The standing crop of goldfish was estimated by first estabhshing a weight for an
individual and then counting the numbers of individuals appearing in 125, 9-m^ grids
placed randomly about tlie pond. The mean number of individuals was converted to mean
weiglit per unit area and proportioned to the entire pond area.

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 631

The biomasses of terrestrial life contacting the pond annually were estimated by using
mean weiglits for designated taxonomic groups and observations of frequency of contact
by these groups. Although it was possible to obtain reUable measurements of mean
weights for these organisms, the measurements of their contact frequencies were much
less precise. It would be desirable to have a more precise understanding of the intensity of
these export activities; however, it is possible to suggest a range within which these
vectors operate based on the ranges of plutonium concentrations found in these
organisms.

Means and 95% confidence limits of plutonium concentrations in pond compartments
were determined on an arithmetic basis. Concentrations, inventories, percentages, and
exported quantities of plutonium are expressed in scientific notation rather than in a
decimal format to draw attention to the order of magnitude rather than to emphasize the
exact quantity as a primary matter of consideration.

Results and Discussion

The Inventor}'

The scope of this study and the resources available for it placed Umitations on the
resolution of compartment-size determination. Although it was possible to determine
with accuracy the weiglits of sediments down to 10 cm and of emergent macrophytes, the
remaining compartment sizes were evaluated with a variety of estimation procedures.
Since much of these data are derived in this way, we feel that they represent only a
reasonable approximation of compartment sizes. Efforts to examine these data
statistically were not redeeming, and it was concluded that statistical confidence intervals
and central tendencies are not appropriate expressions for these results. Instead, these
results are intended to suggest best approximations without indicating ranges within
which the compartment sizes fluctuate.

The biota in U-Pond contain about 1% of the total mass of the pond, including
sediments down to 10 cm (Table 1). Concentrations of plutonium isotopes in the pond's
ecosystem compartments are shown in Table 2. Nonfilamentous algae and sediments
show the higliest mean plutonium concentrations of 2.8 x 10' and 5.0 x 10^ pCi/g,
respectively. Tliis similarly reflects the close association between the two compartments.
Submerged macrophytes and gastropods also have mean plutonium concentrations
exceeding 1 x 10^ pCi/g (1.6 x 10^ and 2.4 x 10^ pCi Pu/g, respectively). Filamentous
algae and emergent insects show mean plutonium concentrations of 8.6 x 10' and
4.6 x 10' pCi/g, respectively, whereas the remaining compartments have mean plutonium
concentrations ranging from 1 to 2 x 10' pCi/g.

U-Pond's eutrophic condition is reflected by its higli rate of primary production,
which occurs as high as 42kg C ha^' day"'. This rate of productivity can also be
expressed as 440 jug C liter" ' hr"' . Verduin (1964) found primary productivity'rates in
two Pennsylvania ponds to range from 120 to 760 /ig C liter"' hr"'. Hence U-Pond's
primary productivity resembles that in ponds not associated with nuclear facilities. Its
rate of carbon assimilation also approaches that of a higlily productive terrestrial
community, a cornfield, which has an average assimilation rate of 63 kg C ha"' day"'
(Robbins,Weier, and Stocking, 1957).

Submerged plant Ufe has most of the total plutonium inventory in pond biomass.
more than 95% (Tables 3 and 4). Submerged flora are composed mainly of algae.

632 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Estimated Quantities of Annual Production

of Biomass in U-Pond

(Weights of existing sediments are also included)

Estimated annual production

Weight, kg (dry) Percent of total

Nonfilamentous algae
Filamentous algae
Submerged macrophytes
Emergent macrophytes
Lower invertebrates
Resident insects
Emergent insects
Gastropods
Goldfish

Sediments

Biota subtotal

Total

8.5 X 10'

5.1 X 10'
1.8 X 10"
9.4 X 10'
6.8 X 10'

2.2 X 10'
1.1 X 10^

7.6 X 10'
2.1 X 10'

4.1 X 10"

3.4 X 10*

2.4 X 10"'

1.5 X 10"'

5.2 X 10-'
2.7 X 10"'

2.0 X 10"'

6.3 X 10-"
3.2 X 10~'
2.2 X 10"'

6.1 X 10"'

1.2 X 10"
9.9 X 10'
1.0 X 10'

TABLE 2 Means and Confidence Limits (95%) of Plutonium Concentrations in the U-Pond System

Concentratio

ns of Plutonium isotopes.

pCi/g(dry)

Lower limit

Mean

Upper

limit

J 3 9.240py

"«Pu

239,240p„

"«Pu

2 3 9 ,2 4 0 pu

2 38pu

Nonfilamentous algae (26)*

5.2 X 10'

7.5 X 10'

1.2 X 10 =

1.6 X 10=

1.9 X 10 =

2.5 X 10 =

Filamentous algae ( 14)

1.3 X 10'

2.3 X 10'

3.4 X 10'

5.2 X 10'

5.5 X 10'

8.1 X 10'

Submerged macrophytes (21)

7.2 X lO"

4.5 X 10°

6.3 X 10'

9.7 X 10'

1.2 X 10=

1.9 X 10=

Emergent macrophytest (18)

2.0 X 10"'

2.0 X 10-'

8.3 X 10°

9.9 X 10°

1.8 X 10'

2.2 X 10'

Lower invertebrates (8)

9.0x10-'

7.2 X 10-'

8.3 X 10°

8.5 X 10°

1.6 X 10'

Resident insects (12)

2.7 X 10°

3.4 X 10°

4.8 X 10°

5.6 X 10°

7.0 X 10°

7.8 X 10°

Emergent insects (84)

1.3 xlO'

1.9 X 10'

1.8 X 10'

2.8 X 10'

2.4 X 10'

3.7 X 10'

Gastropodst (11)

1.2 X 10°

2.0 X 10°

9.1 X 10'

1.5 X 10=

2.6 X 10=

4.3 X 10 =

Goldfish (8)

5.3 X 10°

7.7 X 10°

7.8 X 10°

1.2 X 10'

1.0 X 10'

1.6 X 10'

Sediments (123)

1.9 X 10^

2.0 X 10^

2.3 X 10'

2.7 X 10=

3.3 X 10 =

*Sample numbers are shown in parentheses.

t Lowest observed concentration replaces lower 95% confidence limit since the latter concentration is a negative
number.

including diatoms, Cladophora, Hydrodictyon, and Tetraspora, and the macrophyte
Potamogeton. Among these the diatoms and Potamogeton are the principal components
where plutonium is accumulated, wliich contain more than 99% of the pond's plutonium
inventory accumulated by plants. This suggests that an approximate plutonium inventory
for the biota of a similar ecosystem could be rapidly determined by estimating the
inventories in only a few populations of flora.

Of the invertebrate life emergent insects and gastropods have the higliest mean
plutonium inventories (Tables 3 and 4). Gastropods show an inventory of 1.8 x lO"* nCi,
whereas emergent insects account for 5.1 x 10^ nCi of plutonium. Quantities of
plutonium in emergent insects are of particular interest since they provide the only direct

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 633

route of biological mobilization from the pond. However, they do not contain more than
4 X 10"*% of the plutonium in the pond or more than 1 x 10~^% of the inventory in
pond biota.

Estimates of goldfish production in U-Pond fall within production ranges for suckers
and carp reported by Carlander (1955) for a number of North American lakes and
reservoirs. This goldfish population appears to contain about 4 X 10^ nCi of plutonium,
less than 1 x 10~^% of the plutonium in all the pond biota and less than 1 x 10~^% of
the entire pond inventory. Goldfish are occasionally eaten by herons, coyotes, and
waterfowl.

The bulk of material in the pond is, of course, in the sediments (Table 1). They
contain about 99% of the entire mass of the pond's ecosystem (excluding water).

TABLES Ranges

and Means of the "''"°

Pu Inventory in

U-Pond

Inventoryof '"''^oPu

Lower limit

Mean

Upper

limit

Activity,

Percent

Activity,

Percent

Activity,

Percent

of total

nCi

of total

nCi

of total

Non filamentous algae

4.4 X 10'

6.8 X 10-=

1.0 X 10*

1.3 X 10-'

1.6 X 10*

1.7 X 10-'

Filamentous algae

6.6 X 10*

1.0 X 10-'

1.7 X 10'

2.2 X 10-'

2.8 X 10'

3.0x10-'

Submerged macrophytes

1.3 X 10*

2.0 X 10-=

1.1 X 10*

1.4 X 10-'

2.2 X 10'

2.4 X 10"'

Emergent macrophytes

1.9 X 10'

2.9 X 10-

• 7.8x10*

1.0 X 10-'

1.7 X 10'

1.8 X 10-'

Lower invertebrates

6.1 X 10'

9.4 X 10-<

' 5.6x10'

7.2 X 10-'

1.1 X 10'

1.2x10"*

Resident insects

5.9 X 10>

9.1 X 10-«

' 1.1 X 10'

1.4 X 10-'

1.5 X 10'

1.6 x 10"'

Emergent insects

1.4 X lO'

2.2 X 10-

' 2.0 X 10'

2.6 X 10"*

2.6 X 10'

2.8 X 10"*

Gastropods

9.1 X 10'

1.4x10-

' 6.9 x 10'

8.8 X 10"*

2.0x10*

2.2x10"'

Goldfish

1.1 X 10'

1.7 X 10-

• 1.6x10*

2.0 X 10-*

2.1 X 10'

2.3 X 10"*

Biota subtotal

6.4 X 10'

9.8 X 10-

' 2.4x10*

3.0 X 10-'

4.3 X 10*

4.6 X 10"'

Sediments

6.5xl0»

>9.9x 10'

7.8 X 10»

>9.9 X 10'

9.2 X 10'

>9.9 X 10'

Total

6.5 X 10»

1.0 X 10'

7.8 X 10'

1.0 X 10'

9.2 X 10*

1.0 X 10'

TABLE 4 Ranges and Meai

IS of the ^^ *Pu Inventory in U-Pond

Inventory of " ' Pu

Lower limit

Mean

Uppe

r limit

Activity,

Percent

Activity,

Percent

Activity,

Percent

nCi

ot total

nCi

of total

nCi

of total

Nonfilamentous algae

6.4 X 10'

9.4 X 10*

' 1.4x10*

1.5 X 10"'

2.1 X 10'

1.9 X 10"'

Filamentous algae

1.2 X 10'

1.8 X 10~

' 2.7 X 10'

2.9 X 10~'

4.1 X 10'

3.7x10-'

Submerged macrophytes

8.1 X 10*

1.2 X 10"

' 1.7x10'

1.8x10-'

3.4 X 10'

3.1 xlO"'

Emergent macrophytes

1.9 X 10'

2.8 X 10"

* 9.0 X 10*

1.0 X 10-'

2.1 xlO'

1.9 X 10"'

Lower invertebrates

4.9 X 10'

7.2 X 10"

' 5.8x10'

6.3x10"'

1.1 X 10'

9.9x10 '

Resident insects

7.5 X 10'

1.1 X 10-

' 1.2x10'

1.3x10-'

1.7x10'

1.5 X 10-'

Emergent insects

2.1 xlO'

3.1 X 10"

3.1x10'

3.4 X 10-*

4.1 X 10'

3.7 X 10"*

Gastropods

1.5 xlO'

2.2 X 10"^

' 1.1x10*

1.2 X 10"'

3.3 X 10*

3.0 X 10"'

Goldfish

1.6x10'

2.3 X 10"

2.5 x 10'

2.7 X 10"*

3.4 X 10'

3.1 X 10"*

Biota subtotal

8.5 X 10'

1.2 X 10"

' 3.5 x 10'

3.7 X 10-'

6.2 X 10*

5.6 X 10"'

Sediments

6.8 X 10'

>9.9 x 10'

9.2 X 10»

>9.9 X 10'

1.1 X 10'

>9.9 X 10'

Total

6.8 X 10'

1.0 X 10'

9.2 X 10'

1.0 X 10'

1.1 X 10'

1.0 X 10'

634 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Sediments also have the highest concentrations of plutonium in the pond (Table 2). The
95% confidence interval of plutonium in sediment samples extended from 1.9 to
2.7 X 10^ pCi/g for 239,240p^ ^^^ ^^^^ 2.0 to 3.3 X 10^ pCi/g for ^^^Pu. Hence the
inventory of plutonium in the sediments is more than 99% of the inventory for the entire
system (Tables 3 and 4).

Other studies of plutonium in aquatic systems show sediments playing the dominating
role in the plutonium inventory of their respective ecosystems (Johnson, Svalberg, and
Paine, 1974; Patterson et al., 1976; Trabalka and Eyman, 1976). In nearly all studies that
have reported inventories of plutonium in freshwater systems, the sediments contain at
least 99% of the total plutonium burden. This commonality among ecosystems having
widely different limnological characteristics suggests that the accumulation, retention,
and transport of plutonium is strongly associated with sediment and sedimenting
particles. This idea is supported by results of numerous studies in which particulate
plutonium concentrations were measured apart from those dissolved (or suspended) in
water and interstitial water (Bartelt, Wayman, and Edgington, 1975; Dahlman, Bondietti,
and Eastwood, 1975; Emery, Klopfer, and Weimer, 1974; Hakonson, Nyhan, and
Purtymun, 1976; Johnson, Svalberg, and Paine, 1974; Magno, Reaney, and Apidianakis,
1970; Noshkin, 1972; Singh and Marshall, 1977; Trabalka and Eyman, 1976). These
studies show that plutonium associated with particulates makes up more than 80% of the
total plutonium concentrations in water. This characteristic of plutonium distribution
and transport in freshwaters is the most significant aspect of its envirorrm'^ntal behavior.

The pond is highly enriched with nutrients coming from laundry effluents via U-14
ditch (Fig. 1, Emery, Klopfer, and Weimer, 1974). Tliis nutrient supply supports
luxuriant growths of algae and macrophytes (Table 1) wliich eventually settle to the
bottom and decompose. The result of this process is the formation of a layer of organic
floe that rests on the surface of older floe and sediments. Although this material is
sedimentary, it has several special characteristics. The density of floe approaches that of
water, which causes it to be loosely compacted and easily resuspended. The large quantity
of floe generated each year serves as a source of food for many animal populations in and
around the pond. Perhaps most important, the floe is the primary concentrator of
plutonium in the ecosystem.

Hydrologic considerations of the pond provide additional significance to the
functional role of the organic floe in the ecosystem. Since the pond has no surface
outflow and a short retention time (40 hr), there is a rapid deposition of suspended
material (seston). The sedimentation rate is approximately 1 kgm^^ yr^^ (dry). This
means that about 5.6 x 10'* kg of seston is deposited each year in this loosely compacted
floc/sediment.

If U-Pond has sustained a continuous annual sedimentation rate of 5.6 x 10^ kg since
its formation in 1944, about 1.8 x 10^ kg of sestonic sediments has been deposited. The
total organic production of biomass in U-Pond is about 5 x lO'* kg/yr (Table 1), which is
approximately equal to the annual deposition of suspended matter. If losses of organic
matter in the sediments caused by decomposition are ignored, all sources of sedimentary
materials have deposited 3.5 x 10^ kg since 1944. The weiglit of U-Pond sediments down
to 10 cm is 3.4 X 10^ kg (dry. Table 1). This suggests that U-Pond has not deposited
more than about a 10-cm layer of sediments since its creation. The actual tliickness of
deposition would probably be smaller because of the decomposition of organic matter.
Tliis does not account for wind-blown dust accumulated in the pond.

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 635

10-

0 •••.0

10'

■D

10'

10^

POND PROFILE

■WATER;

ORGANIC FLOC
VJ Ml XTU RE •■■>•■.';<>

.UPPER
5 cm

;-i^;:V-:- SEDIMENT'-' " ;V:::' I LOWE^

COMBINED
10 cm

239,240

238

UPPER
5 cm

LOWER
5 cm

COMBINED
10 cm

Fig. 3 Distribution of plutonium isotopes in the upper 10 cm of U-Pond sediments. The
broken horizontal line indicates the pond-wide mean for both plutonium isotopes in a
10-cm core.

Horizontal distributions of plutonium in the sediments are spatially and temporally
random, but several vertical profiles indicate that plutonium is most heavily concentrated
in the upper 5 cm (Emery et al., 1976). When three sediment cores were analyzed for
plutonium concentrations in the upper 5 cm, lower 5 cm, and combined 10 cm, it was
found that most of the plutonium was located in the upper layer (Fig. 3). In these
samples the plutonium concentration in the upper 5 cm was about 4 x 10^ pCi/g,
whereas that in the lower 5 cm was only about 5 X 10' pCi/g. The combined 10-cm core
showed a plutonium concentration of about 2 x 10^ pCi/g. It should be noted, however,
that the concentrations in these samples did not resemble the mean plutonium
concentrations in samples of 10-cm sediment cores. The difference of plutonium
concentrations between the upper and lower sections of the sediment core in Fig. 3
appears to be exaggerated beyond the normal range, but a vertical reduction of plutonium
in the top 10 cm of sediments is indicated.

This vertical distribution of plutonium appears to be largely the result of the rapid
accumulation of sedimenting seston discussed earUer. Seston has shown the highest
concentrations of plutonium in any subcompartment sample, often greater than 1 X 10'
nCi/g. However, this material settles to the bottom and is captured by a layer of floe. This
accumulation of seston, decomposing plant material and a mixture of older sediments,
contains the largest portion of the pond's inventory.

636 TRANS URANJC ELEMENTS IN THE ENVIRONMENT

The existing mean inventory of plutonium in the sediments (to 10 cm) is 1.7 x 10^
nCi (Tables 3 and 4). Tliis amount of plutonium represents the total accumulation from
all sources minus the losses via the various routes of export. The validity of this can be
examined by calculating a theoretical accumulation of plutonium in the sediments using
annual rates of plutonium deposition via sedimenting seston and annual accumulation of
plutonium by pond biota.

The mean concentration of plutonium in U-Pond seston measured over the study
period is 5.6 X 10° nCi/m'^. Thus the pond's water mass (2.27 X 10'* m^) contains an
average of 1.3 X 10^ nCi of plutonium. This mass passes through the pond's basin at an
average rate of 219 times per year (i.e., 40-hr retention time). If we assume that the
flushing rate and plutonium content of U-Pond water remain constant, then 2.8 X 10^
nCi of sestonic plutonium is deposited as sediments each year. Proceeding with the same
assumptions, the 33-yr total theoretical accumulation of plutonium in the sediments is
9.2 X 10* nCi. The mean amiual accumulation of plutonium by pond biota is 2.4 x 10^
nCi (Tables 3 and 4), which suggests a historic total deposition of 7.9 X 10^ nCi if we
assume that each year biotic accumulation of plutonium is the same. This supply
increases the theoretical accumulation of plutonium in the pond's sediments to 1.0 X 10^
nCi. This quantity is also a theoretical expression of the historic supply of the plutonium
to the pond.

The Export

Percolation. No experiments were undertaken to measure the percolative loss of
plutonium from U-Pond, and defmitive conclusions about the movement of plutonium
from the sediments into the ground below the pond cannot be made. However, there are
indications that nearly all the plutonium that has reached the pond has been retained by
its sediments. In the above discussions, we concluded that the pond has deposited
about 10 cm of sediments since its formation in 1944. This was based on present-day
measurements of sedimentation processes occurring in the pond. It was also theorized
that about 1 Ci of plutonium has reached these sediments during the Ufetime of the pond.
Intensive samphng of the pond's sediments to a depth of 10 cm has shown that about
1.7 Ci of plutonium presently resides there. This agreement between theoretical and
observed accumulation of plutonium in U-Pond sediments suggests that the pond has
received about 1 Ci of plutonium and that most of it has been retained by its sediments.
The downward migration of plutonium in Hanford soils has been studied by several
workers to assess the seepage of reprocessing wastes from crib sites (Crawley, 1969;
Ames, 1974; Price and Ames, 1976). Their findings indicate a vertical reduction of
plutonium concentrations over the upper 10 m of the vadose zone (i.e., soil lying above
the water table). Price and An-^s (1976) found that particulate plutonium (>2 /im) was
deposited within the upper 1 m of the vadose zone. The nonparticulate plutonium (<2
)um), which was less than 0.5% (by weight) of the plutonium entering the vadose zone,
showed deeper penetration and was eventually deposited in association with the silicate
hydrolysis of sediment particles. Brown (1967), studying the vertical migration of other
long-hved radionuclides below disposal facilities, found that more than 99.9% (by
activity) of these materials was deposited within the upper 10 m of the vadose zone. In
addition to this, Myers, Fix, and Raymond (1977) indicate that plutonium concen-
trations in the groundwater below Hanford (at =^50 m) are not significantly different
from those of other areas.

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 63 7

Althougli we have not determined how much plutonium has percolated out of
U-Pond since its formation in 1944, the available evidence points to the retention of
virtually all of it in tlie sediments — probably in the upper several centimeters.
Furthermore, we find no evidence to indicate that plutonium has migrated from U-Pond
into the groundwater below Hanford. Thus we have no reason to beheve that percolation
is a significant route of plutonium export from U-Pond.

Emergent Insects. Insects emerging from U-Pond constitute the only direct route of
biological export. However, if the life cycles of these insects are considered, it appears
that emergence alone does not account for the export of the entire plutonium inventory
contained in this compartment. The cast exoskeletons left in the pond at the final
ecdysial stage prior to emergence may contain a substantial portion of the plutonium
burden of the insects. It is also possible that some of these insects complete their life
cycles without leaving the pond's ecosystem, and their plutonium burdens may ultimately
be returned to the pond. We will not attempt to estimate the fraction of plutonium that
is left in the pond by these processes, but instead we will assume that the entire inventory
of this compartment leaves the pond when these insects emerge.

With the foregoing considerations taken into account, the mean annual export of
plutonium by emerging insects is about 5 X lO^nCi. This quantity is about 9 X 10~^% of
the plutonium inventory of the biota and 6 X 10^^*% of the total pond inventory
(Tables 3 and 4).

Waterfowl. The waterfowl that contact the U-Pond ecosystem are mostly mallards and
an assortment of other ducks and coots. Some of these waterfowl nest along the shoreline
of the pond. Since it is unlikely that these waterfowl contact other locations where they
may be exposed to above-background levels of plutonium, it is assumed that most of the
plutonium found in their gut and tissues came from U-Pond.

Examination of crop and gut contents of ducks collected from U-Pond indicates that
they feed most heavily on the organically rich floe that covers the pond's sediments and,
to a lesser extent, on goldfish and other material. Recall that floe contains most of the
plutonium in the pond's ecosystem.

Concentrations of plutonium in whole bodies (including gut and contents) of four
wild ducks (Anas) ranged from 3 X 10"^ to 3 X 10° pCi/g, with a mean of 4 x 10"^
pCi/g. These ducks were in contact with the pond when they were sampled, and most of
tlieir plutonium burdens were contained in the gut. Less than 5% of their entire
plutonium burdens was contained in the body tissue.

Knowledge of the relationship between the length of time a duck spends in the
U-Pond ecosystem and the amount of plutonium accumulated would be useful.
Information about the contact frequency and duration is not available for waterfowl
sampled from the pond; thus there is no basis to establish this relationship. However, a
short experiment was performed to determine the amount of plutonium accumulated in
ducks (Anas) held on tlie pond in large cages for 5 days and fed a continuous diet of
organic floe (Emery and Klopfer, 1977). These experimental conditions represent the
highest potential for tlie accumulation of plutonium by a duck on a short-term basis.

Results of this experiment suggest that ducks could accumulate about 6x10° pCi
Pu/g (whole duck) in 2 to 5 days of continuous contact with U-Pond (Fig. 4). Under
these conditions the accumulation of plutonium in the gut could be around 7x10^
pCi/g, and tlie body tissue may concentrate about 3 X 10° pCi Pu/g. It is evident that in
the experimental ducks the gut contained most of the plutonium burden (>95%), which

638 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

10'

2 102

1-2

Experimental: After After

2 days 5 days

Whole duck O •

Duck without gut D ■

Gut and contents A A

Fig. 4 Total plutonium accumulated by experimental ducks after 2 and 5 days of
continuous contact with U-Pond compared with minimum, maximum, and mean
concentrations of plutonium in whole waterfowl occurring naturally on U-Pond. Mean
plutonium concentrations in experimental ducks are shown by location of appropriate
symbol within the ranges of data depicted by horizontal lines.

indicates that most of the plutonium accumulated while the ducks were in contact with
the pond would be lost soon after they flew away.

The mean plutonium concentrations of whole experimental ducks after 2 to 5 days of
continuous-contact with U-Pond (~6 X 10° pCi/g) were more than an order of magnitude
greater than the mean plutonium concentrations in whole wild duck samples (~4 X 10~ ^
pCi/g, Fig. 4). However, the maximum plutonium concentrations in whole wild ducks
were higher than the minimum concentrations in whole experimental ducks. This
comparison suggests an upper limit of contact duration, and it is concluded that wild
ducks have a plutonium burden that is less than that obtained from 2 days of continuous
contact with U-Pond.

The estimated total weight of waterfowl contacting U-Pond annually is 1.2 x 10^ kg.
Since the range of plutonium concentrations in whole wild ducks was from 3 X 10~^ to
3x10° pCi/g, the amount of plutonium exported by waterfowl annually could be from
approximately 4 x 10^ nCi to 4 x 10^ nCi. The mean plutonium concentration in whole
wild ducks of 4 X 10"^ pCi/g suggests a mean annual export of 5 x 10^ nCi of
plutonium. These export quantities are approximately four orders of magnitude lower
than the total plutonium inventory for U-Pond.

Birds. Birds observed around U-Pond are mainly sparrows, swallows, blackbirds, doves,
and shorebirds (Emery, Klopfer, and Weimer, 1974). Estimates of contact frequency

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 639

along with a mean weight for each taxon of bird suggest that approximately 5 x 10^ kg
of bird biomass moves through the air mass around U-Pond each year. Greater resolution
of this annual biomass quantity is limited by the lack of information on the magnitude of
bird activity in the pond region.

Samples of sparrow, swallows, and killdeer (total of 6) were analyzed for whole-body
content of plutonium (including gut and contents). Plutonium concentrations ranged
from less than 1 X 10~^ to 5 x 10° pCi/g, with swallows showing the highest
concentrations. This may be associated with their mud-gathering activities involved in
nest construction. The mean of these samples is approximately 2 x 10~* pCi/g.

An estimation of plutonium exported by birds suggests that approximately 1 X 10^
nCi is removed from the pond each year. The maximum export is approximately 2 X 10^
nCi of plutonium, whereas a minimum annual export is less than detectable. These export
quantities are below the estimated amounts of plutonium exported by waterfowl. They
also represent less than 1 X 10~^% of the total plutonium inventory for U-Pond.

Mammals. Mammals (other than man) that contact U-Pond are mice, rabbits, coyotes,
and deer. Only mice were sampled from the mammal population and were collected
within 10 m of the pond's shoreline.

The mouse population density around U-Pond is estimated to be not greater than 1
per 10 m^, or <6 X 10^ mice in a region around U-Pond that is equivalent to the pond's
surface area (~6 X 10^ m^). Since one mouse weighs approximately 2 g (dry), the mass
of this population is about 1 X 10^ kg. Whole-body samples of seven mice analyzed for
total plutonium showed a range of concentrations from 1 X 10~^ to about 1 X 10° pCi/g
(mean, ~5 x 10~^ pCi/g). This suggests that the mouse population may contain a
maximum of 1 X 10~^ nCi of plutonium, or a mean of about 5 X 10° nCi. It is not
known if all the plutonium found in mice came from U-Pond since there are regions
adjacent to the pond that have plutonium concentrations above background levels.

The coyote population on tlie Hanford Site (600 mi^) has been estimated to be
around 300 following spring breeding (Rickard et al., 1977). Occasionally coyotes are
observed feeding on goldfish at U-Pond. A contact frequency by coyotes is estimated to
be not greater than about 1 per day, or about 4x10^ coyote visits per year. If we
assume that each coyote removes 1 goldfish (3 g, dry weight) with each visit, the dry
weight of goldfish removed by coyotes each year is about 1 x 10^ g. The maximum
concentration of plutonium in goldfish is about 3 X 10^ pCi/g (Table 2), which suggests
that as much as 3 x 10^ nCi of plutonium might be exported by coyotes each year.

Rabbits and deer are seldom observed at U-Pond; less than 50 sightings of either
mammal have been made by the study team in over 3 yr. It is hkely that these mammals
visit U-Pond to drink water only; therefore we will consider their annual export of
plutonium to be negligible.

It is interesting to note that the largest plutonium export route among mammals is
the researcher. During a normal study year, approximately 2 x 10^ kg of samples are
taken from the pond, which contains about 2x10^ pCi Pu/g. This "export" quantity of
4 X 10'* nCi/yr as research samples appears to be greater than that caused by all other
mammals combined.

Wind. Attempts to quantify the export of airborne particulate plutonium from U-Pond
via wind were not made. It would be desirable to have data that express the rate of
plutonium movement away from the pond and its shore as airborne particles, but there
are many Umitations to this assessment. Sehmel (1977) reports that airborne plutonium

640 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

concentrations in samples taken at resuspension sites in the U-Pond region have been
significantly higlier tlian fallout levels in other areas (but still less than maximum
permissible concentrations). This region has other sources of airborne plutonium besides
U-Pond, and it was not possible to determine if the plutonium detected in these samples
came from U-Pond, fallout, or some adjacent source.

However, it appears that U-Pond does not contribute significantly to the plutonium
concentrations in the air downwind from the chemical processing areas. This is indicated
when plutonium concentrations in the air downwind from these areas are compared with
tlie concentrations in the air of distant upwind and downwind perimeter communities.
Differences between the plutonium concentrations of the two air masses were not
significant. In 1972, for example, the average plutonium concentration in the air of
distant communities was 1.8 x 10~^ pCi/m^, whereas the mean air concentration of
plutonium downwind from the chemical processing areas was 1.9 X 10~^ pCi/m'^
(Energy Research and Development Administration, 1975).

The premise that U-Pond does not release appreciable amounts of plutonium particles
through wind action is strengthened by an additional consideration. Between 1944 and
1955 U-Pond was occasionally flooded to the limits of its basin and subsequently
overflowed into an auxiHary basin. Since 1955 the pond has remained within its original
shoreline. Plutonium deposited as sediments while these areas were flooded is now
exposed to the movements of air. Although this exposed area is larger than the present
surface area of U-Pond, the plutonium concentrations of downwind air are not
significantly elevated by its presence.

Summary and Conclusions

In its 34-yr history, U-Pond has received an estimated 1 Ci of plutonium. Since the same
quantity presently resides in the sediments, it appears that U-Pond has retained nearly all
the plutonium that has been discharged into it.

In relative terms, sediments, submerged plants, and gastropods have the highest
concentrations of plutonium, ranging from 3.2 X 10° to 6.9 X 10^ pCi/g. Plutonium
concentrations of emergent plants and the remaining fauna range from 4.0 X 10~^ to
6.1 X 10' pCi/g. Emerging insects had the highest plutonium concentrations of the latter
group, ranging from 3.2 x 10' to 6.1 xlO' pCi/g.

The mean plutonium inventory of the sediment is 1.7 x 10' nCi, ranging from
1.3 X 10' to 2.0 X 10' nCi of plutonium (Fig. 5). This essentially represents the total
pond inventory since more than 99% of the plutonium in the pond is found in the
sediments. The mean plutonium inventory for the biota is 6 x 10^ nCi, ranging from
1 X 10^ to 1 X 10*^ nCi (Fig. 5). Among these, biota plant life contains more than 95% of
the plutonium. Diatoms and pondweed (Potamogeton) alone account for more than 99%
of the plutonium in plants. Emergent insects contain less than 1 X 10~'% of the
plutonium in biota and less than 1 X 10"^% of the plutonium in the pond. The inventory
of this compartment has particular relevance since it is the only direct biological route of
export from the pond. Remaining pond biota contain less than 1 X 10~^% of the total
plutonium inventory in the pond and can leave the pond only by the forces of external
export vectors.

If all emergent insects successfully leave the pond, they could export from 3.5 X 10^
to 7x10^ nCi of plutonium. These quantities are more than five orders of magnitude
lower than the total pond plutonium inventory (Fig. 5). Estimated quantities of

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 641

INVENTORY

r5n r^

10'

I-

D
_J
O-

_i
<

H
O

10'

10^

0 —

a.
_i
<

H
O

UJ
CO

1—6— I r-9— I

EXPORT

95%
Mean [-confidence
interval

■^ Lower limit of export
not determined

^^ Range of export not
determined

r^ T

LU
t/0

CO
Q
QC

_i
<

Fig. 5 Plutonium in ecological compartments of U-Pond compared with estimated
quantities that are exported from the pond annually. Export of plutonium by
percolation or wind does not appear to be significant.

plutonium annually exported by waterfowl range from 4 x 10^ to 4 x 10^ nCi, with a
mean annual export of 5 x 10^ nCi of plutonium (Fig. 5). Other birds appear to export
about 1 X 10^ nCi of plutonium each year, with a maximum of 2 x 10^ nCi (Fig. 5).
These export quantities are about six orders of magnitude lower than the total inventory
of plutonium in the pond. Mammals are estimated to export a maximum of 3 x 10^ nCi
of plutonium from the pond (Fig. 5) annually, which is at least five orders of magnitude
lower than the minimum total plutonium inventory of the pond. There is no apparent
significant export of plutonium from the pond via wind or percolation.

In conclusion, U-Pond has been exposed to plutonium longer than any other aquatic
system and has received about 1 Ci of ^^^'^'*°Pu and ^^^I*u. This 14-acre pond provides
a realistic illustration of the mobility of plutonium in a lentic or nonflowing ecosystem.
Although this pond has a rapid flushing rate, is highly enriched with plant nutrients, is
ecologically well established with a natural complexity of populations and diversity of
communities, and has continuous interaction with associated terrestrial life, it appears to
effectively bind the plutonium discharged into it and prevent it from moving significantly
into routes leading to man and other remote life. Fur^ermore, the environmental
behavior of plutonium in U-Pond appears to be quite similar to that of other aquatic

642 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

systems having vastly different ecological character. So long as this pond remains in its
present condition, the Ukelihood of its releasing hazardous quantities of plutonium to
man and his environment is very small.

Acknowledgments

Advice in the interpretation of data provided by Battelle Scientists R. F. Foster, D. G.
Watson, J.M. Thomas, L. L. Eberhardt, K. A. Gano, R. E. Fitzner, and R.J. Serne is
greatly appreciated. We also wish to express thanks to the Atlantic Richfield Hanford
Company for providing assistance in the funding of radioanalysis. This research was
funded by the U. S. Department of Energy, Division of Biomedical and Environmental
Research, under contract No. EX-76C-06-1830.

References

Ames, L. L., Jr., 1974, Characterization of Actinide Bearing Soils: Top Sixty Centimeters of216-Z-9
Enclosed Trench, USAEC Report BNWL-1812, Battelle, Pacific Northwest Laboratories, NTIS.

Bartelt, G. E.,C.W. Wayman, and D. N. Edgington, 1975, Plutonium Concentrations in Water and
Suspended Sediment from the Miami River Watershed, Ohio, in Radiological and Environmental
Research Division Annual Report, January -December, 1974, USAEC Report ANL-75-3 (Pt. 3),
pp. 12-11 , Argonne National Laboratory, NTIS.

Brown, D. J., 1967, Migration Characteristics of Radionuclides Through Sediments Underlying the
Hanford Reservation, m Disposal of Radioactive Wastes into the Ground, Symposium Proceedings,
Vienna, May 29-June 2, 1967, pp. 215-228, STI/PUB/156, International Atomic Energy Agency,
Vienna.

Carlander, K. D., 1955, The Standing Crop of Fish in Lakes,/ Fish. Res. Board Can., 12: 543-570.

Crawley, D. T., 1969, Plutonium-Americium Soil Penetration at 234-5 Building Crib Sites, Report
ARH-1278, Atlantic Richfield Hanford Company.

Dahlman, R. C, E. A. Bondietti, and E. R. Eastwood, 1975, Plutonium in Aquatic and Terrestrial
Environments, in Environmental Sciences Division Annual Progress Report for Period Ending
Sept. 30, 1974, USAEC Report ORNL-5016, pp. 1-8, Oak Ridge National Laboratory, NTIS.

Electric Power Research Institute, 1976, Plutonium: Facts and Inferences, C. L. Comar (Ed.), Report
EPRI-EA-43-SR, Electric Power Research Institute.

Emery, R. M., and D. C. Klopfer, 1977, Ecological Distribution and Fate of Plutonium and Americium
in a Processing Waste Pond on the Hanford Reservation, in Pacific Northwest Laboratory Annual
Report for 19 76 to the ERDA Assistant Administrator for Environment and Safety, B. E.
Vaughan (Ed.), ERDA Report BNWL-2100 (Pt. 2), p. 4.15, Battelle, Pacific Northwest Labora-
tories, NTIS.

, D. C. Klopfer, and W. C. Weimer, 1974, Ecological Behavior of Plutonium and Americium in a

Freshwater Ecosystem. Phase I. Limnological Characterization and Isotopic Distribution, USAEC
Report BNWL-1867, Battelle, Pacific Northwest Laboratories, NTIS.

, D. C. Klopfer, T. R. Garland, and W. C. Weimer, 1976, The Ecological Behavior of Plutonium and

Americium in a I-reshwater Pond, in Radioecology and Energy Resources, Fourth National
Radioecology Symposium, Corvallis, Ore., May 12-14, 1975, C. E. Gushing, Jr. (Ed.), pp. 74-85,
Dowden, Hutchinson and Ross, Inc., Stroudsburg, Pa.

Energy Research and Development Administration, 1975, Final Environmental Statement, Waste
Management Operations, Hanford Resen>ation, Richland, Washington, ERDA Report ERDA-1538
(Vol. 1),NTIS.

Hakonson, T. E., J. W. Nyhan, and W. D. Purtymun, 1976, Accumulation and Transport of Soil
Plutonium in Liquid Waste Discharge Areas at Los Alamos, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Francisco, 1975, pp. 175-189, STI/PUB/410, Inter-
national Atomic Energy Agency, Vienna.

Johnson, J. E., S. Svalberg, and D. Paine, 1974, The Study of Plutonium in Aquatic Systems of the
Rocky Flats Environs, Final Technical Report, Department of Radiation Biology and the
Department of Animal Sciences, Colorado State University.

MIGRATION OF PLUTONIUM FROM FRESHWATER ECOSYSTEMS 643

Magno, P., T. Reaney, and J. Apidianakis, 1970, Liquid Waste Effluents from a Nuclear Fuel
Reprocessing Plant, in Health Physics Aspects of Nuclear Facility Siting, Proceedings of the 5th
Annual Health Physics Society Midyear Topical Symposium, Idaho I'alls, Nov. 3-6, 1970,
pp. 208-220, Health Physics Society.

Myers, D. A., J. J. Fix, and J. R. Raymond, 1911, Environmental Monitoring Report on the Status of
Ground Water Beneath The Hanford Site. January -December, 1976, ERDA Report BNAVL-2199,
BatteUe, Pacific Northwest Laboratories, NTIS.

Noshkin, V. E., 1972, Ecological Aspects of Plutonium Dissemination in Aquatic Environments,
Health Phys.. 22: 537-549.

Patterson, J. H., G. B. Nelson, G. M. Matlack, and G. R. Waterbury, 1976, Interaction of ^^'PuO^
Heat Sources with Terrestrial and Aquatic Environment, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Francisco, 1975, pp. 63-78, STI/PUB/410, Inter-
national Atomic Energy Agency, Vienna.

Price, S. M., and L. L. Ames, 1976, Characterization of Actinide-Bearing Sediments Underlying Liquid
Waste Disposal facilities at Hanford, in Transuranium Nuclides in the Environment, Symposium
Proceedings, San Francisco, 1975, pp. 191-211, STI/PUB/410, International Atomic Energy
Agency, Vienna.

Rickard, W. H., et al., 1977, Densities of Large and Medium-Sized Mammals on the Hanford
Reservation, in Pacific Northwest Laboratory Annual Report for 1976 to the ERDA Division of
Biomedical and Environmental Research, B. E. Vaughan (Ed.), ERDA Report BNWL-2100 (Pt. 2),
p. 4.33, Battelle, Pacific Northwest Laboratories. NTIS.

Robbins, W. W., T. E. Weier, and D. R. Stocking, 195 7, Botany: An Introduction to Plant Science,
p. 204, John Wiley & Sons, Inc., New York.

Schell, W. R., and R. L. Watters, 1975, Plutonium in Aqueous Systems, Health Phys., 29: 589-597.

Sehmel, G. A., 1977, Radioactive Particle Resuspension Research Experiments on the Hanford
Reservation, ERDA Report BNWL-2081, Battelle, Pacific Northwest Laboratories, NTIS.

Singli, H., and J. S. Marshall, 1977, A Preliminary Assessment of 239,24opy Concentrations in a
Stream near Argonne National Laboratory, //£"c/r/2 Phys., 32: 195-198.

Trabalka, J. R., and L. D. Eyman, 1976, Distribution of Plutonium-237 in a Littoral Freshwater
Microcosm, Health Phys., 31: 390-393.

Verduin, J., 1964, Principles of Primary Productivity: Photosynthesis Under Completely Natural
Conditions, in Algae and Man, D. V. Jackson (Ed.), Proceedings of the NATO Advanced Study
Institute, Louisville, Ky., July 22-Aug. 11, 1962, pp. 221-228, Plenum Press, Inc., New York.

Plutonium in Rocky Flats Freshwater Systems

D. PAINE

This study was initiated to determine the behavior of plutonium in the freshwater aquatic
environs at the Rocky Flats Dow Chemical plutonium fabrication plant. Golden, Colo.
The principal study area included four holding ponds for waste solutions generated at the
plant complex.

Samples of biotic and abiotic components were collected from the spring of 1971
through the summer of 1973. These components consisted of sediment, water, seston,
zooplankton, fish, vegetation, and small mammals in close proximity to the aquatic
systems. Laboratory experiments were performed to quantify field results. Owing to the
high variability of plutonium concentrations in the environment, numerous samples were
collected and analyzed by a modified solvent-extraction liquid-scintillation counting
procedure.

Sediments were the major site of^^^'^^^Pu deposition. Coring analysis revealed the
largest concentrations at subsurface-sediment depths, and thus depth-profile data were
used in calculating total inventory. A retention function determined in the laboratory
demonstrated a rapid transfer of plutonium from water to sediment. Pond reconstruction
during the study period resulted in significant increases in mean-surfaceftop 5
cmj-sedimerit concentrations.

Seston contained 30 to 80% of the '^^^''^^^Pu in an unfiltered water sample.
Concentration ratios in seston, ranging from 10^ to 10^ , were higher than those found in
marine studies. No vertical distribution of '^^^'^'^'^Pu was noted in pond water.
Laboratory experiments suggested active uptake by algae rather than by simple surface
adsorption. Zooplankton showed a discrimination against plutonium concentration along
the simple phytoplankton-to-zooplankton food chain. Fish flesh and bone showed no
levels above minimum detectable activity (MDA, 0.03 d/min per 10-g sample for a
100-min count). Vegetation associated with pond sediments contained higher concentra-
tion ratios from sediment to aerial portions of plants than previously observed, ranging
fromlO"^ to 10-\

Although plutonium in the biosphere presently exists at very low concentrations, trophic
biomagnification and possible locaUzed contamination may result in increased plutonium
concentrations in organisms of higher trophic levels. CycHng processes and biological
uptake of plutonium must be understood before environmental releases so that rational
assessment of its potential hazard can be performed. The major concern with plutonium
is its potential hazard to man. Plutonium could enter man either directly through
inhalation of atmospherically suspended material or indirectly through incorporation into
his food chain. The inhalation route is considered the most hazardous mode of entry to
man (Taylor, 1973). However, the concentration of plutonium in sediments or in aquatic

644

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 645

BOULDER
(POP., 66,780)

BROOMFIELD
(POP., 7282)

DENVER METROPOLITAN
AREA (POP., 711,295)

Fig. 1 Rocky Flats installation relative to nearby population centers.

organisms frequently exceeds concentrations in surrounding waters by orders of
magnitude. This concentration process may pose unexpected hazards when considering
food-chain transport.

Before this investigation little information concerning plutonium movement in
aquatic systems was available (Stannard, 1973; Noshkin, 1972). Fallout and marine
studies comprised the bulk of this environmental data, with average concentrations in the
femtocurie range (Pillai, Smith, and Folsom, 1964; Aarkrog, 1971; Noshkin et al., 1971).
In general, all freshwater Siudies have concurred that sediments appear to be the major
reservoir for ultimate plutonium deposition and that relatively insignificant transport of
plutonium through biotic systems to man exists (Emery and Klopfer, 1976; Hakonson,
Nyham, and Purtymun, 1976).

The purpose of tliis investigation was to determine the behavior of plutonium in
freshwater systems at the Rocky Flats Dow Chemical plutonium fabrication plant,
Golden, Colo. The objectives were to (1) investigate the distribution patterns of
plutonium in the biotic and abiotic components of the Rocky Flats freshwater systems,
(2) determine any concentrating processes that were occurring, and (3) determine if any
biological mobilization processes existed. It was the first attempt of its kind at delineating
the cycling processes of plutonium using a holistic systems approach.

Methods and Materials

Figure 1 shows the location of the Rocky Flats area relative to the larger surrounding
metropolitan areas. The plant site itself covers approximately 10 km-^ .

Figure 2 shows the general sampling area at Rocky Flats. The principal study area
included the four holding ponds (B-series ponds) for waste solutions generated at the
plant complex. These ponds were drained by Walnut Creek, which flowed into Great
Western reservoir, the City of Broomfield's municipal water supply. Great Western's water
sources were provided by Walnut Creek (2%), Coal Creek (8%), and Clear Creek
watershed (90%) (Hammond, 1971). The A-series and C| ponds were monitoring ponds
that did not receive routine releases of plutonium waste. Pond Ai had received low-level

646 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Fig. 2 Study area of Rocky Flats environs showing ponds, streams, and reservoirs. Flow
on Woman and Walnut Creeks is from west to east.

plutonium contamination from past nonroutine releases. After the completion of this
study, pond A2 was constructed to handle excessive water runoff. This pond was not
investigated in this study. Pond Cj received runoff from the "pad" area located several
hundred yards due northeast. This area was previously contaminated by plutonium
from leaking 50-gal drums and was subsequently covered with an asphalt pad. The
southernmost pond (pond Ci) drained into Woman Creek, which flowed into Standley
Lake, an irrigation reservoir as well as the municipal water supply for Westminster, Colo.
Pond 7, located several miles northeast of the study area, was used as a control pond for
this study. Samples were periodically collected from pond 7 and used as background
correction.

Primary plutonium waste discharged to the pond complex included laundry wastes
and process waste solutions generated by various phases of the plant's operation. Owing

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 647

to the high variability of plutonium concentrations in the environment, numerous
samples were collected and analyzed by a modified solvent-extraction liquid-scintillation
counting procedure (Keough and Powers, 1970). Counting yield was 96%. Overall
chemical recovery was 90%. The minimum detectable activity (MDA) was 0.30 d/min per
sample for a 100-min count. Modifications and additional analysis information are
presented in a report by Johnson, Svalberg, and Paine (1974). Unless otherwise stated, all
references to plutonium in this chapter include both ^^^Pu and ^'^"Pu plutonium
isotopes because the analytical procedure did not discriminate between plutonium
isotopes.

From 20 to 30 surface-sediment cores (approximately the top 5 cm) were obtained
during each sampling period (~1 per month) from each pond. The procedure followed
that outlined by Hakonson (1972). Additionally, core samples were extracted from the
sediment beds of the pond, and the procedure defined by Johnson, Svalberg, and Paine
(1974) was used to determine vertical distribution of plutonium within the pond
sediments. Incremental samples were composited for analysis. The coefficient of variation
determined from composited sediment samples was approximately 30%.

Surface-water samples were initially taken at each sediment sampUng location from
each pond. Later samples were taken not only at the surface but also at 0.5-m increments
to the sediment- water interface. A mechanical water sampler was used to collect the
subsurface water samples.

Water samples were filtered in a MiUipore filtering apparatus that was modified by
adding a brass screen with a pore size of 250/Lim to the top of the water intake funnel. The
screen removed most of the zooplankton and large organic material from the water
sample. The water was pulled through a glass-fiber filter (Whatman GF/A, 4,7-cm
diameter) with a vacuum pump to remove the remaining suspended material. The residue
was analyzed as a separate component called seston, which included primarily
phytoplankton, detritus, and other suspended soHds. Seston as defined here does include
some small zooplankton. The filtration process was usually carried out within 24 to 48 hr
after the collecfion. Water samples were kept in darkness to inhibit growth until filtration
could be accomplished.

The term "zooplankton" was used collectively for all small planktonic animals
trapped in a number 10 plankton net (160-ium mesh size). These samples contained seston
as well as large aquatic insects. The insects were separated from the samples. It was
assumed that most of the sestonic material was probably smaller than the 160-/im mesh
size and would pass through the net since the zooplankton sample was rinsed in the
collecting net by repeated dunkings in the pond. Zooplankton were identified to species,
but biomass estimates were not determined. A 12. 7 -cm-diameter Clark— Bumpus plankton
sampler was towed behind a boat in an attempt to sample organisms.

Bass (Ictiobus bubalus) and carp (Cyprinus carpio) were collected by seining and
angling. No fish were present in the B-series ponds, but minnows (Hybosis sp.) were
collected in pond Ci and pond Aj with a large collecting net. The fish were too small for
angling, and seining would have disturbed the bottom sediments. The maximum fish
length observed in the ponds was approximately 6 cm. Vegetation (primarily Juncus
balticus, Rumex crispus, and Typha latifolid) was collected in and around the ponds,
streams, and reservoirs throughout the study period. Generally, the aerial portions of the
plant samples were clipped with grass shears, and the roots were extracted separately
owing to excessive sediment— soil contamination.

648 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 649

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650 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Small mammals {Microtus pennsylvanicus modestus and Peromyscus msutus) were
kill-trapped throughout the study period. Mule deer {Odocoileus hemionus) samples were
collected from road kills.

Results

Sediment

Figure 3 shows the average 2 3 9,2 4opjj concentrations in water and surface sediments over
the entire study period. The accidental release in March 1971 resulted when process waste
solutions low in 2 39,24 0pu content, due to be pumped to solar evaporation ponds for
concentration, were accidentally released to the sanitary waste system. The resultant
elevation over ambient conditions is readily apparent in Figs. 3 and 4. Reconstruction of
the pond area had a marked effect on the mean sediment concentrations. Concentration
levels increased significantly during construction and remained high during the rest of the
study period for most ponds (Table 1). Plutonium concentrations also increased in the
sediment sampled at Walnut Creek at Indiana Avenue (baseline reservoir), wliich indicated
that considerable activity escaped the holding-pond system during the period of
reconstruction (Fig. 4).

Pond Bi showed the higliest surface concentrations throughout the study period. A
slight decrease in mean sediment concentrations was noted in ponds B2 and B4 following
peak levels. However, pond B3, which was the first pond to undergo reconstruction and
underwent the most extensive remodeling, showed a marked decrease in sediment activity
after the period of maximum values. This can be attributed to the deposition of
appreciable soil that contained lower concentrations of plutonium during and after the
dam and pond remodeling.

Subsurface sediments were probably mixed when the ponds were refilled. Core
samples contained highest plutonium concentrations at 20- to 30-cm depths. Some minor
construction modifications were made in the effluent bypass system which could have
caused redistribution of high-level plutonium sediments from this area. Sewage-treatment
modifications in May 1972, before reconstruction, could also have resulted in
high-activity flocculate being released to the holding ponds.

Plutonium-239,240 concentrations in pond B3 sediment peaked in late June 1972;
those in pond B2 peaked in July, and those in ponds B, and B4 peaked in August. This
suggests that pond reconstruction played a major role in tlie redistribution of plutonium
since this is the order in which remodeling occurred. In any case it is readily apparent that
mean-surface -sediment values increased markedly during tlie period of pond reconstruc-
tion and remained at liigher levels except in pond B3 .

The clay sediments showed an extremely high affinity for plutonium, and, if left
undisturbed, they appear to be an excellent reservoir for plutonium in an aquatic system.

Water

The mean concentrations of plutonium in unfiltered water samples during the course of
this study (Fig. 3) showed that construction played a major role in the redistribution of
plutonium from pond to pond. The increase in plutonium concentrations was also
detected downstream at the Walnut Creek at Indiana Avenue sampling station (Fig. 4).

The majority of the plutonium in the water component was usually associated with
the filterable fraction (> 0.45 pm) (Table 2). However, ponds Aj and Ci , the monitoring

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 651

100

LLI

^10

• to

■u
l<

I LL
' CC

0.1

1971

,1. I I L

Baseline reservoir

Period of pond
reconstruction

J L

1972

J I \ L

J L

1973

J \ \ l_

J L

100

Q- _

AMJ J ASONOJFMAMJJASONDJFMAMJ J A

TIME, months

0.1

Fig. 4 Mean plutonium concentrations in surface sediment (pQ/g) and mean plutonium
concentrations in unfiltered water (pCi/liter) for baseline reservoir. This sampling station
is located where Walnut Creek crosses under Indiana Avenue.

TABLE 1 Mean-Surface(Top 5 cm)-Sediment ^^^'^'^^Pu

Concentrations During Preconstruction

and Postconstruction Periods

Preconstruction

Postconstruction

Pond

pCi/gt

pCi/g+

200 ± 70

1300+ 350

80 ±30

200 ± 60

30 + 10

200 + 200

20 ± 10

55 + 15

3 + 3

3 + 1

15 + 5

15 ±4

Baseline reservoir

3± 1

10 ±4

Bypass dam

750 ± 150

*n = number of sampling periods.

tMean ± standard error.

jMean of two samples taken in June 1971.

ponds that were separate from the holding-pond chain, received httle contaminated plant
effluent and contained a larger fraction of nonfilterable plutonium. Less suspended
material, including phytoplankton, in ponds Ai and Ci is probably the explanation for
this phenomenon. Because of the shallow nature of the pond systems, no apparent
vertical distribution of plutonium could be determined.

652 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Percent of Plutonium Isotopes

Associated with Filterable Fraction of Water

Samples from Rocky Flats Ponds

Pond

Filterable fraction *

90 ±6

80 ± 12

80 ±8

70 ± 12

c»

30 ± 30

35 ±20

*Mean ± standard error.

Laboratory experiments were performed to study the transfer of plutonium from
water to sediment as a function of time. The function appeared to consist of two
exponential terms and was described by the equation

C(t) = Co (0.75 e-o-^2^ + 0.25 e-°-5«^) (1)

where C(t) is the concentration of plutonium in water at any time t, t is time (in days),
and Co is initial concentration of plutonium in water.

This experimental finding fits remarkably well with actual pond limnological data.
The average S. Walnut Creek flow into pond Bi was measured to be 480 m^/day during
1971. The water volume of pond Bi was calculated to be 1500 m^. Therefore the mean
Ufetime of any parcel of water in pond Bi , if mixing is uniform, can be calculated as
follows:

^ 1500 m^ - ^ ,
t = 7^7; — rr; — = 3.1 days
480 m^ /day

According to Eq, 1, 75% of the plutonium in water delivered to pond Bi should be
exchanged to sediment in an average residence time of 3.1 days,

Plutonium concentration as a function of sediment depth for the ponds is illustrated
in Table 3. These data were plotted and integrated by a planimeter to determine the area
(picocuries-centimeter per gram). This value was then divided by the mean sediment depth
to give the mean sediment concentration of plutonium. When multiplied by the estimated
sediment volume, these data yielded plutonium inventories for the sediment of the ponds.
The same sediment cliaracteristics were assumed for each pond. The variation could be
due to shunting of water past ponds at unknown times. The calculated value also assumed
that no plutonium was being transferred from pond to pond by suspended materials. The
agreement between the calculated and measured inventories is shown for the holding
ponds in Table 4.

Seston

Seston was defined as primarily phytoplankton, some detritus, and some small
zooplankton. Planktonic algae constituted by far the majority of aquatic plant material
found in the holding-pond chain on S. Walnut Creek (B ponds).

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 653

TABLE 3 Distribution of Plutonium

Concentrations in Depth Profiles of Ponds

B,, 82,63,64, and Ci

Plutonium concentration,

(d/min) g '

Depth, cm

0-5

2,000

260

430

5-10

2,200

170

190

10-15

10,900

370

15 -20

32,100

330

20-25

8,800

230

25 -30

1,100

190

30-35

880

390

35-40

900

340

380

40-45

190

100

45 -50

50-55

100

55 -60

60-65

70-75

75 -80

80 - 85

85 -90

90-95

95 - 100

100- 105

105-110

110- 115

115 - 120

TABLE 4 Calculated and Measured
Inventories of ^ ^ ^ '^ "* ° Pu in
Holding Ponds During 1971

Measured

Pond

mCi

Calculated,* %

84.5

62.4

74.8

27.0

19.9

21.9

19.4

14.3

2.9

4.6

3.4

0.4

*Calculated values are from the
retention function obtained from lab-
oratory experimentation.

654 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 5 Concentration Ratios (CR's) in Seston,
Zooplankton, and Crayfish Relative to Filtered Pond Water

No. of sampling

Pond

periods

CR*

Seston

7.5 ± 1.1 (10'')

100 ± 20(10")

1.7 + 2.8(10")

2.5 ±5.7(10")

16 + 3.7(10")

12 + 3.4(10")

Zooplankton

0.14 + 0.02(10")

0.17 + 0.02(10")

Whole crayfish

0.13 + 0.002(10")

0.07 ±0.019 (10")

0.06 ±0.003 (10")

*pCi ' ^ ' '^ " °Pu/g acceptor - pCi ' ^ ' '^ " ° Pu/ml water. Mean ±
standard error.

The transfer of plutonium from water to seston was extremely high (Table 5). The
concentration ratios (CR's) relative to filtered water were of the order of lO'* to 10^.
Concentration ratio is defined as

_ picocuries per gram seston (dry weight)
picocuries per milliliter water (filtered)

These CR's were higher than those previously observed in marine systems. Laboratory
experiments revealed that the mechanisms involved were more than simple surface
sorption (Johnson, Svalberg, and Paine, 1974).

Zooplankton

Although several species of cladocerans, copepods, and amphipods were collected,
sufficient biomasses for analysis were never obtained at any one sampling period. This
necessitated a pooling of the samples over several months. This was especially true for the
B-series ponds, which contained almost no zooplankton throughout the study. Zoo-
plankton showed CR's relative to filtered water in the IC* range (Table 5), These CR's
are similar to those reported in marine studies. If ingestion is the primary route of
transfer in these organisms, then higher concentration factors would be expected from
the simple phytoplankton-to-zooplankton food chain. Since an increase in trophic-level
concentration of plutonium did not occur, there appears to be a selective mechanism that
discriminates against plutonium at this level. This would result in a decreased potential
hazard when considering the transfer of plutonium through ingestion routes.

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 655

TABLE 6 Concentration of ^ ^ ^ '^ ^ ° Pu in Fish Inhabiting
Rocky Flats Environs

Concentration,!

Sample type

Location

Sample

pci/g

Minnow (Hybosis sp.)

Whole

1.7 ± 0.2

Minnow {Hybosis sp.)

Whole

5.1 ± 1.8

Carp {Cyprinus

Great

Whole and

<0.02

carpio)

Western

dissected

Bass {Ictiobus

Pond 7

Whole and

<0.02

bubalus)

dissected

Minnow (Hybosis sp.)

GI tract

Flesh

Head

Skin

Bone

0.6 ± 0.7

<0.02

0.9 ± 0.9

2.3 ± 0.4

<0.02

Minnow (Hybosis sp.)

GI tract

Flesh

Head

Skin

Bone

0.9 ± 0.9

<0.02

2.3 ± 2.2

4.6 ± 4.2

<0.02

*Number of composite fish samples analyzed (5 fish/composite),
f Mean ± standard error.

Crayfish

Crayfish, a large invertebrate common to the pond system, showed CR's relative to
unfiltered water in the range of 320 to 1290 with a mean value of 830 (Table 5). These
values are similar to those found in other studies. Seventy -seven percent of the plutonium
in crayfish was associated with the exoskeleton, even though the crayfish were scrubbed
extensively. The benthic origin of these organisms probably explains the higli plutonium
concentrations associated with the exoskeleton.

Fish

Fish flesh and bone from ponds A and C and reservoirs were never above MDA (0.30
d/min per sample) even when several samples were composited (Table 6). Whole fish,
however, contained measurable amounts of plutonium in the gut contents, the head, and
the outer skin. This suggests that plutonium is being discriminated against at this trophic
level.

Flora

No true aquatic vascular plants and relatively few emergent species existed in the pond
systems at Rocky Flats. Bulrush (Junciis balticus) rooted sporadically within the ponds,
and cattail (Tvpha latifolia) frequently grew with its roots submerged. Dock (Rumex
crispus) was abundant in the riparian area. Concentration ratios for plants associated
closely with pond sediments confirmed the observation that the transfer of plutonium
from sediments to aerial portions through roots is higher than that previously reported in
laboratory experiments (Romney, Mork, and Larson, 1970) (Table 7). Concentration
ratios were in the 10~- to 10~* range. This could suggest that the plutonium associated

656 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Concentrations of ^^^'^'**'Pu in Vegetation

Samples (Juncus halticus, Rumex crispiis. and Typha

latifolia) Associated with the Rocky Flats Pond System

Concentrati

ion,

pCi/g

(dry weight)

Mean

Min.

Max.

Total roots

11.2

0.31

93.2

2.03

Total standing vegetation

5.1

0.01

44.3

2.00

Aj roots

1.69

0.31

4.77

0.76

A, standing vegetation

1.47

0.28

3.68

0.80

Bj roots

45.4

2.46

93.2

0.76

B, standing vegetation

18.9

1.23

44.3

0.80

Bj roots

2.45

0.67

4.39

0.55

Bj standing vegetation

2.M

0.01

8.27

1.26

B3 roots

1.16

1.91

1.41

0.30

B3 standing vegetation

0.45

0.33

0.56

0.36

B4 roots

3.89

1.49

6.22

0.50

B4 standing vegetation

1.28

0.21

3.94

1.06

C, roots

2.84

0.55

7.38

0.55

C, standing vegetation

1.83

0.12

5.66

0.93

*Number of samples.

with the ponds is of a more biologically available form. This appears contradictory to
laboratory experiments which do not include a variety of environmental factors that
could contribute to an increased uptake of plutonium, such as surface contamination.

Fauna

A variety of small and large mammals were opportunistically captured during the course
of this study. The data associated with this compartment were too few except to draw
tentative conclusions. However, it would appear that fauna associated with the Rocky
Flats area, in general, maintained a relatively low systemic body burden of plutonium
(Table 8).

Conclusions

The results obtained in this study were of a very preliminary nature because of the more
general systems approach to the study and to the use of an analytical technique that
provided no isotopic discrimination. Owing to the cost of sophisticated sample analyses, a
majority of activity levels near fallout background and /or analytical detectability, and the
overall complexity of a systems approach, only tentative conclusions can usually be
ascertained for transuranic elements in the environment. However, the tentative
conclusions drawn from this study and others are, in general, the same.

Althougli the various components of the aquatic system at Rocky Flats are
concentrating plutonium to a relatively high degree, there appears to be no direct
evidence that concentrations of plutonium observed will result in a biological hazard to

PLUTONIUM IN ROCKY FLATS FRESHWATER SYSTEMS 657

TABLE 8 Plutonium Concentrations in Some
Animals Collected at Rocky Flats

Odocoileus hemionus (Rocky Mountain mule deer)

Concentration,
pCi/g (dry weight)

Sample

Mean

Min.

Max.

n=*

Spleen

0.03

0.02

0.05

0.42

Kidney

0.08

0.05

0.10

0.41

Lung

0.03

0.01

0.10

0.90

Broncheoles

0.07

Bronchus

0.08

Liver

0.03

0.01

0.09

1.43

Heart

0.01

Hide

0.06

0.03

0.16

0.74

Lymph nodes

(broncheolar)

0.33

Esophagus

0.08

Rumen contents

0.05

0.01

0.15

1.08

Blood

0.02

0.01

0.02

0.30

Muscle

0.33

<0.01
Mean SD

1.80
n

1.05

Internal = 0.12 ± 0.15 7
Exte- .al = 0.06 ± 0.02 6

Microtus pennsylvanicus modestus (meadow mouse)
Sample pCi/g (dry weight)

Liver

0.58

Lungs

5.10

GI tract

0.17

Bone

0.06

Peromyscus nasutus (white-footed deer mouse)

Sample pCi/g (dry weight)

Liver 0.99

Lungs 40.10

1 lesh 0.07

Bone 0.55

Rana pipiens (leopard trog)
Sample pCi/g (dry weight)

Liver

II. (11

Lungs

14.411

llesh

( 1. 1 7

Bone

0.31

Number u| samples.

638 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

man through ingestion routes. This was concluded on the basis that (l)the majority of
plutonium in the system was associated with sediments; (2) plutonium in unfiltered water
leaving the Rocky Flats plant site averaged <10 pCi/liter, even during pond reconstruc-
tion, which was below accepted maximum permissible concentration (1600 pCi/liter,
International Commission on Radiological Protection); and (3) plutonium concentrations
did not increase along simple trophic -level routes to any significant extent.

References

Aarkrog, A., 1971, Radioecological Investigations of Plutonium in an Arctic Marine Environment,

Health Phys., 20: 31-47.
Emery, R. M., and D. C. Klopfer, 1976, The Distribution of Transuranic Elements in a Freshwater

Pond Ecosystem, in Environmental Toxicity of Aquatic Radionuclides: Models and Mechanisms,

pp. 269-285, M. W. Miller and J. N. Stannard (Eds.), Ann Arbor Science Publishers, Ann Arbor,

Mich.
Hakonson, T. E., 1972, Cesium Kinetics in a Montane Lake System, Ph.D. Dissertation, Colorado

State University, Fort Collins, Colo.
- — , J. W. Nyham, and W. D. Purtymun, 1976, Accumulation and Transport of Soil Plutonium in

Liquid Waste Discharge Areas at Los Alamos, in Transuranium Nuclides in the Environment,

Symposium Proceedings, San Francisco, Nov. 17-21, 1975, pp. 175-189, STI/PUB/410,

International Atomic Energy Agency, Vienna.
Hammond, S. E., 1971, Industrial-Type Operations as a Source of Environmental Plutonium, in

Proceedings of Environmental Plutonium Symposium, Los Alamos, N. M., Aug. 4-6, 1971, E. B.

Fowler, R. W. Henderson, and M. F. MiUigan (Coordinators), USAEC Report LA-4756,pp. 25-35,

Los Alamos Scientific Laboratory, NTIS.
Johnson, J. E., S. Svalberg, and D. Paine, 1974, Study of Plutonium in Aquatic Systems of the Rocky

Rats Environs, Final Technical Report, Dept. of Radiology and Radiation Biology and the Dept.

of Animal Sciences, Colorado State University.
Keough, R. F., and G. J. Powers, 1970, Determination of Plutonium in Biological Materials by

Extraction and Liquid Scintillation Counting, .4/;a/. Chem. ,42: 419-421.
Noshkin, V. E., 1972, Ecological Aspects of Plutonium Dissemination in Aquatic Environments,

Health Phys., 22: 537-549.
— -, V. T. Bowen, K. M. Wond, and J. C. Burke, 1971 , Plutonium in North Atlantic Ocean Organisms

Ecological Relationships, in Radionuclides in Ecosystems, Proceedings of the Third National

Symposium on Radioecology, Oak Ridge, Tenn., May 10-12, 1971, D.J. Nelson (Ed.), USAEC

Report CONF-710501, pp. 681-688, Oak Ridge National Laboratory, NTIS.
Pillai, K. C, R. C. Smith, and T. R. Folsom, 1964, Plutonium in the Marine Enviionment, Nature

(London), 203: 568.
Romney, E. M., H. M. Mork, and K. H. Larson, 1970, Persistence of Plutonium in SoU, Rants, and

SmaU Animals, Health Phys., 19: 487-491.
Stannard, J. N., 1973, Chemical and Physical Properties of Plutonium, in Uranium and Plutonium

Transplutonic Elements, pp. 670-686, H. C. Hodge, J. N. Stannard, and J. B. Hursh (Eds.),

Springer- Verlag, New York.
Taylor, D. M., 1973, Chemical and Physical Properties of Plutonium, in Uranium and Plutonium

Transplutonic Elements, pp. 323-347, H. C. Hodge, J.N. Stannard, and J. B. Hursh (Eds.),

Springer-Verlag, New York.

Plutonium in the Great Lakes

M. A. WAHLGREN, J. A. ROBBINS, and D. N. EDGINGTON

Since 1971 plutonium concentrations have been measured annually in Lake Michigan and
Lake Ontario and at less frequent intervals in the other Great Lakes. The concentrations
of plutonium in the water column have decreased only slightly during the 7 yr of
measurement. Tfie residence times for plutonium in the lakes have been estimated by
simple time-concentration models. Tfie apparent sinking rates for plutonium have been
found to be essentially constant in all the Great Lakes, which suggest that the basic
processes that control the concentrations of dissolved plutonium are similar despite
considerable differences in chemical, biological, and physical cliaracteristics of the lakes.
Analyses of plutonium in water, suspended solids, material from sediment traps, and
sediment cores show that considerable resuspension of previously sedimented material
into the hypolimnion occurs throughout a major part of the year. A mechanism is
proposed to account for the seasonal cycling of plutonium in the epilimnion of Lake
Michigan. Recent studies show that plutonium in Lake Michigan (and in the Irish Sea)
exists primarily in the water column as Pu(VI) and on the sediments as Pu(IV). For a
better understanding of the long-term geochemical and biological behaviors of plutonium
in aquatic environments, further study of the limnological factors that control the
chemical forms of plutonium is required.

Approximately 40% of the population of the United States lives in states bordering the
Laurentian Great Lakes (Fig. 1). The economic advantages to the electrical generating
industry of using these lakes for once-through cooling have long been recognized in both
the United States and Canada. The advent of large multiple-unit nuclear plants has led to
the operation of 8 such reactors on Lake Michigan and a total of 16 on the four lower
Great Lakes. The rapid growth of the nuclear power industry has generated considerable
public concern about possible environmental effects of radioactive discharges, whether
routine or accidental, and this concern has been directed primarily toward plutonium.

Very little, if any, plutonium from nuclear power plants has entered the lakes. The
source of plutonium in the Great Lakes is almost entirely stratospheric fallout as a result
of nuclear weapons testing. Because of the very low concentration and consequent
analytical difficulties of plutonium, neither concentrations nor inventories of ■^■'^'■^'*°Pu
or ^^^Pu were measured during the period of maximum fallout. However, excellent
records exist for the deposition of ^°Sr from the worldwide fallout monitoring program
(Environmental Measurements Laboratory, 1978). Therefore, if a general 239,240p^^
^°Sr ratio can be assigned, it is possible to obtain a reasonable estimate of the annual
deposition of plutonium. Measurements of plutonium and strontium in the atmosphere
since 1965 have been summarized by Krey, Schonberg, and Toonkel (1974). who found

659

660 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

Fig. 1 Laurentian Great Lakes, with water sampling sites for comparative studies. O,
May lOto June 6, 1973. A, July 1 to 3, 1974. •, June 30 to July 6, 1976.

that a ^^^•^'*°Pu/^°Sr ratio of 0.017 should give a good estimate of plutonium
deposition. The annual inputs- of plutonium to Lake Michigan, which lies almost wholly
within the 40°N latitude band, are shown in Fig. 2. A knowledge of this source term is
important for the interpretation of residence times in the water column and concentra-
tion profiles in the sediments. Measurements of the cumulative deposition of plutonium
in soil at Argonne (~2.2 mCi/km^ ; Golchert, Duffy, and Sedlet, 1978) are in agreement
with the integral of all the inputs shown in Fig. 2 and the worldwide fallout data
summarized by Hardy, Krey, and Volchok (1973). On this basis the total content of
plutonium in the Great Lakes basin is 1660 Ci (or 25 kg), of which 540 Ci was deposited
directly onto the surface of the lakes. The plutonium content in each lake is summarized
in Table 1. Sprugel and Bartelt (1976) have measured the loss of plutonium from a
typical midwestern watershed to be about 0.05% of the total accumulated deposition per
year.

The presence in the Great Lakes waters of these low concentrations of long-lived
radioactivity from nuclear fallout provides the opportunity to characterize the environ-
mental behavior of these isotopes and to study the biogeochemical and geophysical
processes that determine the residence times of radioactive and stable trace materials
entering the lakes. For plutonium it is of particular importance to determine (l)the
potential pathways to man (food chains and drinking water); (2) a radiological baseline
data set for the Great Lakes; and (3) the likely distribution of possible future inputs
between various compartments of the lake, including the tlnal sinks, if any.

PLUTONIUM IN THE GREAT LAKES 661

1955

1960

1965
YEAR OF INPUT

1970

1975

Fig. 2 Estimates of the annual deposition of 2 3 9,2 4op^j ^^^ ^j^jj ^^^ ^^^j. Lake
Michigan on the basis of monitoring ^°Sr in rainwater at Aigonne, III., and Green Bay,
Wis., the average rainfall, and the monitoring of 2 3 9,2 4npjj ^J^ atmospheric fallout at
Argonne.

TABLE 1 Inventory of Plutonium
in the Great Lakes as of 1977

Plutonium,

Deposited on

Lake

Watershed

lake surface

Water

Sediments

Superior

290

180

175

Michigan

270

130

128

Huron

290

130

128

Erie

130

Ontario

140

50*

*The amount stored in the sediments is greater than that
deposited on the lake surface because of the ^"Pu exported
from Lake Erie down the Niagara River.

The purpose of this chapter is to describe mainly the biogeochemical and physical
processes that appear to determine the behavior of plutonium and other transuranic
elements in the Great Lakes. The roles of recent atmospheric and watershed erosion
inputs, sedimentation and resuspension, and export by outflow as controls of the longer
term availability of plutonium are discussed. Rapid sedimentation from the epilimnion by
association with autochthonous particulate material is shown to account for the seasonal
cycling of plutonium in Lake Michigan. The methods for measuring the transuranic
elements in the lakes have been adequately described elsewhere (Golchert and Sedlet,
1972; Nelson etal., 1974; Wahlgren et al.. 1976).

662 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Long-Term Behavior of Plutonium in the Great Lakes

The earliest measurements of plutonium in any of the Great Lakes were made in 1971
(Maletskos, 1972). Since that time measurements of plutonium have been made every
year in Lake Michigan and Lake Ontario (Bowen and Noshkin, 1972; 1973; Bowen, 1974;
1975; 1 976 ;-l 977 ;Wahlgren and Nelson, 1973; 1974a; 1974b; 1975; 1977a; Wahlgren and
Marshall, 1976; Wahlgren, Nelson, and Kucera, 1977, unpublished) and occasionally in
the other lakes. The annual data for plutonium for the years 1971 to 1977 are
summarized in Table 2. Since it was found in 1972 that the fallout radionuclides ^^^Pu,
^ ^^Cs, and ^°Sr in Lake Michigan have a homogeneous distribution throughout the water
column following the winter mixing period (Walilgren and Nelson, 1973), the
concentrations in early spring or from the hypolimnion during early summer may be
taken to represent the mean for the lake. Therefore the numbers in the table are the best
estimate of the total concentration of each nuclide remaining in the water column each
year. From these data it is possible, knowing the volume of the lake, to calculate the
inventory of plutonium in the water column. As indicated in Table 1, the fraction of the
total amount in each of the lakes in the water column at the present time is very small.
Since very little plutonium can be lost by outflow from the upper Great Lakes (Michigan,
Superior, and Huron), it follows that there must be a very efficient transfer of this
element to the sediments. In the lower lakes (Erie and Ontario) losses due to outflow and
gains from the upper lakes must be considered as well. The mean ^^^Pu concentrations
for the three upper lakes (about 0.6 fCi/liter) are significantly different from those for
Lake Erie and Lake Ontario (about 0.2 fCi/hter). At first this would appear to be due
solely to the greater outflow in the last two lakes. However, the situation is more
complicated and interesting.

To understand the long-term behavior of plutonium in the Great Lakes, one must
consider not only the inputs and losses but also the volume and area of each lake and the
efficiency of the scavenging of plutonium from the water column. To the extent that
each lake is well mixed, the change in concentration of plutonium in the water column is

^ = ;^(A0 + aW+l-S-O) (1)

where V = volume of the lake
A = area of the lake

0 = plutonium flux to the lake surface, femtocuries per square centimeter per year
W = amount stored in the watershed, femtocuries
a = annual fraction lost to the lake

Since the lake is assumed to be well mixed, the amount lost by outflow, 0, is given by
QC, where Q is the mean annual outflow from the lake (cubic centimeters per year). In
addition, 1 is the input (femtocuries per year) from the next higher lake, where 1 = QC,
and S is the amount lost to the sediments (femtocuries per year). Introducing subscripts
i=l to 5, wliich refer, respectively, to Lakes Superior, Michigan, Huron, Erie, and
Ontario, the change in concentration in each lake is given by

dCj 1 Qi Si

— •=-(Ai0i + aiWi + Qi_, Ci_i)-^Ci-;^ (2)

PLUTONIUM IN THE GREAT LAKES 663

c«

<1>

■^^

+-»

♦-

• ^^

-^rf

«4-l

1/3

■*«(

-4i-*

-J

>■

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o ^

d d

+( +(

CX5 VO

in lo

.^ «N

'^ -H

■^—^ Nw^

O <N

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r-H 1-H

d d

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o o

.^ ,

r— *

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d d

d
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o

d
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d

q
d
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d
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d
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o
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c
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■a

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E
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664 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

The ratio Vj/Qj is the residence time of plutonium in the lake with respect to outflow
(Trj). If the rate of loss by sedimentation is proportional to the concentration in the
water column, then the term Sj/Vi can be replaced by Ci/Tk;, where Tr; is the residence
time of plutonium in the water column with respect to losses via sedimentation. In this
case Eq. 2. can be written as

d^i _ Ai0i — 1 '<i-i -1-1 I " " I p (-.^

The mean residence time Tr; of plutonium in each lake is given by

111

^ZJ- (4)

TRi Trj T'rj

For Lake Superior and Lake Michigan, there is no inflow from the other lakes, i.e.,
Ii ~ I2 ~ 0 (Fig. 1); for Lake Huron, however, I3 = Q2C2 + QiCi . The above systen. of
equations (Eq. 3) is equivalent to the concentration— time model described by Lerman
(1972) to describe the behavior of ^°Sr in the Great Lakes. For plutonium, losses by
radioactive decay are, of course, negligible. Since the values of Ai0i, Vj. and Qj are known
(Table 3) and a can be assumed as a first approximation to be zero, the only
undetermined parameter in the model is T'r.

TABLE 3 Physical and Hydrological Data
for the Great Lakes

Area, 10^

km'

Drainage

Lake

Volume,

Outflow,

Lake

basin

surface

10^ km^

km^ yr

Superior

12.8

8.2

Michigan

n.8

5.8

4.9

Huron

13.1

6.0

4.6

157

Erie

5.9

2.6

0.48

175

Ontario

6.0

2.0

1.6

209

The results of evaluating Eq. 3 are shown in Fig. 3. The monthly values of 0j used
were taken from Lerman (1972). For deposition from 1973 to 1977, values measured at
Argonne National Laboratory were used for each lake (Golchert. Duffy, and Sedlet,
1978). Values of T'r were chosen to reproduce the earliest available measured
concentration in each lake. This approach was taken to evaluate subsequent changes in
the value of T'r. In all lakes but Lake Erie, the mean residence time is around 2 to 3 yr.
Under such conditions short-term (monthly) fluctuations in deposition are averaged out
in the water column, and the calculated time— dependence of plutonium levels shows a
smooth variation over the past 25 yr or so. Since plutonium in waters of the Great Lakes
has been measured at the "tail-end" of the concentration— time record, reconstruction of
previous levels in the water column is an exercise in extrapolation. However, the choice of
residence times, T'r, summarized in Table 2, gives a very reasonable prediction of recent
levels in all the lakes as well. In other words, values of T'r, which by definition correctly

O '

2 -

0
1950

PLUTONIUM IN THE GREAT LAKES 665

I I I I I

1 rn \ r

Lake Superior _

Lake Michigan —
Tr = 2.4 yr

Lake Huron
Tr = 2.3 yr

1960 1970

YEAR

1950

1960 1970

YEAR

Fig. 3 Comparison of concentrations of 2 3 9,240py jj^ ^j^^ ^vater column predicted by
the coupled-lakes model with available experimental data points (•). The best estimate
of the residence time for deposition in the sediments is given for each lake.

predict the concentration of plutonium in the early 1970s, give an adequate account of
levels in each lake (except Lake Superior, for which there is only one data point) 3 to 7
yr later. For Lake Michigan, where plutonium has been measured each year since 1971,
the model appears to slightly underestimate concentrations from 1973 onward.
Plutonium levels in the water of Lake Erie show comparatively strong fluctuations over
the past two decades because of the very short mean residence time resulting from rapid
losses to sediments.

The small differences between observed and calculated mean plutonium concentra-
tions in Lake Michigan after 1973 could be due to a combination of small effects because
levels were so low in the 1970s. For example, a 20% increase in the Tr gives a better
least-squares fit to all the mean concentration data. Thus the lower value of Tr resulting
from use of earliest (1971) concentration values alone could be an artifact of the
approach or reflect uncertainties in the estimate of the mean concentration in 1971.
Alternatively, the sliglitly liigher recent value of Tr could reflect an increasing
importance of sediment— water exchange or watershed erosion in the regulation of the
very low plutonium concentrations in Lake Michigan. Unfortunately there are insufficient
data to discuss the other lakes in these terms.

Little is presently known about the inputs of plutonium from watersheds of the Great
Lakes. The few measurements of the concentration of plutonium in the Grand River, one
of the major tributaries to Lake Michigan, suggest total concentrations of 0.5 to 1.0
fCi/hter. If this range is representative of average concentrations, tributary rivers would
contribute up to ~0.1 Ci/yr at the present time compared with ~0.7 Ci/yr from direct
fallout to the surface of Lake Michigan.

The recent results of Sprugel and Bartelt (1976) suggest that watershed contributions
may be more important than previously supposed. They found that 0.05% of the total
plutonium stored on a typical midwestern watershed is lost by erosion each year. If this

666 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

erosion rate applies to the large watersheds of the Great Lakes, the annual input to Lake
Michigan, for example (see Table 1), should be ~0.14 Ci/yr (assuming that there is 270
Ci of plutonium deposited on the watershed). The possible effect of such a 0.05%/yr
erosion rate on plutonium concentrations in Lake Michigan since 1973 is illustrated in
Fig. 4. Equation 3 is evaluated on a monthly time interval starting in 1973, with initial
conditions determined by observed concentrations in the lake, and T'r = 2.4 yr. Montlily
inputs from the rivers are estimated by the crude assumptions: (l)The total inflow of
water to Lake Michigan is proportional to that from the Grand River (U. S. Geological
Survey, 1973 — 1976; U. S. Geological Survey, personal communication, for 1977 data);
and (2) the concentration of plutonium in river water is constant. Clearly the addition of
plutonium from the watershed at a rate of 0.05%/yr (~3 fCi/liter) better reproduces the
data. In years like 1976, when there is little new atmospheric deposition, levels do not
decrease significantly because of continued inputs from the tributaries.

Although there may be many ways to account for the minor variations in mean
concentration of plutonium in the lakes each year, the overall behavior is adequately

0.4

o
E

O-

0.3 —

0.2 —

0.1 —

0.05 —

0.025

0.8

2 0.6

0.4

0.2 —

Atmospheric input

Watershed input
(a =* 0.05%/vr)

Plutonium in

Lake Michigan water

1973 1974 1975 1976

YEAR

1977

1978

Fig. 4 Kvaluation of the possible role of watershed erosion in maintaining the recent

levels of plutonium in Lake Michigan waters. , predicted for unstratified,

well-mi.xed lake, a = 0.05%. , predicted for unstratified well-mi.xed lake, a = 0. •,

measured values of mean concentration.

PLUTONIUM IN THE GREAT LAKES 66 7

described in terms of a single parameter, T'r. The preceding discussion suggests that
uncertainty in the estimated value for Lake Michigan is around 20% and somewhat higher
(~30%) for the other lakes. Thus the variation in the values of T'r from lake to lake by a
factor of 6 is real.

Althougli the variation of values of Tr is large, the apparent settling rate (L/Tr,
where Lis the mean lake depth) is essentially independent of the lake. These values are
given in Table 4. For all lakes except Lake Superior, the apparent settling rate is 35 ± 2
m/yr; for Lake Superior the value is 48 m/yr. However, in view of the uncertainties in
the average plutonium concentration for each lake and therefore in calculating Tr, tliis
value is not significantly different, and the mean apparent settling rate for the five lakes is
37 ± 3 m/yr^ . This result is consistent with the observed loss rate, corrected for

TABLE 4 Residence Time of Plutonium in the Great Lakes

Residence

time,

Apparent settling
rate {LlV^),

Depth

Mean

Outflow

Sedimentation

Lake

(L),in

(Tr)

(T'r)

m/yr

Superior

149

3.1

190

3.1

Michigan

2.4

100

2.4

Huron

2.1

2.3

Erie

0.44

0.52

Ontario

1.8

2.3

degradation, of DDT (Bierman and Swain, 1978) in Lake Michigan and Lake Superior. It
is also similar to the apparent settling rate for total phosphorus, 10 m/yr^ (S. C. Chapra,
Great Lakes Environmental Research Laboratory, personal communication), and for
detrital particles, 36 m/yr^ (D. M. DiToro, Manhattan College, personal communication).

This rate of 37 m/yr^ is comparable to mean particle settling rates inferred from trap
and sedimentation studies. The rate of accumulation of particles in traps placed in
southern Lake Michigan shows both a strong seasonal dependence and a marked increase
v^th increasing water depth (Wahlgren and Nelson, 1977b). However, for most of the
year, the net accumulation rate in near-surface waters is about 0.02 mg cm~'^ day~^.
Increases in flux with increasing depth must result either from resuspension or from
transient effects associated with earlier particle production in surface waters. The low
average downward particle flux from surface waters is comparable to average sedimenta-
tion in the southern basin of Lake Michigan (7 mg cm~^ yr"' = 0.02 mg cm~'^ day"')
estimated by Edgington and Robbins (1976) on the basis of ^ "^Pb and ' ^ ''Cs profiles in
a large number of sediment cores. Since the mean concentration of particles in the water
column is about 1 to 2 mg/liter (Wahlgren and Nelson, 1974b; additional unpublished
data for the period 1973 to 1977), the mean particle settling rate is about 40 m/yr (0.02
mg cm~'^ day"' X 2 mg/liter). That plutonium carrying particles has a net settling rate
that is comparable to the mean rate indicates that plutonium is not selectively scavenged
by an atypical suite of particles in the water column. This idea is also supported by
studies on the distribution of plutonium between particles and water discussed further in
the following text.

668 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

That the apparent sinking rate is essentially invariant from lake to lake is surprising
considering the diversity of limnological characteristics of these lakes. Evidently, the
processes determining the long-term removal of plutonium from the water column must
be similar for each of the Great Lakes, and the rate of removal from the water column to
underlying sediments is determined by the net rate at which the particles scavenging
plutonium sink. Furthermore, this result suggests that the availability of suitable particles
for scavenging plutonium and kinetics of exchange between dissolved and particulate
phases are not presently determijiing the long-term removal rate.

It must be emphasized that the apparent settling rate could be very different from the
actual net downward motion of particles. Vertical turbulence and resuspension of bottom
sediments can significantly alter the apparent particle settling rate. The role of
resuspension will be discussed further in the following text.

It is clear that, since the apparent settling rate is constant from lake to lake, the
processes controlling the removal of plutonium from the water column are similar for
each lake, and, therefore, considering the large differences in productivity, it is unlikely
that association with autochthonous organic matter is the rate determining step. However,
although this generalization is true for the water column as a whole, it will become
evident, from the experimental data discussed in the next section, that association with
biogenic material can explain, at least in part, the more rapid removal of plutonium from
the epilimnion (at least for Lake Michigan) during the period of stratification.

Seasonal Cycling of Plutonium in Lake Michigan

The seasonal cycling of plutonium from surface waters was first observed (Wahlgren and
Nelson, 1974b) at an offshore station 40 km west of Grand Haven, Mich. Since 1973
onward the cycle has been followed in detail in Lake Michigan (Wahlgren and Nelson,
1977a; Walilgren, Nelson, and Kucera, 1977, unpublished data). This seasonal cycling of
plutonium occurs at other stations in Lake Michigan and probably is common to all the
Great Lakes (Bowen, 1975; Alberts, Walilgren, and Nelson, 1977) except Lake Superior.
More than 75% of the total (dissolved and suspended) plutonium is lost from the
epilimnion at station ANL-5 in Lake Michigan during the sununer months and returned
during the fall and winter mixing period each year. The results, summarized in Fig. 5,
contrast the strong seasonal cycle in surface waters with mean levels expected t>om the
concentration— time model discussed in the preceding text. The data shown are primarily
from a single sampling station, ANL-5 (12 km southwest of Grand Haven, Mich.), but the
same trends have been observed at other stations farther offshore, including EPA-18,
which is in the middle of the southern basin. Farther offshore the onset of stratification
occurs later in the spring season, and the initial removal of plutonium from surface waters
is delayed correspondingly. A limited number of cross-lake transects during 1976 show
that the degree of removal uom surface water by September was comparable across the
whole of the southern half o'i the lake.

The change in concentration of plutonium in the well-mixed epilimnion of depth is
given by

dC 0 C

dt Li, T'k

(5)

provided that there is no upward transfer across the thermocline. (The etTect of outfiow
can be ignored.) The results of evaluating Eq. 5 with 0 = 0 as v/ell as with values of 0

PLUTONIUM IN THE GREAT LAKES 669

^'°|l I I I I I I I I I I I M I I I I M M M I M I I I I M I I M I I I I I I I I I I I I I I I I I I I M I I

0.8

^ 0.6

0.4

0.2

" I I M I I I I I I I I I I I I I I I I I I

I I M I

AMJJASONDJFMAMJJASON DJFMAMJJASO NDJFMAMJJASONDJFMAMJJASOND
1973 I 1974 I 1975 | 1976 | 1977

YEAR

JFM

Fig. 5 Comparison of concentrations of 239 ,240 p^ jj^ ^y\q water column predicted by
the coupled-lakes model with experimental measurements of surface-water concentra-
tions in Lake Michigan. The annual cycling of plutonium is clearly evident in this

comparison. , surface-water values predicted by the model. , surface-water

values measured.

determined from monthly measurements of fallout (Environmental Measurements
Laboratory, 1978) are summarized in Table 5. The results of these calculations indicate
that, for the years 1973 to 1977, there is a sUghtly greater variation in the calculated
residence time in the epilimnion when new inputs are ignored (i.e., 0 = 0). For example,
the effect of new fallout on the calculated values was most important in 1977; the
residence time for plutonium was reduced from 0.31 to 0.21 yr. When new inputs are
included, the residence time is almost constant from year to year. The mean residence
time T'rj- = 0.22 yr corresponds to an apparent particle settling rate (coe) of about 90
m/yr. The effect of an additional input of plutonium from the underlying waters would
be to further decrease the estimate of T^e and increase the calculated value of gJe-

The losses of plutonium from the epiMmnion are more rapid than expected on the
basis of the average residence time of this radionuchde in the lake. To the extent that the
epilimnion is isolated from underlying waters and the Tr scales with water depth, the
residence time of plutonium in the epilimnion would be T're = Tr X (Le/L), where Le
is the mean depth of the epilimnion (~20 m). Thus Tre ~ 2.4 X (20/84) ~ 0.6 yr. Thus
the observed value T'r£ is at least 2.5 times lower than that expected from the average
residence time of plutonium in the lake. It is therefore clear that the removal of
plutonium from the epilimnion is not solely due to its isolation from underlying waters.

This increased efficiency of the removal of plutonium from surface waters during the
period just before and during stratification of the lake is probably due to intensified

670 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 5 Parameters Describing the Removal of Plutonium
from the Epilimnion

c,*

6C/6t,t

rR^,*

'/'/Le,*

T're'§

t^E^

Year

fCi/Uter

fCi liter-' yr"'

m/yr

1973

0.45

2.0

0.27

0.19

0.24

1974

0.52

2.3

0.23

0.38

0.22

1975

0.39

1.3

0.28

0.16

0.25

1976

0.32

1.5

0.21

0.16

0.19

105

1977

0.31

0.90

0.31

0.64

0.21

Average

0.27

0.22

*C is the mean plutonium concentration in the epilimnion, May through September.
jsC/St is the mean rate of change in plutonium concentration, May through
September.

$Trp is uncorrected for atmospheric inputs (i.e., 0 = 0 in Eq. 5); tJg = Le/Tr£.

§T'RE=C/6C/6t.

scavenging by particles produced in the epiHmnion. Limnological factors affecting the
cycHng of plutonium in the Great Lakes are summarized in Fig. 6. The major inputs of
particles from external sources (allochthonous) occur during the very early spring and late
fall months and tend to be rapidly distributed throughout the length of the water
column. During the late spring and through the summer and fall, two principal types of
particles, diatoms and calcite, are produced in surface waters. The onset of the decrease in
plutonium levels occurs just before stratification in June (Fig. 5), which coincides with
the end of the major plankton bloom. The reduction is largely completed during the
period of in situ calcite formation during August and September (Fig. 5). If it is assumed
that the thermocline averages 15 m deep over the whole season, then the total clearance
of plutonium is about 1 fCi/cm^.

The initial decrease in plutonium parallels, in time and extent, that reported for
soluble silicon in offshore Lake Michigan waters (Holland and Beeton, 1972). One
possible removal mechanism would therefore appear to be the accumulation of plutonium
by phytoplankton (primarily diatoms) and the subsequent setthng of phytodetritus and
zooplankton fecal pellets from the epilimnion. From a knowledge of the concurrent
decrease in the sihca content of the epihmnion and the uptake of plutonium by net
plankton, it is possible to estimate the removal of plutonium from the epilimnion due to
the production and setthng of diatoms in May and June.

As shown in Fig. 7, after the spring diatom bloom there is a reduction in the con-
centration of particles in the 8- to SO-(j.m size range through June and July followed by
a dramatic increase in the concentration of particles due primarily to the appearance of a
large number in the 3- to 8-//m size range. Since the particulate matter collected in August
and September is predominantly calcium carbonate (up to 75%), it is taken that these
particles result from the in situ production of calcite particles, which agrees with
observations made elsewhere (Brunskill, 1969).

Thus the formation of calcite may also be an important mechanism for efficient
clearance of plutonium from the epilimnion. From July to September the concentration
of calcium in the epilimnion decreases by about 1 to 2 mg/hter. If it is taken that at this
time the epihmnion is 20 m deep, then the formation of calcite could clear from 5 to 10

PLUTONIUM IN THE GREAT LAKES 671

z; <

< y

Major
diatom-
bloom

Dominance
• green, blue-green — »-
algae

Maximum

• zooplankton-

grazing

Minor

•diatom

bloom

CO a

uj O

(- OJ

cc
< —

< _l

z <

Major sediment
input from
watershed

In situ formation
of calcium carbonate

Resuspension of

bottom sediments,

shore erosion

<
UJ
CO

Winter
stratification'

Spring convectlve
mixing period

Thermal bar-

Summer
stratification'

Fall convective
mixing period

Isothermal
mixing period

JAN. FEB. MAR. APR. MAY JUNE JULY AUG. SEPT. OCT. NOV. DEC.

MONTH

Fig. 6 Seasonal processes in Lake Michigan.

I-
<

O 0.1
o

H
to

UJ
CO

0.01

>80 Aim

JAN. FEB. MAR. APR. MAY JUNE JULY AUG. SEPT. OCT. NOV. DEC.

1973

Fig. 7 Seasonal dependence of the concentration and size distribution of particulate
matter in the surface waters of Lake Michigan (ANL-73-3,40 km west of Grand Haven,
MicH.).

612 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 6 Comparison of the AbUity of Authigenic Silica and Calcite
to Remove Plutonium from the EpiUmnion

Net plankton,
May and June

Suspended

particles

Analysis

May and J

une

August and September

Plutonium, fCi/g ash

165

90 to 100

Percent SiO^ ash

-95

-3

Percent CaCOj ash

-10

-75

Average plutonium, fCi/g

SiO^

750

Average plutonium, fCi/g

CaCOa

Clearance from epilimnion,*

pCi/cm^

In relation to loss of SiOj

(May and June)

0.06

0.2

In relation to loss of

CaCOj (August and

September)

0.6 to 1.2

*See text.

mg CaCOa/cm^. From the data in Table 6 and assuming that all the plutonium is carried
on calcite, the clearance from the epilimnion could range from 0.6 to 1.2 fCi/cm^. Thus
the combined scavenging ability of diatoms and calcite could account for the removal of
all the plutonium from the epilimnion. It must be made clear that in these calculations it
is assumed that plutonium is carried exclusively by biogenic silica or calcite. No direct
measurements have been made as yet.

The results of analyses for plutonium and silica in samples of phytoplankton from net
tows are compared with those for suspended particles filtered from the epilimnion and
are summarized in Table 6. Detailed measurements of the concentration of SiOo in the
water column have shown that during the phytoplankton bloom its concentration in the
epilimnion decreases by about 1 mg/liter. If it is taken that the early epilimnion is about
10 m deep, then the total clearance of SiOi is equivalent to about 1 mg/cm^ , and,
therefore, since the concentration of plutonium expressed in terms of Si02 content is 58
fCi/g Si02 , about 0.06 fCi/cm^ of plutonium will be carried with the sinking diatoms.
That is less than 10% of the total removed from the epihmnion during the whole summer.
The concentration of plutonium in the ashed material collected on filters is far higher
than that in ashed phytoplankton. It is not clear whether this additional plutonium is
associated in any way with planktonic detrital silica. However, if plutonium is, in fact,
removed in association with this silica, then about 0.2 fCi/cm^, or about 20%, would be
lost from the epihmnion.

The enhanced removal of plutonium from tlie epilimnion is probably accomplished
by scavenging particles with a short lifetime. An extremely small fraction (<5%) of the
diatoms produced annually is incorporated into permanent sediments (Parker and
Edgington, 1976), and most of the siUca tied up in their frustules is redissolved in the
water column (Parker, Conway, and Yaguchi, 1977), whereas the remainder is
resolubilized within a few weeks after reaching the benthic zone. The concentration of
plutonium in surface water starts to increase again with the breakdown of the

PLUTONIUM IN THE GREAT LAKES 673

thermocline but does not attain its maximum value until sometime after early December.
In contrast, reactive silica, which also undergoes an annual cycle in the water column,
returns to nearly its spring value as early as the end of September. Thus the release of
plutonium from dissolving frustules is delayed in relation to the sihcon (and calcite)
cycles. The production of calcite later in the year may account for this delay, but the
lifetime of calcite particles in the water column is presently unknown for the Great
Lakes. Alternatively, this delay could indicate that, if plutonium is released from
dissolving frustules or calcite particles, it is transferred to other particles in hypoUmnion
or in the vicinity of the sediment— water interface. Alberts, Wahlgren, and Nelson (1977)
have shown that almost all the plutonium in floe collected at the sediment-water
interface is associated with reducible hydrated oxides, such as ferric and manganese
oxides, and not with carbonates or silica.

Since epilimnetic plutonium losses are mainly due to scavenging by particles that
dissolve rapidly, this transient process may not result in appreciable net transfer to
sediments during most of a year. In 1975 the mean concentration of total (dissolved and
particulate) plutonium at station ANL-5 (from the surface to 60 m) shows a seasonal
variation that is consistent with its average residence time in the lake (Fig. 8; the dashed
line in the figure corresponds to T'r = 2.4 yr; a = 0.05%/yr). The net loss to the
sediments during each month of this year is predicted by the coupled-lakes model. Since

0.8

0.6 —

a.
o

CN
O)
CO

0.4

0.2

Measured
total

Particulate

M A M J J A
MONTH (1975)

0 N

Fig. 8 Mean concentrations of plutonium in whole water, in solution, and on particles
in the upper 60 m of the water column of Lake Michigan (ANL-5; 1975; water depth,
67 m). The concentrations expected for a mean residence time (Tr) of 2.4 yr and
watershed erosion (a) of 0.05%/yr are indicated by the dashed curve.

674 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

no enhanced transfer to the sediments is expected during the period of diatom and calcite
production, there appears to be a dramatic exchange of plutonium between dissolved and
particulate phases and temporary loss to the benthic layer (between 60 m and the
sediment— water interface). This complementary behavior can be understood in terms of
the constancy of the distribution coefficient Kp (Wahlgren and Nelson, 1977c). Since Kp
is essentially independent of particle type ('^2 x 10^), the proportion of plutonium tied
up with particles should reflect the amount of particulate matter in the water column.
From a comparison of the total concentration of particles in surface water each month
(Fig. 7) with the amount of plutonium in the water column above the benthic layer
(Fig. 8), it can be seen that the proportion of plutonium above the benthic layer reflects
the time— dependence of particle concentrations in surface waters.

The extent to which plutonium is removed from the dissolved phase before fall
overturn is illustrated in Fig. 9. The seasonal variation in the profiles of dissolved
plutonium from the years 1975 to 1977 shows that the removal from the soluble phase
extends progressively deeper into the water column as the season advances.

With the onset of fall overturn in October, appreciable losses of plutonium from
solution extend essentially to the bottom at this station (ANL-5). The effect was most
pronounced in 1976. On the basis of a series of samples collected from sediment traps
moored at all depths across the southern basin, it appears that this behavior is
characteristic of the deeper waters of the lake as well.* During November 1976 there was
a significant reduction in the total amount of plutonium in the water column at this
station. Provided that horizontal transport is ignored, the subsequent regeneration of
plutonium in the water column would require transfers of plutonium from sources below
the deepest sampling point, which is 7 m off the bottom.

The composition of particles reaching this depth illustrates the complementary
behavior of silica and calcite in regulating the cycling of plutonium. It can be seen
(Fig. 10) that the proportion of silica in the particulate material reflects diatom
productivity, with peak values in May and a secondary maximum in November. In
contrast, the percentage of calcium carbonate (calcite) remains low through the spring
and early summer but rises dramatically in September. There is approximately a 1 -month
delay between the onset of the plankton blooms or formation of calcite and the
appearance of increased concentrations of SiOj or CaCOa in particulate material 7 m
above the bottom. Since it is possible to disfinguish seasonal variations at this depth, it
strongly suggests that there is either Utile resuspension of bottom sediment at 7 m above
the bottom or that the only material that is resuspended is newly deposited detritus. The
sharp drop in proportion of Si02 and CaCOa in November is probably associated with
the input of terrigenic material washed into the lake as a result of shoreline erosion from
severe early winter storms.f Beyond November both the silica and calcite contents of the
seston decline dramatically. The total (dissolved and particulate) plutonium per unit
weight of particulate material exhibits a monthly variation that reflects the combined
effects of silica and calcite scavenging. The total plutonium concentration is high in the
spring, goes through a minimum in June, and then steadily rises to a maximum with the
addition of calcite to near-bottom seston. Considerable plutonium, Uke siHca and calcite,
is lost from the seston and regenerated in the water column during the period from

*Earth Resources Technology Satellite satellite photographs show that the "whiting" of the
surface waters of Lake Michigan is a lake-wide process.

fit has been shown that up to 50% of the total annual erosion can occur in November (E. Siebel,
University of Michigan, personal communication).

PLUTONIUM IN THE GREAT LAKES 675

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§1

'C 4)

cr>

1 ~

—3

.2 13
> ta

1 .2

Ul '

Hid3a

•0

ON 4)

. -o

00 ca

■£ -S

616 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

200

C31

180

()

IfiO

14U

— Total Plutonium (particulate and dissolved)

120

^30

Lu 20
o
cc

!^ 10

0
20

o
cc

(J
cc

H 60

-z.

o 40

°- 20
0

Loss on
ignition
(organic)

Calcium
carbonate

J I \ \ L

J \ \ \ \ L

Terrigenic
(by difference)

J \ \ L_L

M A M J J AS
MONTH (1975)

0 N

Fig. 10 Total (particulate and dissolved) piutonium and principal constituents of
particulate material from filtration (ANL-5, 7 m above bottom).

November to December. The excellent correlation between total plut-onium and the
combination of Si02 and CaCOa is shown in Fig. 11. To the extent that a sample of
particles and water constitute a closed system, it is the total piutonium that should
correlate with this combination of variables if piutonium is primarily derived from the
dissolution of either siUca or calcite particles.

Regeneration of piutonium in the water column would thus appear to result from the
dissolution of calcite and silica, especially during November and December. The role that
resuspension plays in this event is uncertain. It has been suggested (Wahlgren et al., 1976)
that resuspension of bottom sediments and redistribution of piutonium between particles
and water must occur to account for the reappearance of piutonium in surface waters and
for the apparent plateau leveling-off of average piutonium concentrations in the lake from

PLUTONIUM IN THE GREAT LAKES 677

u.o
0.7

—

1 1

0.6

—

en
E

y 0.5

• /•

D
Z

O 0.4

Q-

—

/ •

•

—

-J
<

H 0.3

<r-

—

•

—

0,2

—

0.1

1 1

—

0.1 0.2 0.3 0.4

SiOj + CaC03, mg/mg

0.5

Fig. 1 1 Relation between total (in solution and on particles) plutonium and the content
of SiOo and CaC03 in particles from filtration (ANL-5. 7 m above bottom).

1973 onward. The results of the concentration— time models presented above show that it
is unnecessary to assume that present levels reflect equilibrium between sediments and
water. The preceding arguments indicate that plutonium is released from dissolving
particles that may be localized in the benthic zone, depending on lake dynamics.
Redistribution of soluble plutonium in the water column can readily occur without
particle resuspension (oj^ = 0).

The study of Chase and Tisue ( 1977) indicates that the material most available to be
resuspended (benthic floe residing above consolidated sediments) is probably hydro-
dynamically unsuited to appreciable upward movement in the water column. The benthic
floe consists primarily of organic-mineral aggregates, typically a few tens of microns in
diameter having bulk densities of about 1.05 g/cm^. The aggregates consist of diatom
frustules, calcite and other minerals, and other unidentified detritus. Particles with such
properties (7~ 20 jLtm) remain aggregates on resuspension and have Stokes' settling
velocities co^ of roughly 10~^ cm/sec, or about 500 m/yr (Lerman, Lai, and Dacey,
1974). Under normal conditions (excluding fall overturn) the vertical eddy diffusivity Ky
in the hypolimnion is on the order of 1 cm'^/sec, or ~3 X 10^ m'^/yr (Kullenberg,
Murthy, and Westerberg, 1973). Thus the scale length for the resuspended material flux in
the water column under steady-state conditions is Ky/ws "^ 5 m. The flux of resuspended
material is given by J = Jo e ^^ '^, where h is the height above the bottom and Jq

67 S TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Q.
LU
Q

X Station 5 —
A Station 19
• Station 18 _
■ Station 16

■ — x-

-TTTTTTTJ

kzzd \ \ \ L

0.2 0.4 0.6 0.8 1.0

PARTICLE FLUX (J), mg cm"2 day"'"

1.2

Fig. 12 Particle-settling-rate profiles from sediment-trap studies at several stations in the
southern basin of Lake Michigan (1976).

(milligrams per square centimeter per year) is the downward particle flux at the
sediment— water interface.

This scale length is confirmed by sediment-trap studies conducted at a series of
locations in Lake Michigan. Each of the particle-settling-rate profiles (J in milligrams per
square centimeter per day) shown in Fig. 12 has a "foot" that extends from the bottom
upward to about 10 to 15 m, The rate of decrease in J over the interval is very nearly
exponential and has a scale of about 5 m, which is essentially the same at each location
except at station 5. Therefore it is likely that the effects of resuspension are confined to
the bottom 10 to 15 m in deep waters (over 80 m). It is possible that Umited numbers of
smaller particles of freshly formed aggregates of lower densities in the benthic floe may
be resuspended to greater average heights above the bottom.

However, the monthly series of settling-rate profiles from station ANL-5 (Fig. 13)
suggest that there is a nonconstant settling rate above 20 m that is probably due to
seasonal variations in particle production in the epilimnion. In addition, these profiles
suggest that resuspension may account for most of the settUng material at relatively
shallow stations similar to this.* During the period in August when the nonash

*This presents a problem in that, although the water depth of ANL-5 is representative of the mean
depth of the lake (~84 m), it is close enough to shore (~12 km) to make it highly susceptible to
contributions from shoreline erosion and tributary inputs.

PLUTONIUM IN THE GREAT LAKES 679

I-

till III

1 1

1 1 1 1

' _

20
40

60
0

20
40

- \

—

— Apr. 3 - May 18

1 1 1 1 III

_ July 22 -

Aug. 16

: \

—

— May 18 - June 4 \

— Aug. 16

- Sept. 1 1

1 1 1 1 II F^

1 1

1 1 1 1

N^^

—

- N.

—

_ June 4 - July 22 >w

1 r^

_ Sept. 11

- Oct. 3

0.01

0.1

1.0

0.01

0.1

1.0

PARTICLE FLUX (J), mg cnn-^ day"''

Fig. 13 Seasonal variation in particle-settling-rate profflesfrom sediment-trap studies in
the southern basin of Lake Michigan (ANL-5, 1976).

component (plankton remains) is low, the particle flux profile is very nearly exponential
from depths of 65 up to 10 m. This may indicate vertical mixing by eddy diffusion, at
least below the thermocline. The apparently high sedimentation rate (~365 mg cm~^
yr~^) in traps near the bottom must reflect both the effect of resuspension and
deposition of transient particles, such as diatom frustules and calcite. The mass
sedimentation rate for consoHdated sediments in this region is only 30 mg cm~^ yr~^
(Edgington and Robbins, 1976). In view of these results, resuspension during isothermal
mixing could be of some importance for the return of plutonium to the whole water
column at this inshore station. The effect of resuspension in shallower waters could also
influence plutonium levels in offshore areas, depending on the efficiency of horizontal

680 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 7 Seston Concentrations, Particle Fluxes,
and Apparent Settling Rates in the Upper 10 m of Water

(ANL-5, 1976)

Seston

Particle

Apparent settling

concentration (Q,*

flux (J),

rate (V = J/Q,

Month

mg/Uter

mg cm~^ day"'

' m/yr

April

2.1

0.35

610

May

2.1

0.12

220

June

1.5

0.005

July

2.0

0.006

August

2.7

0.03

September

1.7

0.01

*Average of samples taken at beginning and end of sediment-trap
collection interval.

transport of resuspended particles. The fact that regeneration of plutonium proceeds
concurrently at offshore stations during the winter indicates that either inshore
resuspension is intense and horizontal transport rapid or that dissolution rather than
resuspension is more important in recychng plutonium lost to particle phases during the
earlier sequence of diatom and calcite production.

The particle flux profiles (Fig. 13) provide a crude measure of the seasonal variation
in apparent particle settling rates. Althougli the lower portions of the profile must result,
at least partially, from resuspension, the particle flux measured above the thermocline is
hkely to be unaffected by resuspension. Under conditions of a stable thermocline (July
and August), the upward particle flux into the epilimnion will be negligible. The ratio of
this flux to the estimated concentration of particles in surface waters (Fig. 7) is the
apparent particle settling rate" given in Table 7. Apart from extremely high rates during
the months of the spring convection mixing period and diatom bloom, the rates during
the succeeding months are comparable (22 m/yr; June to September average) to the rate
inferred from the concentration— time model. Thus plutonium is apparently scavenged by
particles that, on the average, are settling at the same apparent rate as most particles in
the water column. Further, it seems likely that resuspension does not have an appreciable
effect on this apparent rate of settling.

Conclusions

The results of extensive measurements of plutonium in the Great Lakes have shown that
for each lake the concentration of plutonium in the water column at the end of the spring
convective mixing period can be described by a simple time— concentration model with
only one variable for each lake, viz., the residence time of plutonium. This retention time
for the loss of plutonium in each lake is controlled by two processes, the outflow of
water to the next lower lake (Tr) and transfer to the sediment on settling particles (T'r).
With the use of a plutonium source term based on the fallout ^°Sr monitoring values
from the Environmental Measurements Laboratory, values of Tr that gave best fit to the
experimental data were obtained from the model. The values of Tr were found to be
proportional to the mean depth of the lakes, which implies that the appaient settling rate

PLUTONIUM IN THE GREAT LAKES 681

of the particles carrying plutonium is the same in each lake. Considering the large
differences in their limnological properties, such as primary productivity, this result is
rather surprising in that it might have been expected that a large proportion of plutonium
would be carried with organic detritus.

The rate of clearance of plutonium from the epilimnion of Lake Michigan between
June and September is far faster than that from the whole water column of the lake and
is constant from year to year. Enhanced removal of plutonium from the epilimnion
results from intensified particle production during the spring and summer months. Most
Ukely, plutonium is scavenged by diatoms and calcite particles, which subsequently
redissolve.

Although it is possible to develop an understanding of what appears to be an
extremely complex system, such as the behavior of plutonium in the Great Lakes, in
terms of a simple model, there are still many unanswered questions. For example, the
uptake of plutonium on biogenic siBca or autochthonous calcium carbonate in the
epilimnion must be a transient process because it is clear that almost all this material
redissolved in the hypoUmnion. Some redissolved plutonium may be taken up by other
particulate material. Presently it is impossible to balance the downward flux of plutonium
on Si02 and CaCOa, measured near the bottom in sediment traps, with that apparently
deposited in the surface sediments each year because of the subsequent horizontal
redistribution of older sediment containing plutonium in the lake.

Since the lakes have very similar general chemical properties, it is possible that the
exchange of plutonium between the water column and sediments is controlled by
chemical reactions. In fact, it has been suggested that the concentration of plutonium in
the water column is largely controlled by chemical equiUbrium between specific species in
the water column and the sediments. If this equilibrium is a major factor in controlling
the concentration of plutonium in the lake, the value of Tr should have increased
significantly during the period of major deposition in the sediments (1963—1970). The
data at present demonstrate that little change has occurred in the past 7 yr since the
present concentration in the water column can be described by the value of Tr calculated
for 1971.

The situation is complicated by the very recent observations that (1) plutonium
in Lake Michigan . (and the Irish Sea) exists primarily in the water column in the VI
oxidation state and on the sediments as the IV oxidation state and (2) Lake Michigan
water can readily extract sorbed plutonium from high-activity pond sediments. Since the
distribution coefficients of Pu(VI) and Pu(IV) v^th sediment are very different, a critical
step in the clearance of plutonium from the water column may be the reduction of
Pu(VI) to Pu(IV) either in the water or at the sediment surfaces. It is clear, however, that,
if there is to be a complete understanding of the long-term behavior of plutonium,
especially from other source terms in aquatic environments, more attention must be paid
to determining its chemical forms.

Acknowledgments

This work was performed under the auspices of the U. S. Department of Energy.

We gratefully acknowledge the contributions of D. M. Nelson, K. A. Orlandini, and
other members of the Ecological Sciences Section to various aspects of the field sampling
program that made this chapter possible.

682 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

References

Alberts, J. J., M. A. Wahlgren, and D. M. Nelson, 1977, Distribution of Fallout Plutonium in the

Waters of the Lower Great Lakes, in Radiological and Environmental Research Division Annual

Report, January -December 1976, ERDA Report ANL-76-88(Pt.3), pp. 50-52, Argonne National

Laboratory, NTIS.
Bierman, V. T., Jr., and W. R. Swain, 1978, DDT Losses in Low-Lipid Fishes of the Great Lakes, in

Abstracts of the 21st Conference on Great Lakes Research, Windsor, Ontario, May 9-11, 1978,

International Association for Great Lakes Research.
Bowen, V. T., 1974, Plutonium and Americium Concentrations Along Fresh Water Food Chains of the

Great Lakes, U.S.A., General Summary of Progress, USAEC Report COO-35684, Woods Hole

Oceanographic Institution, NTIS.
, 1975, Plutonium and Americium Concentration Along Fresh Water Food Chains of the Great

Lakes, U.S.A., General Summary of Progress, USAEC Report COO-3568-8, Woods Hole

Oceanographic Institution, NTIS.
, 1976, Plutonium and Americium Concentration Along Fresh Water Food Chains of the Great

Lakes, U.S.A., General Summary of Progress, USAEC Report COO-3568-13, Woods Hole

Oceanographic Institution, NTIS.
, 1977, Plutonium and Americium Concentration Along Fresh Water Food Chains of the Great

Lakes, U.S.A., General Summary of Progress, ERDA Report COO-3568-18, Woods Hole

Oceanographic Institution, NTIS.
, and V. E. Noshkin, 1972, Plutonium Concentrations Along Fresh Water Food Chains of the

Great Lakes, U S. A., General Summary of Progress, USAEC Report COO-3568-1, Woods Hole

Oceanographic Institution, NTIS.
, and V. E. Noshkin, 1973, Plutonium Concentrations Along Fresh Water Food Chains of the

Great Lakes, U.S.A., General Summary of Progress, USAEC Report COO-3568-3, Woods Hole

Oceanographic Institution, NTIS.
BrunskiU, G. J., 1969, FayetteviUe Green Lake, New York. II. Precipitation and Sedimentation of

Calcite in a Meromictic Lake with Laminated Sediments, Limnol. Oceanogr., 14: 830-847.
Chase, R. R. P., and G. T. Tisue, 1977, Settling Behavior of Lacustrine Organic Mineral Aggregates, in

Radiological and Environmental Research Division Annual Report, January -December 1976,

ERDA Report ANL-76-88(Pt.3), pp. 99-102, Argonne National Laboratory, NTIS.
Conway, H. L., J. I. Parker, E. M. Yaguchi, and D. L. MeUinger, 1977, Biological Utilization and

Regeneration of Silicon in Lake Michigan,/. Fish. Res. Board Can., 34: 537-544.
Edgington, D. N,, and J. A. Robbins, 1976, Records of Lead Deposition in Lake Michigan Sediments

Since \%m, Environ. Sci. Technol, 10: 266-274.
Environmental Measurements Laboratory, \91 %, Environmental Quarterly , Appendix, USDOE Report

EML-342 Appendix, NTIS.
Golchert, N. W., and J. Sedlet, 1972, Radiochemical Determination of Plutonium in Environmental

Water Samples, /?ac^/oc/zeA?z. Radioanal. Lett., 12: 215-221.
, T. L. Duffy, and J. Sedlet, 1978, Environmental Monitoring at Argonne National Laboratory,

Annual Report for 1977, USDOE Report ANL-78-26, Argonne National Laboratory, NTIS.
Hardy, E. P., P. W. Krey, and H. L. Volchok, 1973, Global Inventory and Distribution of Fallout

V\\xXom\xm, Nature, 241: 444-445.
Holland, R. E., and A. M. Beeton, 1972, Significance to Eutrophication of Spatial Differences in

Nutrients and Diatoms in Lake Michigan, L//?i«o/. Oceanogr., 17: 88-96.
Krey, P. W., M. Schonberg, and L. Toonkel, 1974, Updating Stratospheric Inventories to January

1973, Fallout Program Quarterly Summary Report, USAEC Report HASL-281, Health and Safety

Laboratory, NTIS.
KuUenberg, G., C. R. Murthy, and H. Westerberg, 1973, An Experimental Study of Diffusion

Characteristics in the Thermocline and Hypolimnion Regions of Lake Ontario, in Proceedings of

the Sixteenth Conference on Great Lakes Research, Huron, Ohio, Apr. 16-18, 1973, pp. 774-790,

International Association for Great Lakes Research.
Lerman, A., 1972, Strontium-90 in the Great Lakes: Concentration— Time Model, J. Geophys. Res.,

11: 3256-3264.
, D. Lai, and M, F. Dacey, 1974, Stokes' Settling and Chemical Reactivity of Suspended Particles

in Natural Waters, in Suspended Solids in Water, R. J. Gibbs (Ed.), Plenum Press, New York.

PLUTONIUM IN THE GREAT LAKES 683

Maletskos, C. J., 1972, Intercalibration and Intercomparison Study of Radionuclide Concentrations in
Lake Michigan Samples, Progress Report, Oct. 1, 1971-Mar. 31, 1972, USAEC Report COO-3371-1,
New England Deaconess Hospital, NTIS.

Nelson, D. M., E. M. Yaguchi, B. J. Waller, and M. A. Wahlgren, 1974, Radiochemical Methods, in
Radiological and Environmental Research Division Annual Report, January -December 1973,
USAEC Report ANL-8060(Pt.3), pp. 6-17, Argonne National Laboratory, NTIS.

Parker, J. I., H. L. Conway, and E. M. Yaguchi, 1977, Seasonal Periodicity of Diatoms, and Silicon
Limitation in Offshore Lake Michigan,/. Fish. Res. Board Can., 34: 552-558.

, and D.N. Edgington, 1976, Concentration of Diatom Frustules in Lake Michigan Sediment

Cores, Limnol. Oceanogr., 21: 887-893.

Sprugel, D. G., and G. E. Bartelt, 1976, Preliminary Mass Balance of Plutonium in a Watershed Near
Sidney, Ohio, in Radiological and Environmental Research Division Annual Report, January-
December 1975, USAEC Report ANL-75-60(Pt.3), pp. 20-22, Argonne National Laboratory,
NTIS.

U. S. Geological Survey, 1913-1916, Surface Water Records for Michigan.

Wahlgren, M. A., J. J. Alberts, D. M. Nelson, and K. A. Orlandini, 1976, Study of the
Behavior of Transuranics and Possible Chemical Homologues in Lake Michigan Water and Biota, in
Transuranium Nuclides in the Environment, Symposium Proceedings, San Francisco, 1975,
pp. 9-24, STI/PUB/410, International Atomic Energy Agency, Vienna.

, and J. S. Marshall, 1976, The Behavior of Plutonium and Other Long-Lived Radionuclides in

Lake Michigan. I. Biological Transport, Seasonal Cycling, and Residence Times in the Water
Column, in Impacts of Nuclear Releases into the Aquatic Environment, Symposium Proceedings,
Otaniemi, Finland, 1975, pp. 227-243, STI/PUB/406, International Atomic Energy Agency,
Vienna.

, and D. M. Nelson, 1973, Plutonium in Lake Michigan Water, m Radiological and Environmental

Research Division Annual Report, January -December 1972, USAEC Report ANL-7960(Pt.3),
pp. 7-14, Argonne National Laboratory, NTIS.

, and D. M. Nelson, 1974a, Residence Times for ^^'Pu and '^'Cs in Lake Michigan Water,

in Radiological and Environmental Research Division Annual Report, January -December 1973,
USAEC Report ANL-8060(Pt.3), pp. 85-92, Argonne National Laboratory, NTIS.

, and D. M. Nelson, 1974b, Studies of Plutonium Cycling and Sedimentation in Lake Michigan, in

Proceedings of the Seventeenth Conference on Great Lakes Research, Ontario, Canada,
Aug. 12-14, 1974, pp. 212-218, International Association for Great Lakes Research.

, and D. M. Nelson, 1975, Plutonium in the Laurentian Great Lakes: Comparison of Surface

Waters, Verh. Int. Ver. Theor. Angew. Limnol., 19: 317-322.

, and D. M. Nelson, 1977a, Seasonal Cycling of Plutonium in Lake Michigan, in Radiological and

Environmental Research Division Annual Report, January-December 197S, ERDA Report
ANL-76-88(Pt.3), pp. 53-55, Argonne National Laboratory, NTIS.

, and D. M. Nelson, 1977b, Lake Michigan Sediment Trap Study: Preliminary Assessment of

Results, in Radiological and Environmental Research Division Annual Report, January -December
1976, ERDA Report ANL-76-88(Pt.3), pp. 107-110, Argonne National Laboratory, NTIS.

, and D. M. Nelson, 1977c, A Comparison of the Distribution Coefficients of Plutonium and Other

Radionuclides to Those in Other Systems, in Radiological and Environmental Research Division
Annual Report, January-December 1976, ERDA Report ANL-76-88(Pt.3), pp. 56-60, Argonne
National Laboratory, NTIS.

Transport of Plutonium by Rivers

H. J. SIMPSON, R. M. TRIER, and C. R. OLSEN

A number of nuclear facilities are-iocated on rivers and estuaries, and thus it is important
to understand the primary transport pathways of transuranic elements in such systems.
Relatively few field studies of point-source releases of plutonium to river systems have
been made up to now. Information from research on the behavior of fallout plutonium in
rivers can, however, provide some useful insights. The range of variation of soluble-phase
fallout, ^^^'^^^Pu, in freshwaters and estuaries is relatively small (0.3 ± 0.2 fCi/ liter) and
appears to be "buffered" to some extent by the large reservoir of fallout '^^^''^^^Pu in
soils and the relative uniformity of the specific activity on soil particles (~20 pCi/kg).
The Hudson River, Hudson estuary. New York City tap water. New York bight, and
Great Lakes all have reasonably similar concentrations of soluble-phase 239,240^^^
despite the large range of chemical and other characteristics. The distribution of fallout
2-3 9,2 4 Op^ l)QP^QQfi soluble phases and particles in rivers can be approximated by a
partition coefficient of about 10~^. For suspended particle loads of about 10 mg/liter,
which are reasonably typical of low-flow summer conditions for rivers in the northeastern
United States, '^^^-'^^^Pu is transported by both soluble phases and particles in
approximately equal amounts. For higher suspended loads, typical of northeastern rivers
during greater freshwater discharge and of most other large, nontropical rivers, the
transport of fallout ^^^'^^°Pw is clearly dominated by particles (by about an order of
magnitude ). For point-source addition of plutonium to a river, the most important
transport pathway appears to be binding to the suspended load and the mobile portions
of the fine-grain sediments and subsequent downstream movement with the fine particles.
Since the kinetics and downstream transport pathways of fine particles of a particular
river depend on a number of factors peculiar to each system, the most direct approach
would be to exploit the presence of "tracers" already present to define the parameters of
most relevance to transuranic-element transport over various time scales. Nuclear facilities
often release sufficient quantities of fission and activation products during normal
operations which can be used .as indicators of fine-particle transport pathways. The
behavior of these radionuclides cannot be expected to be identical to transuranic
elements in river systems, but those elements with strong particle-phase associations can
provide very useful information for sites of primary interest for transuranic-element
transport assessments.

A number of nuclear power plants are now located on rivers and estuaries, and many
more probably will be in the future. The only major reprocessing faciUty currently
operating in the United States is located on a small tributary of the Savannah River. Thus
knowledge of the transport pathways of transuranic elements in rivers is essential for

684

TRANSPORT OF PLUTONIUM BY RIVERS 685

proper monitoring of the routine operations of these facilities and for developing plans
for dealing with any abnormally large releases of transuranic elements that might occur.

In principle, rivers can carry plutonium and other transuranic elements either in
solution or as part of the suspended load. These two transport pathways are probably
strongly coupled by some type of quasi-equilibrium partitioning between the two phases
and thus cannot really be considered separately. As with many elements that are reactive
in natural waters, the classifications of "dissolved" and "particulate" plutonium are based
largely on operational procedures, such as whether or not material will pass through a
filter of a certain nominal pore size. The actual species distribution of plutonium in
natural waters is probably some kind of continuum from small-molecular-weight
complexes through silt- or sand-size particles. To further compHcate matters, particles can
be transported in suspension or as bed load in a stream or accumulated in depositional
environments and either buried or resuspended at a later time.

There have been relatively few field studies of point-source releases of plutonium to
river systems. Three areas in the eastern United States that have received such attention
are the Savannah River and its tributary downstream of the reprocessing facility in South
Carolina (Hayes and Horton, this volume), the Miami River (a tributary of the Ohio
River) downstream of Mound Laboratory in Ohio (Sprugel and Bartelt, 1978) and streams
near Oak Ridge, Tenn. These river systems are the focus of ongoing research programs
which should provide considerable information about the transport by rivers of
plutonium derived from point sources. This chapter discusses the distribution of fallout
plutonium in a few natural systems, including the Hudson River and estuary, and
attempts to derive some first-order principles by which the transport pathways of
plutonium in other river systems can be predicted. The Hudson estuary is now the site of
three nuclear reactors, and at least half a dozen other units which are planned for this
estuary in the next two decades.

Plutonium in the Hudson River Estuary

The Hudson River discharges into one of the large estuarine systems that dominate much
of the coastal environment of the northeastern United States (Fig. 1). The Hudson has an
unusually long, narrow reach of tidal water (>250 km), most of which is usually fresh.
Saline water intrudes only about 40 km from the coastline during seasonal high
freshwater discharge and reaches as far inland as 120 km during summer and early fall
months of drought years. The near-surface suspended load of the Hudson is relatively low
(10 to 20 mg/Uter), as it is for nearly all the larger rivers in the northeastern United
States, except during maximum spring runoff and following major storms.

From studies of the distribution of fallout nuclides and gamma-emitting nuclides
released from Indian Point, the patterns of suspended particle transport and recent
sediment accumulation in the Hudson estuary have been described (Simpson et al.,
1976; 1978; Olsen et al., 1978). Much of the estuary has relatively little net
accumulation of fine particles, whereas a few areas, such as marginal coves and especially
New York harbor, account for a major fraction of the total deposition of fine particles
containing fallout and reactor nucHdes. The zone of major sediment accumulation is more
than 60 km downstream from the reactor site, and the time scale of transport of fine
particles labeled with reactor nucUdes from the release area to burial in the harbor
sediments varies from probably less than a month to years. At present there is no
evidence in the Hudson sediments, including New York harbor, of releases of reactor

686 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

NARRAGANSETT
BAY

SUSQUEHANNA
Q = 950 ^

POTOMAC,
Q = 340

30 0 30 60

I I I 1

Scale, km

-CHESAPEAKE BAY

Fig. 1 Location map of the major estuarine systems of the northeastern United States.
Q is mean annual freshwater flow.

2 3 9,240py which are resolvable in the presence of the burden of fallout ^^^'^"^^Pu. Thus
the current distribution of ^^^'^"^^Pu in the Hudson appears to be governed primarily by
the delivery of global fallout to the drainage basin mostly more than a decade ago and the
transport processes that have occurred since delivery.

Table 1 shows concentrations of 2 3 9,240p^ ^ large-volume Hudson water samples
that have had the suspended load removed by settling for 24 to 48 hr or by passing
through a continuous-flow centrifuge followed by a 0.45-)um filter. The range of observed
values for samples collected in 1975 and 1976 was 0.12 to 0.88 fCi/Uter; the median
value was about 0.3 fCi/liter. The current annual transport of 2 3 9,24 0p^ ^ ^^^
"dissolved" phase in the Hudson can be estimated to be about 5 x 10~^ Ci if we assume
a concentration of 0.3 fCi/liter and a mean annual river discharge of 550 m^ /sec. This
represents somewhat less than 0.01% of the fallout burden of ^^^'^'*°Pu (~80 Ci) in the

TRANSPORT OF PLUTONIUM BY RIVERS 687

TABLE 1 Dissolved 2 3 9 ,2 4 Op^ j^^ Continental Waters

Sample

Location*

fCi/Uterf

volume, liters

Hudson River (mp 61) (S)

0.32 ± 0.01

660

Hudson estuary (mp 19) (S)

0.88 ± 0.07

625

Hudson estuary (mp 18) (F)

0.47 ± 0.03

490

Hudson River (mp 47) (S)

0.27 ± 0.02

570

Hudson estuary (mp 19) (S)

0.12 ± 0.02

570

Hudson estuary (mp 8) (S)

0.15 ±0.02

570

Hudson estuary (mp 24) (F)

0.30 ± 0.03

1500

New York bight (S)

0.25 ± 0.03

380

New York bight (U)

0.59 ± 0.09

660

New York bight (U)

0.68 ± 0.05

660

New York bight (U)

0.68 ± 0.09

660

New York bight (U)

0.68 ± 0.09

660

New York bight (U)

0.91 ± 0.14

660

New York bight (U)

0.95 ± 0.14

660

New York bight (U)

1.18 ±0.14

660

New York City tap water (1973

to 1975)t

-0.3

Lake Ontario ( 197 3) §

-0.3

Great Lakes (1972 to 1973)11

-0.5

*mp indicates "mile point" upstream of the mouth of the Hudson,
defined as the southern tip of Manhattan Island. The pre treatment
procedure of the large-volume samples is indicated by one of the following
three letters: U, unfiltered; F, filtered after passing through a
continuous-flow centrifuge; S, suspended particles allowed to settle,
usually for 24 to 48 hr, before the clarified water was transferred to
another tank for processing.

fMean ± standard error.

tData from Bennett (1976).

§Data from Farmer et al. (1973).

H Data from Wahlgren and Marshall (1975).

soils of the Hudson drainage basin (~3.5 x 10'* km^). Soluble-phase release of fallout
2 3 9,240p^ from Hudson soils thus has a half-time of the order of 10"^ yr and supplies an
insignificant amount of dissolved 2 3 9,240p^ ^^ ^j^^ coastal ocean compared with that
transported onto the shelf from surface waters of the deep ocean.

The suspended-load activity of 2 3 9,2 4 0pjj ^^^ which dissolved plutonium concentra-
tions are Usted in Table 1 averaged about 20 pCi/kg (18.9 ± 0.9 and 23.4 ± 1.0 pCi/kg).
The distribution coefficient (Kj) of 2 3 9,240p^ between the dissolved phase and
suspended particles for those two samples was about 1.5 x 10"^. Thus the transport of
2 3 9,2 4 0pjj i^y suspended particles equals that in the dissolved phase when the
concentration of suspended particles is about 15 mg/liter, a value that is reasonably
typical of moderate and low freshwater flow periods in the Hudson. During periods of
higher suspended load, the transport of 23 9,24 Op^ ^ ^^^ Hudson is predominantly on
particles.

From the quantities of material dredged annually from New York harbor (~2 x 10^
tons), the downstream transport of particles by the Hudson must be about a factor of 4

688 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

higher than indicated by multiplying typical near-surface suspended-load concentrations
times mean annual flow. The "extra" transport of particles is probably accomplished by
some combination of very high suspended loads coinciding with the highest freshwater
flow rates and bed-load transport, which in the Hudson appears to consist largely of
resuspension and deposition of fine particles in the lowest meter of the water column on
the time scale of a tidal cycle. Thus in the Hudson the total downstream transport of
2 3 9,240pu -g approximately a factor of 4 greater than that in the dissolved phase, which
indicates a half-time for removal of fallout 2 3 9,240p^ from the drainage basin, largely on
particles, of about 10^ years. Similar calculations for the Savannah River (Hayes and
Horton, this volume) and the Greater Miami River (Sprugel and Bartelt, 1978) suggest
drainage-basin removal times of about 2x10^ and 2 x 10"' yr, respectively. Again this
supply term to the coastal ocean is not significant relative to advection of deep-ocean
fallout 2 3 9 ,2 4 op^ Qj^^Q ^j^g ^j^gjj^ Pq^ ^j^g Hudson most of the delivery of " ^ "^ ^ °Pu on

particles to the coastal ocean is accomplished by the dumping of dredge spoils rather than
by estuarine discharge of suspended particles.

Plutonium in the New York Bight

The concentrations of dissolved 2 3 9,240p^ ^^ ^^^ coastal waters off the New York City
area are two to three times those in the Hudson (Table 1). The suspended loads in the
New York bight are almost two orders of magnitude lower than those in the Hudson, and
fine-grain sediments in the bight have activities of 2 3 9.2 4 0pjj comparable to those in the
Hudson. Thus the transport of -3 9,2 4 0p|j ^^ ^^^ shelf environment appears to be largely
in the dissolved phase, in contrast to the situation in the Hudson River and estuary.

Plutonium in Other Freshwaters

Data for the concentration of 2 3 9.2 4 0p|j ^^ ^^^ York City tap water (Bennett, 1976) are
available for the period 1973 to 1975 (Table 1 ). The water supply for New York City is
derived from tributaries of the Hudson and Delaware rivers. The activities ranged from
0.08 to 0.60 fCi/liter, with a mean value of about 0.3 fCi/liter (about 2% of the average
rain activities during the same period). The range and mean value of the tap water
2 3 9,2 4 0p|j concentrations are almost identical with the values observed for the Hudson
River and estuary.

Farmer et al. (1973) have reported '39,240p|^^ activities in Lake Ontario (Table 1)
that are in the same range as the data discussed here for the Hudson River and New York
bight. During the period 1971 to 1973, the average -39,240p^ activity for the entire lake
declined from about 0.8 fCi/liter to about 0.3 fCi/liter. The average 2 3 9,2 4 0p|j ^^tivity in
all five Great Lakes (Wahlgren and Marshall, 1975; WahJgren et al.. 1976) during 1972
and 1973 was about 0.5 fCi/liter (Table 1 ).

Transport of Fallout Plutonium to the Oceans

The data available indicate that the range of variation of soluble-phase 2 3 9,2 4 0p|j j^^
freshwaters is relatively small. The transport by rivers of fallout 239,240p|^j j^^ "solution"
can thus be estimated relatively easily solely on the basis of the rate of freshwater
discharge. The concentrations in freshwaters appear to be "buffered" to some extent by
the large reservoir of fallout -^^■-''Opu jn soils and the relative uniformity of the specific
activity of 2 3 9,2 4 0p|j ^^^ ^^n p^^ticles and river suspended particles (^20pCi/kg). The

TRANSPORT OF PLUTONIUM BY RIVERS 689

distribution of fallout 2 3 9,2 4 0p^j between soluble phases and particles in rivers can
probably be approximated by a partition coefficient of about 10~^. The total delivery of
dissolved fallout 2 3 9,240py ^^ ^^iq oceans by rivers is probably about lOCi/yr, if we
assume a discharge rate for all rivers of about 10^ m^/sec and a concentration of about
0.3 fCi/Uter. Since the global average of suspended load in rivers is about 600mg/hter,
the transport of fallout 2 3 9,2 4 0p^j ^^ rivers will clearly be dominated by particles. If we
assume that the specific activity of all river suspended matter is similar to that of surface
soils, the total delivery of fallout 2 3 9,240p^ ^^ ^^le ocean by rivers is about 5 x 10^
Ci/yr, about 50 times the soluble-phase delivery. The specific activity of particles in rivers
with very high suspended loads is probably somewhat lower owing to the presence of
more large silt- and sand-size particles; so a more reasonable estimate for the total annual
delivery of fallout 2 39,240py ^^ ^j^^ ocean by rivers is probably 1 to 5 x 10^ Ci.

Transport of Plutonium by Rivers Added at Point Sources

The distribution of fallout 2 3 9,24 Op^ provides information about the partitioning of
plutonium between soluble- and suspended-particle phases in rivers and about the
processes by v/hich transuranic-element transport occurs in rivers. For point-source
addition of plutonium to a river, the most important transport pathway appears to be
binding to the suspended load and the mobile portions of the fine-grain sediments and
downstream movement with the fine particles. Since the effective concentrations of
suspended particles, including the upper few centimeters of fine-grain sediment, in a river
will be far greater than 10 to 15 mg/liter, the dominant transport of plutonium would be
in association with particles. The kinetics and downstream transport pathways of a
particular river system will depend on many factors, such as the frequency and duration
of deposition and resuspension episodes for the suspended particles. In the tidal reach of
the Hudson, the downstream movement of fine particles tagged with reactor nucHdes is
distributed such that some particles require several years to move 50 km whereas others
probably require considerably less than a few months. In other rivers, such as the
Columbia, which is above tidal influence, the downstream transport of some portions of
the suspended load is probably similar to the rate of water transport, whereas other
portions of the suspended particles are trapped for long periods, perhaps indefmitely,
behind dams.

The distribution of fallout nucHdes can provide valuable information about which
areas of the bottom in a river system are actively scoured and which portions accumulate
fme-grain sediments rapidly but probably cannot provide a very detailed picture of the
kinetics of downstream transport of fine particles. A tracer added relatively uniformly to
the earth's surface, as vras weapons-testing fallout, is not very powerful for providing such
information. Fortunately the river systems for which the kinetics of fine-particle
movement are most important to understand for predicting transport of transuranic
elements are also the ones for which point-source tracers are available. Many nuclear
power plants and reprocessing facihties release sufficient quantities of fission or activation
products during normal operations which can be used as indicators of fine-particle
transport pathways. The behavior of these radioactive tracers cannot be expected to be
identical to that of transuranic elements in river systems, but some of these tracers are
associated with particles sufficiently to provide very valuable information about the
patterns and kinetics of movement and accumulation of fine particles of most importance
for evaluating the transport pathways of point-source releases of transuranic elements.

690 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

References

Bennett, B. G., 1916, Fallout ''^^•''"'Pu in Diet, USAEC Report HASL-306, pp. 115-125, Health and
Safety Laboratory, NTIS.

Farmer, J. G., V. T. Bowen, V. E. Noshkin, and M. B. Gavini, 1973, Long-Lived Artificial
Radionuclides in Lake Ontario. I. Supply from Fallout, and Concentrations in Lake Water, of
Plutonium, Americium, Strontium-90 and Cesium-137, unpublished.

Olsen, C. R., H. J. Simpson, R. F. Bopp, S. C. Wilhams, T. H. Peng, and B. L. Deck, 1978, A
Geochemical Analysis of the Sediments and Sedimentation in the Hudson Estuary, /. Sediment.
Petrol, 48: 401418.

Simpson, H. J., C. R. Olsen, R. M. Trier, and S. C. Williams, 1976, Man-Made RadionucUdes and
Sedimentation in the Hudson River Estuary, -S'c/e/!ce, 194: 179-183.

, R. F. Bopp, C. R. Olsen, R. M. Trier, and S. C. Wilhams, 1978, Cesium-137 as a Tracer for

Reactive Pollutants in Estuarine Sediments, in First American— Soviet Symposium on
Chemical Pollution of the Marine Environment, L. L. Turekian and A. I. Simonov (Eds.), Report
EPA-600/9-78-038, pp. 102-111, U. S. Environmental Protection Agency.

Sprugel, D. G., and G. E. Bartelt, 1978, Erosional Removal of Fallout Plutonium from a Large
Midwestern Watershed,/. £'«v/>oa Qual, 7: 175.

Wahlgren, M. A., and J. S. Marshall, 1975, The Behavior of Plutonium and Other Long-Lived
Radionuclides in Lake Michigan. 1. Biological Transport, Seasonal Cycling and Residence Times in
the Water Column, in Impacts of Nuclear Releases into the Aquatic Environment, Symposium
Proceedings, Otaniemi, Finland, 1975, pp. 227-243, STI/PUB/406, International Atomic Energy
Agency, Vienna.

, J. J. Alberts, D. M. Nelson, and K. A. Orlandini, 1976, Study of the Behavior of Transuranics and

Possible Chemical Homologues in Lake Michigan Water and Biota, in Transuranium Nuclides in the
Environment, Symposium Proceedings, San Francisco, 1975, pp. 9-24, STI/PUB/410, Interna-
tional Atomic Energy Agency, Vienna.

Biological Effects of Transuranic Elements
in the Environment: Human Effects
and Risk Estimates

ROY C. THOMPSON and BRUCE W. WACHHOLZ

The potential for human effects from environmentally dispersed transuranic elements is
briefly reviewed. Inhalation of transuranics suspended in air and ingestion of transuranic s
deposited on or incorporated in foodstuffs are the significant routes of entry. Inhalation
is probably the more important of these routes because gastrointestinal absorption of
ingested transuranics is so inefficient. Major uncertainties are those concerned with
substantially enhanced absorption by the very young and the possibility of increased
availability as transuranics become incorporated in biological food chains.

Our knowledge of plutonium distribution and retention in the human is based on
human autopsy data and on the extrapolation of a large body of experimental animal
data. These data are undoubtedly more precise than our knowledge of the environmental
exposure pathways that may lead to such deposition and more precise than our
knowledge of the health consequences that may result from this deposition.

There is no positive information on the effects of transuranic elements in either man
or experimental animals at the very low exposure levels with which we are concerned.
Various approaches to the evaluation of this problem are discussed. We can conclude with
some certainty that effects from present fallout levels will never be detected as a
perturbation on normal cancer death rates. The possibility of no cancer deaths from
fallout plutonium cannot be precluded.

The principal focus of this book is the environment, exclusive of man. Our ultimate
concern, however, is for the effect of this environment on man. This chapter, therefore,
reviews briefly the routes by which man can interact with transuranics in the environment
and the possible consequences to man of such interaction. The level of treatment in this
chapter is less detailed than in other chapters. The reader interested more specifically in
effects on man and in the animal studies bearing on that problem can find such detail in
several recent compilations: Hodge, Stannard, and Hursh, 1973; Bair, 1974: Thompson
and Bair. 1972: Jee, 1976; and Wachholz. 1974.

Routes of Exposure

Opportunities for exposure of man to transuranics are of two quite different types: those
resulting from employment in the nuclear industry and those resulting from the general
dispersal of transuranics throughout the environment. It is the latter type which concerns
us in the context of this book. Routes of exposure will differ for the two types, but
effects are presumed to be similar in both cases.

691

692 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Two routes of exposure, ingestion and inhalation, are of significance tor general
environmental dispersal. Of these, inhalation is the better understood because the
mechanisms and kinetics of deposition and retention in the lung are primarily determined
by physical factors, which have been studied for many substances in man. Such data can
be extrapolated with some confidence in predicting the behavior of transuranics in the
lung. The absorption of ingested transuranics from the intestinal tract is of more critical
uncertainty, however, because such absorption is primarily determined by chemical
factors unique to each individual transuranic and because only that fraction which is
absorbed is of primary hazard concern. Unabsorbed alpha-emitting transuranics, while
passing through the gastrointestinal tract, are conservatively assumed by the International
Commission on Radiological Protection (ICRP) to deliver 1% of their decay energy to
sensitive cells of the intestinal wall. On this basis the intestine becomes the critical organ
for ingestion of insoluble transuranics (International Commission on Radiological
Protection, 1960), although it is unlikely that significant damage actually occurs by this
mechanism.

Inhalation

The model usually employed to describe the kinetics of inhaled transuranics in man is
that of the ICRP Task Group on Lung Dynamics (1966) as modified by the ICRP Task
Group on Plutonium and Other Actinides (1972). This model adjusts for three classes of
parficle solubility and for a wide range of particle sizes. To illustrate the general case of
environmental transuranics, we will assume insoluble particles (class Y) oiOA-jim activity
median aerodynamic diameter (AM AD), which were assumed by Bennett (1976) to be
typical of airborne fallout. The model predicts that 32% of such an inhaled aerosol will be
deposited in the pulmonary region of the lung; the remainder will be immediately exhaled
or rapidly cleared from the nasopharynx or tracheobronchial region. Of the 32%
deposited in the pulmonary region, 40% is cleared with a half-Hfe of 1 day; the remaining
60%, which is equivalent to about 20% of the aerosol initially inhaled, is retained in the
lung with a half-life of 500 days — this fraction is responsible for essentially the total
irradiation of the lung. About 6% of the quantity initially inhaled eventually reaches the
bloodstream and is distributed among the systemic organs. Although these fractions will
vary with chemical form, particle size, and other exposure variables, a sizable fraction of
inhaled transuranic will, under any circumstance, be tenaciously retained by man. This is
an efficient route of entry, as compared to ingestion, and atmospheric transport is a
correspondingly hazardous environmental pathway.

Ingestion

Because of the general insolubility of transuranic oxides and hydroxides and the
propensity for more soluble compounds to hydrolyze at physiological pH, one anticipates
httle absorption of these elements from the gastrointestinal tract. The ICRP assumes a
fraction absorbed of 3 X 10~^ for plutonium and 10""* for americium and curium
(International Commission on Radiological Protection, 1960); a value of 10~^ is
suggested as appropriate for plutonium oxide (International Commission on Radiological
Protection, 1972). These values are based on the results of animal studies. Several recent
investigations have indicated somewhat higher absorption than that assumed by the ICRP
and a typically large variability from experiment to experiment (Durbin, 1973; Sullivan
and Crosby, 1975; 1976). Because of diis variability and because there is no direct

HUMAN EFFECTS AND RISK ESTIMATES 693

measurement of gastrointestinal absorption in man, one must be cautious in applying the
animal data to populations ingesting very low levels of transuranics in unknown chemical
forms. Certain complexed forms of plutonium and hexavalent plutonium compounds are
known to be more readily absorbed than other plutonium compounds (Durbin, 1973).

Under present conditions of recent fallout deposition, most ingested transuranics are
likely to be either swallowed following inhalation or consumed as external contaminants
on food (Bennett, 1976). With the passage of time, however, biologically incorporated
transuranics may become a more important factor relative to other forms of ingested
transuranics and relative also to inhaled transuranics. One might expect biologically
incorporated transuranics to be more readily absorbed than inorganic forms, and there are
limited animal data on milk (Finkel and Kisieleski, 1976), meat (Sullivan and Crosby,
1976), and alfalfa (Sullivan and Garland, 1977) which suggest that this may indeed be
true for plutonium. There are also data that suggest the opposite conclusion for
neptunium (Sullivan and Crosby, 1976). The extent to which transuranics may become
biologically incorporated in foods and the gastrointestinal absorbability of such material
are uncertain factors in any evaluation of the impact of environmental transuranics on
man.

Another uncertainty relating to ingestion is the question of enhanced absorption in
the very young. There is now an abundance of data attesting to the fact that neonatal rats
(SuUivan and Crosby, 1975; 1976; Ballou, 1958; Sikov and Mahlum, 1972a; Sullivan,
1978), cats (Finkel and Kisieleski, 1976), and swine (Sullivan, 1978; Buldakov et al.,
1969) absorb a very much larger fraction of ingested plutonium than do adult animals.
Such anomalous absorption by the infant has been reported for many other normally
nonabsorbed substances in many species, including man (Koldovsky, 1969; Sikov and
Malilum, 1972b). How one should extrapolate these animal data on transuranics to man is
not clear either with regard to the magnitude of increased absorption or to the duration
of this effect. In rats absorption drops to near-adult levels by the age of weaning
(3 weeks) (Ballou, 1958). The hfe-style of the infant may protect it from plutonium
ingestion, as compared with the adult, and plutonium deposited at an early age will be
subsequently diluted as a consequence of growth. It is possible, however, that the infant
is more sensitive to the production of deleterious effects from deposited plutonium,
although preliminary reports of studies in rats (Mahlum and Sikov, 1974) and dogs
(Stevens et al., 1978) suggest that this is not the case.

Distribution and Retention of Transuranics in Man

Although few health consequences and no fatalities have been observed to result from
transuranic deposition in man, the distribution and retention of these elements in man
can be measured. Such distribution and retention data permit the calculation of radiation
doses in human tissues which can be compared with tissue doses from other forms of
radiation known to produce effects in man. The similarity of human distribution and
retention data to that measured in experimental animals also lends confidence to the
extrapolation of health -effects data obtained in experimental animals.

Most directly relevant to environmental transuranics are the deposition data for
fallout plutonium in man. An example of such data is shown in Table 1, which hsts 50th
percentile values (50% of individual values are lower) for tissues from more than 170 U.S.
autopsies (Mclnroy et al., 1977). Also shown in Table 1 are computed estimates of
plutonium concentration in tissues (Bennett, 1976), which are based on measured

694 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 1 Concentration of Fallout Plutonium in Man

Plutonium concentration, pCi/g

Computed New York

Autopsy samples*

estimatesf

Organ

(50th percentile values)

1964 1974

Bone (vertebra)

0.50

0.08 0.20

Liver

0.58

0.23 0.54

Lung

0.24

2.48 0.12

Lymph nodes

2.42

43 27

Kidney

0.06

0.03 0.06

Gonads

0.13

0.02t 0.12t

♦Data from Mclnroy et al. (1977).
fData from Bennett (1976).

^Estimated on the assumption that 0.05% of the total body
burden is present in 10 g of ovaries.

airborne plutonium concentrations in the New York area and on the assumption that all
intake by the human is via inhalation according to the ICRP lung model and ICRP
assumptions of tissue distribution and retention (International Commission on Radiologi-
cal Protection, 1972). The autopsies, on material collected from many states, were
performed over a period extending from 1959 to 1976; the estimates are calculated for
New York cumulative exposure to 1964 and 1974. Considering the uncertainties of time
and place, the agreement between measured and computed organ burdens is quite
reasonable for bone, liver, and lung, the organs of principal hazard concern as deduced
from animal toxicity studies. Kidney and gonads also show excellent agreement. This
agreement lends confidence to the ICRP assumptions regarding deposition and subse-
quent redistribution. The lower than predicted lymph node measurements may, at least
partially, reflect the considerable difficulties of sampling pulmonary lymph nodes.

Transuranic distribution data in man are also available from autopsies on occupation-
ally exposed persons (Mclnroy, 1976; Norwood and Newton, 1976) and on intentionally
injected patients considered to be suffering from terminal illness (Durbin, 1972; Rowland
and Durbin, 1976). In these cases tissue levels are much higher, and a larger number of
organs can be analyzed with higher precision. These autopsy data are generally in accord
with the more extensive animal data and with the ICRP assumption that 45% of a
transuranic reaching the blood will deposit in bone and 45% in liver (International
Commission on Radiological Protection, 1972).

Of particular interest are recent human data relating to the distribution of plutonium
within organs. Limited data from sectioned lungs of occupationally exposed persons
suggest that initial distribution of inhaled plutonium is relatively uniform (Mclnroy et al.,
1976) but that, at long time periods following exposure, plutonium concentration is
higher in the periphery of the lung (Mclnroy et al., 1976; Nelson et al., 1972). The
distribution of plutonium among different bones (Larsen, Toohey, and Ilcewicz, 1976)
and the microscopic distribution within bone (Schlenker, Oilman, and Cummins, 1976)
have been studied in autopsy material from a patient who died 17 months after
plutonium injection. This patient was suffering from Cushing's syndrome, and bone
metaboUsm was not normal; nevertheless, the general distribution pattern was encourag-
ingly similar to that which would have been predicted from animal studies.

HUMAN EFFECTS AND RISK ESTIMATES 695

The retention of plutonium in the various organs of man must be known if radiation
doses are to be calculated. We have already noted the assumptions regarding retention in
the lung as postulated in the ICRP lung model. Retention in the systemic organs is known
to be prolonged, as deduced from human plutonium excretion data following intentional
or accidental administration and from much data showing long-term retention in a variety
of animal species (Durbin, 1972). The ICRP has assumed a biological half-Hfe of
100 yr for transuranics in bone and 40 yr in liver; 90% of the plutonium deposited in
lymph nodes is assumed to be retained with a biological half-Ufe of 1000 days, and the
remaining 10% is assumed to be retained without loss (International Commission on
Radiological Protection, 1972). On the basis of a thorough review of the pertinent data,
Durbin (1972) concluded that human bone plutonium might exhibit a shorter retention
half-hfe than Uver plutonium, and, on the basis of recently acquired nonhuman primate
data, Durbin and Jeung (1976) have suggested shorter half-lives for both bone and liver
plutonium.

In summary, it would seem fair to conclude that our knowledge of plutonium
distribution and retention within the human, although uncertain in many details, is
considerably more precise than our knowledge of the environmental and exposure
pathways that lead to this deposition and more precise than our knowledge of the health
consequences that may result from this deposition.

Effects of Transuranics in Man

An unevaiuatable uncertainty attaches to any prediction of specific health effects from
the exposure of humans to transuranic elements at levels contemplated for environmental
dispersal. This uncertainty is due to the absence of any positive information on the
effects of these elements in either man or experimental animals at the exposure levels of
concern. Data are available at much higher exposure levels on the effects of transuranics
in experimental animals and on the effects of certain other forms of radiation in man.
The extrapolation of these data is made difficult by our lack of understanding of the
mechanisms by which these effects occur. In the absence of such understanding, it has
been common practice to extrapolate from the high-dose data by assuming a linear
relationship between radiation dose and biological effect. Such a practice is endorsed by
the Advisory Committee on the Biological Effects of Ionizing Radiations (BEIR) of the
National Academy of Sciences-National Research Council (1972) as "warrant[ing] use
in determining public policy on radiation protection"; in the same sentence they caution
that "explicit explanation and qualification of the assumptions and procedures involved
in such risk estimates are called for to prevent their acceptance as scientific dogma."
Although in this chapter we have used the Unear dose— effect assumption in estimates of
the consequences of human exposure to environmental transuranics, we must emphasize
that these estimated effects, if they occur at all, will be difficult to detect over the
background of indistinguishable effects from other causes.

Experience with Transuranics in Animals

Direct information on the toxicity of transuranic elements is available only from studies
in experimental animals. The radiobiological literature suggests that the effects observed
in such animal experiments will at least qualitatively approximate those which would
occur in man if he were exposed under the same conditions. On the basis of extensive
data from several animal species, it is concluded that the most probable serious effects of

696 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 2 Comparison of Transuranic Health Risk Estimates

Cancer deaths or genetic defects per 10^ organ-rem

Risk

estimates based on data

from humans

BEIR

Risks estimates
based on data

Highf

Lowt

UNSCEARS

MRCH Mays**

from animals

Lung tumors

100

25-50

25 20

60-200tt

Bone tumors

2-51:1:

5 4

10-100§§

Liver tumors

20 10

Genetic

defects (in

SOOHf

50111

HOUH

all subse-

quent

1000***

10***

42***

generations)

*National Academy of Sciences- National Research Council, 1972.

t Relative-risk model with lifetime plateau (U. S. Atomic Energy Commission, 1974).

JAbsolute-risk model with 30-year plateau (U. S. Atomic Energy Commission, 1974).

§ United Nations, New York, 1977.

H Medical Research Council, 1975.

**Mays, 1976.

ft Data from Bair and Thomas, 1976.

1 1 Expressed by UNSCEAR as risk per 10* rads of low-LET radiation to endosteal cells,
which should be roughly equivalent to risk per 10* rem of plutonium alpha radiation averaged
throughout bone.

§§Data from Mays, 1976.

m Specific genetic defects.

***Defects with complex etiology.

long-term low-level exposure to transuranics are lung, bone, and possibly liver cancers.
Most of these data are from experiments with plutonium, but they can probably be
applied to other transuranics with less error than is involved in many other unavoidable
assumptions. Although quantitative extrapolation from animal to man involves consider-
able uncertainty, the animal data suggest cancer risks per 10^ organ-rem of 60 to 200 for
lung (Bair and Thomas, 1976) and 10 to 100 for bone (Bair, 1974; Mays et al., 1976).
These estimates are compared with others in Table 2.

Experience with Transuranics in Man

It is clearly impossible to relate specific observed biological effects in man to the
exposure of man at present levels of environmental plutonium. Some conclusions have
been drawn from the absence of observed effects in the substantial numbers of persons
occupationally exposed to very much higlier levels of plutonium. Cave and Freedman
(1976), investigating the adequacy of present plutonium exposure limits, conclude that,
"total exposure represented by the available human data is not yet large enough to
substantiate fully, on a statistical basis, the value of 0.016 fJiCi for the maximum
permissible lung burden. However, regarded as a 'best estimate' this value should not be
too higli by a factor of more than 15 or by a factor of more than 40 at the 95% upper
confidence level." On the basis of the long-term survival without bone tumors of eiglit
"terminal" patients injected with plutonium, Rowland and Durbin (1976) conclude that,

HUMAN EFFECTS AND RISK ESTIMATES 697

"the bone-tumor risk from plutonium is no greater than that from radium, and might be
less." Certainly it would seem clear by now that occupational exposure to plutonium has
not resulted in the kind of tragedy visited on the radium dial painters or the uranium
miners.

Experience with Natural Radiation in Man

Alpha-emitting elements are a natural part of man's environment. He has lived with these
internally deposited radioelements and with radiation from other natural sources
throughout the history of the species. It is of some relevance to note that inhaled
naturally occurring alpha-emitting radionuchdes contribute an average annual dose of
about lOOmrem to the lung and that naturally occurring alpha emitters in bone
contribute an average annual dose at bone surfaces of about 40 mrem (National Council
on Radiation Protection and Measurements, 1975). Although these doses cannot be
related to any measure of specific effects, they have been at least tolerable on the
evolutionary scale, and therefore sUght increases would not be expected to have
catastrophic effects.

Experience with Other Types of Radiation in Man

Inferences concerning the effects of transuranic elements in man may be drawn from
information available on the effects of other forms of ionizing radiation in man; e.g., data
derived from medical, occupational, accidental, or wartime exposure of humans to
different radiation sources, including external X radiation, atomic bomb gamma and
neutron radiation, and radium, radon, and radon daughters. Such information was
summarized by the BEIR Committee (National Academy of Sciences-National Research
Council, 1972) and, most recently, by the United Nations Scientific Committee on the
Effects of Atomic Radiation (UNSCEAR) (1977). Both groups arrived at comparable risk
estimates. England's Medical Research Council (MRC) (1975), considering much the same
information covered in the BEIR and UNSCEAR reports, derived risk estimates
specifically applicable to plutonium.

Of particular relevance are recently accumulated data on the carcinogenicity of
^^"^Ra in human bone (Spiess and Mays, 1970; 1973); ^^"^Ra has a very short half-Ufe
(3.62 days) and, because of this, irradiates only the surface layer of bone in much the
same manner as transuranics. From these ^^"^Ra data. Mays et al. (1976) have estimated
human bone cancer risks from plutonium; Mays (1976) has also estimated liver cancer
risks, which are based largely on experience with Thorotrast, and lung cancer risks, which
are based largely on data from the Japanese atomic bomb survivors.

Concluding Comments

Table 2 compares estimates of cancer risk from several sources previously discussed. Also
included in Table 2 are estimates of genetic risk as derived in the BEIR and UNSCEAR
reports.

Although of dubious quantitative applicabiUty to the problems of environmental
exposure because of the extrapolation uncertainties discussed previously, the kind of data
presented in Table 2 will inevitably be used to estimate health effects from such
exposure. As an example of such an exercise, in Table 3 we have derived an estimate of
the human health consequences of the environmental dispersal of bomb-test plutonium in

698 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

TABLE 3 An Estimate of Cancer Deaths in the United States
Due to Fallout Plutonium

Risk

Individual

Popula-

factors,t

dose,*

tion dose,t

cancer deaths/

Cancer

Organ

mrem

10* organ-rem

deaths §

Lung

3.2

Bone

6.8

Liver

3.4

Total

34
125

*Cumulative organ dose to year 2000 based on inhalation exposure
from 1954 to 1973. Data from Bennett (1976).

fProduct of individual dose and U.S. population of 2 x 10*.

:t:Risk factors suggested by Mays (1976).

§ Product of population dose and risk factor. Cancer death estimate
is uncorrected for prior death from other causes.

the United States. On the basis of New York air samples, the ICRP lung model, and other
metabolic parameters previously described, Bennett (1976) has calculated cumulative
organ dose rates to the year 2000 for an individual exposed from 1954 through 1973. If
we multiply these doses by 200 million people, we have an estimate of the total man-rem
exposure resulting from plutonium fallout in the United States — an estimate that has
obvious limitations but is probably more accurate than many other factors that go into
the health-effects estimate. Multiplying this population dose by the cancer risk factors of
Mays (1976), we arrive at an estimate of 125 cancer deaths. Because of the difficulty in
defining a genetic effect and uncertainties in regard to the genetically effective dose from
transuranics, we did not attempt an estimate of genetic effects in Table 3; it is generally
agreed that such effects are probably "a minor part of the total" (Medical Research
Council, 1975).

Mays' (1976) risk factors were used in Table 3 as the "best guesses" in our opinion.
Use of the most pessimistic estimates of Table 2 would have led to a maximum cancer
death count about four times higher. Neither estimate would constitute a detectable
perturbation on normal cancer death rates; the possibiHty of no cancer deaths from
fallout plutonium is not precluded.

Acknowledgment

This work was supported by the U. S. Department of Energy under contract
EY-76-C-06-1830.

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700 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

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Radium, pp. 329-341, W. S. S. Jee (Ed.), J. W. Press, Salt Lake City, Utah.

Schlenker, R. A., B. G. Oltman, and H. T. Cummins, 1976, Microscopic Distribution of ^^'Pu
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in Radiobiology of Plutonium, pp. 261-272, B. J. Stover and W. S. S. Jee (Eds.), J. W. Press, Salt
Lake City, Utah.

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Health Phys., 19: 713-729; also Addendum, //eaWz Phys., 20: 543-545 (1971).

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Adults, in Radionuclide Carcinogenesis, AEC Symposium Series, Richland, Wash., May 10-12,
1972, C. L. Sanders, R. H. Busch, J. E. Ballou, and D. D. Mahlum (Eds.), pp. 437-450,
CONF-720505, NTIS.

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the Gastrointestinal Tract of Newborn and Adult Rats, in Pacific Northwest Laboratory Annual
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Administrator for Environment and Safety, ERDA Report BNWL-2100(Pt.l), pp. 137-139,
Battelle Pacific Northwest Laboiatories, NTIS.

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tal Distribution and Biomedical Effects, USAEC Report WASH-1359, NTIS.

Ecological Effects of Transuranics
in the Terrestrial Environment

F. W. WHICKER

This chapter explores the ecological effects of transuranium radionuclides in terrestrial
environments. No direct studies that relate the level of transuranic contamination to
specific changes in structure or function of ecological systems have been carried out. The
only alternative approach presently available is to infer such relationships from
observations of biota in contaminated environments and models. Advantages and
shortcomings of these observations as well as those of the direct experimental approach
are discussed. Searches for ecological effects of plutonium contamination in terrestrial
ecosystems adjacent to the Rocky Flats plant (Colorado) and at the Nevada Test Site
have not positively demonstrated plutonium-induced perturbations. These studies were
carried out in areas containing of the order of 10 to 1000 jjiCi ^^'^Pu/m^ in the upper 3
cm of soil. Simple calculations suggest that ^^^Pu applications on the order of 1 Ci/m^
may be required to cause significant mortality in plant populations. Models and
calculations indicate that over 1 mCi ^^^Pu/m^ would likely be required to produce
subacute mortality in mammals. Additional research applicable to ecological effects is
suggested.

To grasp the ecological implications of transuranium elements in the environment, we
must understand their chemical, physical, and biological behavior through time. We must
also understand the effects on biological systems of these elements when they are
dispersed into the environment. Knowledge of the biological effects is particularly
lacking. This may seem surprising in view of the large research efforts that have been
devoted to the biological effects of plutonium and other transuranics. The lack of
quantitative understanding in the area of ecological effects is not so surprising, however,
when the complexities of the problem are considered. Such complexities include the
environmental behavior of transuranics, which is dependent on the physical and chemical
form of the nuclides as well as on the nature of the ecosystem. Of major importance is
the dose to certain tissues, but dose distribution is especially complex for relatively
insoluble alpha emitters. A high-level application of transuranic may have Uttle radiation
effect if energy is not deposited in critical cells.

Although we know a great deal about the effects of plutonium on experimental
mammals (Bair and Thompson, 1974), we know very little about its effects on the other
classes of animals that have important functions in natural systems and even less about its
effects on plants. Also, very little is known about the general biological effects of the
other transuranics. The effects of X- and gamma radiation on major plant and animal
phyla have been studied in depth, but the extrapolation of X- and gamma radiation

701

102 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

information to insoluble alpha emitters is seriously complicated by dosimetry and the
relative biological effectiveness (RBE) of alpha particles for most actinides. Recent
reviews on plutonium and other actinides in the environment have very little to say about
ecological effects; rather, they dwell primarily on distribution and behavior (Romney and
Davis, 1972; Martell, 1975; Hakonson, 1975; Hanson, 1975).

There are three basic approaches to the study of ecological effects of transuranics:
(1) direct experiments in which radionuclides are applied at various levels to study
systems, (2) observations of populations that occupy contaminated areas, and (3) model-
ing and extrapolation from applicable research data. Each approach has inherent
advantages and shortcomings. The direct experimental approach might enjoy a relatively
high degree of credibility and accuracy, but it has not been used with transuranics for
reasons of safety and lack of public acceptance. Examination of contaminated areas is
quite feasible and has been done at such places as the Nevada Test Site, Enewetak, and
Rocky Flats (in Colorado). This approach is less than ideal, however, because of the usual
lack of good experimental control and the common presence of more than one
potentially toxic substance, which lead to uncertainties in data interpretation. The third
approach can be used when needed with existing data, but accuracy may be poor because
of the complexity and uncertainty associated with parameter values.

Ecosystems can probably tolerate higher levels of radioactivity from most, if not all,
of the transuranics than from the more biologically mobile fission products, such as ^^Sr
and '^^Cs. Low solubility, lack of essential nutrient analogues, and the virtual lack of
penetrating radiations for most transuranics form the basis for this opinion. However,
critical experiments to make this comparison have not been done; there is some
concern that biological incorporation of long-lived transuranics in the environment may
slowly increase with time, and it is known that very low levels may be carcinogenic.

Direct Experiments

The literature dealing with effects of ionizing radiation on plants and animals is massive.
Important reviews and bibUographies include the BEIR report (National Academy of
Sciences-National Research Council, 1972), the UNSCEAR report (United Nations,
1972), and the bibliography by Sparrow, Binnington, and Pond (1958). The vast majority
of this literature, however, is based on laboratory studies with X- or gamma radiation. A
far smaller body of literature exists on radiation effects on natural populations. Whicker
and Fraley (1974) reviewed field studies dealing with the effects of ionizing radiation on
terrestrial plant communities, and Turner (1975) prepared a similar review for native
animal populations. This literature also is restricted primarily to X-and gamma radiation,
but it provides a substantial basis for understanding dose— effect relationships.

A major problem in applving this information to transuranics is that of determining
the equivalent dose to critical tissues which would result from a given level of
contamination. Of the 17 transuranic nuchdes listed as being of some importance in the
nuclear industry to the year 2000 (Energy Research and Development Administration,
1976), 13 are alpha emitters with generally infrequent emission of weak (mostly <0.07
MeV) photons. The other 4 are beta emitters with accompanying weak photon emissions.
Alpha— weak-photon emitters include the particularly important nuclides Pu, Pu,
^"^^ Am, ■^'^^Cm, and ^'*'*Cm. Alphas from these nuclides have energies of 5 to 6 MeV and
ranges in air and biological tissue of roughly 4 cm and 40 jum, respectively (Walsh, 1970),
which lend considerable complexity to the problem of dosimetry.

TRANS URANICS IN TERRESTRIAL ENVIRONMENT 703

From field studies in plutonium-contaminated areas, most of the plutonium
associated with vegetation appears to be surficial and not incorporated within tissues.
Therefore critical tissues (meristem for growth and flower bud for reproduction) may
receive a widely variable dose from surface contamination, depending on the location of
the material and the thickness of epidermal tissue layers. I am not aware of any studies
designed to show the detailed histological distribution of transuranics in and on plant
tissues in contaminated-field environments.

The effective dose to animal tissues is equally difficult to determine. The dose from
inhalation and ingestion of transuranics is subject to many variables. Absorption,
translocation, deposition, and retention are affected by the physical and chemical forms
of the nuclide and physiology of the animal (International Commission on Radiological
Protection, 1972). The environmental chemistry of plutonium is extremely complex
(Wildung et al., 1977), and our overall understanding is inadequate (Dahlman, Bondietti,
and Eyman, 1976).

A few studies have been conducted in which simulated fallout particles containing
beta and beta— gamma emitters were administered to field plots. The studies by Murphy
and McCormick (1973) and Dahlman, Beauchamp, and Tanaka (1973) come closer to the
kind needed for transuranics in that the problems of dosimetry are circumvented by
simply relating effects to the level of fallout simulant applied. Murphy and McCormick
applied ^°Y-coated albite particles to experimental granite outcrop plant communities.
The effects on the reproductive potential of Viguiera porteri treated with 0, 205, and 526
mCi/m^ were measured. Dahlman, Beauchamp, and Tanaka applied ^ ^^Cs fused to silica
sand particles to 100-m^ plots in a fescue meadow. The levels applied (22 mCi/m^)
caused measurable decreases in seed production of Festuca arunduiacea. A similar study
using ^^ Y-tagged sand grains to produce effects on crop plants was conducted by Schulz
(1971). Fallout simulants containing ^ ^^Cs were also apphed to field plots at Oak Ridge,
Tenn., to study the effects on arthropods and small mammals (Auerbach and Dunaway,
1970).

For research findings to be integrated and understood, however, it is highly desirable
to estimate the dose to critical tissues from the levels of simulants applied. In the studies
cited, beta-particle doses were estimated by thermoluminescent dosimeters and various
computations. The Stanford Research Institute developed fallout-particle simulants for
the field studies and measurement and computational techniques for beta dosimetry
(Lane, 1971; Brown, 1965; Mackin, Brown, and Lane, 1971). Similar technology could
probably be applied to alpha emitters for their use in field studies.

I am not aware of any studies in which physically and/or chemically characterized
transuranics have been experimentally applied to field plots at levels sufficient to cause
measurable ecological effects. The safe conduct of such studies would require an area
remote from human habitation and stringent health physics practice and cleanup. Such a
study would be expensive, possibly hazardous, and difficult to justify. A greenhouse
study involving plants growing on soil that has been heavily contaminated with
transuranics is being conducted by A.Wallace and E.M. Romney at the University of
California at Los Angeles. One of the objectives of this study will be the effects of alpha
particles.

Another investigation that bears on the problem of biological effects of transuranics
in the environment is under way at Battelle— Pacific Northwest Laboratories under the
direction of R. E. Wildung. Early results indicate radiafion toxicity from ^^^Pu and
^^^Pu to some strains of soil actinomycetes and fungi at levels of 0.7 /jCi/g (soil)

704 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

(~2.5 X 10^ )uCi/m^). Such toxicity was expressed as a decline in microbial numbers.
Since microbes perform functions in soil that are important to plant growth, indirect
effects to higher plants and animals could be elicited through microbial perturbations
from plutonium in soil.

It appears that large quantities of a transuranic nuclide would be required in the field
to cause obvious ecological effects. Two very crude calculations, one for liigher plants and
one for animals, illustrate the approximate levels of ^^^Pu required to produce, for
instance, detectable mortality.

Plant Communities

Assumptions: A grassland plant community requires a dose rate of about 40 rad/day to
show measurable changes in diversity (Whicker and Fraley, 1974); the effective decay
energy for ^^^Pu is 53 MeV/d, considering an RBE of 10 (International Commission on
Radiological Protection, 1960); a concentration ratio (CR = activity per gram of
plant -i- activity per gram of soil) of lO""* is assumed (Energy Research and Development
Administration, 1976); the '^^Pu is assumed to be uniformly distributed within plant
tissues and uniformly distributed in the upper 3 cm of soil, which has a bulk density of
1 .2 g/cm'' . Surficial contamination is neglected.
Calculations:

Required ^^^Pu concentration in plant tissue

1.5 X 10"^ inCi/g

(40 rad/day )(6.25 X 10' MeV/g-rad) _ , , „ ,^_2

(53 MeV/d)(3.2 X 10^ d/day-MCi)

Required ^■'^Pu concentration in soil
^l.5xlO"'MCi/g

Required -^^^Pu application to soil

= (150MCi/g)(1.2g/cm^)(3cm3/cm2)(10'* cm^/mM=5.4 x 10*^ A/Ci/m^

This value is within an order of magnitude of the soil plutonium levels that appear to
evoke some toxic effects in plants under greenhouse conditions (R. E. Wildung and T. R.
Garland, Battelle— Pacific Northwest Laboratories, personal communication).

Animals

Assumptions: Inhalation of suspended soil is considered the critical route of entry;
human and experimental animal data and standards are used; the maximum permissible
"^ ^^Pu human lung burden of 1 .6 x 10^-'' /^Ci/g is achieved with a mean air concentration
of 10^^ ixQilnv' (International Commission on Radiological Protection. 1950); the
critical concentration of ^^^Pu in the lung required for subacute death is 1 x 10^' M^i/g
(Bair, 1974); and a mean resuspension factor of lO~^/m is assumed.

TRANSURANICS IN TERRESTRIAL ENVIRONMENT 705

Calculations:

Required air concentration

__ (lxlO-^MCi/g)(IO-'MCi/,n3) ^ , 3 ^ ,0^3 ^„/™=
1.6 X 10-5 ^(^j/g

Required ^^^Pu application to soil
_6.3 X lO-^iuCi/m^ _

10-5/m

630MCi/m^

If these calculations approach reality, it is clear that very large applications of ^ ^^Pu
would be required to produce measurable ecological changes, especially in plant
communities. Nevertheless, such studies, if done, would carry more credibility than crude
extrapolations and simplified calculations.

Contaminated Environments

The approach of examining ecosystems that have been accidentally contaminated with
transuranics is feasible and probably desirable. Because of the lack of direct experimental
data and the inherent complexity and uncertainty in computational models, we should
look at areas that have been contaminated to ^^^Pu activity levels that significantly
exceed worldwide fallout levels. Several such areas exist or have existed in the past. These
include Rocky Flats, Trinity, several areas at the Nevada Test Site, and various sites on
Enewetak atoll (in the Pacific). In addition, plutonium releases to the environment have
occurred from nuclear facilities at Oak Ridge, Hanford, Mound Laboratory, Los Alamos,
Savannah River, Idaho National Engineering Laboratory, and from bomber crashes in
local areas in Greenland and Spain.

If sufficiently careful searches for ecological changes in contaminated areas prove to
be negative, then it probably can be concluded that the observed levels had no detectable
consequence. Such data should be examined in the light of laboratory information for
additional assurance. If biological perturbations are discovered in contaminated areas,
then it may or may not be possible to assign causal factors. In many contaminated sites,
more than one toxic substance may be present, or other factors may be responsible for
changes. It may be possible to offset these problems if a proper control area is available,
but this should be determined before the initiation of any search for effects.

A comparison of various biological measurements between two ecologically similar
study areas of substantially differing '^^Pu levels at Rocky Flats was conducted by T. F.
Winsor* and C. A. Littlet of Colorado State University and G. E. Dagle of Battelle—
Pacific Northwest Laboratories (Whicker, 1976; Little, 1976). The ^^^Pu readings from
soil in the principal study areas ranged from 100 to over 20,000 d/min per gram in the
upper 3 cm (2 to 400 )uCi/m^). In addition, comparative data were obtained from control
areas containing only worldwide fallout plutonium of the order of 0.1 d/min per gram

* Present address: Rockwell International, Rocky Flats Plant, Golden, Colo.

t Present address: Division of Health and Safety Research, Oak Ridge National Laboratory, Oak
Ridge, Tenn.

106 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

(0.002 iiCxjm^). Biological measurements were made, including vegetation-community
structure and biomass; litter mass; arthropod community structure and biomass; and
small mammal species occurrence, population density, biomass, reproduction, and whole
carcass and organ masses. In addition, small mammals were examined by X ray for
skeletal sarcomas, microscopy for lung tumors, and necropsy for general pathology and
parasite occurrence.

Although minor differences in certain biological attributes between study areas were
observed, none could be related to plutonium levels. Pathological conditions and parasites
were found in some rodents, but occurrence frequencies between control and
contaminated areas were similar. No evidence of cancers or other radiogenic disease was
found. Tliese observations and measurements, combined with intensive field observations
over a period of 5 years, led to the conclusion that plutonium contamination at Rocky
Flats has not produced demonstrable ecological changes. Furthermore, the levels of
plutonium observed in tissues of plants and animals in contaminated areas were
insufficient to produce the doses that would be required to produce obvious biological
changes.

Subcellular biological changes, such as chromosome aberrations, cannot be ruled out
at Rocky Flats. Even if chromosome aberration frequencies were increased in the more
highly contaminated areas, however, population-level changes would likely not persist
because of the surrounding reservoir of normal genetic information. The possibility of
long-term biological effects cannot be discounted either, although this would appear
highly unlikely, nor can we conclude that a similar level of plutonium contamination
spread over a much larger area would be without ecological consequence. The latter idea,
discussed by Odum (1963), stems from the fact that population effects in a highly
localized area can be readily masked by immigration of unaffected individuals and
propagules from the surrounding area, emigration of affected individuals, and gene pool
mixing between the contaminated and surrounding areas. Such masking might not
operate, at least to the same degree, for a large area. The validity of any future studies of
animals in small-size contaminated areas might be increased if a barrier were erected to
prevent the animals from moving into or out of the study area.

Extensive searches for ecological changes in contaminated areas have been carried out
at the Nevada Test Site (Wallace and Romney, 1972; Allred, Beck, and Jorgensen, 1965;
and Rhoads and Piatt, 1971). In the majority of these studies, however, the
contamination consisted principally of mixed fission products, and, except for the work
reported by Rhoads and Piatt, the more dramatic ecological effects were generally
attributed to nonradiological perturbations. The best opportunities for searching for
ecological effects from plutonium alone exist in a number of areas on or adjacent to the
Nevada Test Site which have been used for "safety shot" tests. These tests involved
detonation by conventional explosives of plutonium in various containment configura-
tions. Some 300 acres containing on the order of 10 iiC\ Pu/m^ exist, and a few acres
have levels exceeding 6000 wCi Pu/m^ (Wallace and Romney, 1975; Martin and Bloom,
1976). Studies of small mammals and grazing cattle in these areas have failed to discover
any evidence of radiogenic pathology (Romney and Davis, 1972; Smith, Barth, and
Patzer, 1976). Varney and Rhoads (1977) have examined shrubs in areas assumed to have
been contaminated primarily with plutonium. Although their data implied that such
shrubs had increased frequencies of chromosomal aberrations in comparison to controls,
the evidence was not conclusive.

TRANSURANICS IN TERRESTRIAL ENVIRONMENT 707

Although, as mentioned, other sites in the world have been contaminated with
plutonium, I am not aware of any specific searches for ecological effects at these sites.
Competent ecologists have conducted studies on plutonium distribution and behavior
within many of these sites, however, and any readily apparent ecological changes would
likely have been reported. I am also not aware of any sites at which other transuranics
have been released at levels greater than existing plutonium levels.

Another approach to the study of contaminated environments is to examine areas
containing above-normal amounts of the naturally occurring radionuclides. Many areas
contain substantial quantities of natural uranium and thorium. These primordial
radionuclides and many of their progeny are alpha emitters. Possibly some inferences to
the transuranics could be made from studies in areas of high natural alpha activity. For
instance, the rodents on Morro do Ferro in Minas Gerais, Brazil, which receive an
estimated lung dose of 10^ to 10^ rem/yr, might provide a good study opportunity
(Drew and Eisenbud, 1966). Major problems with such an approach include differences in
radiation schemes and chemical properties between the naturally occurring and
transuranium radionucHdes. We know something about the relative toxicities of ^^^Pu
and ^^^Ra (Thompson, 1976) but very little about the relative ecological and
physiological behavior and toxicities of transuranics with other naturally occurring alpha
emitters. Pochin (1976) points out some other difficulties inherent in trying to quantify
biological effects of environmental radioactivity at low levels.

On the basis of data summarized by the United Nations (1972), I calculate that the
upper 3 cm of soil in the United States averages roughly 0.3 juCi of natural alpha activity
per square meter. A similar calculation applied to atypically high natural radiation
background areas yielded alpha activities of 7 juCi/m^ in the upper 3 cm near Central
City, Colo. (Mericle and Mericle, 1965), 50 /iCi/m^ in local areas in Brazil (Eisenbud et
al., 1964), and 200 /jCi/m^ in the USSR (Maslov, Maslova, and Verkhovskaya, 1967).

A number of genetic and ecological studies have been done in some of these and
similar areas. Rats occupying a monazite sand area in Kerala, India, had no discernible
phenotypic effects (Gruneberg et al., 1966). There is, however, suggestion of radiation-
induced genetic damage leading to mental retardation of humans who occupy the same
region in India (Kochupillai et al., 1976). Furthermore, Gopal-Ayengar et al. (1977)
report genetic alterations in plants indigenous to the monazite belt in Kerala. Cullen
(1968) reported preHminary findings of a human cytogenetic study in Guarapari, Brazil,
in which an apparently increased incidence of somatic chromosome aberrations in
comparison to a control area was found. A high incidence of multiple-break aberrations
was noted which was thought to be compatible with the presence of internal alpha
emitters. These findings were apparently corroborated more recently by Barcinski et al.
(1975).

Osburn (1961) observed an increased incidence of morphological anomalies and
flower abortion in Penstemon virens growing on the more radioactive sites near Central
City, Colo. However, the chemical toxicity of thorium and possibly other factors cannot
be ruled out as causal. In the USSR, Maslov, Maslova, and Verkhovskaya (1967) reported
various deleterious effects on reproduction, parasite infestation, and the general condition
of small mammals in areas of high natural radiation. Although radiation was implied as
the cause of such effects, it was not the only variable between experimental and control
populations. I am not convinced from these studies that naturally occurring alpha
emitters, even in the unusually high natural background regions of the world, cause
demonstrable ecological consequences. Potential genetic changes in local areas would

708 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

likely be disadvantageous and selected out of populations (National Academy of
Sciences-National Research Council, 1972;Muller, 1950).

Application of Existing Data

Existing data can be used to predict the magnitude of human or ecological hazard from a
given level of transuranic contamination. As mentioned, however, computational models
that reasonably simulate actual environmental and physiological processes require many
parameter values which in themselves vary with circumstances and site. Existing
knowledge of appropriate parameter values for plutonium behavior in extensively studied
areas, such as the Nevada Test Site, the White Oak Creek drainage at Oak Ridge, and
Rocky Flats, appears sufficient to develop models with reasonable credibility. Data that
could be applied to most other terrestrial environments, however, are essentially lacking.
This is especially true for transuranics other than plutonium.

Complexities involved in computational models have been discussed in considerable
detail by Healy (1974), Anspaugh et al. (1975), and Martin and Bloom (1976). Healy
(1974) undertook the difficult task of calculating the levels of plutonium in soil which
might be considered standard or guideline levels for humans residing on and deriving
sustenance from such soils. The standard levels calculated could conceivably result in the
attainment of maximum permissible doses for members of the public. The computations
were general in application and used available experimental data and conservative
assumptions. The conceptual model considered surface soil to be the major reservoir and
source of plutonium and considered processes by which the material might reach the
critical organs of man. These processes included resuspension, atmospheric dispersion,
cloud depletion, deposition, inhalation, ingestion of soil and contaminated foods, skin
absorption, and metabolic behavior following intake. The calculations suggested that
4 X 10'"* iJiCi ^•'^Pu/g or 25 /iCi ^^^Pu/m^ in the top 3 cm of soil was probably a
conservative standard.

Using a similar approach but with site-specific data from the Nevada Test Site, Martin
and Bloom (1976) calculated that 3 nCi ^^^Pu/g (soil) ( 1 70 idCi/nr ) could result m the
nonoccupational maximum permissible dose to the lung (1.5 rem/yr) of a standard man
living over and obtaining food from the soil in question. This model was presented in a
lucid and practical way, and the relative degree of confidence that can be placed on each
parameter used in the model was made clear. The basic approach relates intake rates for
ingestion and inhalation to surface soil concentrations; human metabolic and dose
calculations are based on International Commission on Radiological Protection (ICRP)
models and recommended parameter values.

In the ecological context, it seems important to consider the concept of the "critical
organism." Although our primary concern, is focused on man, the general welfare of the
human population cannot be separated from environmental quality. Legal, moral, and
scientific justification exists for ensuring the protection of species other than man from
environmental contaminants. Indigenous species of plants and animals, by virtue of
proximity and life habitats, will receive substantially higher radiation dose rates than man
at many sites likely to receive transuranic contamination. On the other hand, many wild
species, because of shorter normal life-spans, may not live long enough to develop serious
pathology from chronic low-level exposures. In addition, tbr wildlife, society is generally
concerned about performance of the population, whereas for humans, we are concerned
about the more limiting case of individuals ( Auerbach, 1971 ).

TRANSURANICS IN TERRESTRIAL ENVIRONMENT 709

Historically, assessments of radionuclides in the environment have considered man to
be the critical organism. Tlie assumption has often been made that, if adequate protection
for man is assured, we need not worry about ecological effects. Auerbach (1971) has
addressed this question with the conclusion, "Piesent knowledge based on these and
similar studies of the ecological effects of low-level chronic doses, such as could result
from routine reactor releases under current standards, guidelines, and operational
experience, indicates that any possible biological effects would be undetectable."
Althougli this philosophy generally appears defensible, especially for reactor effluents as
stated, I hope that we do not blindly adopt it for all situations. For example,
nuclear-waste disposal could present unanticipated ecological problems in the future,
possibly without causing hazardous doses to humans.

Present Status and Directions for Future Work

To add clarity, before discussing research needs and possible directions for future work, I
will recapitulate what I think is the status of our knowledge on biological responses to
alpha emitters in the environment. Apparently, transuranics have not been experimentally
applied to study plots in the field. On the basis of limited observations of terrestrial
environments accidentally or inadvertently contaminated with plutonium in the range of
10 to 1000 )uCi/m^, no clear-cut ecological effects attributable to plutonium have been
found. A few investigations have shown biological differences between areas containing
natural alpha radioactivity in the range of 5 to 200 /jCi/m^ in the top 3 cm of soil and
nearby control areas. It is not clear, however, that the differences are caused by variations
in radiation dose. Simulation models and available data imply that humans should be able
to occupy and derive sustenance from land areas containing of the order of 20 to 200 /iCi
^^^Pu/m^ in the top 3 cm of soil without exceeding the nonoccupational maximum
permissible dose to critical organs as recognized by the ICRP. Simplified calculations
suggest that ^^^Pu applications of roughly 1 Ci/m^ may be required in grassland areas to
cause significant mortahty in plant populations. I am not aware of computational models
relating ecological effects to the level of apphcation of transuranics other than plutonium.
The general lack of confidence in the accuracy of our predictive capability at present
appears to justify substanfial research efforts in this area. The shortcomings of the three
general approaches have been discussed; yet I see no other approaches to the problem.
Therefore it seems that enhanced efforts in each area are called for with continual
integration of findings from each.

For direct measurements of the relationship between levels of transuranic application
and ecological effects, such applications would need to be made under controlled
experimental designs. The use of shorter Uved transuranics and engineered barriers to
prevent unwanted dispersal of the radioactive material would reduce the risks from such
an experiment. If such experiments ever become feasible, remote, controlled areas, such
as the Hanford Reservation, the Idaho National Engineering Laboratory, and the Nevada
Test Site, might be considered. In addition, the application of effect-inducing quanfities
of transuranics to terrestrial microcosms might be considered. Although direct-application
experiments seem needed from a scientific viewpoint, I do not necessarily advocate them.

Areas presently contaminated with substantial quantities of transuranics should be
investigated for suitabiHty for long-term study. Areas in which higher levels of
transuranics occur without a previous history of contamination with other materials, such
as fission products, and for which good control areas exist would seem particularly

710 TRANS URANIC ELEMENTS IN THE ENVIRONMENT

valuable for study. Such areas have existed in the past (e.g.. Rocky Flats), but, as a result
of public concern, cleanup operations were judged more expedient than biological
studies. Cleanup decisions are deserving of greater scientific input because in some cases
the operation itself may expose the public to greater risk than leaving the protected
material in place.

Areas that contain notably high levels of naturally occurring alpha emitters seem
deserving of further study, particularly if it can be shown how results might be integrated
with current knowledge of transuranic behavior and effects. Potentially valuable study
areas exist in Brazil, Colorado, Wyoming, and the USSR.

In terms of theoretical efforts, it seems clear that more generally applicable models
are needed. This will require more data from a greater diversity of environments,
however, and a much better understanding of basic transport mechanisms. For example,
we need to know how climate, vegetation, soil, and other ecosystem attributes affect
model parameters that describe such processes as erosion, resuspension, assimilation, and
retention. The substantial quantities of data on the environmental behavior of plutonium
in the Nevada desert or in Colorado grasslands have only limited applicability to
ecosystems in regions of higher precipitation. Resuspension seems to be a particularly
critical process affecting the hazard of deposited transuranics, especially in arid regions.
As a final point, our knowledge of the effects of pure alpha emitters on plants is far less
than our knowledge on animals and is grossly inadequate. Since plants provide stability
and the food base of ecosystems, this deficiency should be corrected.

From a scientific viewpoint, it is clear that additional and redirected research can be
justified for transuranium elements in the environment. Social tolerance of environmental
contamination with radioactive materials, however, appears to be far lower than
biological tolerance. In other words, the level of contamination tliat appears in many
cases to prompt cleanup efforts is considerably lower than that which might be expected
to elicit obvious biological cliange. Tliis argument might be used against continued
funding for environmental transuranic research. If this is to be the case, scientists in the
field may need to provide stronger justification for their work in the future.

Acknowledgments

Preparation of this manuscript was made possible tlirough support from the U. S. Energy
Research and Development Administration under Contract No. EY-76-S-02-1 156 with
Colorado State University. I am indebted to a number of colleagues who assisted in the
development of the manuscript. In particular, I wish to cite the exceptional help of A. W.
Alldredge, R. O. Gilbert, D. C. Hunt, C. A. Little, W. A. Rlioads, R. C. Thompson, T. F.
Winsor, and M. R. Zelle.

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Dosimetry and Ecological Effects

of Transuranics in the Marine Environment

WILLIAM L. TEMPLETON

Radiation doses received by aquatic organisms as a result of exposure to transuranics in
the ertvironment are comparable to those received from natural radionuclides even in
known contaminated areas. At these levels it is doubtful whether experimental studies in
the field, or in the laboratory at similar levels, could reasonably be conducted which
would offer some degree of success in determining radiological effects on individuals,
populations, or ecosystems. Some of the mechanisms of recruitment to exploited fish
populations are considered, and these mechanisms suggest that any radiation-induced
effects would probably be compensated for by density-dependent responses in highly
fecund species. In species with low fecundity, increased stress would clearly increase the
chances of diminishing these populations; however, from the dose-rate estimates, present
levels of radiation are unlikely to provide any additional stress in comparison to
exploitation of some of these species by man. Although little quantitative genetic
information is available for aquatic populations, it appears unlikely, from estimated
mutation rates, that significant deleterious genetic effects due to radiation would be
produced at the present low levels in the environment.

Over the past quarter of a century, the open oceans and coastal waters of the world have
received quantities of artificially produced transuranic elements. These elements have
been distributed globally in the atmosphere as a result of nuclear weapons testing
conducted by the United States, the United Kingdom, the Union of Soviet Socialist
Republics, France, India, and China and the burnup of a U. S. space sateUite (SNAP-9A)
in 1964. Estimates are that about 325 kCi of ^^^'^"^^Pu and about 8 kCi of "^Pu have
been deposited over the globe by weapons testing and about 1 7 kCi of ^ ^ ^ Pu by the SNAP
failure (Hardy, Krey, and Volchok, 1973). On a more local scale, transuranic elements
have been, and continue to be, introduced to the marine environment at nuclear weapons
testing grounds, at nuclear-fuel reprocessing plants, and, to a much lesser degree, by
nuclear power facilities and occasional nuclear-device accidents. In the Pacific the total
inventory of transuranic elements in the, Bikini and Enewetak atolls at the Pacific Proving
Grounds probably is as high as 10 kCi, with a reported net flux from the Bikini lagoon to
the North Equatorial Current of about 6 Ci of ^^^'^"^^Pu per year and 3 Ci of ^'** Am
per year (Nevissi and Schell, 1975). During the period 1960 to 1974 (Hetherington et al.,
1975; 1976), the nuclear-fuel reprocessing plant at Windscale (United Kingdom)
discharged approximately 10 kCi of ^^^'^^^'^''^Pu to the northeast Irish Sea. In recent
years the average transport out of the Irish Sea has been 40 Ci/yr.

The major world inventories of transuranics are, of course, contained in reactors,
weapons stockpiles, reprocessing plants, and waste-storage systems, and only a very small

714

TRANSURANICS IN MARINE ENVIRONMENT 715

fraction of the transuranics produced has been or will be released or disposed of to the
environment. We still, however, have only a scant knowledge of the processes that control
the behavior and fate of the transuranics in the environment, and hence our abihty to
predict and assess the potential effects is hampered.

In the past, and this is still partly true today, the primary consideration has been
anthropocentric, and research priorities have been particularly directed to the assumed
primary pathway of inhalation by man. Less attention has been given to secondary
pathways, wliich result in chronic long-term exposures to man through the food web and
to biota in the ecosystem. Although the primary short-term hazard to man from an
atmospheric release may be via the inhalation pathway, it behooves us to give increased
consideration to the long-term hazard potentials because of the potential time lag in
transfer and long-term persistence in natural reservoirs.

The behavior and fate of transuranics in the marine environment were given very little
attention before the mid-1960s. One reason for this was the lack of methodology and
instrumentation to determine the ultralow levels that existed in the aquatic environment.
In fact, it was not until Pillai, Smith, and Folsom (1964) determined the levels of
2 3 9,2 4 0p^j -j^ marine organisms from weapons tests that any data were published other
than total-alpha measurements from some selected sites. Since that time data have
become available for plutonium isotopes and americium from weapons tests, SNAP -9 A,
and some reprocessing plants in particular in air, freshwater, seawater, sediment, and
biological materials. Much of the available published data was reviewed by Noshkin
(1972). Although intensive monitoring and research studies have been conducted more
recently at the Pacific test sites (U. S. Atomic Energy Commission, 1973), in the
northeast Irish Sea (Hetherington et al., 1976), at La Hague in France (Frazier and Guary,
1976; Guary and Frazier, 1977a; 1977b), and in Lake Michigan (Wahlgren et al., 1976),
by far the greatest amount of data on plutonium isotopes generated in the late 1960s was
applicable only to the determination of the residence times of these materials in the
oceans. Laboratory and field studies on plutonium kinetics in marine ecosystems have
been very limited until recently, and even today Httle research has been conducted on
transuranic elements other than plutonium (International Atomic Energy Agency, 1976).
Data for americium, neptunium, and curium are sparse, and none have been published for
berkelium and californium. This chapter discusses one aspect of transuranics in the
marine environment: the potential effects of radiation from these materials on organisms
in the marine ecosystem.

One way to assess whether the present levels of transuranics, more particularly
plutonium, in the aquatic environment have a potential to result in somatic or genetic
damage to aquatic organisms is to compare the radiation dose rate from plutonium, both
from weapons-test fallout and from selected sites contaminated to a relatively higher
level, with that from natural radiation to which organisms, populations, communities, and
ecosystems have been exposed for near- geological time for their life-spans. These
calculated dose rates can also be compared with experimentally determined data on
effects. Finally, where the availabiUty of data for the latter are lacking, we need to assess,
and perhaps hypothesize, on the basis of other related data on radiation effects,
ecological interactions, and population dynamics, whether or not these dose rates in the
environment could result in somatic and genetic effects that would be detrimental to the
maintenance of aquatic populations. This chapter draws heavily on a recent review and
assessment of the ecological effects of radiation in the marine environment (International

7/6 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Atomic Energy Agency, 1976) and reviews some recent experimental data on the effects
of plutonium on aquatic organisms.

Dose Rates from Transuraiiics and Natural Radionuclides in Natural Waters

A few recent papers have dealt with dose rates received by aquatic organisms exposed to
plutonium in natural waters, and they are compared here with dose rates from the natural
radionuclide ' ' *^Po.

Till, Kaye, and Trebalka (1976), Till and Franks (1977), and Till (1978) have
estimated the dose rate during embryogenesis for six species of marine and freshwater
fishes exposed to plutonium in a variety of natural waters (northeast Irish Sea; Wliite Oak
Lake, Oak Ridge National Laboratory; U-Pond, Hanford; Enewetak lagoon; and Lake
Michigan). They estimated that the dose rates ranged from 7 X 10^^ /irad of plutonium
per hour for plaice (Pleuronectes platessa) eggs exposed at 1 pCi of plutonium per liter of
water in the Irish Sea to 3 X 10~^ /Jrad/hr for carp (Cyprinus carpio) eggs exposed at
1 X 10"^ pCi/liter in Lake Michigan. These dose rates are less than those from the
natural radionuchdes ^ ' °Pb and ^' °Po (Woodhead et al., 1976).

Hetherington et al. (1976) estimated the dose rates to the embryos of plaice, adult
plaice, and other organisms from the northeast Irish Sea where plutonium concentrations
in water are of the order of 1 pCi/Hter and the concentrations of 238,2 39,240p^ -^^ ^^
sediment out to 10 km from the discharge point are of the order of 40 pCi of plutonium
per gram of sediment. The ratio of ^'^^ Am to 2 3 8, 2 3 9,2 4 op^^ tj-^^ ^-^^^^ stations had a
mean value of 1.3 ± 0.2. The estimated dose rates to the developing plaice eggs that had
been exposed to 1 pCi/Uter were in the range of 0.09 to 0.47 Atrad/hr, which is somewhat
less than the 0.7 /urad/hr from natural '^^K and less than tlie dose received by
zooplankton from ^^^Po (Woodhead et al., 1976). Dose rates from ^^^Pu, "^'^"^^Pu,
and ^'^^ Am have also been estimated for young (Group III) plaice, crab (Cancer pagiims),
and mussels (Mytilus edulis). The total dose rates (Table 1) are of the same order as those
from the natural background from ^ ' °Po except for the soft tissues of mussels and gills
of crab. In these exceptions there is a possibility that the tissues were contaminated by
sediment.

Effects Studies

There is a paucity of data in the literature on the effects of transuranics on aquatic biota.
Studies of the effects of ^^'^Pu on the eggs of carp have been reported by Auerbach,
Nelson, and Struxness (1974). No observable effects were seen either in tlie rate of
hatching or in the frequency of abnormalities when eggs were exposed to concentrations
ranging from 5x10^ to 5 x 10^ pCi/liter, which is many orders of magnitude above
concentrations known to exist in the natural environment. Till, Kaye, and Trebalka (1976)
have reported on the doses that produced effects on hatching, survival, and abnormalities
in carp and fathead minnow ('P/>??£'p/za/£'spn)«?^/flSy' eggs (Table 2). Accumulated doses that,
over the period of embryogenesis, first produced a significant effect on hatching and
survival were in excess of 2000 rad for both species. Abnormahties were first produced
when the accumulated dose exceeded 750 rad. These doses would be significantly greater
in dose-equivalent units, where for alpha radioactivity the dose (in rad) is multiplied by a
quality factor. These data indicate that these fish eggs are relatively insensitive to the
effects of alpha radiation from ^^^Pu. The data also infer that concentrations of 1 ^Ci of

TRANSURANICS IN MARINE ENVIRONMENT 71 7

TABLE 1 Total Dose Rates (jurad/hr) to Biota in the Northeast
Irish Sea from Internal Plutonium, Americium, and Polonium*

Tissue

Dose rate, Ai'rad/hr

Species

238py

2 3 9 ,2 4 Op„

24 1

Total

""Po

M. edulis

Visceral
mass

4.7

MoUusca: 4.5 to 12.0
Crustacea:

C. pagimis

Muscle

0.35

1.1

6.9

8.4

Whole, 4.5 to 18

GUI

0.9

130

Hepatopancreas, 140

Digestive

0.58

2.2

gland

Fish:

P. plate ssa

Skin

9.4 X 10-'

3.5 X

10"

-2

0.23

0.27

Bone

4.4 X 10-'

0.15

1.3

1.5

0.2 to 2.5

GUI

0.14

2.4

2.5

Gut

4.3 X 10-'

0.16

1.8

2.0

2.2 to 2.9

Muscle

1.2 X 10-'

4.4 X

10"

-3

2.3 X

10-'

2.8 X 10-'

5 X 10-' to 1.5

Liver

8.7 X 10-'

0.32

6.7

7.1

2.2 to 10

Kidney

0.5

0.96

6.1

*Basedon data from Hetherington et al. (1976) and Woodhead et al. (1976).

TABLE 2 Summary of Estimated* Doses of "^ ^ ^Pu to C. carpio and
P. promelas Eggs Which Produced Effects on Hatching, Survival,

and Abnormalitiesf

2 38

Pu dose,rad

Eggt

Hatching §

Survival

Abnormalities

Number abnormal
Sample size

C. carpio
P. promelas

1.57 X 10"
(8.17 X 10')

9.71 X 10'
(5.60 X 10')

8.19 X 10'
1.94 X 10'

4.27 X 10'
5.68 X 10'

5/377
15/454

*Except where noted, the dose is for the concentration that first produced a
significant effect on hatching, survival, and abnormalities.

t Based on data from TUl et al. (1976).

^Period of embryogenesis (C. carpio, 3 days; P. promelas. 7 days).

S Number in parentheses is the dose at which little or no effect was observed
on hatching and may be considered as an estimate of the threshold dose to affect
hatching.

^^^Pu per liter of water may result in a synergistic effect between chemical and
radiological toxicity because of the high concentration of plutonium mass present.

These doses are within the range of e.xperimental dose rates reported in the
Uterature that result in damage and are clearly many orders of magnitude above
those experienced even in the most higlily contaminated areas (International Atomic
Energy Agency, 1976).

718 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Potential Impact of Ionizing Radiation on Aquatic Resources

The available data on the effects of ionizing radiation on individual organisms indicate
tliat, when experimental and field exposure dose rates are less than 1 rad/day, it is
difficult to demonstrate effects that are not already within the inherent variation already
present. It is pertinent, however, to discuss some of the concerns and potential problems
to provide an improved basis for the assessment, understanding, and acceptance of the
degree of risk tliat aquatic populations are exposed to as a result of the introduction of
radionuclides into the aquatic environment. A valid perspective can be developed by
comparing potential effects with losses caused by natural mortahty and fishing. This
aspect has been discussed in some depth by an international panel (International Atomic
Energy Agency, 1976).

The available evidence suggests that (1) fish are the most radiosensitive component;
(2) developing gametes, fertilized eggs, and larvae are the sensifive stages; and (3) any
damage that might occur to a fishery resource would most Hkely arise from the direct
effects of radiation on the fish rather than from effects from disturbances or changes in
the food web. Some of the earlier literature from laboratory experimentation observed
that effects, seen as the number of abnormal larvae hatched, resulted from exposure to
radionuclides at extremely low concentrations. These findings were extrapolated to
suggest that, if fish were exposed to these concentrations, the yield from commercial
fisheries would be adversely affected. However, consideration needs to be given to (1) the
nature of the stock and recruitment relationships for highly fecund fish where large
numbers of eggs are produced at each spawning (10^ to 10^ per female); (2) the small
number of eggs required to survive to maintain stock at equilibrium; (3) the high
mortality that occurs during the larval stages such that, say, only 1 in 10,000 survives;
and (4) the fact that a direct relationship does not necessarily exist between spawning
stock size and recruitment to the reproductive stock. These factors suggest that, even if
mortality at this stage of development is being enhanced by the low dose rates presently
existing in the aquatic environment, recruitment in highly fecund species of marine fish
would unlikely be adversely affected unless these stocks were already at risk because of
severe overexploitation. The mechanisms controlling recruitment in invertebrates appear
similar, except that environmental factors probably play a more important role.

For species with low fecundity (rays, sharks, dogfish, and marine mammals), most of
whom produce live young, recruitment is related to adult stock size, and, for the marine
mammals, the relationship may be almost direct. Although it is reported that the
fecundity of baleen whales (Laws, 1962) and elasmobranch stocks (Holden, 1973) has
increased as a result of exploitafion, there is an obvious upper limit; available data would
suggest that this upper limit lias been reached by the exploited stocks. Further stress
would cleariy increase the chances of extinguishing the populations. However, although
there are no data on the radiosensitivity of these low-fecund species, the estimates of the
dose rates received by aquatic organisms in the natural waters of the world are rarely of
the same order as, and generally less than, the limits recommended for humans. Hence
they are unhkely to provide any additional stress in comparison to man's continued
exploitation of some of these stocks.

The possible increase in mortality rates due to radiafion in exploited fish stocks can
also be compared with mortality rates experienced by those fisheries presently exploited
by man. Rates of exploitation as high as 50% a year on all classes recruited to a fishery
are common. In addition, mortality due to natural causes occurs; thus a heavily exploited

TRANS URANICS IN MARINE ENVIRONMENT 719

stock may survive even thougli it is subjected to a total mortality rate of 60 to 70% per
year. Any additional mortality, say, due to radiation, would reduce the stock but would
not necessarily affect the ability of the stock to replace itself. Additionally, any mortality
caused by radiation would probably not be detectable as such. There are only a very few
studies on natural populations that have been subjected to low-level irradiation; but, in
those reported (Donaldson and Bonliam, 1964; 1970; Bonham and Donaldson, 1966;
Blaylock, 1961 ; Templeton et al, 1976), there is no evidence that low levels of radiation
have any adverse effects on these populations.

There is virtually no pubUshed data on the genetic effects of irradiation on fish
populations. Although studies have indicated that chromosomal abnormalities can occur
in irradiated eggs and larvae of aquatic species (Ophel et al., 1976), there is no evidence to
indicate that these abnormalities have been detrimental to the population. Predictions
therefore can be based only on studies conducted with other species, e.g., Drosophila. It
can be argued from those data that modest increases in mutation rates with concomitant
enhancement in the genetic variability may even lead to improved fitness (Neel, 1972).
Additionally, the large amount of genetic variability revealed by recent biochemical
techniques may challenge the consensus that mutations are always detrimental in nature
and emphasize the importance of understanding the dynamics of newly introduced
mutants in tlie gene pool and selective processes. These developments have increased the
difficulty of assessing the potential long-term genetic implications of the irradiation of
natural populafions (Neel, 1972; International Atomic Energy Agency, 1976). In this
connection, Woodhead (1974) has conservatively estimated, on the basis of very limited
data, that, if all mutations are dominant lethals resulting in nonviable zygotes, then less
than 1 of every 1000 fish embryos would be eliminated as the result of an integrated dose
of 0.5 rad received by each of the parents.

Research Needs

Many assumptions used in considering the somatic and genefic effects of radiafion on
populations in the aquatic environment are to some degree speculative. Since this is also
true for assessments of the effects of any energy-related pollutants that enter the aquafic
environment, the recommendations for future research made by the International Atomic
Energy Agency (1976) could equally be appUed to pollutants other than radiation.
It is recommended that prime consideration be given to:

• Comprehensive studies on a sufficient spatial and temporal scale to determine the
significance of changes in populations, communities, and ecosystems resulting from
low-level chronic exposure to pollutants, singly and in combination. Emphasis should be
given to determining the rates of change, the rates of recovery from various degrees of
damage, and tlie rates of repopulation in decimated areas.

• Comparative studies of mutation rates induced by pollutants, singly and in concert,
on a wide range of marine organisms, including species with both high and low
fecundities. Emphasis should be given to both genetic damage (gene mutation,
chromosomal aberrations, recombination, etc.) and effects on population size, biomass,
fecundity, and fitness components.

• Studies designed to provide an understanding of the role of genetic variation,
expressed as discrete polymorphisms and quantitative variations of individual species, in
the maintenance of aquatic communities. Emphasis should be given to clarifying the
significance of the response of these polymorphisms to varying physical, chemical, and
biological parameters.

120 TRANSURANIC ELEMENTS IN THE ENVIRONMENT

Summary

Since the dose rates received by aquatic organisms as a result of exposure to transuranics
are comparable to those received from natural radionucUdes, even in known contami-
nated areas, it is apparent that there are few experimental field studies that reasonably
could be conducted which would determine whether radiological effects are occurring in
the environment as a result of present levels of radionuclides. The comparisons drawn
here between the estimated doses from plutonium and americium on the one hand and
naturally occurring polonium on the otlier would be more accentuated if the total dose
rates from all natural radionuclides were computed.

Consideration of some of the mechanisms of recruitment to exploited fish
populations would suggest that any effects as a result of chronic exposure to low-level
ionizing radiations would probably be compensated for by density-dependent responses
in higliJy fecund species. Effects due to radiation therefore would not likely be
distinguishable from those due to natural fluctuations in aquatic populations. Although
httle quantitative genetic information is available for aquatic populations, it is unlikely,
on the basis of predicted mutation rates, that significant deleterious genetic effects due to
radiation at the low levels present in the environment today would be produced in
aquatic populations.

References

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Report, C. I. Taylor (Ed.), USAEC Report ORNL-5016, p. 10, Oak Ridge National Laboratory,

NTIS.
Blaylock, B. G., 1961, The Fecundity of Gambusia affinis affinis Population Exposed to Chronic

Environmental Contamination, y?a^/ar. Res., 37: 108.
Bonham, K., and L. R. Donaldson, 1966, Low Level Chronic Irradiation of Salmon Eggs and Alewives,

in Disposal of Radioactive Wastes into Seas, Oceans, and Surface Waters, Symposium Proceedings,

Vienna, 1966, p. 869, STI/PUB/126, International Atomic Energy Agency, Vienna.
Donaldson, L. R., and K. Bonham, 1964, Effects of Low Level Chronic Irradiation of Chinook and

Coho Salmon Eggs and Alevins, Trans. Am. Fish. Sac, 93: 333.
, and K. Bonham, 1970, Effects of Chronic Exposure to Chinook Salmon Eggs and Alevins to

Gamma Irradiation, Trans. Am Fish. Soc, 99: 112.
Frazier, A., and J. C. Guary, 1976, Recherche d'indicateurs biologiques appropri6s au controle de la

contamination du littoral par le plutonium, in Transuranium Nuclides in the Environment,

Symposium Proceedings, San Francisco, 1975, pp. 679-689, STI/PUB/410, International Atomic

Energy Agency, Vieima.
Guary, J. C, and A. Frazier, 1977a, Influence of Trophic Levels and Calcification on the Uptake of

Plutonium Observed In Situ in Marine Organisms, Health Phys., 32: 21.
, and A. Frazier, 1977b, Etude compar^e de teneurs en plutonium chez divers mollusques de

quelques sites littoraux francais. Mar. Biol, 41 : 263.
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?\\x\om\xm. Nature, 241: 444.
Hetherington, J. A., D. F. Jeffries, and M. B. Lovett, 1975, Some Investigations into the Behavior of

Plutonium in the Marine Environment, in Impacts of Nuclear Releases into the Aquatic

Environment, Symposium Proceedings, Otaniemi, 1975, p. 193, STI/PUB/406, International

Atomic Energy Agency, Vienna.
, D. F. Jeffries, N. T. Mitchell, R. J. Pentreath, and D. S. Woodhead, 1976, Environmental and

Public Health Consequences of the Controlled Disposal of Transuranic Elements to the Marine

Environment, in Transuranium Nuclides in the Environment, Symposium Proceedings, San

Francisco, 1975, pp. 139-154, STI/PUB/410, International Atomic Energy Agency, Vienna.
Holden, M. J., 1973, Are Long-Term Sustainable Fisheries for Elasmobranchs Possible? /?flpp. P.- K.

Reun. Cons. Int. Explor. Mer., 164: 360.

TRANSURANICS IN MARINE ENVIRONMENT 721

International Atomic Energy Agency, 1976, Effects of Ionizing Radiation on Aquatic Organisms and

Ecosystems, Technical Reports Series No. 172, STI/DOC/10/172, International Atomic Energy

Agency, Vienna.
Laws, R. M., 1962, Some Effects of Whaling on the Southern Stocks of Baleen Whales, in The

Exploitation of National Animal Populations, E. C. LeCueu and M. W. Holdgate (Eds.), p. 137,

Oxford University Press.
Neel, J. v., 1972, The Detection of Increased Mutation Rates in Human Populations, in Mutagenic

Effects of Environmental Contaminants, H. E. Sutton and M. 1. Harris (Eds.), p. 39, Academic

Press, Inc., New York.
Nevissi, A., and W. R. Schell, 1975, Distribution of Plutonium and Americium in Bikini Atoll Lagoon,

Health Phys.. 28: 539.
Noshkin, V. E., 1972, Ecological Aspects of Plutonium Dissemination in Aquatic Environments,

Health Phys., 22: 537.
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Radiation on Aquatic Organisms and Ecosystems, Technical Reports Series No. 172, Chap. 2,

p. 57, STI/DOC/10/172, International Atomic Energy Agency, Vienna.
PiUai, K. C, R. C. Smith, and T. R. Folsom, 1964, Plutonium in the Marine Environment, Nature,

203:568.
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in Effects of Ionizing Radiation on Aquatic Organisms and Ecosystems, Technical Reports Series

No. 172, Chap. 3, STI/DOC/10/172, International Atomic Energy Agency, Vienna.
Till, J. E., 1978, The Effect of Chronic Exposure to ^^*Pu(IV) Nitrate on the Embryonic

Development of Carp and Fathead Mmnow Eggs, Health Phys., 34: 333-343.
, and M. L. Franks, 1977, Bioaccumulation, Distribution and Dose of ^"^ Am, ^'"' Am and ^^*Pu

in Developing Fish Embryos, in Proceedings of the Fourth International Congress of International

Radiation Protection Association, Paris, Vol. 2, pp. 645-648.

, S. V. Kaye, and J. R. Trebalka, 1976, The Toxicity of Uranium and Plutonium to the Developing

Embrvos of Fish. ERDA Report ORNL-5160, Oak Ridge National Laboratory, NTIS.
U. S. Atomic Energy Commission, \91?>, Enewetak Radiological Survey, USAEC Report NVO-140,

Vols. I-III, Nevada Operations Office, NTIS.
Wahlgren, M. A., J. J. Alberts, D. M. Nelson, and K. A. Orlandi, 1976, Study of Behavior of

Transuranics and Possible Chemical Homologues in Lake Michigan Water and Biota, in

Transuranium Nuclides in the Environment, Symposium Proceedings, San Francisco, 1975,

pp. 9-24, STI/PUB/410, International Atomic Energy Agency, Vienna.
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Envixonments, and the Assessment of the Possible Consequences, in Population Dose Evaluation

and Standards for Man and His Environment, Symposium Proceedings, Portor6z, 1974,

STI/PUB/375, International Atomic Energy Agency, Vienna.
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International Atomic Energy Agency, Vienna.

Index

Americium-24 1
airborne at Hanford, 263-264
^ " ' Am/.^ 3 9 ,2 4 0 py jatios in Enewetak sedi-
ments, 583
in Bikini Atoll sediments, 545-551
biokinetic behavior in marine biota, 529-530
chemical reactions in soil, 309
concentration ratio (CR), 343-349, 363-368
concentrations in arctic animals, 455
distribution in plants, 343-344, 368
effect of soil concentration on, 340-342,

365-368
in effluent ponds at Hanford, 612-613
field experiments with, 361-368
in-powth from ^'" Pu, 54, 73-75, 90
inventory at Enewetak and Bikini, 588-589
released to White Oak Creek, 384-385
uptake by plants, 340-346, 361-368
variability in sampling in soils, 191-193

Aquatic ecosystems
freshwater, 644-658
general, 7-9, 15-24
Great Lakes, 659-681
Hanford, 625-642
Hudson River, 685-689

Availability of ^ '" Pu compared with ^ ^ ' Pu
in agricultural crops, 398-399
effects of several factors on, 340-344
long term, 349

Behavior of transuranics
in Great Lakes, 662-680
in marine ecosystems, 524-536

Bikini Atoll, transuranic elements in marine
environment of, 541-576

Biological half-life of ^ ^ ' Pu in marine mussels,

527

Biological uptake of transuranic elements
by cattle (model), 479-480
by marine organisms, 525-531

Chemistry of transuranic elements, 3-4, 9-12,
146
in Bikini Atoll seawater, 569
Coefficients of variation (CV)
components of, 165-167
effect of particle size on, 171-172
in predicted values of simulation models,

517-519
for^'^Pu, ^3»Pu, and^^°Pu, 165

in soil, 165-166
in Rocky Flats ecosystem samples, 429-430

Complexation of transuranic elements in soils,
309-312

Concentration ratio (CR)
of^'^'Am, 343-349, 363-368

in marine biota, 529-530
of ^^^ Cm, 365-368

complications in marine ecosystems, 531-532
difficulties in interpretation, 180, 188-189,

343-346
factors influencing in plants, 349-357
in Oak Ridge and Los Alamos biota, 376
of Plutonium in agricultural crops at

Savannah River and White Oak Creek,

398-399
for Plutonium and americium in aquatic

biota, 613-617
of Plutonium in Rocky Flats samples,

433435,654-658
of ^^ ''Pu in marine biota, 526-529
of "" Pu, " ^ '^ "o Pu, and ' ^ ' Cs in lichens,

452453
fQj 2 3 8 ,2 3 9 ,2 4 0 py jj^ southwestcm U. S.

vegetation, 41 1-413, 473477, 479480
of '^»Pu, 363-368

on respirable particles, 253-254
of 2 3 9 ,2 4 0 py jj^ Alaskan biota, 452
Qf 2 3 9 ,2 4 0 py jj^ Enewetak algae, water, and

sediment, 593
fQj 2 3 9 ,2 4 0 pu j„ Q jggt Lakes, 6 1 6-6 1 7
Qj- 2 39 ,24opu jn Rocky Flats biota, 654-658
in terrestrial and aquatic food webs, 26-33

122

INDEX

723

of transuranics under agronomic conditions,
363-368, 398-399
Curium-242

chemical reactions in soil, 309

in Enewetak Atoll water, 583

fallout tracer at Bikini, 542

precursor to ^ ^ ^ Pu, 542
Curium-244

behavior in environment, 14

biokinetics in marine ecosystems, 530

chemical reactions in soil, 309

concentration ratio (CR), 365-368

distribution in plants, 368

field experiments with, 361-368

released to White Oak Creek, 384

undetected in Bikini sediments, 583

uptake by plants, 363-368
effect of soil concentration on, 365-368

Deposition
estimates of fallout on arctic ecosystems,

445^46
estimates of ^ ^ ' -^ ^ <> Pu fallout in Great

Lakes, 659-660
of particles on plants, 290-292
velocities under field conditions, 466-467
Distribution

of ^ "' Am and ^ ^ ' '^ "o Pu in plants, 348

in humans, 694-695

vertical, in Bikini Atoll sediment, 551-559

in Enewetak Atoll sediment, 589-590
Distribution coefficient (K^^j)
in aquatic environments in general, 19-21
in aquatic systems, 615, 618
in Enewetak Atoll sediments, 597-598
in Hudson River components, 687
for neptunium in sodium-saturated clay,

147-148
in suspended sediment, 615-616
Distribution and retention
in aquatic organisms, 617-619
in channel catfish, 6 1 8
in humans, 693-695
Dose to man from plutonium-contaminated
foodstuffs, 394, 489-508, 621-622

Ecological effects
in terrestrial ecosystems, 701-710
from contaminated environments, 705-708
by direct experiments, 702-705
from natural radioactivity, 707-708
predictive models, 708
of transuranic elements in general, 34-36
Ecosystem
arctic, 441-456
deciduous forest, 513-522
Enewetak Atoll, 578-600

grassland, 421-439

terrestrial, ecological effects in, 701-710
Effects of transuranic elements
cancer risks in animals, 695-697
dose rates to aquatic organisms, 715-717
in ecosystems, 702-704
animal communities, 705-708
direct experiments, 702-705
plant communities, 704-705
on soil microorganisms, 315-319
health effects in experimental animals,

695-696
health effects in man, 696-697
Enewetak Atoll
inventory of transuranic elements at, 588-589
nuclear test history at, 579-581
as a transuranic source term, 578-581
Enrichment factor
use in ^ 3 ' Pu particle research, 1 24
use in resuspension calculations, 229
Environmental inventory of ^ ^ ' '^ "" Pu
from accidents, 89
from laboratories and weapons-fabrication

facilities, 89-90
from nuclear weapons tests, 88
from "safety shots," 88-89
in surface soU, 88
Environmental monitoring, 45-49
general aspects of, 47-50
guidance for, 48^9

Foliar absorption of elements, 294-297
Foliar absorption and translocation of pluto-

nium, 294-297,470-471
Food chains

aquatic, 654-656

arctic terrestrial, 444

Savannah River estuary, 609-610

Trophic Transfer Factor (TTF) in, 620
Fractionation of ^ '* ' Am and plutonium in

marine waters, 533
Frequency distributions of transuranic element

concentrations in Rocky Flats samples,

429-431

Great Lakes
estimates of ^ ^ « -^ '^ » Pu faUout in, 659-660
Michigan, plutonium concentrations in, 616,

668-680
Ontario, 2 3 9 ,2 4 o p^ concentrations in, 6 1 7
plutonium in, 670-680
ratio of ^ 3 ' .2 4 0 py/9 0 gj. ^^ 659-660

residence time of ^ ^ ^ '^ '' ° Pu in, 664
sediments, remobihzation of transuranic
elenientsin, 673-674

724

INDEX

Hanford

^ "* ' Am in effluent ponds at, 6 1 2-6 1 3

plutonium in U-Pond components, 631-640

transuranic elements in air and soil, 245-263
Hudson River and estuary, distribution of

plutonium in, 685-689
Human exposure to transuranics

dose rates from, 697-698

effects on humans, 695-698

ingestion, 487-488, 692-693

inhalation, 486-487,692

possible routes, 486-488, 691-693

respirable Rocky Flats contamination,
260-263
Hypothesis testing, 204-205

Industrial wastes

from Light Water Reactor fuel cycle, 93-103
Inventories
of plutonium, in biotic components, 5-6
in deciduous forest components, 377,

519-521
in Hanford U-Pond, 631-636
in New Mexico ecosystems, 377, 413-416
pf 2 3 9 ,2 4 0 py^ jj^ Enewetak sediments, 594
in Rocky Flats ponds, 652
in soils and alluvium at Thule, Greenland,
and in Alaska, 447-449
Ionizing radiation, effects on aquatic organisms,

716-719
Isotopic ratios of ^ ^ « Pu/^ 3 9 ,2 4 o p^
in aquatic systems, 613
in arctic biota, 449-454
in arctic soils, 448
in fallout, 70

in Rocky Flats soils and air, 253-260
in Savannah River estuary sediments, 608-609
in world soils, 71-73

Kj {see Distribution coefficient)

Logarithmic transformed data used in models,

183
Long-term behavior, 33-34
Los Alamos, New Mexico

biota, ^^^■^^'•^^"Pu in, 407-414

rates of plutonium release, 372

soils, "«-" ^-^""Pu in, 409-414

Marine ecosystems, transuranic element

behavior in, 20-24
Mass loading of soil contamination in air,

226-228, 231,468
Microorganisms
role in transuranic element, distribution in
plants, 326-328
solubility in soils, 303, 312-315

Modeling methods, simulation techniques,
COMEX, 517

Modeling results
comparison of predicted and observed values

in Enewetak waters, 598-600
power-function model to describe plutonium

in Rocky Flats soil, 427
predicted release of plutonium by fire in

deciduous forest, 520-521
simulation of plutonium dynamics in

deciduous forest, 517-521
sources of variation in predictions, 520-521

Models
categories, 181

closed system, descriptive, 515-517
to describe plutonium cycHng in deciduous

forest ecosystem, 513-521
for ingestion of transuranics, 490-492
for inhaled transuranics, 50, 493-504, 692
for predicting transuranic element concen-
trations in Enewetak waters, 597-598

comparison of predicted and observed
values, 598-600
^ ^ ^ Pu transport and dose estimation, 459-

508
research needs, 182

theoretical resuspension, 280
resuspension, 210, 467-468
sampling for, 181-184

Mound Laboratory, chemical nature of
plutonium in canal, 155

Natural radiation
compared to other sources, 697-698
dose to humans, 697-698
dose rates to aquatic biota, 697, 715
dose rates to terrestrial biota, 707
health effects in humans, 697
human exposures to, 697-698

Neptunium-237
chemical properties, 14
distribution in plants, 368
in Enewetak Atoll samples, 582
field experiments with, 361-368
in marine mussels and shrimps, 530-531
uptake by plants, 361-368
effect of soil concentrations on, 365

Neptunium-239, ratio to "*°Ba in Chinese
nuclear test debris, 78

Nevada Test Site, solubility of plutonium in
soils, 154

New Mexico ecosystems, inventories of

plutonium in, 377, 519-521
Nuclear fuel reprocessing plants, 382

INDEX

725

Oak Ridge National Laboratory

chemical nature of plutonium in ponds and
floodplain, 155

rates of plutonium release, 372
Oxidation states

of actinide elements, 3-4, 9, 146-148
in soil, 303-312

effect of Eh and pH on, 10-12, 147-148

effect on sorption, 147-154

of plutonium in Great Lakes, 681

Particles
CaMoO^ , used as tracer in resuspension

studies, 276
»^^Cs, at Hanford, 272-273
deposition of, on plants, 290-292
plutonium, airborne flux at Hanford and
Rocky Flats, 245-250, 266-267

attachment to Hudson River sediments,
688-689

behavior on leaf surfaces, 292-294

interception of wheat foliage, 388

^ ^ * Pu and ^ ^ ' Pu in respirable and non-
respirable, 260-263, 272

resuspension of, 240-250

in or on soil, 156-160, 165-166
produced in nuclear tests at Bikini Atoll,

543-544
^^«Pu, study of, 107-142
settling rates in Great Lakes, 667-668, 680
zinc sulfide, used as tracer in resuspension

studies, 276
Pathways for human exposure, 461-463, 691
ingestion, 50-51, 487^88, 692-693
inhalation, 50, 486-487, 692
Physical/chemical forms of transuranic ele-
ments
in Bikini Atoll marine environment, 562-569
in marine ecosystems, 581, 595-597
partitioning in Enewetak waters, 597-600
Plants
distribution of transuranics in, 368
effect on resuspension, 218, 285
interception factor, 470
uptake of transuranics by, 365-368

role of microorganisms, 326-328
Plutonium (general)
in Bikini Atoll sediments, 545-559
chemical reactions in soil, 303-309
comparison with behavior of uranium and

thorium in soils, 156-157
complexity of adsorption to soils, 154
distribution of, in Hudson River and estuary,

685-689
export from Hanford U-Pond by biota, 637-

639

external contamination on agricultural plants

at Oak Ridge, 391-394
foliar absorption and translocation, 294-297,

470^71
in Great Lakes, 659-681
in Los Alamos canyon soils, 374-375
prediction of airborne concentrations, 270-

273
rates of release at Los Alamos, New Mexico,

372
solubility of, in soils at Nevada Test Site, 154
in White Oak Creek floodplain soil, 374-375
Plutonium-237, use in labeling experiments,

525-529,617-618
Plutonium-238
in arctic ecosystems, 441-456
concentration ratio (CR), 363-368
concentration of SNAP-9A debris in

Antarctic air samples, 78
concentrations on airborne soil, 250-253
decay product of ^ " ^ Cm at Bikini, 542
distribution, in Enewetak Atoll environs, 582

in plants, 368
field experiments with, 363-368
foliar retention by plants, 291-293
isotopic differences in fallout vs. reprocessing

plant sources, 382
from nuclear weapons fabrication plants,

89-90
from nuclear weapons tests, 54

2 3 8 py/2 3 9 ,2 4 0 py j^^Jq^^ ^^ g^J^j^J ^^^^

sediments, 551-554
in Enewetak sediments, 586-588
in/on nonrespirable particles at Rocky
Flats, 283
released to atmosphere at Savannah River

plant, 382-384
released to White Oak Creek, 382-384
resuspension of, 247-250
in Rocky Flats biota, 431-433
in soil depth profiles at Rocky Flats, 425-426
from spacecraft power systems, 54, 83-85
uptake by plants, 363-365

effect of soil concentration on, 365
variation in soil, 166-172
in wheat and corn at Savannah River plant,
388-391
concentration ratio (CR), 388-391
field grown compared with glasshouse
grown, 388-391
Plutonium-238, -239, and -240
in biota at Los Alamos, 407-414
in effluent ponds at Hanford, 612-613
in soils at Los Alamos, 409-414
Plutonium-239
airborne, at Hanford, 245-260
resuspension factors, 268-269

126

INDEX

at Rocky Flats, 242-260
concentration ratio (CR), 363-368
distribution in plants, 365
in particles at Savannah River, 107-144
transport and dose estimation model, 459-

508
uptake by plants, 363-368
variation in soil, 166-172
Plutonium-239, -240
in arctic ecosystems, 441-456
concentration ratio (CR) in Great Lakes,

616-617
concentrations in agricultural crops, 390-394
concentrations in Bikini Atoll waters, 560-

561
concentrations in Nevada Test Site soils, 46 1-

462
distribution in Enewetak sediments, 583-589
distribution in plants, 365
in Enewetak coral growth increments, 597-

600
in Enewetak sediments, 594
in Enewetak zooplankton, 596-597
field experiments with, 363-368
fractionation between Bikini Atoll sediments,

551-559
in Hudson River and estuary, 685-688
from nuclear weapons tests, 88-89
'""Pu/^^Pu ratios at Enewetak, 582-583
released to environs of Savannah River plant,

381-384
released to White Oak Creek, 384-385
in Rocky Flats aquatic systems, 650-656
in Rocky Flats biota, 431-437
in Rocky Flats ponds, 652
in Savannah River water, 606-607
in soil depth profiles at Rocky Flats, 425-430
from spacecraft power systems, 83
uptake by plants, 344-346, 363-368, 390-394

effect of soil concentration on, 365
Plutonium-241
2 4 1 py/2 3 9 +2 4 0 p^j ^^^j^^ -^^ Encwctak water

samples, 583
as a source of ^^ ' Am in nuclear weapons
debris, 90
Plutonium particles, formation in nuclear tests,
87-88

Radiation dose rates
to humans from Savannah River biota, 609-

610
to marine organisms, from natural radio-
nuclides, 716
from transuranic elements, 716
Radiation effects on marine organisms, 716-
718
compared to other mortality, 718-719

comparison of plutonium and ^ ' " Po, 7 1 6

research needs, 719
Radiological assessments, 45-51

requirements for, 46-49

standards for plutonium in soils, 161-162
Ratios

Am/

2 3 9,240

Pu, in Enewetak sedi-

ments, 583

in marine waters, 5 33
concentration (CR), 26-33, 189-207
concentration ratios and inventory ratios

of plutonium in Oak Ridge and Los

Alamos, 377-379
inventory (IR), 4-7, 184-207

comparison by profile analysis, 203
mathematical considerations in use of, 187-

207
of ""-^ ^°Pu/' 2^ Cs in fallout, 68-70
Qf 2 3 9 ,2 4 0 p^j/9 0 g^ ^^ Q^^^^ ^akcs, 659-660

of transuranic elements in airborne vs.

surface-soil solids at Rocky Flats and

Hanford, 283
types of, 187

Remobilization of transuranic elements

in Enewetak Atoll environment, 593-599

in Great Lakes sediments, 673-674
Research needs

in aquatic studies, 622

in ecological effects, 709-710

general, 37

in marine studies, 535-536

in plutonium availability to native animals,
417

in plutonium resuspension. 231, 250, 280-
281

radiation (pollutant) effects, 709-710. 719

in sampling for models, 182

soil-plant chemistry, 330-331

in transport and dose modeling, 462-463
Residence time

of nuclear weapons test debris in atmo-
sphere, 62-66

oj- 2 3 9 ,2 4 0 py j,^ Q^^^^j Lakes, 664

Resuspension

definition, 209-210, 237-238

mechanical, 224-226

of plutonium from Great Lakes sediments,
674-678

of plutonium by wind at Trinity Site, 416

rates, 213, 278
factors important in, 284

wind, 213-224
Resuspension factors

used in deciduous forest model, 515

definition and range, 210-212, 237-238, 268

at Hanford, 268

at Nevada Test Site, 467-468

INDEX

121

Retention of particles

factors intluencing, 289

half-times, 290

by plants, 289-294
Risk estimates of ecological hazard, 708-709
Rocky Flats

plutoniiim in aquatic systems of, 644-658

Plutonium in sediments of, 650

Plutonium in soils, 154-155

2 3 8 py/2 3 9 ,2 4 0 p^j jsotopj^, ^.^^^^^ }„ SOllS

and air, 253-260
^^'Pu contamination, 705-706

resuspension of, 241-245
Routes of exposure
ingestion, 692-693
inhalation, 692

Samphng designs
analytical sampling, 1 77-1 78
Battelle large volume water sampler used at

Bikini Atoll, 562-569
descriptive (survey), 176-177
descriptive sampling, 1 76
estimate of sampling error, 167
vs. experimental design, 1 74-178
importance in radiological assessment, 51
for modeling, 1 78
random vs. systematic, 183
relationship of concentration : aliquot size

in, 194-196
spatial pattern, 177
stratification in, 179

Savannah River Plant, 382-384

Seasonal cycles of ^ ' ' -^ ^ " Pu in Lake Michigan,
668-680

Sediments
Plutonium in Great Lakes, 670-680
Plutonium in Hanford U-Pond. 634-637
Plutonium in Hudson River, 685-687
Plutonium in Rocky Flats, 650
size fraction at Enewetak Atoll, 589
association of transuranics uitli, 589-593

Soil
airborne, concentrations of ^ ■' •'* Pu on, 250-

253
behavior of plutonium in, 13
cliemical reactions of - " ' Am in, 309
chemical reactions of " "Cm in, 309
cliemical reactions of plutonium in, 303-309
complexation of transuranic elements in,

309-312
extraction of plut'miuni from, 159, 398
horizontal and vertical distribution of
2 3«.2.9..4op^jj^ Los Alamos, 408

iniplicalions of particle size, 161-162
ingestion b\' cattle, 481-482
movement in resuspensiiin, 213-232

plutonium concentration in, effects of culti-
vation on, 395-398
Enewetak AtoU, 588
as function of particle size, 160-162, 250,

429
Rocky Flats, 250,425-430
plutonium in/on airborne vs. surface, 280-

281
plutonium index, 229
plutonium transport by, 24-26, 414-416,
429-430
Soil factors, effect on transuranic nuclide

uptake by plants, 302-315, 337-357
Solubility of transuranic elements
in Bikini Atoll marine water, 568-569
factors governing, 10-12
general considerations, 9-10
microbial alteration of, 312-315
Source terms
'"'Am, 90
classification according to solubility in soil,

302-303
decommissioning of nuclear power plants,

103-106
Enewetak Atoll, 579-581
general considerations, 2-3, 404-408, 605
nuclear fuel reprocessing plants, 89-90,

92-103, 241-245, 382-384, 605, 646-647
nuclear weapons accidents, 89
nuclear weapons fabrication plants, 89-90,

241, 404-406, 420-439, 626, 646-647
nuclear weapons tests, 54-78, 266, 542-544
^•^«Pu, 83
'-''•""Pu, 83

spacecraft power systems, 83-85
Stability constants, 10-12
Statistics in sampling
aliquots. 182-183
counting statistics, 182
multivariate vs. univariate, 200-203
random vs. systematic, 183
ratios, constraints on, 197-200
"zero" or nondetectable values, 182

Transfer of transuranics
effect of complexation in aquatic systems,

617-618
effect of complexation on, 326-330
between ecosystem model compartments,

461-463
to man from aquatic ecosystems, 621-622
through Rocky Flats aquatic systems. 654-

656
from soil to plants, 159
from water to sediment, 652

Transformation of transuranic elements b)'
microorganisms in soil, 319-3 26

128>

INDEX

Transport
of Plutonium in Hudson River, 688-689
of Plutonium in Savannah River, 607
of Plutonium by soil, 24-26, 414-416,429-

430
of transuranic elements by biota in marine

systems, 532-536
of transuranic elements by water, 26-27
Transport ratios (TR) for roots and seeds of

plants, 296-297
Trophic Transfer Factor (TTF)
in food chains, 620

used in place of concentration ratio (CR),
620

Uptake
of transuranic elements by animals, 24-33

of transuranic elements by plants, 28-29,
363-365,474475
Uranium-235,-237, -238
in Chinese nuclear test debris, 78
in nuclear devices at Bikini, 78, 542

Variability of transuranic elements in soil

sampling, 167-171, 191-193
Variation, sources of, in soil sampling, 165-166,

191-193

Weathering (ageing) of transuranic elements
effects on assimilation by plants, 399
half-Ufe of airborne radionuclides, 278
removal from plant surfaces, 471-472

White Oak Lake, behavior of plutonium in
biota of, 617

DISCLAIMER

This book was prepared as an account of work sponsored by an agency of the
United States Governnnent. Neither the United States Government nor any agency
thereof, nor any of their employees, makes any warranty, express or implied, or
assumes any legal liability or responsibility for the accuracy, completeness, or
usefulness of any information, apparatus, product, or process disclosed, or
represents that its use would not infringe privately owned rights. Reference herein
to any specific commercial product, process, or service by trade name, trademark,
manufacturer, or otherwise, does not necessarily constitute or imply its
endorsement, recommendation, or favoring by the United States Government or
any agency thereof. The views and opinions of authors expressed herein do not
necessarily state or reflect those of the United States Government or any agency
thereof.

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Full text of "Transuranic elements in the environment : a summary of environmental research on transuranium radionuclides funded by the U.S. Department of Energy through calendar year 1979"

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