^EPA
United States
Environmental Protection
Agency
Environmental Research
Laboratory
Duluth MN 55804
EPA-600/9-80-034
July 1980
V
Research and Development
Proceedings of the
Third USA-USSR
Symposium on the
Effects of Pollutants
Upon Aquatic
Ecosystems
PB 80-2246S7
W H 0 1
DOCUMENT
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Technical Informa-
EPA-600/ 9-80-034
July 1980
PROCEEDINGS OF THE THIRD USA-USSR SYMPOSIUM
ON THE EFFECTS OF POLLUTANTS UPON AQUATIC ECOSYSTEMS
Theoretical Aspects of Aquatic Toxicology
July 2-6, 1979
Borok, Jaroslavl Oblast
USSR
WHO/
DOCUMENT
COLLECTION
Edited by
Way land R. Swain
and
Virginia R. Shannon
ENVIRONMENTAL RESEARCH LABORATORY-DULUTH
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
DISCLAIMER
This report has been reviewed by the Environmental Research Laboratory-
Duluth, U.S. Environmental Protection Agency, and approved for publication.
Mention of trade names or commercial products does not constitute endorse-
ment or recommendation for use.
n
FOREWORD
These Proceedings result from the third symposium held by Project 02.02-
13 under the aegis of the US-USSR Joint Agreement in the Field of Environ-
mental Protection, established in May, 1972.
Both broad review and narrowly specific papers were presented by parti-
cipants from both countries in an effort to continue the joint procedural,
technological and methodological exchange and familiarization begun at the
two preceeding symposia in 1975 and 1976. Learning does not occur de novo
and subsequent understanding and application must be based on a foundation
of fact. The atmosphere of mutual interest, candor and respect which sur-
rounded this symposium enabled another series of steps in the learning pro-
cess. Perhaps the philosphy underlying this symposium, and the project it-
self is best expressed by an old saying, which transliterated from the
Russian approximates: Vyek zhee-vee, Vyek oo-chee. Live a lifetime, learn
a lifetime.
Norbert Jaworski , Ph.D
Director
Environmental Research Laboratory-
Duluth
m
PREFACE
This volume contains the papers presented at the Third US-USSR Symposium
on the Effects of Pollutants on Aquatic Ecosystems entitled, "Theoretical
Aspects of Aquatic Toxicology". All of the papers were presented in English
or Russian with simultaneous translations into the corresponding language at
Borok, Jaroslval Oblast, USSR during July 2-6, 1979, at the Institute for
the Biology of Inland Waters of the USSR Academy of Sciences.
Professor N.V. Butorin, Director of the Institute and Project Leader for
the Soviet side, served as official host for the American delegation and has
assumed the responsibility for the publication of these proceedings in the
Russian language. This joint bilingual publication represents a reaffirma-
tion of the continuing commitment pledged by both countries to cooperative
environmental activities.
TV
INTRODUCTION
The Joint US-USSR Agreement on Cooperation in the Field of Environmental
Protection was established in May of 1972. These proceedings result from
one of the projects. Project 02.02-13, Effects of Pollutants Upon Aquatic
Ecosystems and Permissible Levels of Pollution.
As knowledge related to fate and transport of pollutants has grown, it
has become increasingly apparent that local and even national approaches to
solving pollution problems are insufficient. Not only are the problems
themselves frequently international, but an understanding of alternate
methodological approaches to the problem can avoid needless duplication of
efforts. This expansion of interest from local and national represents a
logical and natural maturation from the provincial to a global concern for
the environment.
In general, mankind is faced with very similar environmental problems
regardless of the national of political boundaries which we have erected.
While the problems may vary slightly in type or degree, the fundamental and
underlying factors are remarkably similar. It is not surprising, therefore,
that the interests and concerns of environmental scientists the world over
are also quite similar. In this larger sense, we are our brother's brother,
and have the ability to understand our fellowman and his dilemma, if we but
take the trouble to do so. It is this singular idea of concerned scientists
exchanging views with colleagues that provides the basic strength for this
project. While our methods may vary, our goals are identical, and therein
lies the value of such a cooperative effort.
Wayland R. Swain, Ph.D., and
Richard A. Schoettger, Ph.D.
Co-Project Leaders, U.S. Side
CONTENTS
Foreword iii
Preface iv
Introduction v
Figures ix
Tables xiv
Acknowledgment xvii
1. A Research Strategy for Anticipating Contaminant Threats
to Aquatic Resources
Richard A. Schoettger and J. Larry Ludke 1
2. Principles of Estimation of Normal and Pathologic States
of Reservoirs with Chemical Pollution
N.S. Stroganov 18
3. Theoretical Aspects of the "Normalcy and Pathology"
Problem in Aquatic Ecotoxicology
L.P. Braginsky 34
4. Trends in Aquatic Toxicology in the United States:
A Perspective
Foster L. Mayer, Jr., Paul M. Mehrle, Jr. and
Richard A. Schoettger 44
5. Comparison of Principles of Development and Use of Water
Quality Standards in the USSR and USA
L.A. Lesnikov 60
6. Chlorinated Hydrocarbons as a Limiting Factor in the
Reproduction of Lake Trout in Lake Michigan
Wayne A. Willford 75
7. Organophosphorus Pesticides and Their Hazards to Aquatic
Animals
V.I. Kozlovskaya and B.A. Flerov 84
8. Monitoring Contaminant Residues in Freshwater Fishes in
the United States: The National Pesticide Monitoring
Program
J. Larry Ludke and C.J. Schmitt 97
vn
9. Accumulation and Metabolism of Persistent Pesticides in
Freshwater Fish
F.Ya. Komarovskiy and A.Ya. Malyarevskaya Ill
10. Some Factors Affecting the Toxicity of Ammonia to Fishes
Robert V. Thurston 118
11. The Prediction of the Effects of Pollutants on Aquatic
Organisms Based on the Data of Acute Toxicity Experiments
O.F. Filenko and E.F. Isakova 138
12. Age Specifics of Sensitivity and Resistance of Fish to
Organic and Inorganic Poisons
V.I. Lukyanenko 156
13. Synergistic Effects of Phosphorus and Heavy Metals
Loadings on Great Lakes Phytoplankton
E.F. Stoermer, L. Sicko-Goad and D. Lazinsky 171
14. Reversibility of Intoxication and Factors Governing It
I.V. Pomozovskaya 187
15. Aspects of the Interaction Between Benthos and
Sediments in the North American Great Lakes and
Effects of Toxicant Exposures
John A. Robbins 202
16. Recent Advances in the Study of Nitrite Toxicity to
Fishes
Rosemarie C. Russo 227
vn 1
FIGURES
Section Page
Major steps and some sources of input for a research
approach to assessing contaminant threats to fish
and wildlife resources 3
Laboratory evaluation of chronic effects of contaminants
on fish in a flow through diluter system 7
A comprehensive analytical schematic for the separation
and analysis of organic contaminants 8
Locations of CNFRL field stations and their associated
watershed areas of concern 10
Storm tracks which indicate where acids and metals are
deposited by precipitation in poorly buffered lakes
and streams of New England 12
One of many coal -fired power plants under construction in
the Northern Great Plains of the U.S 13
Researchers collecting water containing waste oil from a
drilling operation in Wyoming 15
Extensive clearing of irreplaceable bottomland hardwood
forests 16
2 Main functional groups in aquatic ecosystem 23
2 Surmiarized graphs of the main links in self-purifica-
tion 25
2 Relationship between degree of purification, pollution,
number of species and disturbance of aquatic eco-
system 28
2 Degradation of aquatic communities 30
4 Computerized treatment of residue data from fathead
minnows exposed to 3.7 ng/1 of Kepone 53
IX
Section Page
4 Schematic diagram of the environmental hazard evalua-
tion process 55
6 Commercial production of lake trout in Lake Michigan .... 76
6 Mortality of fry of Lake Michigan lake trout exposed to
DDE and PCBs at concentrations simulating those found
in water and plankton of Lake Michigan and at concen-
trations 5 and 25 times higher 81
7 Acetycholinesterase in nervous ganglia of molluscs with
varying resistance to Dylox 88
7 Inhibition by Dylox of acetycholinesterase in nervous
ganglia of Limnaea stagnalis and Planorbis corneas .... 89
7 Change in the activity of acetylcholinesterase in perch
brain after exposure to Dylox 91
7 Densitograms of the molecular form of acetylcholine-
sterase in carp and the snail unexposed and exposed
to Dylox 92
8 Map of the United States illustrating the National
Pesticide Monitoring Program stations where freshwater
fish are collected for routine contaminants analyses . . . 100
8 Geometric mean total DDT residues in freshwater fish,
1969-1976/77 104
8 Percent occurrences of polychlorinated biphenyl (PCB)
residues in freshwater fish, 1976/77 108
8 Occurrence of toxaphene residues exceeding 1.0 mg/kg in
freshwater fish (1976-1977) 109
10 Effect of prior ammonia acclimation on the acute toxicity
of ammonia to rainbow trout 123
10 Effect of reduced temperature on the acute toxicity of
ammonia to fathead minnows 125
10 Acute toxicity of ammonia vs. temperature for fathead
minnows 126
10 Effect of dissolved oxygen on the acute toxicity of
ammonia to fathead minnows and rainbow trout 128
Section Page
10 Effect of dissolved oxygen on the acute toxicity of
armionia to rainbow trout: LC50 vs. D.O. at 5 time
intervals 130
10 Acute toxicity of ammonia to rainbow trout: 96-hour
LC50 vs. pH 132
10 Acute toxicity of ammonia to fathead minnows: 96-hour
LC50 vs. pH 133
11 The relationship of the number of dead Daphnia magna with
time under the influence of various concentrations of
trimethyl tin chloride 140
11 Daphnia magna mortality with time as a result of expo-
sure to organic tin compounds and some other compounds . . 143
11 Daphnia magna mortality with time as a result of expo-
sure to trimethyl tin chloride in a concentration of
1 mg/1 144
n The relationship of time of death of 25 percent of
Daphnia magna with the concentration of trimethyl tin
chloride 150
11 Graphical determination of acceptable concentrations of
trimethyl tin chloride for Daphnia magna 153
13 Outline map of the southern Lake Huron showing the dis-
tribution of the eutrophication tolerant diatom
Frag il aria capucina Desm. in the waters of Lake Huron
outside Saginaw Bay in early June 1974 176
13 Transmission electron micrograph of a cross section of
Frag il aria capucina 177
13 X-ray spectrum of a polyphosphate contained in vacuole
of Fragilaria capucina 177
13 Outline map of Saginaw Bay, Lake Huron showing the abun-
dance populations containing polyphosphate bodies in
different segment of the bay 178
13 Transmission electron micrograph of Anacystis containing
large polyphosphate bodies 180
13 Transmission electron micrograph of Scenedesmus sp.
showing large polyphosphate bodies in the vacuole .... 180
XI
Section Page
Light micrograph of Scenedesmus sp. stained for poly-
phosphates by the technique of Ebel et al- (1958) .... 180
Light micrograph of Fragilaria crotonensis Kitton
stained for polyphosphate by the technique of Ebel
et ai. (1958) 180
Transmission electron micrograph of cytologically normal
Plectonema boryanum 183
Transmission electron micrograph of Plectonema boryanum
treated with 0.1 yg - at/S, Pb 183
Transmission electron micrograph of Plectonema boryanum
treated with 0.1 yg - at/Jl Zn 183
The dynamics of the survival rate of salmon larvae 192
The dynamics of the survival rate of roach 195
Distribution of benthos and cesium-137 in a core from
Lake Erie 204
The radiotracer scanning system 206
The actual and measured distribution of activity from a
submillimeter line source 207
15 Effect of tubificid worms on the distribution of cesium-
137 208
13
13
13
13
13
14
14
15
15
15
15 Location of the peak activity versus time
209
15 Effect of amphipods (Pontoporeia hoyi) on the distribu-
tion of cesium-137 210
15 Time-dependence of the optics-corrected activity profile
width 211
15 Effect of adding very high levels of NaCI on the rate of
sediment reworking by the Oligochaete worm, Limnodrilus
hoffmeisteri 213
15 Response of the sediment reworking rate to additions of
sulfate (Na2S04) for two species of Oligochaete worms . . 214
15 Activity of cesium-137 and sodium-22 in a control cell
and in a cell with tubificid worms after an elapsed
time of about 200 hours 216
xn
Section Page
15 Concentration of soluble reactive silicon in water over-
lying sediments stored without disturbance in a core
liner collected from Saginaw Bay, Lake Huron 217
15 Relationship between the flux of Si from sediments and
the density of Chironomid larvae in a series of repli-
cate cores taken from Saginaw Bay, Lake Huron, on two
separate cruises in 1978 220
15 Flux of dissolved silicon from a sediment core collected
from northern Lake Huron before and after exposure to
a sterilizing dose of gamma radiation 222
16 LC50 vs. average fish weight for nitrite bioassays on
rainbow trout (Salmo gairdneri) 232
16 LC50 vs. average fish length for nitrite bioassays on
rainbow trout (Salmo gairdneri) 233
16 Toxicity curves showing effect of chloride on nitrite
toxicity to rainbow trout (Salmo gairdneri) 234
16 Effect of chloride on nitrite toxicity to rainbow trout
(Salmo gairdneri) 235
16 LC50 (as NO2-N) vs. pH 237
16 LC50 (as HNO2-N) vs. pH 238
xm
TABLES
Section Page
4 Maximum Acceptable Toxicant Concentrations (MATC) From
Partial and Complete Life-Cycle Toxicity Tests with
Fish as Compared with MATC'S Derived From Embryo,
Larvae, and Early Juvenile Toxicity Tests 46
5 Relationship of LT50 (mg/liter) of Chlorophos for
Current Year's Brood of Fish as a Function of Time
of Exposure 62
5 Relative Toxicoresistance of Fresh-Water Test Organisms
Used in Toxicologic Experiments in the USSR and USA ... 63
5 Reversibility of Intoxication in Perch 64
Organophosphorus Pesticides and Their Hazardous to
Aquatic Animals 84
Persistence of Selected Organophosphorus Pesticides
in Water 85
Persistence of Selected Organic Pesticides in Soil 86
Toxicity of Organophosphorus Pesticides to Aquatic
Animals 86
Dylox Toxicity for Selected Aquatic Organisms 87
Changes in the Acetyl Cholinesterase Activity of the
Perch Brain in the Minimum Tolerable Concentrations of
Dylox with Subsequent Washing in Freshwater 90
7 Cholinesterase Activity in Perch Brain as a Result of
Periodic Additions of Dylox to the Exposure Chamber ... 93
8 National Pesticide Monitoring Program Network: A List
of Environmental Components and the Respective
Agencies Responsible for Monitoring Contaminant Trends
in Each 98
XI v
Section page
8 Freshwater Fishes Recommend for Collection for Tissue
Contaminant Residue Determinations (NPMP), Listed by
Category, Habitat and Species 101
8 Contaminant Residues Measured and Detected in NPMP
Freshwater Fish Samples, 1967 Through 1976-77 102
8 Geometric Mean Residues of Organochlorine Compounds at
74 Selected NPMP Stations, 1970-1976/77 105
8 Percentage of 74 NPMP Stations Where Detectable Residues
of Important Organochlorine Compounds Were Found, 1970-
1976/77 106
n Daphnia Magna Relationships of Percent Mortality in
Daphnia Magna, as Calculated by Various Equations,
With Duration of Experiment 141
11 The Correlation of Experimental and Calculated Relation-
ships Between Mortality and Duration of Exposure of
Daphnia Magna to Trimethyl Tin Chloride Using Various
Equations 146
11 The Date of Death of 25 Percent of Daphnia Magna Exposed
to Various Compounds as Calculated From Experimental
Studies of Varying Duration 147
11 The Relationship of the Time of Death of 25 Percent of
Daphnia Magna with Concentrations of Trimethyl Tin
Chloride Calculated by Different Functions 151
11 Acceptable Concentrations of Compounds for Survival of
Daphnia magna Calculated with Equations of Power
Function 152
13 Morphometric Results of Nutrient Treatments 182
14 Reversibility of Intoxication Caused by Effluents in
Juvenile Salmons 190
14 Reversibility of Intoxication in Juvenile Salmon During
Four Exposures to Effluents Diluted in a Ratio of 1 :1 . . 193
14 Reversibility of Intoxication in Perch Caused by
Effluents 196
14 Reversibility of Intoxication in Juvenile Salmon Caused
by Effluent From a Heat-and-Power Station 197
XV
Section Page
14 Reversibility of Intoxication in Juvenile Salmon Caused
by Effluent Water From a Heat-and-Power Station 197
14 Reversibility of Intoxication in Roach Larvae Caused by
Waste Water From Boiling Shop 198
14 Reversibility of Intoxication in Salmon Larvae Caused by
Water From Aerator-Tank 199
14 Reversibility of Intoxication in Juvenile Fish of Various
Species Caused by Undiluted Waste Water 200
15 Benthos Density and Silicon Flux: Saginaw Bay, Lake
Huron 218
15 Correlations Between Nutrient Fluxes and Organisms
Densities 219
15 Effects of Selected Treatments of Silica Release From
Sediments 223
16 Chemical Characteristics of the Dilution Water Used in
Bioassays 230
16 Acute Toxicity of Nitrite to Rainbow Trout (Salmo
Gairdneri) Under Uniform Water Chemistry Conditions ... 231
XVI
ACKNOWLEDGMENTS
In any project of the scope and complexity of this effort, the Project
Officers become increasingly indebted to a large number of individuals who
contribute their time and effort with no thought of personal gain. Unfor-
tunately, the list of persons who materially aided the effort is too exten-
sive to allow a complete discussion. However, while those persons who made
outstanding contributions to the success of this project are acknowledged
below, the editors also wish to thank all those others, both Soviet and
American, whose efforts and assistance smoothed the way to a satisfactory
completion of this phase of the project.
Sincere thanks are extended for the considerable efforts, patience and
support of Gary Waxmonsky and Jean MaGuire of the U.S. Executive Secre-
tariat of the US-USSR program. Their assistance and prompt attention to
the details of translations of texts, movement of equipment, international
cable traffic and travel clearances enabled the meetings of the U.S.
personnel with Soviet counterparts, and facilitated the preparation of this
report.
The many contributions of Ms. Nina Ivanikiw to the preparation of both
the visit to the Soviet Union and to the coordination and preparation of
materials for this publication are remembered with deep appreciation.
The substantial contributions and tireless efforts of Ms. Debra Caudill
to the preparation of these proceedings are gratefully acknowledged.
To the many Soviet colleagues, friends, and acquaintances who labored
so diligently to make the Borok symposium such a success, and the visit of
the eleven participants to Siberia and Lake Baikal so memorable, we offer
profound thanks, Bo/ibuioa CnacH5o!
xvi 1
SECTION 1
A RESEARCH STRATEGY FOR ANTICIPATING CONTAMINANT THREATS
TO AQUATIC RESOURCES
Richard A. Schoettger and J. Larry Ludke^
The Environmental Contaminant Evaluation Program of the United States
Fish and Wildlife Service (USFWS) is emphasizing a predictive approach to
identify potential contaminant problems and preventing or ameliorating ad-
verse effects of contaminants on ecological systems. The primary objective
is to protect fishery and wildlife resources from the impacts of contami-
nants before the effects become irreversible, or reversible only with great
difficulty and at high cost. Predictive research has long been a priority
objective of USFWS work with environmental contaminants. For example, DDE
was shown to cause reduction in avain populations; exposure to this chemical
resulted in thinned eggshells, which decreased the production of offspring.
Although these effects were repeatedly demonstrated in laboratory experi-
ments, regulatory action to remedy the problem was not taken for several
years.
Contaminant problems of the 1970's, however, overwhelmed the research
capability to address them, and predictive research fell behind in the midst
of pressures to solve current problems. A new thrust was initiated in 1977
to increase USFWS capability to anticipate contaminant threats to the
nation's fishery and wildlife resources. The intent of this renewed empha-
sis is to increase the base of knowledge and thus assist natural resource
managers in anticipating and addressing future or suspected contaminant
problems before they reach catastrophic proportions.
Because manpower and scientific resources are limited, we in the envi-
ronmental research community must emphasize the necessity of placing priori-
ties on our fishery and wildlife resources. We must judge on the relative
importance of different species and habitats on the basis of uniform and
meaningful guidelines, and focus our efforts on protecting the most impor-
tant ones first. Such an effort necessarily involves a multidisciplined
approach with a goal of anticipating contaminant threats of the future.
The Columbia National Fisheries Research Laboratory (CNFRL) has em-
ployed a strategy that accentuates the anticipation of new or previously un-
^Columbia National Fisheries Research Laboratory, U.S. Department of the
Interior, Fish and Wildlife Service, Route #1, Columbia, Missouri 65201,
recognized pollution problems, while continuing to address old problems that
remain a concern (Figure 1). The approach draws upon a number of different
sources to assist in the identification of present and potential contami-
nant effects. It is actually little more than application of the logic of
the scientific method. Information and data that relate to topics of con-
cern are reviewed by scientists and resource managers to develop an over-
view of a problem and to determine data needs. A research design is then
formulated to provide information on the real or potential effects a con-
taminant may have on aquatic organisms or ecosystems. From the results of
such research, we may often be able to make remedial recommendations. Cor-
rective or preventive alternatives that include one or more of the following
may then be recommended:
a) legislative action to regulate or prohibit the manufacture,
use, or disposal of a chemical,
b) modification of management techniques or practices to protect
fish or other aquatic resources from the contaminant,
c) changes in the development, use or application of certain
chemicals,
d) suggested substitute chemicals which prove less harmful,
e) selection of a less harmful activity or process over one that
is proven deleterious.
Our strategy insures that resource managers are involved in the process
of problem identification and formulation of research design, so that the
objectives and results are applicable to the actual environmental problems
that confront the aquatic resources. It also assures consideration of the
most vulnerable resources that may be impacted by a contaminant.
The key to applying this strategy successfully at the national level is
to simultaneously identify the most critical resources of concern and the
activities and contaminants most likely to adversely affect those resources.
Limited funds and manpower dictate the necessity of identifying the most
critical or vulnerable biota and habitat that may be affected by any con-
taminant or polluting activity of man. This identification requires that
we develop a comprehensive inventory of resources and habitat under our
protection. We must distinguish between localized problems and those that
are widespread. Problems of short duration (e.g., one-time occurrences) or
those which are in the process of remediation must be recognized, but re-
search emphasis must be oriented toward long-term contaminant problems that
have potentially devesting impacts in the foreseeable future.
It has been estimated that the number of potential chemical contaminants
that may pollute U.S. lakes and streams could exceed 87,000. There are 129
priority toxic substances listed by the U.S. Environmental Protection Agency
(EPA) for immediate assessment of production, distribution, disposal, toxi-
city, fate within the environment, and ecological impacts. Hundreds more of
these chemicals are awaiting ecological hazard evaluation. Though some of
Essential Research Process for Environmental Contaminant Evaluation
Public
Fish & Wildlife Managers
I
Assess Relevant Contaminant — Fish
and Wildlife Resource Interactions
Government Agencies
Research Scientists
Industry
Monitoring Program
Status of Fish & Wildlife Populations
Chemical-Physical Properties
Chemical Production and Use
I
Define Scope of Problem
and Research Tasks
Survey Current Research Activities
Literature Review
Habitat Status
Chemical Distribution and Disposal
Postulated Biological Impacts
Fisheries and Wildlife Biology
Ecology
Ethology
Microbiology
Pollution Abatement and
Regulatory Recommendations
Technological Improvements
and Methods Development
1
Design and Conduct Research
1
\
Interpretation and Application of Results
Biochemistry and Physiology
Toxicology
Statistics
Analytical Chemistry
Data Reports and
Scientific Publications
Identify Additional
Research Needs
Hazard Evaluations and Management Recommendations
Figure 1. Major steps and some sources of input for a research approach
to assessing contaminant threats to fish and wildlife resources.
the needed information is available for hazard assessment, as stewards of
the nation's biological resources, the USFWS must increase its efforts in
determining which of the many pollutants are reaching or may reach the re-
sources we are charged with protecting. To make this determination, the
Service is developing its priorities, emphasizing the resources that can
least afford to be lost. If a contaminant or polluting activity is not
likely to affect a priority resource, we need not waste valuable time and
effort in studying it.
We now have all of the components for a framework to address environ-
mental contaminant impacts on living resources. Implementation of the
approach requires that the components be placed together in a logical se-
quence to achieve proper perspective, set priorities, and then act. Con-
ceptually, we progress through a logical continuum of four steps: a) prob-
lem identification, b) definition of scope of problem, c) research to pro-
vide data or fill information gaps, and d) interpretation and application
of results.
Information elucidating potential contaminant problems that threaten the
well-being of fish and wildlife resources come from a variety of sources:
1. Resource managers - Federal and state management personnel
identify contaminant problems, often from obserservation of
mortality of fish or wildlife in the environment. Declines
in populations may be observed and reported. Through resi-
due surveys or in concert with USFWS Research monitoring ac-
tivities, "hot spots" are identified. Follow-up research
studies are initiated to elucidate the full scope and effects
of observed problems.
2. Other government agencies - The most obvious source of input
suggesting contaminants of concern comes from the EPA. Under
the Toxic Substances Control Act, EPA is charged to "regulate
conmerce and protect human health and the environment by re-
quiring testing and necessary use restrictions on certain chem-
ical substances...". The total of 129 priority compounds now
on the EPA toxic substance list for environmental hazard evalu-
ation includes the following major chemical families: chlori-
nated benzenes, chlorinated naphthalene, haloethers and halo-
methanes, nitrophenols, phthalate esters, nitrosamines, poly-
nuclear aromatic hydrocarbons, organochlorine pesticides, poly-
chlorinated biphenyls, and selected metals.
3. Research scientists - Scientists who are expert in research on
environmental contaminant effects have particularly valuable
insight regarding contaminants that require further study. Ob-
servations and results obtained through carefully planned re-
search often provide the researcher only with parts of the ans-
wer being sought. New questions arise which may be answered by
further research to provide additional insight.
4. Industry - Industry can be, and often is, a contributing parti-
cipant in identifying potential contaminants that must be
assessed before they are marketed. CNFRL has worked closely
with one chemical company in initial toxicity assessment of com-
pounds that are being considered as PCB replacements. Results
have been encouraging, and we believe this working relationship
between government and industry to be highly desirable.
There are other sources from which we get leads or indications as to the
contaminants of highest potential concern (e.g., academia, monitoring pro-
grams, conservation groups, etc.).
The important point is that there is no paucity of contaminants and con-
taminant problems. The possibilities far exceed our potential in manpower,
funds, and time to address them in detail. So it is incumbent upon us to
identify and locate the populations and habitats that are most important to
us, whether they be highly vulnerable and pristine, threatened or endan-
gered, or of sport, commercial or aesthetic value. Only by ordering our re-
sources into categories of priority can we assess the relevancy and scope of
contaminant-resource interactions, and thereby make more meaningful manage-
ment and research decisions. It does not matter whether the potential con-
taminant is an organophosphate, a dioxin, toxaphene, or crude oil. What
does matter is whether that substance will adversely affect, directly or in-
directly, a valued resource.
Traditionally, we have oriented our efforts toward studying the chemical
and its effects under highly controlled conditions. Emphasis has been on
anticipating contaminants which may have highly detrimental effects because
of their toxicity, distribution, or disposal. We are now putting greater
emphasis on assessing the resource-contaminant interaction. We want to
better consider the potential availability of the toxic contaminant to the
fish and wildlife resources that have been identified as being of high
priority.
PCBs are known pollutants of the Upper Mississippi River, and in some
areas their residues are alarmingly high. In 1971, commercial fishermen
harvested 31.5 million pounds of fish from this productive stream. The ex-
tent, distribution and ecological significance of PCB residues in prime
fish and diving duck habitats of the Upper Mississippi River have not yet
been determined. Our field laboratory at LaCrosse, Wisconsin, is under-
taking studies to describe the movement and fate of PCBs in productive
fishery and wildlife habitat downstream from a major municipal 'source.
Toxicity and bioconcentration of PCBs in aquatic biota is being studied to
assess the relative hazard of these contaminants in the environment.
Through contact with fish and wildlife management personnel, our field
research scientists are focusing on several broad areas of concern with
respect to contaminant problems. Some of the topics relate to energy,
including petroleum pollution, but numerous non-energy related contaminant
poblems also require attention.
Ongoing work at CNFRL includes considerable effort in continued acute
and chronic toxicity testing (Figure 2), monitoring and surveillance of con-
taminants in the environment, and continued methods development in analyti-
cal chemistry to better enable us to identify and quantitate a wide spectrum
of contaminants in the environment. We are placing additional emphasis on
ecosystem approaches, behavior studies, highly sophisticated analytical ap-
proaches to identify unknown contaminants in the environment, and assessment
of biological or biochemical indicators of contaminant stress.
Contamination of the aquatic environment by agricultural and industrial
chemicals, oil spills, mine effluents, and other forms of pollution has been
recognized for many years. Evaluating the impact of the many contaminants
on aquatic organisms has been limited mainly to short-term laboratory
studies. Only recently have long-term laboratory studies been used to
evaluate growth, reproduction, mortality and residue dynamics in relation to
the environment. Although these studies strongly indicate safe toxicant
concentrations, their disadvantages include the length of time required to
complete partial and chronic toxicity studies, cost, and the limited number
of aquatic species that can be cultured in laboratory or artificial environ-
ments. Much of the laboratory research lacks field verification, and the
true impact of chemical contaminants on aquatic organisms in the natural
environment is poorly understood. New techniques are needed that can be
used as biological indicators or predictors in both laboratory and field in-
vestigations for estimating the health or status of a particular resource.
Development and validation of analytical capabilities must accompany
laboratory studies dealing with the toxicological effects of contaminants.
New analytical procedures have been implemented for di-2-ethylhexyl phtha-
late, pentachlorophenol , mirex, and Kepone in water and fish, and for mixed
Arochlors (PCBs) in sediments from the Upper Mississippi River. The use of
adsorbents has greatly facilitated the analysis of certain trace organics
and led to the development of a new multichromatographic material that may
permit one-step purification of many aromatic compounds, including dioxins
and dibenzofurans.
Routine methods currently used in monitoring and surveillance programs
enable us to measure fewer than 50 kinds of residues in fish. Thus, it is
essential to develop a comprehensive strategy to detect and measure contami-
nants in fish and other sample material. Recent advances in chemical detec-
tion, sample extraction, and clean-up procedures make it possible to iden-
tify and quantitate a greater number of the components that make up the com-
plex contaminants in aquatic systems.
Techniques are under development to fractionate complex mixtures of con-
taminants present in samples from aquatic environments into classes of
chemicals to simplify the detection and to provide more comprehensive resi-
due data (Figure 3). By using advanced scientific instruments, such as the
mass spectrometers and the inductively coupled plasma emission spectrophoto-
meter, we are gaining the ability to perform comprehensive analyses with
much greater precision and accuracy. Separations of contaminants into
classes, combined with new instrumentation, have helped identify several
Figure 2. Laboratory evaluation of chronic effects of contaminants
on fish in a flow through diluter system.
Comprehensive Scheme for Cleanup, Fractionation,
and Analysis of Environmental Contaminants
itCH, Reverst f~
Pesticides
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Epoxides
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Figure 3. A comprehensive analytical schematic for the separation
and analysis of organic contaminants.
previously unknown contaminants. Once contaminants are identified, needed
toxicity data can be gathered to assess their impact on resources.
We have recently added to our professional staff eight fishery biolo-
gists who are located in major watershed regions of the United States (Fi-
gure 4). These scientists are working with toxicologists at our laboratory
and with federal and state fishery and wildlife resource managers to iden-
tify present contaminant problems and potential contaminant threats of the
future. The field biologists have been working to place contaminant prob-
lems of the present and future into perspective for planning and accomplish-
ing research needed to assess contaminant hazards to natural resources.
Contaminant problems associated with new or intensified activities of the
future are undoubtedly numerous.
Many possible threats exist to wildlife and fish from activities in
energy development. Although many of the activities are not new, their pro-
jected intensity is far greater than once expected. We have much to learn
about the impacts of these activities on the environment.
The development, transport, or use of gas, coal, oil and oil shale could
have substantial impact on the environment, particularly in the western
United States where ecosystems have a low resiliency to ecological perturba-
tion. Any material present in the crude energy source or used in the con-
version to usable energy is a potential pollutant. Projected coal gasifica-
tion and liquefaction plants and oil shale retorting facilities of the
1980's will result in a new area of contaminants associated with energy pro-
duction. At this point, we can speculate on the identity of some of these
potential contaminants, on the basis of existing technology in the analysis
of crude oil and the by-products of conventional coal combustion. Toxic
phenols, cresols, and water-soluble aromatics are high on the list of po-
tential troublemakers. Certain aromatics of higher molecular weight (e.g.,
benzo-pyrene, benzanthracene, and naphthalene) are known carcinogens. A new
generation of organometallics will be associated with coal conversion.
During exploratory drilling and production at petroleum wells, large
amounts of water must be disposed of. In addition to metallic salts, the
water contains numerous organic compounds derived from underlying petro-
leum pools. Much of this waste water is being dumped into freshwater
streams and estuaries.
The "shopping list" of contaminant problems associated with energy is
extensive. The Columbia National Fisheries Research Laboratory has ini-
tiated research in energy-related subjects that have been identified as
being of high priority.
In many parts of the world, precipitation is becoming polluted with
strong acids, trace elements, and complex organics. The major sources of
these contaminants appear to be combustion of fossil fuels. Trace elements
and organic compounds have not been routinely sampled in the past. However,
some 450 organic contaminants including PCBs, DDT, polycylic aromatic hydro-
carbons, and others, have been detected in precipitation.
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Prevailing weather patterns are such that the northeastern U.S. is sub-
ject to extensive fallout of acid and metals in precipitation (Figure 5).
Most of the acid apparently originates over the industrial Midwest. Trace
elements are higher in precipitation in the Northeast and Midwest or West
than elsewhere. Halogens, mercury, selenium, arsenic and antimony are vola-
tilized during coal combustion and many of the organic compounds identified
in precipitation are the same as those found in some fuels.
Direct addition of acid from precipitation has caused a marked decline
in pH of lakes and streams in Scandanavia; Ontario, Canada; and the
Adirondak Mountains of New York. In many lakes in the Adirondaks, where the
water is poorly buffered, pH ranged from pH 6.0-7.5 in the 1930' s, but is
commonly less than 5.0 today. Lowered pH renders most heavy metals more
soluble and potentially more toxic to aquatic biota. Concentrations of mer-
cury, copper, cadmium, nickel, lead and zinc have been shown to be higher in
lakes affected by polluted precipitation than in others. Lowered pH also
promotes increased leaching of naturally occurring metals (e.g., aluminum)
from soils.
Surveys of lakes indicate that fish populations are virtually absent in
waters with a pH below 5.5. Recent evidence indicates that lowland lakes
are decreasing in buffering capacity and small headwater streams may be af-
fected, particularly during spring melts.
There is a critical need for more information about the extent and dis-
tribution of polluted precipitation and its effects on lakes and streams.
There is currently a lack of information on the chemistry and fish popula-
tions of vulnerable lakes in New England. The CNFRL field research unit at
Orono, Maine, is beginning a study to correlate the pH, and metal content of
lakes believed to be impacted in the northeastern United States. Diatom
analysis will be used to document the history of pH changes. Fish popula-
tions will be surveyed for species composition and age distribution. Fish
will be subjected to analysis for aluminum, arsenic, cadmium, copper, lead,
silver, zinc, antimony and mercury.
Our objectives are (a) to determine recent history of pH and metal con-
tent of selected New England lakes, (b) to determine the chronology of fish
population changes, (c) to correlate the heavy metal content with acid pol-
luted lakes, and (d) to determine water quality changes in headwater streams
in northern New England at spring thaw.
The United States has vast coal reserves in the West. Most of the re-
serves used over the next 20 years will be taken by surface mining. Some of
it will be transported to points throughout the country, where it will be
converted to usable energy. However, much of it will be converted to elec-
tric power at coal-fired power plants near mining sites, and the electri-
city transported to the user (Figure 6). The energy output of coal -fired
facilities in Montana, Wyoming and the Dakotas will increase almost three-
fold between 1977 and 1985. The distribution and effects of airborne con-
taminants on aquatic and terrestrial systems are largely unkown. Questions
that need to be answered include such items as the manner and degree that
trace inorganics and organics cycle in the environment; the kinds of trans-
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Formations elements undergo as they cycle from air into water and biota; and
the availability and toxicity of the trace contaminants that do penetrate to
the aquatic system.
The Field Research Station at Victoria, Texas, in cooperation with the
Texas Parks and Wildlife Department, conducted acute toxicity tests of oil-
produced brine water to several estuarine fishes. Brine water from oil
wells located near coastal areas of Texas are generally discharged into es-
tuaries. An increase in the concentration of brine was followed by an in-
crease in death rates of test organisms. Organisms tested in synthetic sea
salt at the same salinity as the brine concentration showed a much lower
death rate. Evidently some toxic component of the oil is dissolved in the
brine, or the brine is interacting with the oil to increase toxicity.
Further research at Victoria will include testing the effects of oil-pro-
duced brine water to standing crops and diversity of stream organisms. In-
creased salinity in Oklahoma streams has been traced to improperly capped
wells and faulty injection casings; field research is planned to assess the
impact of the increased salinity.
The pressures of oil shortages and deregulation of oil prices will re-
sult in additional exploration and development of new oil reserves and in-
creased production from existing ones. Public lands in the mountainous
areas of the western U.S. have been targeted as sites for new production.
In active oil fields, large volumes of water are produced with crude oil.
Water is separated from the oil and then reused or discharged. The limit of
"oil and grease" discharge allowable is 10 parts per million (ppm) (Figure
7). No information has been generated to allow a proper hazard evaluation
of these tolerated levels.
The CNFRL Field Research Laboratory at Jackson, Wyoming, conducted 90-
day exposures of cutthroat trout to water soluble components of Wyoming
Green, one of the major crude oil types produced in that area. At test
concentrations of 0.5 ppm (less than one-tenth the allowable effluent con-
centration) trout mortality was 48% and growth was reduced by 88%. Growth
of trout treated with as little as 0.1 ppm was reduced by 20%, and extensive
fin erosion occurred. Avoidance studies have demonstrated that cutthroat
trout are attracted to oil concentrations in water that also result in re-
duced growth and survival.
Numerous other contaminant threats to important aquatic resources have
been identified. Some problems are of concern because they are ubiquitous,
whereas others may affect specific isolated resources that are highly valued
and especially vulnerable to contaminant stresses.
Millions of acres of riparian habitat have been degraded or destroyed by
water resource projects over the past 50 years (Figure 8). Much of the des-
truction results from restriction of annual overflows of natural wetland
areas. Overflow restriction has encouraged extensive land clearing and dis-
rupted the normal hydrologic regime and water fluctuations in headwaters and
backwater lakes and swamps. Flood control practices have destroyed hardwood
forests and degraded once productive aquatic habitats, allowing these areas
to be cleared and used for agriculture. Sediments and associated contami-
14
Figure 7. Researchers collecting water containing waste oil from a
drilling operation in Wyoming. This discharge has been
shown to be toxic to cutthroat trout.
15
16
nants further degrade lakes that become surrounded by agricultural land.
Even systems receiving annual overflow are being degraded by agricultural
pollutants stemming from land-use activities. Though the literature is re-
plete with qualitative information expounding the value of wetland systems,
there is a paucity of quantitative information describing the effects of re-
duced overflow and contaminant effects on these ecosystems. Such informa-
tion is needed to verify and document the effects of flood control activi-
ties (damming, channelization, diking, levee construction, etc.) and agri-
cultural chemical impacts resulting from land use changes.
Other environmental contaminant problems of potentially serious conse-
quence include the following:
a) Impact of contaminants in irrigation return waters on the
anadromous fishes of the San Joaquin and Sacramento Rivers
in the Central Valley of California;
b) Widespread toxaphene contamination of freshwater fisheries
from increased use and atmospheric transport of the chemical;
c) Extensive use of herbicides in agriculture and silviculture;
d) Accumulation and chronic toxic effects of relatively unstudied
industrial contaminants;
e) Continuing contamination of the environment by PCBs, dibenzo-
furans, and dioxins.
The proper evaluation of contaminant impacts of living resources in-
volves a multidisciplined approach with input from scientisits, resource
managers, industry, and academia. Matching the locations of more serious
contaminant problems with areas of high resource value can serve as a guide-
line for directing limited research resources to properly assess contaminant
threats or hazards to the environment. Researchers and resource managers
can then work together to recommend approaches to identify and avoid or mit-
igate serious contaminant impacts on the environment.
17
SECTION 2
PRINCIPLES OF ESTIMATION OF NORMAL AND PATHOLOGIC STATES
OF RESERVOIRS WITH CHEMICAL POLLUTION
N.S. Stroganov^
A need has been demonstrated for giving hydrobiologic principles
priority over other principles in the evaluation of the status of a reser-
voir. The starting point for development of principles for evaluation is
the need to preserve pure water in the reservoir, in which valuable commer-
cial organisms can exist for long periods of time, and for fresh reservoirs,
suitable also for supplying potable water. A reservoir which has water of
this quality can be considered normal, one which does not have these
qualities must be considered pathologic. Unless man's use of the water is
brought into the picture, there is no foundation for speaking of the degree
of normality of reservoirs.
The degree of pathology may differ. Selection of the species of aquatic
organisms to be protected by man will be determined primarily by the func-
tional significance of the species in the cycle of matter in the aquatic
ecosystem, assuring good water quality and high productivity of valuable
commercial species.
For water toxicology, theoretically, scientific determination of the
limits of permissible changes in hydrobiologic processes in an organism is
of great importance.
The increase in man's effect on nature (Bernadskiy 1967), including sur-
face reservoirs and streams, has set for mankind a number of new problems
which must be solved as quickly as possible. Man began influencing nature
long ago. Ecologic crises have occurred in the past (Budyko 1977), but they
have become particularly striking in certain regions since the 1940s. The
situation has deteriorated to the point that the outlook of many toward the
relationship of man and nature is quite pessimistic. We hear predictions of
ecologic catastrophes (Douglas 1975), and various plans are set forth to
avoid such catastrophes (Medouz, et al^. 1972), and thus, the ecologic crises
are denied for the present time (Budyko 1977). The disruption of equili-
brium between man and nature is real. While it should not be drawn in emo-
tional terms, there are rational means for solution of the problem. Probably
the greatest of all problems with which society has ever wrestled (Oldak
^Moscow State University, Biology Faculty, Lenin Hills, B-234 Moscow, USSR.
18
1979), must be addressed. Degradation of the environment and the advent of
the ecologic catastrophe must be prevented. The biosphere is a single,
integral system (Bernadskiy 1967).
The surface waters of rivers, lakes, reservoirs, seas and oceans receive
tremendous quantities of various chemical compounds today, for which no pre-
cise accounting can be made. Apparently, there are several thousand such
substances, and each year increasing numbers of substances are dumped, cre-
ating chemical pollution of the environment. The powerful inflow of pollu-
tants changes the environment of aquatic organisms, as a result of which
the quality of water decreases and the biologic productivity of commercial
organisms is reduced. It is quite obvious that mankind cannot simply con-
tinue polluting his waters unchecked, but it is also impossible to exclude
reservoirs and streams from the circle of human economic activity. The only
proper path for establishment of the interrelationship of society with
nature is efficient utilization of nature, designed to continue over many
years. We must not simply protect or simply utilize without control the
waters of surface reservoirs and streams, but rather we must utilize them
efficiently and in a multiple use fashion, i.e., by many water users. In
connection with these new tasks, the need arises to develop principles for
estimation of water quality in reservoirs and evaluation of their normal
state.
All reservoirs and streams undergo changes over a period of years in ac-
cordance with changes in climate, geologic-geographic variation and other
changes, not related to the effects of human factors. Therefore, we must
develop criteria which can be used to maintain reservoirs and streams in a
state satisfying the needs of man. If man is not considered, any body of
water is in its normal state, i.e., it corresponds to the surrounding con-
ditions. Only man, based on his own needs, makes an evaluation as to
whether the reservoir is in a normal or pathologic state. The time has
come for regulated interrelationships between human society and nature.
The need has arisen to develop principles and standards for estimating the
quality of reservoirs, establishing limits of permissible changes in water
quality and, finally, formulating requirements for man - that which he must
not do with natural water.
Noted elsewhere (Stroganov 1977), in a work on the concepts of the norm
and pathology in water toxicology, is a new approach to the solution of the
problem at hand. Hydrobiologists cannot limit themselves to a simple
description of what occurs in a reservoir following chemical pollution. An
"engineering" method of thinking is required, i.e., we must first formulate
how the body of water should be, then how this end can be achieved.
In order to formulate how a body of water should be, we must select
principles, in accordance with which we can develop the necessary water
quality indexes.
Based on the historic relationships between the abiotic medium of reser-
voirs and the hydrobiologic processes occurring in them, to which man has
now been added, several principles can be formulated. These principles
must lie at the base of the development of standards regulating the quality
19
of water in reservoirs. It seems that theoretical problems of water
toxicology should be solved in the aspect of development of principles.
In estimating the qualitative state of a reservoir, one can obtain
varying answers, depending on our requirements, i.e., the initial stand-
point. Among the many water users, the highest demands for water quality
are those of but two: fishermen and those who drink the water. Therefore,
all of the questions which are stated can be answered in terms of satis-
faction in the reservoir of the condition of high productivity of commercial
species and good quality of drinking water. If these standards are met, we
must call this body of water a normal one; if they are not met, it must be
considered an anomalous or even pathologic body of water. This last term is
used by hydrobiologists, although it is not really quite applicable to
bodies of water.
As the economy becomes increasingly industrialized and "chemicalized", a
situation arises in which the need for fresh water of good quality increases
greatly, both for various branches of the economy and for water supply for
the population. However, the quality of fresh water is continually reduced,
a situation which has led to great difficulties in water supply.
The Soviet Union has tremendous reserves of fresh water, but their dis-
tribution does not correspond to the needs of the regions with the greatest
concentration of industrial entities, agriculture and other branches of the
economy. Redistribution of fresh water over the territory of the country is
quite expensive, and furthermore has great effects on the ecology of large
areas. Therefore, various steps must be taken to preserve good quality of
fresh water (purification of industrial wastes, improvement of the tech-
nology of production in order to decrease the consumption of water and
dumping of wastewater into reservoirs, transition to closed cycles and dry
technologies). In order to preserve the water quality which is needed, it
is necessary to first of all limit the discharge of pollutants into reser-
voirs, i.e., standardize or regulate the discharge of chemical pollutants.
Various indexes characterize the level of pollution in water: chemical,
bacteriologic, hydrobiologic and the MPC's for individual toxins. The
chemical and biologic factors are the most widely used, the MPC's being less
frequently used and hydrobiologic indexes being quite rarely used. However,
it is hydrobiologic processes in reservoirs which play the decisive role in
the formation of water quality. Aquatic organisms, on the one hand, develop
their vital activity on the basis of hydrochemical and hydrologic modes;
water for their habitation and, on the other hand, the predominance of
various species of aquatic organisms determines the direction of hydro-
biologic processes and thereby determines the nature of formation of water
quality.
This interrelationship of water quality and hydrobiologic processes in a
reservoir causes definite difficulties in standardization of the discharge
of chemical pollutants into reservoirs and in the production of water qual-
ity. The necessity has arisen for indicating hydrologic principles which
must form the basis for development of standards for the protection of good
water quality, and for estimation of normal and pathologic states of reser-
20
voirs. To do this, let us discuss the main elements of the problem, in
order to note paths for their solution.
In each reservoir, the quality of water is formed by all aquatic or-
ganisms. They pass through their bodies the entire mass of water of the re-
servoir, enriching it by many products of their metabolism and, simulta-
neously, changing the gas and mineral composition of the water. In the
cycle of matter, some organisms play a determining role while others play a
subordinate or even hardly noticeable role. Bacteria, protozoa, algae, and
all invertebrate animals - the filter feeders - play a significant role.
A reservoir is a multicomponent system, consisting of living organisms
and the water itself, containing various chemical substances in the mole-
cular and supermolecular states, as well as the bottom, which contains a
number of organisms and silt particles. The number of species is usually
several hundred or even thousands in such reservoirs as Lake Baikal, while
the number of individual substances is not precisely known, but it must be
assumed that there are also several hundreds, or perhaps even thousands.
For example, some of the large rivers pick up along their way not only
several hundreds of different chemical compounds and ions, depending on the
geochemical status. of the watershed, but also several hundreds of chemical
compounds from industrial enterprises, cities and population centers, water
transport, and atmospheric precipitation. The complete chemical composi-
tion of such waters is unknown. We know indirectly that it includes a long
list of substances.
This tremendous number of components in the water system is in total
interaction and interrelation. The quality of water is a resultant of these
many interrelationships. It is practically impossible to consider them all
at the present time. Therefore, we must distinguish the most important
determining components. This approach to determination of the regularities
of behavior of an aquatic system is simplified, but is necessary in order to
solve the problems of standardization of water quality which have been set
before us.
Among aquatic organisms, three main functional groups must be distin-
guished: 1) producers - organisms which create organic matter in their
bodies by the process of photosynthesis, utilizing mineral substances dis-
solved in the water (salts and gases); 2) consumers, transformers - organ-
isms which construct their bodies by consuming organisms of groiup 1. This
group includes phytophages and organisms which feed on the phytophages,
i.e., predators; and 3) reducers. A large group of organisms (bacteria,
protozoa, fungi) decompose the waste substances from the vital activity of
other organisms as well as dead organisms, to mineral substances once more.
In each of these groups there are many species which follow each other
in a regular sequence during the seasons of the year. The specific composi-
tion of each functional group changes depending on the specifics of the re-
servoir, its geographic position,, climate, nature of bottom, hydrologic and
hydrochemical modes. For the full cycle of matter in the reservoir, the
specific composition of the functional groups (1-3) is of no great signifi-
cance, while for commercial organisms (their nutrition, growth, breeding),
21
the specific composition, particularly of organisms of the first group, may
be of decisive significance. For direct consumption by man (commercial or-
ganisms), some organisms of the second functional group are of great
significance.
Of the many hydrochemical components, substances defining the overall
characteristics of the water (carbonate system, relationship of calcium and
magnesium, sodium and calcium, chlorine and sulfate), as well as dissolved
organic matter and biogenic elements (nitrogen, phosphorus, iron) and micro-
elements (manganese, boron, copper, cobalt, etc.) are quite significant. To
this normal composition of natural water, we must now add chemical pollu-
tants, consisting of many different compounds, the chemical nature and
biologic activity of which are not fully known. We do not know in what form
they are present in the water and what are the paths of their transforma-
tion. We note that they always influence hydrobiologic processes in the
reservoir. As a rule, this influence is not desirable for man and his
activity. The aquatic organisms of each functional group have differing
sensitivities to the effects of toxic substances which, with pollution,
leads to restructuring of the specific composition within each group and
among species from various groups. Toxic substances, depending on their
chemical nature and concentration, suppress and reduce the population of
some species while others are stimulated and increase their numbers, while
still others are indifferent, i.e., retain their previous status (Stroganov
1978).
A change of dominance (predominant species) may not change the quantita-
tive aspect of a functional group. It will play its role in the cycle of
matter in a reservoir. However in the formation of good water quality and
the creation of high productivity of commercial organisms, these changes in
hydrobiologic processes may be undesirable. Therefore, we must limit the
delivery of chemical pollutants to a body of water if we desire to use it
for fishing purposes or for the supply of drinking water.
The interrelationships between functional groups in a reservoir can be
drawn in the form of a diagram (Figure 1).
An actual body of water is an open system for both matter and energy.
Therefore, reducers must process not only the substances which are trans-
formed from primary organic matter by the producers, but also substances
which enter the body of water from without. Usually, as organic matter in
the water increases, the number of organisms which mineralize it also in-
creases, but this process always involves some delay.
If we represent primary producers as P, all consumers and transformers
as C and reducers as R, in the ideal case P = C •»■ R. However, reducers can-
not mineralize all dissolved organic matter completely, and some of it falls
to the bottom sediment, while some remains in the dissolved state. Since
there are sediments accumulated in past eras in all reservoirs, we can con-
clude that reducers have never been capable of mineralizing all of the dead
organic matter in reservoirs. Consequently, the actual relationship has
been: P+A=C+R+0, orP+A=C+R+B+0, where P is the primary
organic matter of producers; A is that entering from without (allochthonic
22
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23
matter); C is the organic matter in consumers; R is the organic matter in
reducers and broken down by them; 0 represents bottom sediment and B is the
catch of commercial species and insects which migrate out of the system.
At the present time, the situation is complicated by the fact that com-
ponent A consists not only of organic matter washed away from the surface of
the land, but also many toxic substances in industrial waste, residential
sewage and flood water. If a reservoir is used for commercial purposes
(fishing, catching of crabs and mollusks), some of the organic matter is re-
moved from the reservoir in the form of commercial species. All industrial
reservoirs are populated, particularly around their shores, with insect lar-
vae, which leave the reservoir in the imago stage, thus carrying away a por-
tion of the organic matter from the reservoir.
Chemical pollution acts on the entire aquatic ecosystem (living and in-
direct) and due to the variety in quality and sensitivity of living compo-
nents of the system, restructures it in the direction of greater agreement
to the new quality of the environment. This restructuring almost never
satisfies the needs of humans. This is because processes of self-purifica-
tion are suppressed. Reducers cannot process all of the matter polluting
the water in such a short period of time. Water quality decreases and com-
mercial species disappear.
Reducers function in a definite sequence (biologic oxidation, nitrifica-
tion in two phases) and if the toxin breaks some link, the entire chain of
processes of mineralization is broken.
We have studied the effects of many toxins of various chemical natures
(metals, organometallic compounds, pesticides, antiseptics) and in all cases
a common law is observed, as the concentration of the toxin increases, there
is a delay in the development and an increase in the population of sapro-
phytes and nitrifiers. The delay may be so long that self-purification is
practically absent for 2-4 months. Figure 2 shows the variation of the
several links of self-purification with concentration of toxins and time of
action.
If this delay in mineralization processes occurs in a river, the pol-
luted water flows downstream for 1000-1500 or more kilometers from the
source of pollution. Quite naturally, the river carries traces of the ef-
fects of the chemical pollutant over this entire distance. Various filter
feeders, particularly bivalve mollusks and Cladocera crustaceans, play a
great role in processes of self-purification of water. However, they are
sensitive to chemical pollution and their population drops quite rapidly,
leading to a decrease in the self-purifying capability of the aquatic eco-
system with subsequent death of many species. The aquatic ecosystem is
simplified to a small number of species and, if the chemical pollution
continues to increase, the entire ecosystem may approach zero. This trend
in aquatic communities is reported elsewhere (Stroganov 1978).
Of course, under today's conditions there are no surface natural bodies
of water which have responded to pollution by complete death, but some small
areas near industrial production facilities have approached this state.
24
4 8 12 16
TIME, days
4 8 12 16
TIME, days
2.4 -
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-
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TIME, days
16
Figure 2. Summarized graphs of the main links in self-purification.
Figures at the curves denote increasing concentrations.
25
Therefore, the entire picture of change is quite clear, the flora and fauna
disappear.
The disappearance of valuable commercial species (which are usually
sensitive to chemical pollution) has been described for some time in the
literature. However, the scale of pollution and the variety of pollutants
have increased greatly in the present century and particularly since the
1940s. Therefore, maintenance of reservoirs in a state desirable for man
has become much more difficult.
We must see clearly that the struggle for pure water of good quality and
containing valuable organisms is a difficult task, a long-term task requir-
ing significant effort of the entire state and of intergovernmental organi-
zations as well .
In terms of preservation of hydrobiologic processes in reservoirs, which
assure the required quality of water and productivity of commercial species,
we must limit the arrival of toxic substances into bodies of water. Of
course, it would be quite good if we could completely eliminate any pollu-
tion (from the atmosphere, soil, waste and flood waters), but this is un-
realistic, at least for the foreseeable future. Therefore, regulation and
protection of reservoirs from toxic substances is a task of primary impor-
tance.
In developing specific indexes to be used to limit toxins, it is usually
noted that, if a reservoir has a capacity for self-purification, it should
be used, or allowed to purify all the discharge dumped into the reservoir.
It is said that this is quite economical. This means of solution of the
problem is quite favorable to the industry doing the polluting, but not to
the nation, since other water users will be restricted or even denied the
ability to use the polluted water. Our laws and constitution note that na-
tural waters belong to the state and are used in a combined matter, i.e., by
various water usei^s.
Yet another suggestion has been heard to ease the burden on industry.
Before waste waters are dumped into a reservoir, they should be diluted with
pure water, thus accelerating self-purification of the water. Actually, as
the concentration of organic substances and toxins decreases, the rate of
self-purification increases. However, from where is this pure water to be
taken for dilution at a time when the water consumption of industry is great
and increasing rapidly? Furthermore, studies which we have performed show
that the wastewaters of some chemical combines would have to be diluted by a
factor of 200-500 to eliminate their toxicity (Stroganov, et al^. 1978).
There is not enough pure water for this purpose, and the waTer, which would
be used, is not completely pure. Therefore, even the water in the deltas of
large rivers is not completely pure, not completely suitable for drinking
and fishing purposes. What is the answer?
The only effective answer to this problem is to decrease the quantity of
toxins entering bodies of water. The achievements of science and tech-
nology, all technical progress, allow this to be done, but economic diffi-
culties arise. The techniques needed to decrease the concentration on
26
toxins in wastewater are expensive. No matter how expensive it may be, man
must pay the price. The relationship between the cost of purification of
water, the number of species of hydrobionts living in the water for a given
level of pollution, and the degree of disruption of aquatic ecosystems can
be expressed by the graphs of Figure 3.
A decrease in the purity of waste water (sewage and flood water, water
polluted by water transportation, etc.) leads to a sharp decrease in the
number of species; perhaps, first of all, a significant decrease in commer-
cial species and, along with this, a significant increase in disruptions in
the aquatic ecosystem. Money saved in reduced purification leads to money
lost due to disruption of the normal (favorable for man) aquatic ecosystem.
Limitations of chemical pollution by means of the MPC significantly im-
prove the situation, but do not guarantee complete safety. We must assume
that: 1) the ecosystem includes more sensitive organisms than those which
have been used in biologic testing to establish the MPC. Elimination of
these species from the community may have an influence on the entire eco-
system. 2) Long-term after-effects may result from the influence of chemi-
cal pollutants on various vital processes of aquatic organisms. However,
these two questions must now be stated as issues for the future. Even if
all industrial enterprises, cities and large population centers purified
their waste water to harmless concentrations for aquatic organisms, toxic
substances would still reach reservoirs from the atmosphere and with water
running off the surface of the land. We must assume that the body of water
can handle this quantity of pollutant. If the self-purifying capacity of a
body of water is somewhat greater than is currently being used, this excess
amounts to a reserve of strength in the aquatic ecosystem. At the present
time, many reservoirs cannot cope with the large quantities of chemical com-
pounds entering them. They are functioning beyond the limits of the normal
(useful for man) capacity of self-purification. As a result of this, any
new addition of toxins to a body of water only increases the harmfulness of
the water system for organims which are useful to man. As is shown in
Figure 1, an aquatic ecosystem consists mainly of three functional groups of
organisms, which perform vital processes at different rates. The rates are
determined not only by the specifics of the organisms, but also by the en-
vironment (temperature, gas and salt composition and presence of toxins).
Therefore, we must always consider that, for example, self-purification pro-
cesses do not occur as rapidly as we would like, so that commercial species
disappear. This disagreement between rates of self-purification and quanti-
ties of chemical pollution leads to long-term disruption of all hydrobiolo-
gic processes characteristic for pure reservoirs.
Based on the requirements of a reservoir in terms of preservation of
hydrobiologic processes assuring pure water of good quality and productivity
of valuable commercial species, the following four principles should be used
as a basis for standardization of water quality in fresh surface bodies of
water:
1. The principle of priority in the use of reservoirs. All large and
medium sized reservoirs are used by many users, whose requirements for water
quality vary greatly. The highest requirements for water quality are those
27
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DEGREE OF POLLUTION
Figure 3. Relationship between degree of purification, pollution, number
of species and disturbance of aquatic ecosystem. 1 -Expenditures for
treatment of waste waters, flood waters, and other pollutions;
2-Number of species in ecosystem; 3-Degree of disturbances
in aquatic communities and ecosystems.
28
of fishing and drinking water supply. Only a few industries require water
containing \jery low contents of salts. Such water users perform special
water preparation measures on the water taken from the reservoir. There-
fore, priority in the use of water is quite significant in the protection
of water. In our water law it is noted that priority in the use of water
must be given to organizations supplying water for drinking purposes and to
fishing. Evaluations of the quality of the water and testing of water are
performed by the Health Ministry and the Fishing Industry Ministry. This
principle essentially lies at the basis of our water law, adopted in
December 1970 (see sections 10, 15, 28, 31 and 37). Considering the great
sensitivity of many species to chemical pollution, the formation of pure
water of good quality by various species of aquatic organisms, and also con-
sidering the high sensitivity of valuable commercial species (fish, crabs,
mollusks), priority should be given first of all to the fishing industry,
with all of the results which follow from this (evaluation of water quality,
testing and development of quality standards of discharge, etc.).
2. The principle of sufficient self-purification. This important
principle is the basis of all subsequent principles. It means that all of
the chemical pollutants which enter a reservoir must be mineralized to
limits of concentration such that the species forming pure water of good
quality and the species which are valuable commercial organisms can con-
tinue to exist. This means that for each region, climatic zone, the upper
limit of self -purifying capacity of the water of a reservoir, which must not
be exceeded, is the point of introduction of a greater quantity of pollu-
tants than the body of water can process. Increasing the load of chemical
pollution on a body of water above the limit of its self-purifying capacity
leads to disruption of the principle of sufficient self-purification, lead-
ing to pollution of the body of water and degradation of the entire ecologic
water system.
Processes of self -purification always occur (Figure 4), but not always
with sufficient speed and completeness to assure the subsequent principles
(i.e., 3 and 4). Therefore, self -purification may be sufficient for insen-
sitive commercial species, but not sufficient for highly sensitive species
and not sufficient to assure good quality of drinking water (principle 4).
Consequently, the sufficiency of self-purification is evaluated on the basis
of principle 1 (priority). For some water users, the requirements for water
purity are lower and they may be satisfied with incomplete pur.if ication of
water. Fishing and drinking water supply require water of the highest
purity. Each water user can establish his own level of sufficiency of self-
purification. We shall analyze it on the basis of the priority indicated
earlier.
The quantitative indicators used to evaluate sufficient self-purifica-
tion cannot be limited to BOD, COD AND O2 content. Since we must always ex-
pect toxins to be present in water, we must determine the rate of processes
of nitrification in both phases. As was noted earlier (Stroganov 1978),
toxins decrease the rates of these processes, thus delaying the time of suf-
ficient purification. In addition to these indexes, we must also have in-
formation on the toxicity of water for organisms. In most cases, nitrify-
ing organisms are more sensitive to toxins than are saprophytes, while most
29
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30
aquatic invertebrates and fish are still more sensitive than the nitrifiers.
Therefore, we can evaluate water on the basis of the line of sufficient
self-purification. More complex analysis, than is currently used, is re-
quired. We must also include toxicologic testing.
Certain toxic substances do not break down (e.g., metals) or break down
poorly (some pesticides, detergents, etc.). In these cases, toxicologic
testing will reveal their presence above impermissible concentrations.
Chemical analysis is important and necessary for an overall description of
the quality of water, but the indexes of self-purification and toxicity
reflect another aspect, yery important for the course of normal hydro-
biologic processes.
3. The principle of assurance of conditions of life for commercial
species. This principle falls entirely in the area of human evaluation. In
addition to pure water of good quality, man also needs biologic resources
found in reservoirs, particularly commercial species as a source of food and
industrial raw materials. Valuable commercial organisms react sensitively
to chemical pollution. They decrease their population or disappear as a re-
sult of death and migration to other water areas. Assurance of the condi-
tions of life means the presence of water of a quality such that commercial
species can continue to exist throughout their entire life cycle and do not
lose their valuable qualities (growth rate, fertility, maintenance of high
population, nonaccumulation of substances harmful to man, e.g., metals,
pesticides, hydrocarbons, detergents, etc.). Chemical pollution may have
both a direct effect on commercial organisms and an indirect effect through
their food and the water in which they live.
It might be thought that, if the second principle is fulfilled, the
third is not needed. However, the problem is more complex. Valuable com-
mercial organisms and their sources of food are more sensitive than microor-
ganisms participating in the decomposition of organic matter in water.
Therefore, even if the second principle is fulfilled, though it is quite
important, it is not sufficient to assure the third.
The qualitative and quantitative characteristics of this third principle
are: the specific composition of commercial species, their population and
biomass, ichthyofauna and the dimensions of the catch. Usually, the catch
of aquatic organisms is the first sign of deterioration in water quality for
commercial species, at a level at which the processes of self-purification
reflect no danger.
At the present time, the importance of this principle is great, since
the catch of aquatic organisms will become increasingly concentrated in in-
land bodies of water and the littoral waters of the oceans and seas in the
near future. Hydrobiologic analysis encompasses essentially the entire
ecologic system and catch and, therefore, most completely characterizes a
given ecosystem with respect to its suitability for effective and complete
utilization in the national economy.
4. The principle of suitability of water for drinking. The estimation
of the suitability of water is usually performed by sanitary organizations.
31
We include this principle in hydrobiologic analysis because water quality is
formed by aquatic organisms. What is the required quality of drinking
water? In accordance with State Standard GOST 2874-73, water should be
transparent, colorless and odorless, pleasant to taste, should contain no
pathogenic organisms or toxic substances above the established MPC.
In analyzing water in accordance with the third principle, we find at
times that water has long-term after-effects on aquatic organisms. They are
manifested as changes in fertility, time of maturation, decreased dimensions
of progeny and other deviations from characteristic parameters for the
species. Determination of all of these problems forces medical and veteri-
nary workers to ask the question of possible equivalent or similar in-
fluences on man and domestic animals using the same water for drinking pur-
poses.
Summing up what we have said, it must be noted that evaluation of the
quality of water in reservoirs from a broad hydrobiologic standpoint more
reliably characterizes quality than other existing approaches. Chemical,
physical and bacteriologic analyses cannot completely describe the quality
of surface water today. The proposed hydrobiologic principles will help in
developing a better scientific foundation for standardization of the quality
of water of surface reservoirs. These principles are oriented toward devel-
oping standards for water quality in various regions and types of reservoirs
used for fishing and drinking purposes.
The principles which we have set forth for estimation of normal and
pathologic states of bodies of water suffering from chemical pollution are
not new principles. They have been used and considered in the development
of criteria for water quality. What is new is that the principles formu-
lated are presented as a system for determination of the suitability (nor-
mality) or unsuitability (abnormality) of an aquatic ecosystem for the most
demanding water users. These principles can serve as a basis for develop-
ment of measures for standardization of water quality in reservoirs.
The principles formulated should assist in the development of standards
for aquatic ecosystems based on the requirements of man's economic activity
and life support. The criterion of the ecologic norm of a given reservoir
might be the completeness with which the second, third and fourth principles
are fulfilled. If these principles are excluded, evaluation of an aquatic
ecosystem is senseless.
Under all conditions, man is the main standard for evaluation of the
normality or abnormality of a body of water. The quality of water is be-
coming increasingly important for him. Therefore, evaluation of an aquatic
ecosystem occurs primarily along the line of quality evaluation. It is not
simply the number and variety of species, but rather useful species and
their population and productivity; not simply the stability of the system,
but rather the stability of the required quality of the system. Any eco-
system with time will reach stability given the surrounding conditions and
becomes stable. An aquatic ecosystem is stable both with polysaprobic
pollution, and with ol igosaprobic pollution. In either case, it is stable,
but the stability of the various qualities of water have different effects
32
on man. Preference is given to the oligosaprobic state of a reservoir over
the polysaprobic state. Aquatic organisms, as we know, are given preference
in accordance with their physiology and biology. For them, a normal body of
water is that which best corresponds to their physiologic and biologic pecu-
liarities. A polysaprobic organism cannot live in pure water (oligosapro-
bic) and vice versa. Evaluation of what is normal in a reservoir can be
performed by man, based on the principles outlined above.
Each organism also evaluates the quality of water in a reservoir. Can
it live or not? Based on this evaluation, we can evaluate the usefulness of
the ecosystem for man. Otherwise, we fall into unanswerable questions.
REFERENCES
Budyko, M.I. 1977. Global'naya ekologiya (Global ecology). Mysl' Press,
pp. 4-316.
Douglas, William 0. 1975. The Three-Hundred Year War. A chronicle of
ecologic struggle. Progress Press, Moscow, pp. 5-238.
Meadows, D.H., D.L. Meadows, J. Randers, and W.W. Behrens. 1972. The
limits to growth: A report for the Club of Homes Protection the Predi-
cament of Mankind. New York, pp. 2-72.
Oldak, P.G. 1979. A global strategy. Khimiya i zhizn'. No. 5, pp. 11-1?.
Stroganov, N.S. 1977. The meaning of the concept of norm and pathology in
water toxicology. Norma i patologiya v vodnoy toksikologii, Baykal'sk,
pp. 5-11.
Stroganov, N.S. 1978. Pressing problems of water toxicology in connection
with preservation of reservoirs from chemical pollutants. Elementy
vodnykh ekosistem. Nauka Press, Moscow, pp. 150-73.
Stroganov, N.S, A.I. Putintsev, Ye.F. Isakova, and V.I. Shifin. 1979. A
method of toxicologic testing of wastewater. Biologicheskiye nauki. No.
2, pp. 90-96.
Vernadskiy, V.I. 1967. Biosphere. Mysl' Press, Moscow, pp. 225-359.
33
SECTION 3
THEORETICAL ASPECTS OF THE "NORMALCY AND PATHOLOGY" PROBLEM
IN AQUATIC ECOTOXICOLOGY
L.P. Braginsky^
During the rather short period of development of aquatic toxicology as a
scientific trend, attention was mainly focused on the influence of toxicants
upon selected aquatic organisms. The fundamentals of general toxicology
established while investigating warm-blooded animals were the guiding prin-
ciples in this research. Life, however is diverse and complex, and biology
is multifaceted. That is the reason such an approach is insufficient. It
does not include many of the consequences the influences of toxicants on the
living matter of the hydrosphere.
In medicine and veterinary science, many variations from certain stand-
ard average values, characterizing vital manifestations and considered as
"the norm", are usually defined by the concept "pathology". Continuing
further with this analogy, medicine, veterinary science, phytopathology and
ichthyopathology in solving particular problems of diagnosis and treatment
of various human, animal and plant diseases, are based on general pathology,
the disease theory. However, even in such a highly developed science as
medicine, which for many centuries has accumulated information about human
organism functioning, the concepts of "norms" or "standards" are highly in-
definite. Only yery recently has a special science related to healthy
humans, normology, begun to develop in medicine. In both veterinary science
and ichthyopathology this problem remains completely unsettled.
Our knowledge about the biological, physiological, and biochemical pro-
cesses of aquatic organisms is so poor and insufficient, that in every
separate case it is necessary to start a toxicological investigation from
the study of the norm, and then to draw conclusions about various pathologi-
cal effects as a result of studying the responses of known test-organisms to
toxicants, while the number of aquatic species amount to hundreds of thou-
sands, or even millions.
For these reasons, aquatic toxicology and data storage needs tend to de-
fine existing concepts of the normalcy and pathology of aquatic organisms
under toxic environmental conditions. Recently, Soviet scientists have
given much attention to this problem. However, as analysis of the present
^Institute of Hydrobiology, Ukraine Academy of Science, Kiev, USSR.
34
information has shown, primary attention is given to the analysis of
normalcy and pathology at the organism and suborganism levels. Meanwhile,
aquatic life specificity lies in the fact that aquatic organisms live in
communities of different rank, and only their combined activity is of deci-
sive importance in the formation of those aquatic ecosystem characteristics
which are of interest to man, i.e., biological productivity and the
maintenance of proper water quality.
Mass biological processes are of considerable importance for understand-
ing the processes of water quality formation. It is these processes which
lead to community structure transformation and the disturbance of balance in
ecosystems, i.e., the processes at the supra-organism level, which are ob-
jects of ecological/hydrobiological investigation, not the individual re-
sponses of organisms to a toxicant.
A new trend in ecology, ecotoxicology, which has been recently
developed, and has already won world-wide recognition, deals not with the
individual organism response to toxic effects, but with the response of the
community and ecosystem, as well as the transformation of toxicants in
natural ecosystems. That is why it is necessary to understand the concepts
of normalcy and pathology at the supra-organism level of life organization.
What is a normal population? What is a population in the state of
"pathology"? What is a normal and a "pathological" biocenosis? What is a
"normal" and "unhealthy" ecosystem? Finally, what is an "unhealthy" body of
water, or "Krankenzee" described by German authors?
It is not easy to answer these questions, especially considering the ex-
treme lack of knowledge of the consistencies of supra-organism system func-
tioning. At the same time, it is clear that analysis of this problem cannot
be guided by those initial concepts by which medicine, veterinary science
and ichthyopathology operate, since the processes taking place at the supra-
organism level are inadequate for the organism level processes.
In this report the question of "normalcy" and "pathology" of the supra-
organism system is discussed from the points of view of demographic ecology
and synecology.
POPULATION LEVEL
One of the major criteria of conditions favorable to populations is the
ratio between birth and death. It is ^ery difficult to consider this factor
under natural conditions, but it may be characterized rather accurately in
experiments with synchronized test-cultures of short lived invertebrates.
In chronic toxicity tests with cultures of various Cladocera, after a series
of 5-6 generations a decline in fecundity of females as well as offspring
survival is observable. Similarly, an increase in mortality and a subse-
quent diminution of population can be noted.
One of the "pathology" indices at the population level, which can well
estimate statistically and interpret graphically is the potential produc-
tivity value. This value is calculated by an equation, which connects the
35
main biological parameters of the Cladocera, including lifetime of female,
the number of litters during a lifetime, intervals between litters, juvenile
numbers per litter, duration of maturation period duration prior to the
first litter, with the value of potential population productivity (Pigaiko
1971). If potential population productivity is reduced from generation to
generation, then it is a visual indicator of its pathological state, and the
increase of potential productivity, or its maintenance at a stable state,
are indicative of well-being, i.e., of the relative norm (Braginsky, et al .
1979).
Apparently, a number of biological productivity methods of assessment of
aquatic animals, established for general hydrobiology (Vinberg 1968) with
proper ecological and toxicological interpretation can be used in an analo-
gous way to demonstrate the pathological state of a population of aquatic
animals under toxic environment conditions.
For parthenogenetic invertebrates, i.e., Cladocera, Rotatoria, a switch
to sexual reproduction and laying of subitan eggs (ephippia) indicate un-
favorable conditions. However, under the influence of toxicants, this re-
sponse is not always observed. Thus, the shift to sexual reproduction and
formation of ephippia in Daphnia is absent in those cases exposed to chronic
additions of low concentrations of phenyl urea derivatives, triazine, heavy
metals, and surfactants. However, other pathological phenomena such as the
appearance of dwarf males and parthenogenesis in specimens of half the size
of the controls are observed.
The most frequent manifestation of pathological disturbances in Clado-
cera is egg abortion and the appearance of embryonic malformations. While
these disturbances may be considered as a change at the organism level,
their mass manifestation influences the fate of populations considerably.
Fluctuations in the number of aquatic populations in nature are highly
diverse, and depend upon many factors for which it is difficult to account.
Thus, knowledge of causes and mechanisms of these fluctuations is still ex-
tremely scanty. For this reason it is better to confine present activities
to the concept of developing model laboratory investigations.
In the conduct of aquatic toxicological experiments it is necessary to
resort to the study of laboratory "mini-populations" or "pseudo-popula-
tions". An elementary estimation of the median lethal concentration is
made on the population model. If an experimental group of warm-blooded
animals or fishes is impossible to consider as population, and the LC50
value obtained from invertebrates is interpreted as an individual mean, then
the analogous group of invertebrate offspring are derived from the same
parent and may be considered as an extract of a single population. As
experience shows, conclusions drawn from studying such test-culture are in
generally valid for aquatic ecosystems where the same species may be repre-
sented by a rather numerous population.
It is useful to consider the significance to the population the crite-
rion LC50. A wide utilization of this toxicometric criterion means that
36
the death of the test-organism is recognized as the most authentic indicator
of the toxic action of a substance.
It is a criterion which is beyond the concept of normalcy and pathology,
since death represents a leap to a new quality to which no characterizable
biological concepts can be applied. In this case the biological essence of
death is disregarded, and the result of an experiment is considered as
simply the answer to the question: is the substance toxic or not? But at
the population level, the essence of this question is different. The LC50
criterion itself means that any population is heterogeneous in relation to
its sensitivity to the toxicant. It suggests that there are resistant and
tolerant individuals within it, and, therefore, the toxicant functions as a
factor of natural selection with regard to the fate of the population.
Mortality as an ecological and evolutional factor controlling population
numbers has appeared together with life, and it would disappear only
together with it. If death means an awful and final defeat in the struggle
for existence for an individual, then for a population mass death is only
the elimination of the less adaptative, the survival of the more adaptative
incorporates some form of "reorganization", the essence of which is that
the population number declines abruptly first, then as resistant forms ap-
pear, a population numbers outbreak is observed. A health experience with
insecticide application is evidence of this phenomena. As a result of wide
utilization of strong insecticides, the insects not only survived but on the
contrary reproduced intensively. Aquatic animals are no exception to this
phenomena. It is known for instance, that mosquito fish resistant to DDT
have recently appeared (Holden 1973). In another case, a Cladocera test-
culture appeared to be killed in the V-VI-th generation under the influence
of toxicants, however, the XY-XYI-th generation "suddenly" revived and began
to breed rapidly. Finally, an algae culture almost killed under the in-
fluence of algaecide preparations was able to recover, and new cell genera-
tions grew. In principle, all these phenomena mean that the population has
latent resources to aid in elimination, and with the decrease of environ-
mental toxicant concentration, it can function as stimulative factor for re-
production of the organisms inhibited by it, in accordance with the law of
phase reactions.
Aquatic organisms, in contrast to warm-blooded animals, have other
latent resources, namely the ability to survive unfavorable conditions in a
resting stage, i.e., the statoblasts of moss animals, turions of aquatic
animals, ephippia of Cladocera, spores and cysts of Protozoa, the closing of
mollusk, shells, and the resting stage of algae. All these forms of life
exist in sediments, and are not susceptible to toxic effects. The adapta-
tion to very severe conditions in water is a rather good protection against
toxic agents, and it serves to guard populations from destruction by toxic
substances. In contrast to poikilothermal aquatic species, homothermal or-
ganisms are physiologically only accessible to poisons under conditions of
optimal temperature. At temperatures below 15°C, their biologic processes
are so inhibited, and exchange with environment is so reduced the the pre-
sence of a toxicant in their environment is of no serious danger to them.
Thus, the toxicity of a substance, and the even higher values of the LC50
obtained in the experiments with actively functioning individuals is not
37
necessarily evidence of its danger to a population. These factors serve only
to warn about toxic effects under conditions of optimal temperature. When
the temperature of water is raised to 30°C, the toxicity of a given substance
for organisms can be increased by hundreds, thousands, and tens of thousands
times. This has been demonstrated in experiments with cadmium on Daphnia
magna (Braginsky and Scherban 1978). Therefore, the question of the "patho-
logical" reactions of aquatic populations to toxic effects is inseparably
linked with ambient temperatures.
The existence of populations, as opposed to individuals, is in itself
protective, since an irregular distribution of a toxic agent within popula-
tion predetermines the possibility of preserving some quantity of resistant
individuals. This was noted in natural communities of the blue-green algae
treated with algaecide preparations. Luminiscence microscopy data showed
that from 0.5 to 20 percent of the total quantity of algae was unaffected by
algaecides. In experiments with aquatic invertebrates, uneven mortality of
test organisms was observed, although it was not possible to connect this
phenomenon directly with the level of toxicant accumulation in the animals'
body.
An irregularity of toxicant distribution among fish populations was con-
firmed analytically by gas chromotography for extracts of DDT in organs and
tissues. When studying accumulation levels of this pesticide in fish popu-
lations, fluctuations in cerebral fat tissue from 0 to 40 mg/kg were ob-
served, consistent with a normal distribution range. It is natural that
fish with DDT levels exceeding the critical values (3 mg/kg of cerebrum
weight) are in a state of deep pathology; a cumulative intoxication which
does not affect the entire population (Braginsky, et al^. 1979).
All analogeous phenomena are undoubtedly similar, and subject to the law
of survival of the species since the history of the earth, toxic factors are
not new. They probably functioned constantly in the early stages of the de-
velopment of the planet, with respect to high concentrations of ammonia,
methane, phosphorus and other toxic agents in water. The "chemical weapon"
is of importance to interspecies relations, and where this weapon was used,
protective measures were created. Apparently these measures are also ef-
fective with respect to toxicants of anthropogenic origin. Whatever the me-
chanism is for populations reaction to toxic effects, the ultimate result
should be a decrease of population abundance. Occasionally, the population
may even increase, when concentrations promoting reproduction are favored.
In any case, the question where "normalcy" ends and "pathology" begins is a
controversial consideration. It must be noted that deceleration or accele-
ration of a population's reproduction rate, or fluctuations in its range of
abundance are not something fatal or unfamiliar. Sequential sigmoid fluctu-
ations of population quantity are characteristic of life on earth; there-
fore, it is hardly appropriate to speak about pathology in the same sense in
which the term is used in medicine.
38
THE LEVEL OF THE COMMUNITY AND THE ECOSYSTEM
The most greatest problem of the present, the problem of clean water, is
connected not with the processes of individual and population levels, but
with the synecological processes, since water quality is a function of the
combined living activity of aquatic organisms. Therefore, the final crite-
ria in assessing toxicant effects on an aquatic population as a whole, i.e.,
criteria of "normalcy" and "pathology", are the processes taking place with-
in complex biological formations, the community and the ecosystems.
Toxicant inputs into a natural ecosystem leads to a rather specific
situation, the major features of which may be characterized as follows:
1. The toxicant is directed not towards a single target
organism as it is under experimental conditions in aquarium,
or in the whole in vitro system, where the isolated "toxicant-
organism" relationship is artificially created, but rather,
the toxicant effects on variety of targets;
2. As a result of spectrum of action, its concentration is dis-
persed and the real dose per organism is not equivalent to
the present projected concentration;
3. The toxicant quantity per biological organism depends on popu-
lation density, biomass, species diversity, the presence of
the most susceptible organisms consuming the given toxicant,
and on many other factors;
4. Immediately after entering an ecosystem, the toxicant is at-
tacked by active lower organisms, begins to undergo biodegrada-
tion by various exoenzymes, and is intercepted by species sus-
ceptible to accumulation;
5. A decrease in concentration as a result of the process of de-
toxication, dispersion, physico-chemical destruction, and sorb-
tion of the toxicant promotes phase reactions, which may be
responsible for both inhibition and stimulation of vital activ-
ity of aquatic organisms.
Thus in an aquatic ecosystem, the toxicant encounters the system func-
tioning as a whole: it is a negatively eutropic system, and the toxicant is
an entropic factor destroying life. Between the entropic factor, and the
system inclined toward negative eutrophy, a struggle starts. In the system
a counteraction grows in an effort to destroy the entropic factor. This
creates its specific quality buffering, described in the works of M.M.
Kamshilov (1973). The system consumes and transforms the toxicant, but only
within certain limits. When this potential of resistance is exhausted, a
toxic effect is manifested.
Because of this situation, bodies of water with varying trophic status
have varying degrees of resistance to toxicants, and varying rates of
transition to the state of disturbed balance. Generally, the richer in life
39
a body of water, and the more diverse this life quantitatively and qualita-
tively, the slower is the transition from normalcy to pathology. This sug-
gests that eutrophic systems should be less liable to the effects of toxic
substances than oligotrophic and distrophic ones. In this connection the
unstudied problems of toxicity criteria (normalcy and pathology) at the
sapra-organism level of life organization arise. The difficulty of their
formulation lies in the fact that the scientific fundamentals of functional
community studies are not established, and the present knowledge of commu-
nity structure is mainly the knowledge of morphology, composition, quantity,
biomass, occurrence, and various indices or relationship between the major
components in the structure. It concerns planktonic as well as bottom com-
munities, and also the other less studied group of aquatic animals.
Nevertheless, even the morphological approach and related experimental
investigations permits discovery of some of the specific features of commu-
nity reactions to toxic effects. To understand these reactions, it is
necessary to use the concepts of dominant, subdominant, and "shelf" forms.
The results of ecological investigations show that in ecosystems not in-
fluenced extensively by man, the structure of communities and the character
of seasonal changes are rather stable, and may be of the same type over a
period of many years. In waters polluted by toxic substances, or in eco-
systems under conditions of experimental influence, characteristic features
become visible, including a shift of the dominant forms. Occasionally,
shifts are very abrupt and conditioned by the fact that the dominate forms
are inhibited or eliminated completely, whereas forms of minor importance
reach the maximum of abundance and biomass (Braginsky 1975; Braginsky, et
al. 1979). The shift of other community components may be observed, anB~
tFese changes occur spasmodically as well as slowly in accordance with the
degree of toxic effect, toxicant concentration, selectivity of action, com-
munity specific composition, and many other factors. Moreover, there is a
change in total numbers, and in biomass of organisms, as well as an exchange
of roles in the structural components of biocenosis, i.e., a change in
hierarchical relationships. Under the influence of \/ery strong toxicants,
the community may be completely destroyed, and then the system becomes non-
structural. Apparently, the latter may be considered as an indicator of ob-
vious pathology, whereas the shift of dominant forms is not a pathological
process, but represents a form of community stabilization under new condi-
tions. The second case is the typical manifestation of degradation, the
mechanism of which has been studied in detail (Stroganov 1974).
Experimental investigations and mathematical modeling had demonstrated
that aquatic communities, generally speaking, may exist in three stable
states: 1) initial, 2) functionally and structurally reversibly altered,
and 3) irreversibly altered. The second level of change is characterized as
ecological fluctuation, the third as a shift of dominant forms. These do
not represent pathology, but simply the normal range of community vari-
ability related to adaptational changes. Apparently, "pathology" begins
when the system passed the third level of stability and approaches the non-
structural level. In mathematical models this process is shown by a para-
bola and indicates the approach of ecological catastrophe.
40
The structure, i.e., regularity, is characterized by the presence of a
reserve of negative entropy. "Destructuring" indicates the development of
processes of entropy, a movement in the direction of "chaos" (Hilmy 1968).
This is a physical indication of the process promoted by the influence of
toxicants in ecosystem. However, as it was previously noted, the system as
a whole is a complex of factors, among which microorganisms and Protozoa
play a chief role to counteract entropy (Kamshilov 1973; Braginsky 1975;
Geptner 1977). The toxicant is "dispersed" in ecosystem and under the in-
fluence of microorganisms its concentration decreases. In the end, it
determines ecosystem buffering, its ability to consume and transform a cer-
tain quantity of toxicant (Kamshilov 1973).
Buffering may be considered the degree of negative entropy of the system
as a major factor of preservation of its normal life. The transition to
"pathology" begins when the buffering limit is reached, and the system is
unable to withstand this toxic effect.
Now we approach the main question of the problem of clean water: what
is a "pathological" waterbody or ecosystem, and how does it differ from a
"normal" one? In the light of the previous discussion, it appears as if the
answer should be: an ecosystem in a "pathological" state is a body of
water with a disturbed buffer system, in which the detoxification potential
is suppressed and negative entropy processes yield to the entropic pro-
cesses, i.e., degradational ones.
One of the manifestations of such a state is an increased mortality
within community populations, particularly among highly organized life
forms; differing, as a rule, by a greater tolerance to toxicants. As a re-
sult of the increased death rate, population dynamics, age and sex ratio
changes, community structure changes correspondingly, and the system shifts
to a qualitatively different state. This state may be rather stable, parti-
cularly if the population which is resistant to toxicants becomes predomi-
nant, or unstable, with the tendency to further degradation, if this popula-
tion also is rather tolerant to toxicants. In certain individuals (as in
the intermediate stage between the normal state and death) various patho-
logical disturbances appear, which may be considered indicative of unfavor-
able conditions in the system. Symptoms may include disturbances in enzyme
systems and other biochemical changes corresponding functional disturbances,
structural pathohistological changes, alterations of conditioned reflex
activity, and behavioral reactions studied by toxicologists on the organism
and suborganism levels.
Recently, it is difficult to tell what relationship exists between dis-
turbance of various functions and the structure of some organisms, including
fish. Of particular concern are the lethal concentrations of toxicants and
their threat to aquatic life at supra-organism levels. Critical, then, is
the extent that clear and evident pathological changes at organism level re-
flect the "pathology" of supra-organism level, i.e., the community or the
ecosystem, since every lower level of organization is less resistant to
toxic factors than the next higher one, and the ecosystem is in danger of
catastrophe only when all of the buffer systems at lower levels are des-
troyed.
41
The notions of normal and pathological states of aquatic ecosystems are
closely associated with the whole complex of other ecological concepts such
as preservation of homeostasis, transformation of community structure, a
shift of dominant forms, disturbances of bio-geochemical cycles, system buf-
fering, detoxification potential and, finally, with the concept of entropy
and negative entropy system.
From this point of view we consider the study of the general problems of
pathology of aquatic ecosystems in the light of the second principle of
thermodynamics. The consideration of the problem of detoxification of
waters should then be from the view of life as a negatively entropic pro-
cess, evoked by our planet to retain energy, and to prevent its dispersion
into space.
In the same way that consideration of the flux of substances and energy
in aquatic systems from a position of the law of conservation of energy pro-
moted fruitful solution of many problems in productional hydrobiology, the
analysis of aquatic ecosystem responses to toxicants effects in the light
of the second principle of thermodynamics may significantly stimulate our
understanding of the destructive and reduction processes and factors, deter-
mining the stability and degradation of aquatic ecosystems, and the hydro-
biosphere as a whole.
REFERENCES
Braginsky, L.P. 1975. An ecological approach to the investigation of me-
chanisms of the activity of toxicants in the aquatic environment. In:
Formation and Control of the Quality of Freshwater, Vol. 1, Water Toxi-
cology. Published by "Science Thoughts", Kiev, pp. 5-15.
Braginsky, L.P., V.D. Byeskaravaynara, and E.P. Shchyerban. 1977. Reaction
of freshwater phyto- and zooplankton to waterborne pesticides. Pub-
lished by Academy of Sciences of the USSR, 10 p.
Gepther, V.A. 1977. Influence herbicides (Monoron, Dioron and Kotopan)
microhabitat collector, drainage-irrigation systems Turkmen and Uzbecki-
stan. Degree Candidate of Biological Sciences. Dissertation. Moscow
State University.
Helme, G.F. 1968. The basis of physics of the biosphere. Hydrometeriolo-
gist, Leningrad, 299 p.
Holden, A.V. 1972. Contamination of freshwater by persistant insecticides
and their effects on fish. Ann. Appl . Biol., 55, pp. 332-335.
Kamshilov, M.M. 1973. Buffering of living systems. Journal Social
Biology, 34, No. 2, pp. 174-194.
Kamshilov, M.M. 1977. Norms and pathology in a functional aquatic eco-
system. In: Norms and Pathology in Aquatic Toxicology. Thesis report.
All Union Symposium, Baikalsk, pp. 13-16.
42
Pedgayko, M.L. 1971. A comparison of production-biological cultivation
methods in investigating the toxicity of pesticides for zooplankton.
In: Methods of Biological Investigation in Aquatic Toxicology.
Science, Moscow, pp. 169-172.
Stroganov, N.S. 1973. Theoretical basis of action of pesticides on water
organisms. In: Experimental Water Toxicology. Published by
"Benatnyeh", Riga, pp. 11-37.
43
SECTION 4
TRENDS IN AQUATIC TOXICOLOGY IN THE UNITED STATES:
A PERSPECTIVE
Foster L. Mayer, Jr., Paul M. Mehrle, Jr. and Richard A. Schoettgerl
The need for toxicology testing has increased during the 1970's. It was
expanded for pesticide registration; many of the same requirements for
pesticide registration will be required for toxic .substances approval; and
acute and some chronic toxicity testing are being required for ocean dumping
permits. Research approaches are changing from acute toxicity testing and
residue analysis to more complex and integrated research involving chronic
toxicity, clinical chemistry, and ecosystem concepts. These approaches are
resulting in assessments of the environmental hazard of contaminants, some-
times even before they enter the environment, rather than in the production
of acute toxicity and residue data of only limited value. Also, the inte-
grated approach is providing basic scientific concepts that are essential in
the prediction of environmental hazards.
Developmental research is providing better interpretation and shortcuts
in toxicology. In ecosystem studies, scientists are determining what really
must be measured to assess the type and degree of pollution; biochemical
techniques are decreasing the time required for chronic toxicity studies;
and organisms other than fish (plants and invertebrates) are being recogn-
ized for their importance to fish and aquatic ecosystems and are being
tested accordingly. Recognition of the complexity of aquatic contaminant
residues has led to increased emphasis on the development of integrated
strategies for their detection and analysis.
Research emphasis has shifted from the problems of persistent organo-
chlorine pesticides to the prediction of problems that may arise as mining,
smelting, and coal conversion are increased, new methods of sewage dis-
posal, petroleum and detergent use expands, and pesticides use changes in
forest, range, and agricultural practices. The increasing concern of indus-
try with environmental problems is resulting in joint industry-government
research, not only to assess hazards, but to further define less hazardous
substitutes. A new interest is emerging in metals and other inorganics.
Although the literature contains abundant research on organics, much of it
^United States Fish and Wildlife Service, Columbia National Fisheries
Research Laboratory, Route #1, Columbia, Missouri 65201.
44
is unusable, and it is difficult to predict the environmental impact of
energy development and the associated inorganic contaminants. There is a
rapidly increasing trend toward use of larger quantities and greater vari-
eties of herbicides in agriculture. New forest management techniques call
for control of scrub and hardwood vegetation over vast acreages; no-till
farming practices require greater uses of herbicides and herbicide mix-
tures; and conversion of riparian vegetation into agricultural uses results
in herbicide and insecticide run-off. All of the problems with persistent
organochlorine pesticides are not gone, however. Decisions concerning some
of them still await a stronger factual base; others merely require monitor-
ing and surveillance to pinpoint problem areas and insure that the residue
trends continue downward.
Specific research advances and developments in aquatic toxicology in the
United States are presented here.
TOXICITY TESTING
Acute Toxicity
Toxicologists are well aware of the virtues and limitations of the acute
toxicity measure; yet, there are probably few measurements that have been as
misunderstood in evaluating hazard or safety of a chemical to aquatic life
as the LC50 (concentration lethal to 50 percent of the organisms within a
given period--usually £96 h). Users of any acute toxicity data must bear in
mind that the LC50 measures only one biological response — a lethal one.
Its main value is to provide a relative starting point for the evaluation,
along with other measurements (e.g., water solubility of the chemical, its
partition coefficient, its degradation rate), of environmental hazard. In
addition, the acute toxicity test provides a rapid, cost efficient way to
measure relative toxicity of different forms and formulations of a chemical,
its toxicity in different types of water (acidic, basic, hard, cold, warm),
and its toxicity to organisms representing different trophic levels. Until
other techniques can be shown to be equal or more meaningful to aquatic
toxicologists, the acute toxicity test is here to stay.
Chronic Toxicity
Partial and complete life-cycle toxicity tests with fish have become
commonplace, and provide data on survival, growth, reproduction, and other
sublethal responses. However, these tests can be expensive, high-risk in-
vestigations that may require up to a year to conduct. Recent evaluations
(Eaton 1974; Macek and Sleight 1977; McKim 1977) have shown that 30- to 60-
day toxicity tests on embryos and larvae may provide data as sensitive as
that observed in partial and complete life-cycle tests. The maximum accept-
able toxicant concentrations (MATC) derived from tests with embryos and lar-
vae, or juveniles were usually equal to, but never exceeded a factor of 3
times the MATC values derived with partial or complete life-cyle tests
(Table 1).
45
TABLE 1. MAXIMUM ACCEPTABLE TOXICANT CONCENTRATIONS (MATC) FROM
PARTIAL AND COMPLETE LIFE-CYCLE TOXICITY TESTS WITH FISH AS COMPARED
WITH MATC'S DERIVED FROM EMBRYO, LARVAE, AND EARLY JUVENILE TOXICITY TESTS'
Partial /complete
Embryo- 1
arval/
life-cycli
e MATCs
juvenile
MATCs
Toxicant
Fish Species
(yq/1)
(vg/i)
Pesticides
Acrolein
Fathead minnow
11 -
42
11 -
42
Atrazine
Brook trout
60 -
120
120 -
240
Trif luralin
Fathead minnow
2.0 -
5.1
5.1 -
8.2
Endosulf an
Fathead minnow
0.20 -
0.40
0.20 -
0.40
Endrin
Flagf ish
0.22 -
0.30
0.22 -
0.30
Heptachlor
Fathead minnow
0.86 -
1.8
0.86 -
1.8
Diazinon
Flagfish
54 -
88
54 -
88
Fathead minnow
6.8 -
14
6.8 -
14
Guthion
Fathead minnow
0.33 -
0.51
0.70 -
1.8
Malathion
Flagfish
8.6 -
11
8.6 -
11
PCBs
Aroclor 1242
Fathead minnow
5.4 -
15
5.4 -
15
Aroclor 1248
Fathead minnow
1.1 -
3.0
1.1 -
4.4
Aroclor 1254
Fathead minnow
1.8 -
4.6
1.8 -
4.6
Aroclor 1260
Fathead minnow
2.1 -
4.0
2.1 -
4.0
Metals
Cadmium
Flagfish
4.1 -
8.1
8.1 -
16
Fathead minnow
37 -
57
37 -
57
Chromium
Fathead minnow
1,000 -
3,950
1,000 -
3,950
Copper
Brook trout
9.5 -
17
9.5 -
17
Fathead minnow
11 -
18
11 -
18
Lead
Brook trout
58 -
119
58 -
119
Flagfish
31 -
62
62 -
125
Nickel
Fathead minnow
380 -
730
380 -
730
Zinc
Flagfish
2d -
51
51 -
85
Fathead minnow
30 -
180
30 -
180
^Condensed from McKim (1977)
46
other research being conducted that involves short-cut methods to
chronic toxicity studies has been highlighted by the U.S. Environmental Pro-
tection Agency's Environmental Research Laboratory-Duluth (1977-1979) and
includes the following advances:
1. Measurement of ventilatory patterns of fish with a microcomputer
monitoring system.
2. Use of fish cough frequency as an estimate of chronic toxicity.
3. Development of a rapid toxicity test in which the fingernail
clam is used.
4. Monitoring liver aryl hydrocarbon hydroxylase induction in fish.
5. Changes in steroid hormone metabolism in fish.
6. Saltwater tolerance and smoltif ication in salmon.
Aquatic Plants
The effect of point and non-point source contaminants on submersed
rooted vegetation is little known. The contribution of submersed rooted
aquatic macrophytes to the ecological support of fishery and wildlife re-
sources can be separated into three general categories:
1. Numerous species of mammals and waterfowl are directly depend-
ent on macrophytes as food. For example, the stems, leaves,
seeds, and rootstock of sago pondweed constitute up to 50 per-
cent of the diet of migratory ducks and geese. Submersed rooted
macrophytes are also required by fish for forage, cover, and
spawning; furthermore, they provide an important substratum for
invertebrates eaten by fish.
2. The overall metabolism of aquatic systems (lakes and streams)
supporting fisheries is dependent to a major extent on the
detritus components of dead, dissolved, and particulate organic
carbon which form the primary source of biological energy.
Beds of submersed, littoral, rooted macrophytes contribute a
large part of the organic detritus in all but a few aquatic
systems.
3. Littoral vegetation also modulates the flow of inorganic nutri-
ents from the watershed to the limnetic area and stabilizes and
controls the magnitude of planktonic photosynthesis in lakes.
In addition, contaminants deposited in bottom muds may be taken up by
plants and passed along a detrital food chain, ultimately to fish, water-
fowl, and other organisms closely associated with aquatic ecosystems. To
estimate the effects of contaminants on rooted aquatic vegetation, we are
examining the following variables for inclusion in chronic laboratory tests
with appropriate species: growth, reproduction, photosynthesis, nutritive
47
value, and residues. The transfer of residues through food chains of which
the exposed vegetation is a part is also being investigated
CLINICAL (DIAGNOSTIC) TESTS
The use of diagnostic tests in hazard assessment procedures can decrease
the time required for safety evaluation of chemicals, define no-effect ex-
posure concentrations more adequately, and provide a better understanding of
the mode of action of chemicals. Routine diagnostic tests are frequently
not available to aquatic toxicologists because biochemical and physiological
research has been minimal in aquatic toxicology, which is a relatively new
field of science, as compared to such fields as human medicine (Mehrle and
Mayer 1979). The "state of the art" of physiological, biochemical, and his-
tological tests in aquatic toxicology held at Pellston, Michigan (Macek et
a1. 1978). The participants rated the relative utility of eleven toxicity
tests, using the criteria of ecological significance of effects, scientific
and legal defensibility, availability of acceptable methods, utility of test
results in predicting effects in aquatic environments, the general applica-
bility to all classes of chemicals, and the simplicity and cost of the test.
In terms of present utility for use in assessing the hazard to aquatic envi-
ronments, acute lethality tests were rated highest, followed by embryo-
larval tests, chronic toxicity tests measuring reproductive effects, and
residue accumulation studies. Histological tests ranked ninth, and physio-
logical and biochemical tests tenth in overall and present relative utility
because of the inability to relate the results of these tests to adverse
environmental impacts.
Physiological and biochemical tests are generally not conducted for two
reasons: (1) it is felt that they are mainly useful in evaluating the mode
of action of chemicals (Brungs and Mount 1978); or (2) there is not enough
basic information known about fish physiology and biochemistry to ascertain
the ultimate effects, since alterations in these processes do not neces-
sarily indicate a disadvantage to the survival and success of the organisms.
The analytical techniques and instrumentation are well developed for
performing clinical analyses, and considerable research on physiological and
biochemical responses induced by chemical toxicants has been conducted, but
useful biological or diagnostic indicators have not been developed. In our
opinion, the main reason for this lack of progress has been the lack of a
comprehensive, integrated approach in toxicological studies with fish. To
overcome this problem, researchers must conduct biochemical, physiological,
and histopathological investigations in conjunction with toxicity studies
that measure important whole-animal responses. Establishing the relation-
ship of organism to sub-organism responses will help insure development of
pertinent diagnostic indicators of fish health. The choice of whole-animal
responses to evaluate in toxicity studies with fish depends on the purpose
of the toxicology program, but in most aquatic toxicology programs, emphasis
is given to toxicant effects on survival, growth and development, reproduc-
tion, and adaptability.
48
To adequately assess the influence of contaminants on the aquatic envi-
ronment and to overcome the avoidance of biochemical and physiological test-
ing, investigators should develop techniques that can serve as biological
indicators in the field as well as predictors in the laboratory to estimate
the "health" of a particular aquatic resource. However, biochemical and
physiological changes must be viewed in light of the degree and duration of
change to determine whether the organism can adapt or whether the changes
lead to irreversible homeostatic disturbances and finally to the death or
debilitation of the organism.
BEHAVIOR
Any alteration in the ability of an organism to perceive and respond to
its environment will affect its survival and may increase ecological morta-
lity. Reports on behavioral changes induced by toxicosis cover an array of
behaviors, and diverse techniques have been used to study these. The extent
to which these methods can be applied in toxicological investigations de-
pends on the economy of the procedure as well as on the accuracy with which
behavioral changes can be quantified. Two contaminants, or even two concen-
trations of the same contaminant may affect different behavioral responses,
and behavioral alterations caused by a substance may vary among species.
Thus, toxicological studies should rely on multiple behavioral responses.
The following behavioral responses are being evaluated as routine screening
tests for the effects of various contaminants.
1. Avoidance - Aquatic organisms avoid certain comtaminants and
are attracted by others. When a contaminant is introduced
through either arm of a Y-maze, avoidance reactions have been
shown to occur in mosquitofish (Gambusia affinis) to insecti-
cides (Kynard 1974), in rainbow trout (Salmo gairdneria) to
herbicides (Folmar 1976), in shrimp and mosquitofish to PCB's
(Hansen et al_. 1974) and in Atlantic salmon (Salmo parr) to
heavy metals (Sprague 1964).
2. Predator-prey relationships - Various contaminants also dis-
rupt predator-prey relationships by changing locomoter res-
ponses such as swimming or activity levels, or by disorienting
the organism or by impairing its ability to perceive a preda-
tor or prey. Several studies have shown that the certain con-
taminants may increase the prey organism's vulnerability to
predation (Goodyear 1972; Kania and O'Hara 1974; Tagatz 1976;
Farr 1977; and Sullivan et al. 1978).
3. Feeding and swimming activities - The survival of recently
hatched fry or invertebrate larvae depends in part on the time
at which specific behavioral patterns develop. Delayed or in-
hibited behaviors such as feeding or swimming have been shown
to occur as a result of contamination (Dill 1974).
Specific behavioral effects caused by contaminants are being correlated
with other biological characteristics such as pathology, biochemical aber-
49
rations, or reproduction, as well as with the survival of aquatic organisms
in natural systems. Also, the mechanism through which behavior has become
altered in aquatic organisms exposed to pollutants is being examined.
ECOSYSTEMS
Field Studies
One of the least explored areas of either ecology or environmental toxi-
cology is the ability of ecosystems to withstand contaminant stress. The
use of pesticides in environmental management and the deposition of indus-
trial contaminants in natural aquatic ecosystems has created a need for
studies on the effects of these materials on biological communities. Labo-
ratory studies can provide data on the effects of particular pesticides or
contaminants on many species of organisms under various environmental condi-
tions. However, such information may be of limited value at times in pre-
dicting the effects of pesticides and other contaminants on changes in
biological communities where many species interact. Contaminants may modify
these species interactions by affecting non-target organisms or be ecologi-
cally restructuring the biological community. These cause and effect ecolo-
gical interactions in natural aquatic communities can be estimated by mea-
suring certain characteristics such as primary productivity, standing crop,
species diversity, community respiration, nutrient cycling, etc. in con-
trolled lentic environments. Although chemical damage to a variety of eco-
systems is at least partially documented, and, in fact, has constituted a
major public and scientific concern in recent years, the facility with which
ecosystems may resist or recover from the action of toxic compounds has re-
ceived remarkably little attention.
The presence of a contaminant in an ecosystem, however, does not in it-
self imply toxicity. The contaminant must first be biologically available
(Pavlou et al_. 1977). Toxicity is the characteristic of an individual or-
ganism's response to a chemical at a particular concentration or dosage for
a specific period of time. The effect of a contaminant on a community or
ecosystem will depend, therefore, upon the summation of all individual re-
sponses within affected populations. Even though toxicity is generally most
evident at the organismic and population level, community and ecosystem re-
sponses to organic contaminants can hypothetically be assessed directly or
indirectly. The indirect approach is more probably within the present know-
ledge base of ecology and toxicology and involves the determination and
monitoring of critical ecosystem processes. This approach is analogous to
the medical one where the disease or malfunction is ascertained by a set of
symptoms. Symptoms are functional evidences of disease, and the observance
and measurement of symptoms may be far removed f>"om the actual affected
organ(s) or system.
Evaluation of the impact of contaminants on aquatic organisms has been
limited mainly to laboratory studies. Much of the laboratory research lacks
field verification and the true impact of contaminants on aquatic organisms
in the wild is poorly understood. The classical field approach involves
laborious age, growth, and population dynamics studies of fish and extensive
50
surveys of other flora and fauna (species diversity) that would probably be
applicable to that time and place only. Also, field studies are somewhat
limited to effects evaluation after contamination has occurred and can pro-
vide only limited predictability (Brungs and Mount 1978).
One of the main objectives of recent research has been to establish the
necessary measurements essential to predicting pesticide and other contami-
nant effects on lentic ecosystems (Boyle 1979a, b). In experimental ponds
exposed to herbicides (2,4-D DMA, dichlobenil, and fenac), one to seven
characteristics were sufficient to explain 80-90 percent of the differences
observed. The seven characteristics found to be most important were pH,
alkalinity, turbidity, total dissolved nitrogen, total phosphorus, chloro-
phyll a, and zooplankton density.
Biochemical Characteristics of Ecosystem Stress
The onset of environmental change in aquatic systems due to stress im-
posed by man is often difficult to discern. Even after severe ecological
damage has occurred, substantiation requires the collection and evaluation
of voluminous amounts of data. Train (1972) has pointed to the need for
usable indicators of environmental quality. Indicators of ecological stress
would be especially useful if they could be applied at the beginning of
ecological disasters, rather than proof that extensive ecosystem change has
already occurred. Although there is no well developed literature on this
subject, several studies indicate the possibility of using chemical and bio-
chemical characteristicss as indicators of ecological stress. Woodwell
(1972) cites three qualities of stressed ecosystems, (1) simplification of
structure; (2) shifts in the ratio of production to respiration; and (3)
loss of inorganic nutrients. Some marine studie-^ have linked specific bio-
chemical characteristics with ecological change (Jefferies 1972; Jefferies
and Alzara 1970), but similar references are not apparent in the literature
in freshwater. The changes in some chemical variables, such as concentra-
tion and location of inorganic nutrients, total organic matter and bio-
chemical diversity, seem to offer an opportunity to construct a set of
symptoms for early detection of ecological contamination. Interpretation of
the significance of field-measured changes, however, requires realistic
physiological and biochemical studies under experimental conditions. It
also requires development and adaptation of chemical methods for measurement
of contaminants in biota, sediment, and water.
RESIDUE DYNAMICS AND BIOCONCENTRATION
Factors that control the flow of contaminants through an ecosystem have
been classified into four major areas: (1) Physical transport and spatial
distribution; (2) Interfacial processes; (3) All noninterfacial chemical
transformations exogenous to the biota; and (4) Biotransformations (Pavlou
et al. 1977).
The physical transport and spatial dispersion are ecosystem specific and
depend on the circulation and flow dynamics associated with the dispersive
51
medium. These aspects have been discussed extensively by Gillet et a1 .
1974).
Interfacial processes can be broken down into two categories: (1)
Interfacial interactions not involving changes of the contaminant, but which
result in the exchange of the compound with the dispersive medium (soil,
water and air), and (2) all chemical reactions, abiotic or biotic, that al-
ter the chemical structure of the compound. Interfacial interactions not
involving changes of the comtaminant include volatilization, dissolution and
sorption (adsorption and absorption), molecular associations such as chela-
tion, hydrogen bonding, ionic interactions, etc. These physico-chemical
interactions are important because contaminants may not only be immobilized,
but that can also mediate mobilization and transport as reported by Ogner
and Schnitzer (1970). Also, the interactions art amenable to classical
physico-chemical treatment and interpretations. In addition, chemical
structure is a crucial aspect, not only as a flow-factor, but also in toxi-
city (Addison and Cote 1973; Cohen et al . 1974; Kapoor et al_. 1973;
Kopperman et al_. 1974; Sugawara 197^ VTlceanu et al_. 1972; Wildish 1974).
Studies on abiotic noninterfacial transformation reactions (photode-
gradation, hydrolysis, etc.) have been conducted for only a few organic com-
pounds (Crosby and Leitis 1973; Crosby and Moilanen 1973; Crosby and
Moilanen 1974; McGuire et al^. 1970; Pope et al^. 1970; Pope and Zabik 1970;
Ruzo et ^. 1972; Zabik et ^. 1971). Consequently an assessment of their
importance to ecosystem transport and availability is virtually impossible.
However, the results obtained from certain toxicological investigations in-
volving pesticides suggest that biotransformations may activate or deacti-
vate the parent compound to more or less toxic metabolities (O'Brien 1967;
O'Brien and Yamamoto 1970). Since the biological availability of organic
chemicals is of critical importance to evaluating toxicity, and thereby po-
tential ecosystem malfunction, the development of useful transformations
and interfacial exchange features has been undertaken.
The degree of bioaccumulation as a function of the available concentra-
tions in the medium can be predicted. Recent studies by Neeley et al .
(1974) have shown that the octanol/water partition coefficients for organic
chemicals are linearly correlated with bioaccumulation in fish. Correlating
the octanol/water quantities and environmental concentrations for a series
of chemicals may prove useful in providing a rapid screening technique for
predicting environmental concentrations. In addition, computerized treat-
ment of residue data from aquatic organisms continuously exposed to contami-
nants is actively being developed. The uptake phase is usually 28-56 days
and the elimination phase is 28 days (Figure 1). Accelerated bioconcentra-
tion tests of only 4 days have been used with some chemicals to predict bio-
concentration under longer exposures (Branson et a]^. 1975).
ENVIRONMENTAL HAZARD EVALUATION
The Toxic Substances Control Act of 1976 clearly indicates that an "un-
reasonable risk" of injury to health or the environment caused by manufac-
ture, distribution, use, or disposal is needed to establish a chemical as
52
/
• • ••
1
• •
MM I I I I L
00
O
CO
« I
UJ
CO
CO
CN
CN
CM
o
o
o
B/Bu'S3naiS3y 3NOd3>l
Figure 1. Computerized treatment of residue data from fathead minnows
exposed to 3.7 ng/1 of Kepone. Fish were continuously exposed for
56 days and placed in uncontaminated water for 28 days.
Parameter estimates:
Time to reach 90% of steady state 43 days
Bioconcentration factor 15,053
Time for 50% elimination 13 days
53
hazardous. Hazard evaluation is a probability assessment that adverse
ecological effects will result from environmental releases of a given con-
taminant. It involves a sequential and integrated approach to predict the
safety or hazard of the contaminant, and includes information on (1) chemi-
cal production, use, and disposal patterns; (2) acute and chronic toxicity;
(3) residue dynamics and bioconcentration; (4) environmental fate and moni-
toring; and (5) field studies (Figure 2). A hazard evaluation is not a one-
time estimate, and additional evaluations must be made as the data base ex-
pands. Useful assessment schemes have recently been proposed by Kimerle et
al. (1978), Duthie (1977), Stern and Walker (1978), and the American
Institute of Biological Sciences (1978). However, no scheme or procedure
can eliminate the need for sound scientific judgement. The evaluation, in
its essence, is a scientific judgement of the potential for environmental
effects (toxicity tests) with measured (or estimated) environmental con-
centrations. The degree of confidence in the evaluation is greatest with a
reliable estimate of environmental concentrations and with effects data
which includes studies on representative species under conditions simu-
lating those of natural aquatic environments.
REFERENCES
Addison, R.F. and R.P. Cote. 1973. Variation with chain length in acute
toxicity of alkylthydroxamic acids to salmon (Salmo salar) fry. Lipids
8:493-497.
American Institute of Biological Sciences. 1978. Criteria and rationale
for decision making in aquatic hazard evaluation. j_n Estimating the
Hazard of Chemical Substances to Aquatic Life, ASTM STP 657, J. Cairns,
K.L. Dickson, and A.W. Maki, Eds., Am. Soc. Testing and Materials, pp.
241-273.
Boyle, T.P. 1979a. Effects of the aquatic herbicide 2,4-D DMA on the
ecology of experimental ponds. Environ. Pollution (in press).
Boyle, T.P. 1979b. Responses of experimental lentic aquatic ecosystems to
alterations of macrophyte communities. Proc. Efficacy and Impact of
Intensive Plant Harvesting in Lake Management Symposium. Institute for
Environmental Studies Center for Biotic Systems, Univ. of Wisconsin,
Madison (in press).
Branson, D.R., G.E. Blau, H.C. Alexander, and W.B. Neely. 1975. Bioconcen-
tration of 2,2', 4,4'-tetrachlorobiphenyl in rainbow trout as measured
by an accelerated test. Trans. Am. Fish. Soc. 104: 785-792.
Brungs, W.A., and D.I. Mount. 1978. Introduction to a discussion of the
use of aquatic toxicity tests for evaluation of the effects of toxic
substances. ln_ Estimating the Hazard of Chemical Substances to Aquatic
Life, ASTM STP 657, J. Cairns, K.L. Dickson, and A.W. Maki, Eds., Am.
Soc. Testing and Materials, pp. 15-26.
54
Data
Actions
I
Production, Use and
Disposal Information
Exposure
Estimates f Concentrations
Evaluation
Decision
Alternatives
JZ
Minimal Hazard
Review
I
Stop Testing
mil
USE
I
Substance Properties
& Fate Data
Hazard Evaluation
Uncertain Hazard
Identify Further
Data Needs to
Define Hazard
I
Biological Test Data
Excessive Hazard
Review
Stop Testing
ABANDON
Added Tests as
Needed
RESTRICT
Figure 2. Schematic diagram of the environmental hazard evaluation process
(modified from American Society of Testing and Materials Hazard
Evaluation Task Group, J.R. Duthie, Chairman).
55
Cohen, J.L., W. Lee, and E.J. Lien. 1974. Dependence of toxicity on mole-
cular structure: Group theory analysis. J. Pharm. Sci. 63: 1068-1076.
Crosby, D.G. and E. Leitis. 1973. The photodecomposition of trifluralin in
water. Bull. Environ. Contam. and Toxicol. 10: 237-241.
Crosby, D.G. and K.W. Moilanen. 1973. Photodecomposition of chlorinated
biphenyls and dibenofurans. Bull. Environ. Contam. and Toxicol. 10:
372-377.
Dill, P. A. and R.C. Saunders. 1974. Retarded behavioral development and
impaired balance in Atlantic salmon alevins hatched from gastrulae ex-
posed to DDT. J. Fish. Res. Bd. Canada 31: 1936-1938.
Duthie, J.R. 1977. The importance of sequential assessment in test pro-
grams for estimating hazard to aquatic life, hi^ Aquatic Toxicology and
Hazard Evaluation, ASTM STP 634, F.L. Mayer ?»nd J.L, Hamelink, Eds.,
Am. Soc. Testing and Materials, pp. 17-35.
Duthie, J.R. In preparation. Standard practice to evaluate hazard to
aquatic organisms. American Society for Testing and Materials Committee
E-35 on Pesticides and Subcommittee E-35.21 on Safety to man and Envi-
ronment.
Eaton, J.G. 1974. Chronic cadmium toxicity to the bluegill (Lepomis macro-
chirus Rafinesque). Tans. Am. Fish. Soc. 103: 729-735.
Farr, J. A. 1977. Impairment of antipredator behavior in Palaemonetes pugio
by exposure to sublethal doses of parathion. Trans. Amer. Fish. Soc.
106: 287-290.
Folmar, L.C. 1976. Overt avoidance reaction of rainbow trout fry to nine
herbicides. Bull. Environ. Contam. and Toxicol. 15: 509-514.
Gillett, J.W., J. Hill, IV, A.W. Jarvinen, and W.P. Schoor. 1974. A con-
ceptual model for the movement of pesticide through the environment: A
contribution of the EPA alternative chemicals program. U.S. Environ-
mental Protection Agency, Ecol . Res. Ser. No. EPA-660/3-74-024. Wash.,
D.C. 79 p.
Goodyear, C.P. 1972. A simple technique for detecting effects of toxicants
or other stresses on prey-predator interactions. Trans. Amer. Fish.
Soc. 101: 367-370.
Hansen, D.J., S.C. Schimmel, and E. Matthews. 1974. Avoidance of Aroclor
1254 by shrimp and fishes. Bull, Environ. Contam. and Toxicol. 12:
253-256.
Jefferies, H.P. and L. Alzara. 1970. Dominance-diversity relationships of
the free amino acids in coastal zooplankton. Comp. Biochem. Physiol.
37: 215-223.
56
Jefferies, H.P. 1972. Fatty acids ecology of a tidal marsh. Limnol.
Oceanogr. 17: 433-440.
Kania, H.J. and J. O'Hara. 1974. Behavioral alterations in a simple pre-
dator-prey system due to sublethal exposure to mercury. Trans. Amer.
Fish. Soc. 1974: 134-136.
Kapoor, I. P., R.L. Metcalf, A.S. Hirwe, J.R. Coats, and M.S. Khalsa. 1973.
Structure activity correlations of biodegradability of DDT analogs. J.
Agr. Food Chem. 21: 310-315.
Kimerle, R.A., W.E. Gledhill, and G.J. Levinskas. Environmental safety
assessment of new materials. ln_ Estimating the Hazard of Chemical Sub-
stances to Aquatic Life, ASTM STP 657, J. Cairns, K.L. Dickson, and A.W.
Maki, Eds., Am. Soc. Testing and Materials, pp. 132-146.
Kopperman, H.L., R.M. Carlson, and R. Caple. 1974. Aqueous chlorination
and ozonation studies. I. Structure-toxicity correlations of phenolic
compounds to Daphnia magna. Chem. -Biol. Interactions. 9: 245-251.
Kynard, B. 1974. Avoidance behavior of insecticide susceptible and resist-
ant populations of mosquitofish to four insecticides. Trans. Amer.
Fish. Soc. 103: 557-561.
Macek, K.J. and B.H. Sleight, III. 1977. Utility of toxicity tests with
embryos and fry of fish in evaluating hazards associated with the
chronic toxicity of chemicals to fishes. ln_ Aquatic Toxicology and
Hazard Evaluation, ASTM STP 634, F.L. Mayer and J.L. Hamelink, Eds.,
Am. Soc. Testing and Materials, pp. 137-146.
Macek, K.J., W. Birge, F.L. Mayer, A.L. Buikema, and A.W. Maki. 1978.
Discussion session synopsis of the use of aquatic toxicity tests for
evaluation of the effects of toxic substances. XH Estimating the Hazard
of Chemical Substances to Aquatic Life, ASTM STP 657, J. Cairns, K.L.
Dickson, and A.W. Maki, Eds., Am. Soc. Testing and Materials, pp. 27-32.
McGuire, R.R., M.J. Zabik, R.D. Schuetz, and R.D. Flotard. 1970. Photo-
chemistry of bioactive compounds. Photolysis of 1 ,4,5,6,7,8,8-hepta-
ch^oro-3a-4, 7a-tetrahydro-4,7-methanoindene (Cage formation vs. Photo-
dechlorination) J. Agr. Food Chem. 18: 319-321.
McKim, J.M. 1977. Evaluation of tests with early life stages of fish for
predicting long-term toxicity. J. Fish. Rs. Bd. Can. 34: 1148-1154.
Mehrle, P.M., and F.L. Mayer. 1979. Clinical tests in aquatic toxicology:
State of the art. Proc. Soc. Toxicol. Aquatic Toxicol. Symp., New
Orleans, Louisiana (in press).
Neely, W.B., DR.R. Branson, and G.E. Blau. 1974. Partitioning coefficient
to measure bioconcentration potential of organic chemicals in fish.
Environ. Sci. Tech, 8: 1113-1115.
57
O'Brien, R.D. 1967. Insecticides - action and metabolism. Academic Press,
New York. 332 p.
Ogner, G. and M. Schnitzer. 1970. Humic substances: Fulvic acid - dialkyl
phthalate complexes and their role in pollution. Science. 170: 317-
318.
Pope, B.E., M.F. Para, and M.J. Zabik. 1970. Photodecomposition of 2-(l,3-
dioxolane-2-yl )-phenyl-N-methyl caramate. J. Agr. Food Chem. 18: 490-
493.
Pope, B.E. and M.J. Zabik. 1970. Photochemistry of bioactive compounds.
Photochemistry of selected 2-chloro and 2-methylthio-4,6-di (alkyla-
mino)-s-triazine herbicides. J. Agr. Food Chem.
Pavlou, S.P., R.N. Dexter, F.L. Mayer, C. Fischer and R. Haque. 1977.
Chemical -ecoystem interface. ^1 Ecosystem Processes and Organic Con-
taminants: Research Needs and an Interdisciplinary Perspective, J.M.
Neuhold and L.F. Ruggerio, Eds., Nat. Sci. Found. U.S. Govt. Printing
Office, Wash., D.C.
Ruzo, L.O., M.J. Zabik, and R.D. Schuetz. 1972. Polychlorinated biphenyls:
photolysis of 3,4,3' ,4'-tetrachlorobiphenyl and 4,4'-dichlorobiphenyl in
solution. Bull. Environ. Contam. Toxicol. 8: 217-218.
Sprague, J.B. 1964. Avoidance of copper-zinc solutions by young salmon in
the laboratory. J. Water Pollut. Control Fed. 36: 990-1004.
Stern, A.M. and C.R. Walker. 1978. Hazard assessment of toxic substances:
Environmental fate testing of organic chemicals and ecological effects
testing. In Estimating the Hazard of Chemical Substances to Aquatic
Life, ASTM STP 657, J. Cairns, K.L. Dickson and A.W. Maki, Eds., Am.
Soc. Testing and Materials, pp. 81-131.
Sugawara, N. 1974. Toxic effect of a normal series of phthalate esters on
the hatching of shrimp eggs. Toxicol. Appl . Pharmacol. 30: 87-89.
Sullivan, J.F., G.H. Atchison, and A.W. Mcintosh. 1978. Changes in the
predator-prey behavior of fathead minnows (Pimphales promelas) and
largemouth bass (Micropterus salmoides) caused by cadmium. J. Fish.
Res. Bd. Canada 35: 446-451.
Tagatz, M.E. 1976. Effect of Mirex of predator-prey interaction in an
experimental estuarine ecosystem. Trans. Am. Fish. Soc. 1976: 546-
549.
Train, R.E. 1972. The quest for environmental indices. Science. 178:
1009.
U.S. Environmental Protection Agency. 1977. Rapid methods research high-
lights. Environmental Research Laboratory, Duluth, Minn. 8 p.
58
U.S. Environmental Protection Agency. 1979. Rapid methods research high-
lights. Environmental Research Laboratory, Duluth, Minn. 11 p.
Vilceanu, R.Z. Szabadai, A. Chiriac, and Z. Simon. 1972. Multiple struc-
ture toxicity correlation of organic phosphorus compounds. Stud.
Biophys. 34: 1-6.
Wildish, D.J. 1974. Lethal response by Atlantic salmon parr to some poly-
oxyethylated cationic and nonionic surfactants. Water Res. 8: 433-
437.
Woodwell, G.M. 1970. Effects of pollution on the structure and physiology
of ecosystems. Science. 168: 429-433.
Zabik, M.J., R.D. Schuetz, W.L. Burton, and B.E. Pope. 1971. Photo-
chemistry of bioactive compounds. Studies of a major photalytic product
of endrin. J. Agr. Food Chem. 19: 308-313.
59
SECTION 5
COMPARISON OF PRINCIPLES OF DEVELOPMENT AND USE OF WATER QUALITY
STANDARDS IN THE USSR AND USA
L.A. Lesnikov^
Practically all nations, which have experienced the negative influence
of pollutants from industry and agriculture on bodies of water, have arrived
at the need to establish certain standards for these substances which are
considered safe for the use of bodies of water (McKee and Wolf, 1963).
However, in developing biological well-founded standards, a primary dif-
ficulty arises: the development of sufficiently well-founded standards is
quite cumbersome, while the number of pollutants which may enter bodies of
water is quite great. As we learned on a visit to the USA, the "bank of
substances" at one laboratory in Cincinnati includes some 25,000 substances.
In our country, about 600 sanitary-hygienic maximum permissible concentra-
tions (MPC) have been developed for harmful substances, as well as 210 fish-
ing industry MPC's. In the USA, judging from the literature which we have
examined, reports have been published on the degree of harm of a similar
quantity of substances, though as yet this information has primarily been
obtained from short-term experiments. Large numbers of substances have been
studied in both the USSR and the USA. Summing up all the information which
we have available at present, we know of the effect of only about 1,000 sub-
stances.
The following system is used in the USSR. MPC's are the same for all
bodies of water in the country, but there are two systems of MPC's: sani-
tary-hygienic, approved by the USSR Public Health Ministry, and fishing
standards, approved by the Fishing Industry Ministry, USSR. These standards
must be maintained by enterprises, beginning at a "measurement line" and be-
yond it. For the sanitary-hygienic MPC's, the "measurement line" is 1 km
upstream from the nearest point of water use in the case of rivers, or 1 km
distant from the nearest point of water use for reservoirs and lakes. For
the fishing standards, the "measurement line" is established no more than
0.5 km from the source of pollution.
For each specific enterprise, "discharge norms" or, as they have come to
be called in recent years, "maximum permissible discharges" (MPD) are esta-
^State Scientific Research Institute of Lake, River and Fishing Management,
Leningrad, USSR.
60
blished, i.e., the calculated quantity of any polluting substances, both as
to concentration and as to total volume, which can be discharged without
disrupting the MPC at the measurement line.
Sanitary-hygienic MPC's are not the subject of the present report, but
we note that, as they are developed, both short-term and long-term effects
of substances on the sanitary condition of bodies of water are considered
(the oxygen regime, content of substances capable of decomposition, capacity
of the water for stagnation and self-purification, number of microorganisms,
etc.), on the organoleptic properties of water, on the health of the local
population (toxicity, pathogenic organisms, etc.) (Cherkinskiy, 1971). In
the past decade, the stability of the pollutants and their cumulative pro-
perties have also come to be considered.
The fishing MPC's require study of: the stability of the pollutant, its
influence on the sanitary status of the reservoir (transparency, color of
water, pH, oxygen regime, BOD, etc.); the organisms of phytoplankton,
aquatic microorganisms, zooplankton, zoobenthos, spawn, larvae and mature
fish; cumulation of the substance by fish; and the influence on the quality
of fish flesh. Approximate times of experiments were presented by us in our
previous report (Lesnikov, 1976).
In analyzing the materials which we have received from our American col-
leagues, we at first thought to compare all available materials, but then
decided to concentrate our attention on research on fresh-water organisms,
since water toxicologic studies on marine organisms have not yet been suffi-
ciently developed in the USSR (Patin, 1977) to speak of the relative toxi-
city resistance of species. Therefore, the results of USA studies on marine
organisms shall be included only as is convenient.
In the USA, the degree of danger of a substance for fish and other
aquatic organisms, as determined experimentally, is summed up in the inte-
gral indicator "water quality criterion". According to McKee and Wolf
(1963), this indicator is considered in the establishment of "water quality
standards" for specific areas of bodies of water. The specifics of use of
the body of water and relative toxicity resistance of the species which in-
habit it are considered.
In order for one nation to use data obtained by another nation, it is
necessary to gain some idea concerning the relative toxicity resistance of
test organisms. Naturally, representatives of local aquatic fauna are used
both in the USSR and in the USA.
In our country it is the usual practice to divide organisms into four
groups in terms of their relative toxicity resistance (oligotoxobes, beta-
mesotoxobes, alphamesotoxobes and polytoxobes) (Lesnikov, 1976). We shall
attempt to classify the test organisms used for toxicologic research in both
the USSR and USA from this standpoint. It must be considered that this
classification is somewhat arbitrary, since the toxicity resistance of or-
ganisms varies for various toxic substances. It is more correct to speak
only of trends. The relationship of sensitivity also varies as a function
of the duration of exposure. We shall present here data obtained by the
61
ichthyopathologist of our laboratory, O.N. Krylov (1973) on the influence
of chlorophos (Dipterex) on fish (see Table 1).
TABLE 1. RELATIONSHIP OF LT50 (mg/liter) OF CHLOROPHOS FOR CURRENT
YEAR'S BROOD OF FISH AS A FUNCTION OF TIME OF EXPOSURE
Exposure
Coregonus
peled
Salmo
irideus
Gasterosteus
aculeatus
Cyprinus
carpio
96 hours
25 days
0.24
0.031
0.78
0.062
6.0
0.25
282.0
2.0
With an exposure of 96 hours, Coregonus peled was 1200 times more sensi-
tive to chlorophos than Cyprinus carpio, while with an exposure of 25 days,
it was only 64 times more sensitive. As a rule, the longer the exposure,
the less the difference is between sensitivities of species.
Our ideas concerning the relative sensitivity of test organisms to toxic
substances are presented in Table 2. The relative sensitivity of the test
organisms used in the USSR is estimated on the basis of studies of the
GosNIORKh Water Toxicology Laboratory (Lesnikov, 1976, 1973; Krylov, 1973;
Alekseyev and Lesnikov, 1977; Stroganova, 1971), while the relative sensi-
tivity of test organisms used in the USA is based on the works of McKee and
Wolf, 1963, Mayer et a^., 1975; Meerle and Mayer, 1975; Sanders, 1977;
Sanders et al_., 1973; Mayer etaX., 1976, 1977; Carlson, 1972; Hermanutz et
al., 1973; Macek et aj_., 1976; Sauter et al., 1976; Snarski et a^., 1976;
Allison and Hermanutz, 1977; Pickering et aj_., 1977; Christensen et al .,
1977; Eaton etal., 1978; McKim, 1977; McKimet al., 1976; Benoit et al.,
1976; Carwell et al., 1977; Spehar, 1976, Spehar et al., 1978; Hermanutz,
1977; McKimm et al., 1978; Lloyd, 1976; Lloyd et al. , 1976. Of course, this
table must be considered a first approach to the problem. We can see from
the data presented that some organisms, e.g., Salmo irideus, Cyprinus carpio
and Daphnia magna, are used in both countries, while the others are similar
in their sensitivity. At the present time, neither country uses the most
toxicoresistant species. Consequently, the data compared using today's test
organisms are comparable.
The experimental differences are small in most cases, significant in a
few cases.
EXPERIMENTS ON FISH
In the USSR, experiments are performed on eggs, larvae, current year's
brood and second year fish, less frequently on older fish. The usual dura-
tion of acute experiments is not over 15 days. As in the USA, the LC50 is
determined for 96 and 120 hours, and the curve of median lethal time as a
function of substance concentration is studied. Subacute experiments, al-
lowing the boundary of chronic lethal effect to be determined and sublethal
effects to be revealed, last up to 3 months (90 days). Chronic experiments,
62
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64
performed to answer questions similar to those answered by subacute experi-
ments, last up to 6 months or more.
The influence of the substance on survival, growth in length and weight,
development of eggs and larvae are all considered. The pathoanatomic and
pathohistologic changes in the organs and tissues (liver, kidneys, gut,
brain, sometimes spleen, gills, blood - hemoglobin, formed blood elements,
sometimes blood protein) are also considered.
In the USA, experiments are also performed on eggs, larvae, current
year's brood and mature fish. Furthermore, experiments have been undertaken
modeling the spawning of fish, extending over three generations: sexually
mature fish, the production of eggs and larvae which mature to the reproduc-
tive state themselves, observations on eggs and the larvae which they pro-
duce. In many cases, the experiments extend over 2-3 months and may be com-
pared to the "subacute experiments" in the USSR, but in many cases the
length of these experiments is greater than for chronic experiments in the
USSR - up to 1-3 years. Most experiments, however, last 90-150 days, i.e.,
comparable in length to those conducted in the USSR.
The same indexes are considered as in the USSR: survival rate, growth
in length and weight, development of eggs and larvae, but also the influence
of the substance on spawning of the fish is determined. Similar studies
should be organized in the USSR as well. Furthermore, in the USA a success-
ful "proportional diluent" scheme has been developed (Brungs and Mount),
which is quite convenient in the performance of chronic experiments. In the
USSR, new solutions are regularly prepared and manually replaced. Develop-
ment of a standard diluent is desirable for our country.
Of the histopathologic analyses, we found only one work in the USA
(Couch, 1975) which included information on changes in the liver of fish.
Thus, the results of ichthyotoxicologic experiments in the USSR and USA
are basically comparable,
EXPERIMENTS ON ALGAE
In the USSR, the most commonly used test organism of algae is Scenedes-
mus quadricauda, sometimes Chlorella vulgaris, with other species used only
in special studies (Khobot'yev and Korol', 1971; Khobot'yev et aj_, , 1971;
Kohbot'yev and Kapkov, 1971; Mosiyenko, 1974a, 1974b; Pain and Tkachenko,
1974; Vislyanskaya and Vedyagina, 1974; Lisovskaya et aj_, , 1968). Due to
the difficulty involved in replacement of the medium (difficulty in separa-
tion of algae from the liquid), the substance being studied is introduced to
the medium once, or a portion of the medium is replaced with fresh solution,
with an additional quantity of the toxicant introduced. The usual duration
of experiments is 25-30 days. Indexes recorded include: dynamics of popu-
lation of algae, settling rate, influence on pH of medium, on liberation of
oxygen, sometimes on absorption of radioactive carbon.
65
In the USA, toxicologic experiments are performed on Selenastrum capri-
corneum (Bartlett et al_., 1974; Ferris et^ aj_, , 1974), Chlamidomonas sp. (de
la Cruz and Nagvi, 1973); we found more detailed experiments on marine algae
(Walsh, 1972; Walsh et ^., 1977), judging from which the indexes considered
are the same as in experiments performed in the USSR, but the duration of
exposure is shorter--7-10 days. Considering the differences in experimental
duration, the results of the experiments are quite comparable.
EXPERIMENTS ON ZOOPLANKTON ORGANISMS
The main test organism in both countries is Daphnia magna. In the USSR,
experiments are performed in two variants:
1. According to the system of Professor N.S. Stroganov, on three
or more successive generations of Crustacea, the experiments
with each generation lasting 20-21 days (Stroganov, 1971;
Stroganov and Kolosova, 1971; Lesnikov, 1973). The indexes
observed are: survival rate, growth, intensity of reproduc-
tion and quality of progeny. In addition to these indexes,
the nature of processes of oogenesis and embryogenesis, body
color, accumulation of droplets and fat and their color, degree
of filling of the gut and color of its contents and others are
sometimes considered (Lesnikov, 1971).
2. According to the system of Lesnikov, using populations of
daphnia. This differs from the previous method in that the
young which are born are counted but are not removed from the
experimental vessels (the most convenient capacity of which is
1 liter). The duration of the experiments is until the maximum
biomass is obtained in the control and in the vessel containing
the substance being tested at the lowest concentration, usually
20-30 days; sometimes experiments are continued until the se-
cond or third peak of biomass (usually 50-60 and 70-120 days).
The indexes considered are the same as in experiments on series
of generations and, furthermore, consideration of biomass of the
daphnia and the change of parthenogenetic reproduction to bi-
sexual reproduction. Incidentally, it has been determined that
the influence of sublethal concentrations of a number of sub-
stances is manifested in that the daphnia do not go over to the
bisexual method of reproduction at the usual time or defective
latent eggs are formed which later burst.
In the USA, experiments on Daphnia magna are performed according to a
plan quite similar to that of N.S. Stroganov (Sanders, 1977; Sanders et al . ,
1973; Carwell et aj_., 1977). The time of experiments on one generation is
21-28 days; in experiments on series of generations, the times are approxi-
mately the same for each generation (Macek et aj_. , 1976).
The results of the experiments are fully comparable.
66
EXPERIMENTS ON BENTHIC INVERTEBRATES
In experiments with this group of organisms, a great variety of test or-
ganisms is used in both countries, the USSR and the USA.
In the USSR, various species of fresh-water gammaridae are used
(Gammarus pulex, G. lacustris, Pontogammarus robustoides, etc., Asollus
aquaticus), of the insects - Chironomidae, most frequently Chironomus dor-
sal is, for which a method has been developed of year-round cultivation under
laboratory conditions (Konstantinov, 1958). Remaining species of the mol-
lusks, ephemeroptera and odonata are less frequently used.
Experiments with gammaridae are performed over a period of approximately
a month, considering survival, intensity of cannibalism, growth and multi-
plication of the Crustacea and their feeding rates.
Experiments with Chironomidae extend from emergency of the larvae to
flight of the imagoes. Survival rate of larvae, pupae and imagoes are noted
(Bugayeva, Puzikova, 1974).
In experiments on other invertebrates, survival rate and growth are
usually noted, sometimes breeding rate as well.
In the USA, similar groups of benthic organisms are used. One specific
factor is the use of several ephemeroptera (Baetis vagans. Ephemera simi-
lans, Hexagema lineata), species which are rather sensitive to toxins. How-
ever, differences are observed. Our experiments with Baetis sp. (species
not precisely defined) have shown that this form was tolerant to methylni-
trophos, sevin and cobalt chloride. The American species (Baetis vagans),
judging from the results of experiments, has at least moderate sensitivity
(experiments of Lloyd et al_., 1976). In the USA, experiments are performed
on the larvae of Plecoptera (Pteronarcis californica, Acroneura pacif ica)
(Sanders and Cope, 1968). Judging from the figures they present, these
species are moderately, possible highly sensitive to toxins. Of the Chiro-
nomidae, Tanytarsus is used in the USA (in the laboratory at Duluth). Ac-
cording to GosNIORKh, Tanytarsus is somewhat more sensitive, at least to
chlorophos, than is Chironomus.
Thus, there are no basic differences in the methods used in experiments
on benthic organisms in the USSR and USA, and there are no great differences
in the relative sensitivities of the test organisms used.
The greatest differences are observed in methods of estimation of the
influence of pollutants on microorganisms and on the hydrochemical mode.
INFLUENCE OF POLLUTANTS ON AQUATIC MICROORGANISMS
In the USSR, experiments are performed in aquaria, to which fixed con-
centrations of the substances studies are added (once), then the dynamics of
the population of microorganisms are observed (total count on membrane fil-
ters, population of saprophytes growing on MPA) as well as the numbers of
67
specific groups of microorganisms which may be encountered, judging from the
nature of the substances studied, e.g., cellulosolytic bacteria for the sew-
age of cellulose-paper plants, petroleum oxidizing bacteria when studying
petroleum-containing waste water or specific petroleum products, etc. Ex-
perimental durations are 21-30 days (Mosevich, 1973). These experiments
have been included in a large system of studies, mainly performed in labora-
tories of the GosNIORKh systems, though other water toxicology laboratories
do not always include them, since they do duplicate hydrochemical experi-
ments to some extent i
It has been found that when water from natural bodies of water is placed
in aquaria, during the first four days a significant increase in the popula-
tion of microorganisms is observed, after which the number of organisms
varies within limits characteristic for the conditions in question. During
this time, the water from the natural body of water becomes aquarium water.
The effects of pollutants may result in an increase in the total popula-
tion of microorganisms, or of certain specific groups, or may suppress bac-
teria processes.
In the USA, based on the articles available to us, only one work
(Duthrie et aj_., 1974) is similar in methodology to works in the USSR: ex-
periments to determine the effect of diuron on microbial processes were per-
formed in experimental tanks. In the Laboratory for Study of Environmental
Pollutants at Gulf Breeze, Florida, a basically different system of studies
in "microcosms" (glass pipes containing water and soil) is used (Bourquin,
1977; Bourquin et al_. , 1977). The duration of these experiments is also 20-
30 days, but the results are basically different. Each experimental system
has its advantages and disadvantages; therefore, the comparability of re-
sults of these studies requires further checking.
HYDROCHEMICAL EXPERIMENTS
Studies are performed according to two main systems.
1. Estimate of intensity and nature of breakdown of pollutants.
2. Influence of pollutants on hydrochemical regime of bodies of
water, particularly processes of self-purification from sub-
stances other than the pollutant itself.
Studies of the breakdown or the fate of the pollutant in the water
system have been undertaken in both the USSR and USA to varying degrees in
almost all experiments. In laboratories of the GosNIORKh system, chemical
determination of the eventual fate of the pollutant are always accompanied
by biological toxicologic tests, usually using Daphnia magna. Frequently,
the products of decomposition of the substance are more toxic than the sub-
stance itself. For example, experiments in our laboratory have determined
that in solutions of chlorophos (Dipterex) in natural water, during the
first 2-5 days, the mean survival time of Daphnia decreases to half; this
elevated toxicity is retained for 1.5 months in open vessels and up to 2
68
months or more in closed vessels. This phenomenon can be attributed to DDVP
(dimethyldichlorovinylphosphate), a product formed upon decomposition of
chlorophos (dimethyloxytrichloroethylphosphonate) . An increase has been
found in toxicity during the first week in solutions of orthoxylene, though
the mechanisms of the process itself is not clear.
Of works of this type performed in the USA, we would like to note an ex-
ceptionally interesting study by Mancy and Allen (1977), on the influence of
environmental factors on the toxicity of heavy metal ions.
A second trend is estimation of the influence of a pollutant on the
hydrochemical processes in a body of water. This type of experiment is an
obligatory component of all water toxicology studies in the USSR. We found
no analogous studies in the USA. In these experiments, water is taken from
a natural reservoir and placed in an aquarium for study. In our laboratory,
water is taken from a reservoir with hard water (e.g., the Strelka River)
and another bo^dy of water with soft water (e.g.. Lake Ladoga). A series of
concentrations of the pollutant, usually 6-7 gradations, is created, with
pure water serving as the control. Analysis of pH, dissolved oxygen, BOD5,
BOD2o» permanganate and bichromate oxidizability, forms of nitrogen (am-
monia, nitrates, nitrites) are regularly analyzed, and changes in the con-
centration of the pollutant are observed. Many substances cause a decrease
in the content of dissolved oxygen and an increase in BOD, increasing the
saprobic nature of the medium. Toxic substances may significantly suppress,
either temporarily or throughout the experiment (usually 25-30 days) pro-
cesses of self-purification. Most frequently, processes of oxidation nitro-
gen are first suppressed, i.e., processes of formation of nitrites from am-
monia compounds and oxidation of nitrites to nitrates. In many cases, an
increase is found in the content of nitrites which cannot be explained by
oxidation of ammonia compounds and can be attributed only to denitrif ication
processes.
Summing up all that we have said, we note that, with the exception of a
small number of tests used in one country and not in the other, the studies
in the two countries, the USSR and the USA, generally follow the same goals,
and at the present time are performed according to basically similar
methods, which is determined as we compare works performed in the two coun-
tries. The most difficult question is that of the maximum permissible
standardization of a minimum program of these investigations.
It is hardly necessary to change the forms of application of the stand-
ards developed in one or the other of the countries--they are determined by
the specifics of our individual national systems. We can simply state that
the MPC system used in the USSR is equivalent in the nature of its scienti-
fic foundation to the concept of the "water quality criterion" used in the
USA, while the water quality "standards" used in the USA are more or less
equivalent to the "discharge norms" or "maximum permissible discharges"
(MPD) used here.
The system of distribution of test organisms in terms of their relative
sensitivity to pollutants represents some difficulty, since the relationship
of sensitivity of species to various substances differs somewhat. At one
69
time. Professor N.S. Stroganov suggested that the relative sensitivity of
test organisms be estimated on the basis of the ratio to that of Daphnia
magna; this can be done in works in both countries, since this species is
used in experiments in both the USSR and the USA. In any case, the system
which we have proposed (Table 2) should be looked upon as simply a first ap-
proach to the problem and should be refined as data are accumulated.
Thus, there is a firm basis for successful cooperation of both nations
in the development of specific means for protection of bodies of water from
pollution.
REFERENCES
Alekseyev, V.A. and L.A. Lesnikov. 1977. Pesticides and their influences
on aquatic organisms. Izv. GosNIORKh, Vol. 121, pp. 8-94.
Allison, D.T. and 0. Hermanutz. 1977. Toxicity of diazinone to brook trout
and fathead minnows. EPA-600/3-77-050.
Bartlett, L., F.W. Rabe, and W.H. Funk. 1974. Effects of copper, zinc and
cadmium on Selenastrum capricornutum. Water Research, Vol. 8, No. 3,
179-185.
Benoit, D.A., E.N. Leonard, G.M. Christensen, and J.T. Fiandt. 1976. Ef-
fects of cadmium on three generations of brook trout (Salvelinus fonti-
nalis). Transcr. Amer. Fish. Soc, Vol. 105, No. 4, pp. 550-560.
Bourquin, A.W. 1977. Effects of malathion on microorganisms of an artifi-
cial salt marsh environment. J. Environ. Qual., Vol. 6, No. 4, 373-78.
Bourquin, A.W., M.A. Hood, and R.L. Garnas. 1977. An artificial microbial
ecosystem for determining effects and fate of toxicants in a salt marsh
environment. Devel. Industr. Microbio., Vol. 18, 185-91.
Brungs, W.A., R.W. Carlson, W.B. Horning, J.H. McCormich, R.L. Spehar, and
J.D. Young. 1978. Effect of pollution on fresh-water fish. J. Pollut.
Contr. Fed., Washington, D.C., 20016, 1582-1637, June.
Bugayeva, L.N. and N.B. Puzikova. 1974. Method of determination of the in-
fluence of toxic substances on Chironomidae. Izv. GosNIORKh, Vol. 98,
pp. 135-37.
Carlson, A.R. 1972. Effects of long-term exposure to carbaryl (sevin) on
survival, growth and reproduction of the fathead minnow (Pimephales
promelas). J. Fish. Res. Board of Canada, Vol. 28, Ne, 5, 583-87.
Carwell, R.D., D.G. Foreman, Th.R. Payne, and D.J. Wilbar. 1977. Acute and
chronic toxicity of chlordane to fish and invertebrates. EPA-600/3-77-
019.
70
Cherkinskiy, S.N. 1971. Sanitary conditions of discharge of sewage into
reservoirs. Izd. 4-e.M., 208 pp.
Christensen, R., G.E. Hunt and J. Fiandt. 1977. The effect of methylmer-
curic chloride, cadmium chloride and lead nitrate on six biochemical
factors of the brook trout (Salvelinus fontinalis). Toxicol. Appl.
Pharmacol., Vol. 42, 523-30.
Couch, J. A. 1975. Histopathological effects of pesticides and related
chemicals on the liver of fish. The pathology of fishes, by W.E.
Ribefin and G. Migski, Madison, Wisconsin, USA.
Cruz, A. A. de la and S.M. Nagvi. 1973. Mirex incorporation in the environ-
ment, uptake in aquatic organisms and effects on the rates of photosyn-
thesis and respiration. Arch. Environ. Contam. Toxicol., Vol. 1, No. 3,
255-264.
Eaten, J.G., J.M. McKim, and G.W. Helcombe. 1978. Metal toxicity to embryo
and larvae of seven freshwater fish species. 1. Cadmium. Bull. En-
vironm. Contam. Toxicol., 95-103.
Ferris, J. J., S. Kabayashi, and N.L. Clesceri. 1974. Growth of Selenastrum
capricornutum in natural waters augmented with detergent products in
wastewaters. Water Research, Vol. 8, No. 12, 1013-1020.
Guthrie, R.K., D.S. Cherry, and R.N. Fereboe. 1974. The effects of diuron
on bacterial populations in aquatic environments. Water Resource Bull.
Amer. Water Resource Assoc, Vol. 10, No. 2, 304-310.
Hermanutz, R.O., L.H. Mueller, and K.D. Kempfert. 1973. Captan toxicity to
fathead minnows (Pimephales promelas), bluegills (Lepomis macrochirus)
and brook trout (SalvelinUs fontinalis). J. Fish. Res. Board of Canada,
Hermanutz, R.O. 1978. Endrin and malathion toxicity to flagfish (Jorda-
nella f loridae) . Arch. Environ. Contam. Toxicol., Vol. 7, 159-68.
Vol. 30, No. 12, Part 1, 1811-17.
Khobot'yev, V.G. and V.M. Korol'. 1971. Change in active reaction (pH) of
a medium as an index of the status of algae upon exposure to toxic
chemical substances. Metody biologicheskikh issledovanii po vodnoy
toksikologii , Moscow, Nauka Press, pp. 106-107.
Khobot'yev, V.G., V.I. Illarionova, and T.N. Yunasova. 1971. Method of
determination of living and dead cells of blue-green and green algae by
means of dyes. Metodiki biologicheskikh issledovanii do vodnoy
toksikologii, Moscow, Nauka Press, pp. 181-82.
Khobot'yev, V.G. and V.I. Kapkov. 1971. Cultivation of green algae and
their use in toxicologic experiments. Metodiki biologicheskikh issle-
dovanii po vodnoy toksikologii, Moscow, Nauka Press, pp. 219-31.
71
Konstantinov, A.S. 1958. Biology of the Chironomidae and their breeding.
Jr. Saratovskogo otd. GosNIORKh, Vol. 5, Saratov.
Krylov, O.N. 1973. Change in certain physiologic indexes of fish, eggs and
larvae under the influence of chlorophos, entobacterin and the lepidop-
ter nuclear polyedrosis virus. Eksper. vodn. toksikologiya, Vol. 5,
Riga, USSR, pp. 97-116.
Lesnikov, L.A. 1971. Method of estimating the influence of water from
natural reservoirs on Daphnia magna Straus. Methods of biological
studies of water toxicology, Moscow, Nauka Press, pp. 157-155.
Lesnikov, L.A. 1973. Methodologic instructions for establishment of maxi-
mum permissible concentrations of harmful substances in bodies of water
used for fishing. Leningrad.
Lesnikov, L.A. 1975. A research system for developing fisheries standards
for water quality, considering the peculiarities of transferring experi-
mental data to natural water bodies. Proc. Second USA-USSR Symposium on
the Effects of Pollutants Upon Aquatic Ecosystems, June 22-25, Borok,
USSR, EPA-600/3-78-076.
Lisovskaya, E.V., L.V. Grigor'yeva, and Z.I. Zholdakova. 1968. Sanitary-
hygienic and toxicologic prerequisites for use of monuron to combat
"blooming" of water in drinking reservoirs. Tsveteniye vody (Blooming
of water), Kiev, Naukova dumka Press, pp. 294-300.
Lloyd, L.S., Jr. 1976. Chronic effects of low levels of hydrogen sulfide
on freshwater fish. Proc. of First and Second USA-USSR Symposium on
the Effects of Pollutants Upon Aquatic Ecosystems, EPA-600/3-78-076,
113-21.
Lloyd, L.S., Jr., D.M. Oseid, I.R. Adelman, and S.J. Brederius. 1976. Ef-
fect of hydrogen sulfide on fishes and invertebrates. Part 1. Acute
and chronic toxicity studies. EPA-600/3-75-062a.
Macek, K.J., M.A. Lindberg, S. Sauter, K.S. Baxton, and P. A. Cesta. 1976.
Toxicity of four pesticides to water fleas and fathead minnows, acute
and chronic toxicity of acrolein, heptachlor, endosulfan and trifluralin
to water flea (Daphnia magna) and fathead minnow (Pimephales promelas).
EPA-600/3-76-099.
Mancy, K.H. and H.E. Allen. 1977. A controlled bioassay system for mea-
suring toxicity of heavy metals. EPA-600/3-77-037.
Mayer, F.L., P.M. Mehre, and W.P. Dwyer. 1975. Toxaphene effects on re-
production, growth and mortality of brook trout. EPA-600/3-75-013.
Mayer, F.L. 1976. Residue dynamics of di-2-ethylhexylphthalate in fathead
minnows (Pimephales promelas). J. Fish, Res. Board Canada, Vol. 33,
2610-13.
72
McK
ee, J
.E.
and
H.W.
, Wolf.
1963.
Res.
Ag.
, of
Cali
if., Publ
. No.
Mayer, F.L., P.M. Mehrel, and R.A. Schoettger. 1977. Collagen metabolism
in fish exposed to organic chemicals. In: R.A. Tubb, Ed., Recent ad-
vances in fish toxicology: A symposium. Ecol. Res. Ser. No. EPA-600/3-
77-085.
Mayer, F.L., P.M. Mehrel, and W.P. Dwyer. 1977. Toxaphene: Chronic toxi-
city to fathead minnow and channel catfish. EPA-600/3-77-069.
Water quality criteria. Second Ed., The
3A.
McKim, J.M., G.F. Olson, and G.W. Holceme. 1976. Long-term effects of
methylmercuric chloride on three generations of brook trout (Salvelinus
fontinalis) : Toxicity, accumulation, distribution and elimination. J.
fish. Res. Board Canada, Vol. 33, No. 12, 2726-39.
McKim, J.M. 1977. Evaluation of tests with early life stages of fish for
predicting long-term toxicity. J. Fish. Res. Board Canada, Vol. 34, No.
8, 1148-54.
McKim, J.M., J.G. Eaton, and G.W. Helcombe. 1978. Metal toxicity to em-
bryos and larvae of eight species of freshwater fish: 11. Copper.
Bull. Environ. Contam. Toxicol,, Vol. 19, 608-16.
Mehrle, P.M. and F.L. Mayer. 1975. Toxaphene effects on growth and bone
composition of fathead minnows (Pimephales promelas).
Mosevich, M.V. 1973. Methodologic instructions on microbiologic studies to
determine the influence of pollutants and for experimental determination
of the course of bacterial processes of self-purification in water.
GosNIORKh Press, Leningrad, 20 pp.
Mosiyenko, T.K. 1974a. Methodologic instructions for conduct of toxico-
logic experiments on algae. GosNIORKh Press, Leningrad, 16 pp.
Mosiyenko, T.K. 1974b. Influence of refinery wastewaters after physical-
chemical and biologic purification upon subsequent holding in settling
ponds on algae - Scenedesmus quadricanda (Tukhr) Breb. Izv. GosNIORKh,
Vol. 98, pp. 55-60.
Patin, S.A. and V.N. Tkachenko. 1974. The radiocarbon method in toxico-
logic studies of marine and fresh-water phytoplankton. Izv. GosNIORKh,
Vol. 98, pp. 141-143.
Patin, S.A. 1977. Ecologic toxicity and biogeochemistry of pollutants in
the world ocean. Auth. Abst. Dr. Diss., Moscow.
Pickering, G.W., W.A. Brungs, and M. Gast. 1977. Effect of exposure time
and copper concentration on reproduction of the fathead minnow (Pime-
phales promelas). Water Research, Vol. 11, 1079-83.
73
Sanders, H.O. and O.B. Cope. 1968. The relative toxicities of several
pesticides to naiads of three species of stone flies. Limnol. Oceanogr.
Vol. 12, No. 1.
Sanders, H.O., F.L. Mayer, and D.F. Walsh. 1973. Toxicity, residue dynam-
ics, and reproductive effects of phthalate esters in aquatic inverte-
brates. Environ. Res., Vol. 6, 84-90.
Sanders, H.O. 1977. Toxicity of the molluscocide Bayer 73 and residue
dynamics of Bayer 2353 in aquatic invertebrates. Invest. Fish Control,
Vol. 78, 1-7.
Sauter, S., K.S. Buxten, K.J. Macek, and S.R. Petrocelli. 1976. Effects of
exposure to heavy metals on selected freshwater fish. Toxicity of cop-
per, cadmium, chromium, and lead to eggs and fry of seven fish species.
EPA-600/3-75-105.
Snarski, V.M. and F.A. Publisi. 1976. Effects of Aroclor R 1254 on brook
trout (Salvelinus fontinalis). EPA-600/3-76-112.
Spehar, R.L. 1976. Cadmium and zinc toxicity to Jordanella florida. EPA-
600/3-76-096.
Spehar, R.L., R.L. Anderson, and J.T. Fiandt. 1978. Toxicity and bioac-
cumulation of cadmium and lead in aquatic invertebrates. Environ.
Pollut. No. 15, 198-208.
Stroganov, N.S. 1971. Method of determination of toxicity of water media.
Metodiki biologicheskikh issledovanii po vodnoi toksikologii , Moscow.
Stroganov, N.S. and L.V. Kolosova. 1971. Laboratory culture and determina-
tion of fertility of daphnia over a number of generations. Metodika
biologicheskikh issledovanii po vodnoy toksikologii, Moscow Nauka Press,
pp. 210-16.
Vislyanskaya, I.G. and S.V. Vedyagina. 1974. Influence of cellulose-paper
industry wastewater on phytoplankton under laboratory modeling condi-
tions. Izv. GosNIORKh, Vol. 98, pp. 149-52.
Walsh, G.E. 1972.. Effects of herbicides on photosynthesia and growth of
marine unicellular algae. Hyacinth Control J., Vol. 10, 45-48.
Walsh, G.E., K. Ainsworth, and A.O. Wilson. 1977. Toxicity and uptake of
kerone in marine unicellular algae. Chesapeake Sci., Vol. 18, No. 2.
74
SECTION 6
CHLORINATED HYDROCARBONS AS A LIMITING FACTOR IN THE
REPRODUCTION OF LAKE TROUT IN LAKE MICHIGAN^
Wayne A. Willford^
THE FISHERY
From about 1890 until 1945, the lake trout (Salvelinus namaycush) was
the most valuable and sought-after commercial species in Lake Michigan. The
annual commercial catch averaged 8.2 million pounds (3,700 metric tons [t])
from 1890 to 1911, 7.0 million pounds (3,200 t) from 1912 to 1926, and 5.3
million pounds (2,400 t) from 1927 to 1939. The catch increased slightly to
an annual average of 6.6 million pounds (3,000 t) during 1940 to 1944, but
then began to decline precipitously in 1945 and had fallen to only 342,000
pounds (155 t) by 1949 (Figure 1). In 1954, the catch was a mere 34 pounds
(15 kg), and by 1956 the species was probably extinct in Lake Michigan
(Wells and McLain 1973).
The gradual decline in the commercial harvest of lake trout from 1893 to
1938 is believed to have resulted from excessive exploitation (Van Oosten
1949; Wells and McLain 1973). Although the commercial harvest of lake trout
continued into the early 1950's, the apparent extinction of the species in
about 1956 is believed to have been caused directly by the predatory sea
lamprey (Petromyzon marinus), an exotic species that became firmly estab-
lished in Lake Michigan in the decade following its first reported presence
there in 1936 (Wells and McLain 1973).
Early attempts to control the sea lamprey consisted of installing elec-
trical and mechanical barriers, which blocked the spawning runs of adults.
Between 1953 and 1958, barriers were constructed across 65 tributaries
flowing into Lake Michigan. At about the same time (in the late 1950's) a
successful lampricide, 3-trif luoromethyl-4-nitrophenol (TFM), was discovered
and developed by scientists at the Hammond Bay Biological Station of the
U.S. Fish and Wildlife Service (USFWS). This compound was soon being used
to kill larval sea lampreys (ammocoetes) in tributary streams before they
could metamorphose and migrate downstream into the lake. Most barrier
operations were discontinued in 1960 in favor of TFM treatments, thus set-
^Contribution 545, Great Lakes Fishery Laboratory.
^U.S. Fish and Wildlife Service, Great Lakes Fishery Laboratory, 1451 Green
Road, Ann Arbor, Michigan 48105.
75
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76
ting the stage for the highly successful sea lamprey control program which
followed. This program and the ongoing lake trout restocking program, which
began in 1955 in Lake Michigan when about 1.3 million yearling lake trout
were planted, have been coordinated by the Great Lakes Fishery Commission.
In 1965-78, an average of over 2 million fin-clipped lake trout were planted
in the lake each year (data provided by the Great Lakes Fishery Commission)
as part of an effort to restore lake trout stocks to self-sustainability.
By the early 1970's, the lake trout were once again considered abundant
in Lake Michigan and spawning activity was widespread each fall (Wells and
McLain 1973; Great Lakes Fishery Laboratory 1974). Nevertheless, no
naturally produced fingerling or older lake trout (recognizable by their
lack of clipped fins) have been found in the lake during routine assessment
sampling (Great Lakes Fishery Laboratory 1978). Therefore, little progress
has been made toward the goal of rehabilitating self-sustaining stocks of
lake trout, even though the lake contains a large population of mature fish
that should be ^capable of reproducing naturally.
REHABILITATION PROBLEMS
Following the reports of widespread spawning of lake trout in the early
1970's, concern deepened about the apparent failure of the fish to produce
surviving progeny. Numerous theories have been proposed to account for this
reproductive failure, including the following:
1. Contamination of the water and fish by toxic substances such
as pesticides and industrial chemicals;
2. Deterioration in bottom conditions on spawning reefs as a re-
sult of eutrophi cation and possibly increased sedimentation;
3. "Abnormal homing" of planted trout as spawning adults to their
planting sites—generally shallow, inshore areas that offer
little suitable spawning substrate and are vulnerable to sedi-
mentation or scouring action by waves and ice;
4. Predation on, or feeding competition with, young lake trout by
the now abundant, introduced species, rainbow smelt (Osmerus
mordax) and alewife (Alosa pseudoharengus);
5. Artificial selection, extensive inbreeding, or physiological and
behavioral conditioning of hatchery fish which somehow resulted
in their inability to spawn successfully or to produce young
that are capable of surviving in the wild; and
6. Insufficient "critical mass" or numbers of mature lake trout in
the lake to permit the realistic expectation of successful re-
production in the early 1970's.
Various studies addressing these theories were soon initiated by the
Michigan Department of Natural Resources (MDNR) and the USFWS Great Lakes
77
Fishery Laboratory (Rybecki and Keller 1978). Of greatest concern initially
was the problem of toxic substances. The fish were known to contain sub-
stantial residues of DDT and its metabolites and of PCBs (Reinert 1970;
Stalling and Mayer 1972). Concentrations of each of these contaminants ex-
ceeded 10 yg/g in adult lake trout (Willford 1975) and 4 ug/g in their eggs
(Reinert and Bergman 1974). Published reports on the effects of DDTs (DDT,
DDD, and DDE) and PCBs indicated that the concentrations of these contami-
nants in lake trout and their eggs were sufficient to interfere with repro-
duction. For example, Burdick et a_l_. (1964) reported that concentrations of
DDTs in excess of 2.9 yg/g in the eggs of lake trout resulted in increased
mortality of fry. This effect was later confirmed by Macek (1968) who
studied brook trout (Salvelinus fontinalis) fed DDT. Unusually high morta-
lity of fry of coho salmon (Oncorhynchus kisutch) hatched from eggs of Lake
Michigan fish, and possible correlation of that mortality with elevated
levels of DDTs and other chlorinated hydrocarbons were also reported
(Johnson and Pecor 1969; Willford et aj[. 1969). In addition, reduced hatch-
ability of salmon eggs in Sweden was reported as correlated with elevated
PCB residues (Jensen et a/[. 1970), Nevertheless, hatchery records showed
that when eggs of planted Lake Michigan lake tro"t were manually stripped,
fertilized, and hatched, and the fry were reared in hatcheries, survival was
"normal" or "satisfactory" (Stauffer 1979).
HATCHABILITY OF EGGS
In 1972-73, researchers at the Great Lakes Fishery Laboratory performed
studies to investigate further the hatchability of eggs from Lake Michigan
lake trout under three sets of incubation conditions: normal hatchery con-
ditions; a thermal regime similar to that of winter and spring in Lake
Michigan; and the thermal and chemical conditions characteristic of water
from the Hammond Bay Biological Station's intake on Lake Huron. Related
studies were carried out by the MDNR at the Marquette State Fish Hatchery,
at the Thompson State Fish Hatchery, and at two locations (in egg-holding
enclosures) in Lake Michigan's Grand Traverse Bay from 1973 to 1976
(Stauffer 1979). In all of these studies, the survival of contaminated eggs
and fry from Lake Michigan lake trout was compared with that of relatively
uncontami nated eggs and fry from hatchery brood stock. Although occasional
differences in survival were noted between groups of eggs and fry reared
under the various experimental conditions, no consistent relation between
hatching success and the concentrations of DDTs or PCBs in the eggs was ap-
parent. The conclusion reached in the studies performed at the several lo-
cations by the two agencies was that existing levels of DDTs and PCBs in
eggs of Lake Michigan lake trout did not significantly affect survival in
eggs or of early stages of the fry.
The reproductive failure of lake trout in th" lake was nevertheless
still apparent in the mid 1970's. We then speculated that although the eggs
could hatch and the fry survive in a clean (hatchery or laboratory) environ-
ment, the additional chronic exposure to PCBs and DDE in the water and food
of Lake Michigan might reduce the stamina, strength, or wariness of the fry
sufficiently to preclude their survival in the rigorous lake environment.
78
SURVIVAL OF FRY
To test this hypothesis, we began a 6-month study in the winter of 1975-
76 on the effects of chronic exposure of fry of Lake Michigan lake trout to
PCBs and DDE. In addition to routine observations on mortality and growth
of the fry, we also evaluated methodology for, and made measurements of,
their temperature preference, swimming performance, predator avoidance, and
metabolism. About 27,000 eggs were manually stripped and fertilized with
milt from lake trout (about 10 females and 20 males) gillnetted in south-
eastern Lake Michigan near Saugatuck, Michigan in the fall of 1975. Con-
taminant levels in adult lake trout from this area had been monitored for
several years and the fish were known to contain average whole-body concen-
trations of about 22 yg/g PCBs, 7.5 yg/g total DDT, and 0.3 yg/g dieldrin
(Great Lakes Fishery Laboratory, unpublished data). Our analysis of eyed
eggs sampled from those collected for this study revealed 7.6 yg/g PCBs and
4.7 yg/g total DDT. Samples of 1 -day-old sac fry hatched from these eggs
and analyzed at the USFWS Columbia National Fishery Research Laboratory were
shown to contain 3.8 yg/g PCBs (Aroclor 1254), 2.3 yg/g total DDT, 0.06 yg/g
dieldrin, 0.12 yg/g cis-chlordane, and about 5.7 yg/g of a chemical re-
sembling toxaphene. Later analysis showed that the toxaphene-like residue
was actually composed of several chlorinated camphenes of undetermined ori-
gin.
The fry were then exposed for 6 months to 10.0 ng/1 PCBs (Aroclor 1254)
and 1.0 ng/1 DDE in water, and 1.0 yg/g PCBs and 0.1 yg/g DDE in food.
These values approximate the exposure received by fish in the lake as
determined by analyses of water and plankton collected offshore in south-
eastern Lake Michigan. Concentrations 5 and 25 times these values were also
tested to allow dose-effect interpretation and prediction of potential
effects on fry hatched in the more contaminated, nearshore areas of the
lake.
About a week after the eggs hatched, grossly deformed fry were discarded
and the rest were equally divided among 30 tanks (650 fish per tank) in a
constant-flow bioassay system. Serial diluters supplied the appropriate
concentrations (Ix, 5x, 25x, and control) of the contaminants singly or in
combination in 9 C well water. The experimental design thus provided for 10
different treatments (including the controls) and three replicates of each.
Following 11 days of exposure, the fry began to exhibit feeding behavior and
were fed the corresponding dosage of either or both contaminants that had
been added to their food. Analyses of water during the study showed that
the actual average exposures received by the fry corresponding to Ix, 5x,
25x were 20.8, 64.7, and 327 ng/1 PCBs and 1.8, 6.3, and 32.7 ng/1 DDE.
Analyses of the food showed that actual concentrations were all within 28%
of agreement with nominal concentrations.
During the first 16 days of exposure to the three levels of PCBs, DDE,
and PCBs plus DDE in water, the percentages of fry that died ranged from 1.9
to 3.7% across all treatments. Mortalities of fry among the nine exposed
groups were not significantly different from the percentage that died among
the controls. During the next 40 days (days 17-56), when exposed fry began
receiving contaminants in their food as well as from the water, the morta-
79
lity rate in the simulated Lake Michigan exposures (Ix) ranged from 2.2 to
3.9%, that in the 5x exposures ranged from 3.5 to 5.9%, and that in the 25x
exposures ranged from 7.5 to 24.2%. The mortality rate of control fry
(7.3%) was higher during this period than that of fry in the Ix or 5x expo-
sures.
During the second 40-day period (days 57-96), which began about 2 weeks
after completion of yolk absorption, mortality of fry increased signifi-
cantly (P^<0.01) in both the exposed and control groups. This increase was
most dramatic, however, among the exposed groups of fry. Mortality rates
for all nine exposed groups during this period (19.0 to 35.4%) were signi-
ficantly higher (P <0.01) than in the controls (11.2%). By the end of the
third 40-day period (days 97-136), the rates of mortality decreased in all
treatments when compared with the previous period but mortality rates in all
nine exposed groups (4.5 to 13.4%) nevertheless remained significantly
higher (P <0.01) than in the controls (1.3%). Mortality rates further
leveled off during the fourth 40-day period (days 137-175), but the final
cumulative mortality for each of the nine exposed groups was significantly
higher (£ <0.01) than that for the controls. The average total cumulative
mortality on day 176 in each of the exposed groups ranged from 30.5 to
46.5%, whereas that in the control group was only 21.7%.
Especially noteworthy was the final cumulative mortality of fry in the
Ix combination exposure of PCBs and DDE (simulated Lake Michigan exposure) —
40.7% or nearly double the final cumulative mortality of the controls (Fi-
gure 2). This result suggests that if lake trout in Lake Michigan spawned
successfully and their eggs hatched, nearly twice as many of the resulting
fry would die within the first 6 months than would have died if these con-
taminants had not been present. In nearshore areas, where contaminant
levels are generally higher, the potential impact on fry mortality would be
expected to increase. At the highest combined level of PCBs and DDE tested
(25x), 46.5% of the fry died.
PHYSIOLOGY OF FRY
In addition to observations on the mortality of fry during the chronic
exposure, observations were made periodically on the growth, swimming per-
formance, predator avoidance, temperature preference, and metabolism of the
fry. In general, the exposed fry showed no significant physiological
effects attributable to the exposure. Although occasional differences were
noted in the swimming performance and in certain metabolic measurements such
as oxygen consumption rates and whole-body lactate concentrations after
swimming, the results were inconclusive because the variability of the data
was high. Procedural difficulties prevented the testing of temperature pre-
ference at the Ix and 5x exposures; nevertheless, fry exposed to 25x DDE and
25x DDE and PCBs in combination for 4 months preferred significantly lower
(P^ <0.05) temperatures (9.8 C and 8.7 C, respectively) than did the controls
(11.2 C). Because of the inconclusiveness of the observations on the
general condition or performance of the fry, together with the inherent dif-
ficulty of interpreting the impact of these sublethal effects on the pro-
80
60
50
40
^ 30
3 20
O
^ 10
25 X
CONTROL
40
80
1 20
160
200
DAYS OF EXPOSURE TO DDE & PCBs
Figure 2. Mortality of fry of Lake Michigan lake trout exposed to
DDE and PCBs at concentrations simulating those found in water
and plankton of Lake Michigan (Ix) and at concentrations 5
and 25 times higher.
81
ductivity of fish populations, the increase in mortality was clearly the
most sensitive and meaningful observation of effect measured in the study.
CONCLUSIONS
The significant increase in mortality of lake trout fry during 6 months
of exposure to levels of DDE and PCBs in food and water similar to those in
Lake Michigan strongly suggests that these chlorinated hydrocarbons are a
limiting factor in the reproduction of lake trout in the lake. Whether
these two contaminants are the sole or even major cause for reproductive
failure of the lake trout is unclear. Other factors such as the presence
of exotic species and the spawning behavior of planted fish undoubtedly play
a role. The known presence, however, of additional chlorinated hydrocarbons
such as dieldrin, chlordane, and chlorinated camphenes, as well as of
several other organic and inorganic contaminants in the water and biota of
the lakes, raises serious questions about the potential additive or syner-
gistic effects of these multiple contaminants. Regardless of the ultimate
answer to these questions, the current levels of PCBs and DDE in the lake
appear sufficient to impede the restoration of self-sustaining populations
of lake trout in Lake Michigan.
ACKNOWLEDGEMENTS
The studies and conclusions reported here resulted from the dedicated
and professional effort of the entire staff of the Section of Physiology
and Contaminant Chemistry, Great Lakes Fishery Laboratory, Special credit
goes to Robert E. Reinert for initially identifying chlorinated hydrocarbons
as a potential problem in Lake Michigan and for directing the early studies
on hatchability of lake trout eggs. Principal investigators in the studies
I discussed were William H. Berlin, Roger A. Bergstedt, Robert J.
Hesselberg, Michael J. Mac, Dora R. May Passino, and Donald V. Rottiers.
The assistance of Lawrence W. Nicholson and James R. Olson in providing
chemical analyses for most of the studies, and of Neal R. Foster and Thomas
L. Baucom in editing this report is gratefully acknowledged.
REFERENCES
Burdick, G.E., E.J. Harris, H.J. Dean, T.M. Walker, J. Skea, and D. Colby.
1964. The accumulation of DDT in lake trout and the effect on repro-
duction. Trans. Am. Fish. Soc. 93(2): 127-136.
Great Lakes Fishery Laboratory. 1974. Great Lakes Fishery Program. D]_
Sport Fishery and Wildlife Research 1972, pp. 22-32, U.S. Department
of the Interior, Bureau of Sport Fisheries and Wildlife. 124 pp.
Great Lakes Fishery Laboratory. 1978. Great Lakes Fisheries. ln_ Sport
Fishery and Wildlife Research 1975-76, pp. 46-57, U.S. Fish and
Wildlife Service. 140 pp.
82
Jensen, S., N. Johansson, and M. Olsson. 1970. PCB--Indications of effects
on salmon. PCB Conference, Stockholm, September 29, 1970. Swedish
Salmon Research Institute-Report LFI MEDD 7/1970.
Johnson, E., and C. Pecor. 1969. Coho salmon mortality and DDT in Lake
Michigan. Trans. N. Am. Wildl. Nat. Resources Conf. 34: 159-166.
Macek, K.J. 1968. Reproduction in brook trout (Salvelinus fontinalis) fed
sublethal concentrations of DDT. J. Fish. Res. Board CarT 25(9) : 1787-
1796.
Reinert, R.E. 1970. Pesticide concentrations in Great Lakes fish. Pestic.
Monit. J. 3(4): 233-240.
Reinert, R.E., and H.L. Bergman. 1974. Residues of DDT in lake trout
(Salvelinus namaycush) and coho salmon (Oncorhynchus kisutch) from the
Great Lakes. J. Fish. Res. Board Can. 31: 191-199.
Rybicki, R.W., and M. Keller. 1978. The lake trout resource in Michigan
waters of Lake Michigan, 1970-1976. Mich. Dept. Nat. Resour. Fish.
Res. Rep. No. 1863. 71 pp.
Stalling, D.L., and F.L. Mayer, Jr. 1972. Toxicities of PCBs to fish and
environmental residues. Dn Environmental Health Perspectives, Experi-
mental Issue Number One, April 1972, Douglas H.K. Lee and Hana L. Falk,
Eds., pp. 159-164. National Institute of Environmental Health Sciences,
Research Triangle Park, N.C.
Stauffer, T.M. 1979. Effects of DDT and PCBs on survival of lake trout
eggs and fry in a hatchery and in Lake Michigan 1973-1975. Trans. Am.
Fish. Soc. 108: 178-186.
Van Oosten, J. 1949. A definition of depletion of fish stocks. Trans. Am.
Fish. Soc. 76: 283-289.
Wells, L., and A.L. McLain. 1973. Lake Michigan: Man's effects on native
fish stocks and other biota. Great Lakes Fishery Commission, Technical
Report No. 20. 55 pp.
Willford, W.A., J.B. Sills, and E.W. Whealdon. 1969. Chlorinated hydro-
carbons in the young of Lake Michigan coho salmon. Prog. Fish-CuU.
31(4): 220.
Willford, W.A. 1975. Contaminants in Upper Great Lakes fishes. In Plenary
Sessions, Upper Great Lakes Committee Meetings, Appendix V, Milwaukee,
Wisconsin, March 25-25, 1975, pp. 31-39. Great Lakes Fishery Commis-
sion, Ann Arbor, Michigan.
83
SECTION 7
ORGANOPHOSPHORUS PESTICIDES AND THEIR HAZARDS TO AQUATIC ANIMALS
V.I. Kozlovskaya and B.A. Flerov^
Recently, as replacements for DDT and other persistent organochlorine
insecticides, a variety of organic phosphorus compounds have been synthe-
sized. At present, world wide utilization of organophosphorus pesticides
involves more than 150 compounds (Melnykov, et a]_. 1977). As a result of
their large-scale production and use, this group of toxicants requires
investigating.
Pesticides enter the water bodies with the industrial wastes, with the
flows from water collectors, with the waters from drainage systems, and from
the runoff and overcarriage of the spraying of fields from airplanes.
Organophosphorus pesticides were found in the Kuban River in 7 out of 8
sites examined. Their concentrations varied from 0.04 to 0.3 mg/.' (Table
1). In 224 water samples obtained in ponds and rivers of different regions
TABLE 1. THE AVERAGE WEIGHT OF ORGANOPHOSPHORUS PESTICIDES AT STATIONS
IN THE KUBAN RIVER (1967-1974)
■
■
Name of Observation Point
Concentration, mg/l
Karatshayevsk
_
Tsherkassk
0.218
Nevynnomyssk
0.087
Armavir
0.294
Kropotkin
0.246
Krasnodar
0.037
Temryuk (the Petrushkin arm)
0.067
Atshuyevo (the Protok arm)
0.205
■
of the Ukraine, organophosphorus compounds were present in 73. Similarly,
they were found in 30 out of 216 samples of bottom deposits (Kostovetsky,
et al_. 1976). In reservoirs of the southland west regions of Slovakia,
malathion and sumithion found in amounts of 0.5 - 1 mg/2. (Bilikova 1973).
1 Institute of Biology of Inland Waters, Academy of Science, USSR, Borok,
Nekouz, Jaroslavl, 152742, USSR.
84
Since organophosphorus pesticides are easily dissolved in water, heavy
rains contribute to their intensive runoff from agricultural fields to re-
servoirs. For example, after a rainfall of 2.1 vm, the phosalon content of
the water body located near an orchard treated with this chemical exceeded
the permissible concentrations by 7 to 9.6 times, and after a rainfall of
21.1 nm a 12-fold excess was reported (Ivantshenko 1978).
Decomposition of organophosphorus compounds in water compared with or-
ganochlorine compounds occurs very rapidly (Table 2). The time of degrada-
TABLE 2. PERSISTENCE OF SELECTED ORGANOPHOSPHORUS PESTICIDES IN WATER
Pesticide
Type
Concentration
mg/£
Period of Complete
disappearance in
days
Reference
Metaphos
0.02
'0.2
1-2
2.5
3-5
8-14
55
160
Kostovetsky, et al. 1976
Kostovetsky, et al . 1976
Ulyanova, et aT. 1979
Ulyanova, et al . 1979
Dylox
0.05
0.5
1
10
Kostovetsky, et al. 1976
Kostovetsky, et al. 1976
Malathion
0.1
0.5
14
6-11
Drevenkar, et al . 1975
Kostovetsky, et al . 1976
Bazudin
0.6
6.0
60.0
16
21
35
Boyko and Pulatov, 1977
Boyko and Pulatov, 1977
Boyko and Pulatov, 1977
DDVP
0.1
n
Drevenkar, et al ., 1975
tion depends on the concentration of hydrogen ions, and temperature
(Melnykov, et al_. 1977); and it is dependent upon the number of bacteria de-
composing these compounds (Ulyanova, et al_. 1979).
Both the intensity and duration of effects upon water bodies are pri-
marily determined by the length of time that pesticides stay in the soil of
catchment areas. Depending on the type of soil, humidity, and pH, pesti-
cides may be retained for extended periods of time and with surface water
flows, enter reservoirs (Table 3).
Organophosphorus pesticides in concentrations most commonly found in
water bodies, show a high toxicity for aquatic animals, especially for
planktonic invertebrates and aquatic insects (Table 4). In 48-hour expo-
sures to 0.001 mg/l solutions of malathion, Simocephalus vetulus became less
mobile and died after being placed in freshwater for recovery. The 48-hour
LC50 for the eggs of carp is approximately 0.01 mg/£, but for their larvae
the value is ten times as high (Prokopenko, et aj_. 1976). Eight day larval
forms of freshwater invertebrates demonstrate depression changes after three
85
TABLE 3. PERSISTENCE OF SELECTED ORGANIC PESTICIDES IN SOIL
Period of complete
Pesticides disappearance in days Reference
Metaphos 10-150 Korotova and Demtshenko, 1978a
Kostovetsky, et a_l_. , 1976
Yurovskaya and Jhulinskaya, 1974
Dylox 4-45 Korotova and Demtshenko, 1978b
Kostovetsky, et a]_., 1976
Yurovskaya anH~Jhulinskaya, 1974
Malathion 7-60 Kostovetsky, et al_,, 1976
Keazney, et aj^. , 1969
Novozhylov, et £l_., 1974
Diazinon 85 Keazney, et al_., 1969
Takase, 1976
Phosalon 18-90 Manko, et al, 1974
TABLE 4. TOXICITY OF ORGANOPHOSPHORUS PESTICIDES TO AQUATIC ANIMALS
(From Water Quality Criteria, 1972, EPA-R-73-033, 1973)
96-hour LC
, mg/?
Animal species
Guthion
Malathion
Parathion
Dylox
Gammarus lacustris
0.00015
0.001
0.0035
0.04
Gammarus fasciatus
0.0001
0.00076
0.0021
-
Asellus brevicaudus
0.021
3
0.6
-
Daphnia pulex
-
0.0018
0.0006
0.00018
Pteronarcys dorsata
0.0121
-
0.003
-
Pteronarcys californica
0.0015
0.01
0.036
0.069
Acroneuria lycorias
-
0.001
-
-
Acroneuria pacifica
-
-
0.003
0.0165
Salmo gairdneri
0.014
0.17
-
-
Salmo trutta
0.004
0.2
-
-
Oncorhynchus kisutch
0.017
0.101
-
-
Lepomis macrochirus
0.0052
0.11
0.065
3.8
Pimephales promelas
0.093
9
1.41
109
86
exposures to malathion in concentrations of 0.002 and 0.02 mg/£. Chironomid
and mayflies also decrease considerably (Kennedy and Walsh 1970).
Lesnikov (1974) suggests that the most sensitive indicator of the ef-
fects produced by organophosphorus compounds is an increase of both popula-
tion and biomass of aquatic organisms (Table 5).
TABLE 5. DYLOX TOXICITY (mg/£) FOR SELECTED AQUATIC ORGANISMS
Animal species
Acute
Toxicity
Chronical
Effect on the increase
of biomass
Daj)hnia longispina
Gammarus pulex
Salmo irideus
0.0005
0.5
0.121
0.0001
0.1
0.06
0.00002
0.03
0.004
The hazards of organophosphorus pesticides are even greater since ani-
mals demonstrate poor avoidance reactions to these chemicals. Some inverte-
brates do not avoid Dylox at all (Hirudo medicinalis), or some such as
Asellus aquaticus and Stephocephalus torvicornis avoid it only in concentra-
tions of 250 to 1000 times higher than their 48-hour LC50 (Flerov and
Lapkina 1976; Tagunov and Flerov 1978; Flerov and Tagunov 1978). Guppies
demonstrate avoidance reactions to Dylox at concentrations equal to the 48-
hour LCioo (Flerov 1979).
Shrimp (Palaemr"^^-'-:, pugio) fail to avoid malathion (Hansen, et a1 .
1973), and mosquito fish avoid it only in acutely toxic concentrations
(Hansen, et al_. 1972).
The toxicity of organophosphorus compounds is attributed to their
ability to inhibit acetyl cholinesterase irreversibly, which in turn, de-
pends upon the particular enzyme system in the animals.
Thus, the two species of gastropods (Limnaea stagnalis and Planorbis
corneus) differ in resistance to Dylox by 100 times. The nervous ganglia of
these forms contain enzymes of the acetyl cholinesterase type, that
hydrolyze the same substrates, but differ in quantity, electrophometic
mobility and sensitivity to the toxicant. In vitro experimentation with
the sensitive species (Limnaea stagnalis) showed concentrations of 10"^
to 10"^M Dylox completely inhibited enzyme activity, lO'^M inhibited by
97 percent, and lO'^M inhibited by 61 percent. In the resistant species,
Planorbis corneus, even concentration of lO'^^M of toxicant did not inhibit
the enzyme completely, although the enzyme content in ganglia of this
species is much lower than in Limnaea stagnalis (Figure 1 and 2).
The correlation between the resistance of organism to the toxicant and
the sensitivity of an enzyme to it has also been observed in fish. The
roach, Rutilus rutilus, and the blue bream, Abramis ballerus, are poorly
resistant to Dylox. Their blood sera contains an enzyme of the cholinester-
ase type which is absent in more resistant fish, such as the carp, Cyprinus
87
0.050-
>
d 0.195
OQ
O
o
LU
cc
O
X
o
cc
H
O
LLI
_1
LU
>
UJ
CC
e
■ . 48-Hour LC50,
Mo/ use species ^ ,j ^^
1 L. stagnalis
2 P. corneus
0.5
50.0
Figure 1. Acetychol inesterase in nervous ganglia of molluscs
with varying resistance to Dylox.
1 - Limnaea stagnalis, LC50 " ^-^ "^S/l 48-hrs. exposure,
2 - Planorbis corneus, LC50 - 50 mg/1 48-hrs. exposure.
88
100
>
80
>
1-
. — .
O
O
<
1-
4-'
LU
c
o
60
^
o
>
o
N
■»-•
LU
>
40
H
i2
LU
— '
o
DC
LU
90
Q_
^u
0
I I Limnaea stagnalis
Planorbis corneas
10-4 10-5 10-6
DYLOX CONCENTRATION, mg/
Figure 2. Inhibition by Dylox of acetychol inesterase in nervous
gang! ia of Limnaea stagnalis and Planorbis corneus.
89
carpio and the bream, Abramis brama. The latter contains an enzyme of the
acetyl cholinesterase type (Kozlovskaya and Tshuyko 1979).
As intoxication by organophosphorus pesticides advances, the animals ex-
hibit a progressive decline in the level of cholinesterase, although in
dying animals the enzyme may not be entirely inhibited. Such facts are
cited in a number of reviews (0' Brian 1964; Rosengart and Sherstobitov
1978).
After acute exposure of perch (Perca f luviatilis) to Dylox (48-hour
LCioo 0^ 5 mg/ ; 48-hour LC50 of 0.62 mg/£) fish were assayed immediately
after death, 8 and 33 hours of the experiment, respectively. The cholin-
esterase activity in these cases was partially retained (up to 25 percent).
In fish which were left in the toxic environment after death for a few
hours, the enzyme was inhibited to a greater extent (Figures 3a and 3b).
Similar results were obtained in experiments with carp (Carassius carassius)
and pond snails (Limnaea stagnalis) . Densitometry of electrophorograms
showed that not all molecular forms of the enzyme were completely inacti-
vated (Figures 4a and 4b). It appears that the toxicant interacts with
vitally important forms of the enzymes.
The inhibitation of AChE in the brain of perch has been also observed at
sublethal concentrations, although the external symptoms of poisoning were
absent (Table 6). Upon placing the animals in freshwater, the gradual re-
activation of enzymes took place.
TABLE 6. CHANGES IN THE ACETYL CHOLINESTERASE ACTIVITY OF THE PERCH BRAIN
IN THE MINIMUM TOLERABLE CONCENTRATIONS OF DYLOX (0.12 mg/J) WITH
SUBSEQUENT WASHING IN FRESHWATER
Enzyme of Activity
Number of % of the
Exposure samples yM AChE g/h control
Exposure in the Dylox 1-10 427.9 + 0.84 87.2
Solution 5-10 339.3+0.79 67.6
One day exposure in fresh- 1-15 358.6 + 1.03 73.5
water after 5-days expo- 5-11 507.1 +0.43 97.4
sure in Dylox
The periodic addition of Dylox to the test system causes increases in
inhibition of AChE with each dose. Fish mortality occurs at a total concen-
tration of 0.36 mg/l, considerably below the minimum lethal concentration
(Table 7).
Similar results have been obtained with experiments on roach. Daily ex-
posure to one-tenth of the 48-hour LCiqo ^^^ to a greater toxic effect
than the exposure in concentrations equal to the full 48-hour LC]oo.
90
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CN
LU
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LO
CN
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00
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CN
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91
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-z.
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Carassius carassius
Limnaea stagnalis
0^
ro
EXPOSED 1 r
^ - 1
t . * 1
""UNEXPOSED Ir
1 1
/^ 7
1 ' ' ' .1
. • ' ' '
/ ^7
\ \ 1
/ /
, /
w /
• /
/ /
• /
/ /
f /
/ /
• /
1 I
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1 1
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1 1
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1 /
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/ 1
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1 1
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WAVELENGTH, nm
Figure 4. Densitograms of the molecular form of acetylcholinesterase
in carp (Carassius carassius, 1) and the snail (Limnaea stagnalis,
27 unexposed and exposed to Dylox.
92
TABLE 7. CHOLINESTERASE ACTIVITY IN PERCH BRAIN AS A RESULT OF PERIODIC
ADDITIONS OF DYLOX TO THE EXPOSURE CHAMBER
Concentration
mg/£
Enzyme Activity
Days
of
observation
Number
of
samples
pM acetyl
choline
gm/hr
Percent
the cont
fishes
of
rol
Percent
mortality
1
0.12
-
-
-
0
5
, 0.12
10
339.3 + 0.94
67.6
0
10
0.12
10
207.7 + 0.42
42.3
2
11
-
-
-
-
15
12
-
10
147.6 + 0.64
31.4
52
13
-
9
140.2 + 0.58
28.5
76
14
-
-
-
-
94
15
-
-
-
-
100
93
Cholinesterase has been inhibited more in the first case (Kozlovskaya
and Novichikova 1979). Organophosphorus pesticides on continued chronical
exposure prove no less dangerous than with acute intoxication.
CONCLUSIONS
Organochlorine pesticides have been replaced with organophosphorus on
the assumption that as a result of lower persistf^nce in the aquatic environ-
ment those compounds will be of little danger to aquatic organisms. Organo-
phosphorus pesticides have proven to be highly toxic to the majority of
species of aquatic invertebrates. The data provided in this study demon-
strates that there are concentrations in reservoirs, which greatly exceed
lethal levels for sensitive species.
The intensive application of organophosphorus pesticides as a part of
agricultural practices results in a periodic influx of these pesticides into
water bodies. In natural waters, pollution levels are produced which cause
chronic effects upon aquatic animals. This is especially dangerous because
organophosphorus compounds possess an additive effect, and are poorly
avoided by aquatic animals.
An indicator of the effects of organophosphorus pesticides in an inhibi-
tion of cholinesterase in both acute and chronic intoxication. In patholog-
ical processes, the inhibition of cholinesterase as a target enzyme un-
doubtedly plays a leading role, although death occurs when the inhibition of
enzymes is still incomplete.
REFERENCES
Bilikova, A. 1973. Pesticides in Slovene surface water. In: Water
Management, 21, No. 10, pp. 261-263.
Boiko, I.B. and B.A. Pulatov. 1977. Materials on the hygienic reasoning
behind the maximum acceptable concentrations in waste water. Gigiena i
Sanitariya, No. 8, pp. 106-107
Drevenkar, V., K. Fink, M. Stipcevic, and B. Stengl. 1975. The fate of
pesticides in aquatic environment. 1. The persistence of some organo-
phosphorus pesticides in river water. Archives of Council on Hygiene
and Toxicology, 26:4, pp. 275-266.
Flerov, B.A. and L.N. Lapkina. 1976. The avoidance of certain toxic solu-
tions by the medical leech. Informational Bulletin, 30, pp. 48-52.
Flerov, B.A. and V.B. Tagunov. 1978. Analyzing the response to avoid
toxic substances in the Branchipod Streptocephalis tovicorni. Informa-
tion Bulletin, 40, pp. 68-71.
94
Flerov, B.A. In press. Comparative study of the reaction to avoid toxic
substances among water animals. In: Physiology and Parasitology of
Animals. Leningrad.
Hansen, D.J., S.C. Schimmel and Ir.J.M. Keltner. 1973. Avoidance of pesti-
cides by grass shrimp (Palaemonetes pugia). Bull. Environ. Contam. and
Toxicol., 9:3, pp. 129-133.
Ivanchenko, V.V. 1978. A study of the dynamics of the penetration and
preservation of phosalone in soil and the ability of insecticides to
migrate in the plant-water-soil cycle with precipitation in conditions
of the Saratov Oblast. Reports from the Institute for Experimental
Meteorology, 9:82, pp. 68-72.
Korotova, L.G. and A.S. Demchenko. 1978. The effect of various factors on
the process of metaphos dispersion in the soil and its washout by sur-
face runoff. Volume A. Hydrochemical Materials, Gidrometeoizdat,
Leningrad, 71, pp. 34-40.
Korotova, L.G. and A.S. Demchenko. 1978. The rate of chlorophos dispersion
in chestnut soils and its removal by surface runoff. Volume B. Hydro-
chemical Materials, Leningrad, 71, pp. 41-48.
Korovin, V.I., N.I. Sekushenko and A.V. Korovin. 1976. Phosphoro- and
chloro-organic pesticide runoff in the Kuban Region. Theses from re-
ports of the All -Union Scientific and Technical Conference on the Pro-
tection of Water from Contamination by Toxic Chemicals and Fertilizers,
Moscow, pp. 88-91.
Kostovetskiy, Ya.I., S.Ya. Nayshteyn, G.V. Tolstopyatova, and G.Ya.
Chegryanen. 1976. Hygienic aspects of using pesticides in the catch-
ment areas of reservoirs. Water Resources, I, pp. 167-172.
Kozlovskaya, V.I. and N.S. Novchkova. In press. The effect of chlorophos
and polychloropinene on the carbonic acid esterases in the blood serum
of carp. Informational Bulletin, The Biology of Inland Waters,
Leningrad, USSR.
Kozlovskaya, V.I. and G.M. Chuyko. In press. Blood serum cholinesterases
in fish of the genus Cyprinidae with variable resistance to chlorophos.
In: Physiology and Parasitology of Fresh-Water Animals. Leningrad,
USSR.
Lesnikov, L.A. 1974. Characteristics of the action of chlorophos and endo-
bacteria in various groups of water organisms. Reports from the State
Scientific Research Institute for the Lake and River Fishing Industry,
98, pp. 14-19.
Manko, N.N., Ye.G. Malozhanova, D.N. Polishchuk, et al_. 1974. Phosalone
materials for the toxicological and hygienic evaluation of new pesti-
cides. Moscow, 94 p.
95
Melnikov, N.N., A.I. Volkov and S.A. Kortkova. 1977. Pesticides and the
environment. Moscow, 240 p.
Novozhilov, K.V., V.A. Volkova and V.N. Rozova. 1974. Dynamics of the
dispersion of the phosphomide in plants into the soil. Chemistry in
Agriculture, 3, pp. 39-41.
O'Brian, R. 1964. The toxic esters of phosphoric acid. Moscow, 631 p.
Prokopenko, V.A., L.D. Zhiteneva, N.P. Sokolskaya, T.I. Kalyuzhnaya, V.P.
Zavgordnyaya, L.N. Isayeva, and Z.N. Kopylova. 1976. The toxicity of
carbophos for certain water bionts. Hydrobiology Journal, 12:5, pp.
47-52.
Rozengart, V.I. and O.Ye. Sherstobitov. 1978. Selective toxicity in
phosphoro-organic insecticides. Leningrad, 173 p.
Tagunov, V.B. and B.A. Flerov. 1978. The reaction of avoidance of toxic
substances in the water primrose. Informational Bulletin, Biology of
Inland Waters, 39, pp. 80-84.
Takase, Ivao. 1976. The dynamics of phosphoro-organic pesticides in water,
Sekubutsu Boeki, 30:8, pp. 302-306.
Ulyanova, I.N., L.Ya. Kheifetz, N.A. Sabina, and M.M. Kovrevskaya. 1979.
Metaphos breakdown in ground water. Materials from the Sixth All-Union
Symposium on Contemporary Problems Spontaneous Purification of Reser-
voirs for Regulating Water Quality, Tallin, pp. 123-125.
Yurovskaya, Ye.M. and V.A. Zhulinskaya. 1974. The behavior of phosphoro-
organic insecticides in soil. In: Chemistry in Agriculture, 5, pp.
38-41.
96
SECTION 8
MONITORING CONTAMINANT RESIDUES IN FRESHWATER FISHES IN THE
UNITED STATES: THE NATIONAL PESTICIDE MONITORING PROGRAM
J. Larry Ludke and C.J. Schmitt^
INTRODUCTION
The National Pesticide Monitoring Program (NPMP) originated in the mid
1960's as a cooperative effort by members of national agencies of the
Federal Committee on Pest Control. In 1972 the overall responsibility for
NPMP activities was given to the United States Environmental Protection
Agency (EPA). EPA then developed a comprehensive National Monitoring Plan
for Pesticides, which describes and sets broad guidelines for various other
federal agencies cooperating in monitoring pesticide trends in soil, water,
air, man, plants and animals (Table 1). Each participating agency monitors
chemical residues in the one or more segments of the environment which it is
charged with protecting or regulating. In recent years chemical contami-
nants other than pesticides, such as polychlorinated biphenyls (PCBs) have
been added to the list of chemical residues that are routinely analyzed.
For the purposes of the NPMP, monitoring can be defined as the repeti-
tive observation of one or more segments of the environment according to a
prearranged schedule in space and time. The overriding objective of the
NPMP is to ascertain on a nationwide basis, the levels and temporal trends
of selected contaminants in the environment.
A secondary objective of the NPMP is to identify areas where unusually
high residues may occur (i.e., problem areas) and which therefore may re-
quire more intensive study to determine potential contaminant sources and
possible detrimental effects. Data may also be used to initiate or evalu-
ate management and regulatory actions.
U.S. FISH AND WILDLIFE SERVICE SUBPROGRAMS
The U.S. Fish and Wildlife Service is responsible for the fish and wild-
life subprogram of the NPMP, the primary objective of which is to ascertain
^Columbia National Fisheries Research Laboratory, U.S. Department of the
Interior, Fish and Wildlife Service, Route #1, Columbia, Missouri 65201.
97
TABLE 1. NATIONAL PESTICIDE MONITORING PROGRAM NETWORK: A LIST OF
ENVIRONMENTAL COMPONENTS AND THE RESPECTIVE AGENCIES RESPONSIBLE
FOR MONITORING CONTAMINANT TRENDS IN EACH
Environmental Component
Agencies
Soils
Water and Sediment
Oceans, Bays, and Estuaries
Marine Fauna
Atmosphere (pilot program)
Avian Wildlife
Freshwater Fishes
Food and Feed
Environmental Protection Agency (EPA)
Environmental Protection Agency
U.S. Geological Survey (USGS)
National Oceanic and Atmospheric Agency
(NOAA)
Public Health Service
Environmental Protection Agency
U.S. Fish and Wildlife Service (FWS)
U.S. Fish and Wildlife Service
U.S. Department of Agriculture (USDA)
Food and Drug Administration (FDA)
98
on a nationwide basis, and independent of specific treatments, the levels
and trends of selected environmental contaminants in freshwater fishes and
selected bird species. In addition to monitoring trends in contaminants,
the Fish and Wildlife Service also investigates the sources and impacts of
contaminants on natural resources. The Columbia National Fisheries Research
Laboratory (CNFRL) is responsible for monitoring residue trends in fresh-
water fishes and Patuxent Wildlife Research Center, Laurel, Maryland is re-
sponsible for monitoring residues in tissues of selected waterfowl and star-
lings (Sturnus vulgaris).
FRESHWATER FISH FROM LAKES AND STREAMS
Monitoring contaminants in freshwater fish has undergone a series of
changes since collections began in 1967. At first, fish were collected
from 50 sampling stations in the Great Lakes and major rivers throughout the
United States (Stations 1-50, Figure 1). Five adult fish of each of three
predominant species were collected in the spring and again in the fall of
both 1967 and 1968. In 1969, and each year since then, collections have
been made only in the fall. In 1970 the number of collection stations was
increased to 100 with the addition of Stations 51-100 (Figure 1). Deter-
minations have always been based on composited, whole-body samples of five
fish each. From 1967 through 1971 all sample analyses were contracted to a
private laboratory; in 1972, for economic and administrative reasons, the
analytical work was shifted to the Fish and Wildlife Service Laboratory in
Denver, Colorado; and in 1976 the program was relocated to CNFRL, where it
remains today. Collections were suspended for one year in 1975 when fresh-
water fish monitoring was undergoing an internal review and reorganization.
There are now 117 stations in the United States where fish are collected
for analysis of contaminant residues (Figure 1). About half of the stations
are sampled in the Fall of even-numbered years and the other half during
odd-numbered years. At each trend monitoring station three samples of five
fish each are taken: two samples of a predominant bottom-dwelling species
and one sample of a predator species. The preferred species to be collected
vary geographically and according to habitat (Table 2).
The number of contaminants studied has increased over the years from
eight in 1967 to more than 20 today (Table 3). At CNFRL there is a strong
research emphasis on improving methods and developing the technology neces-
sary to quantify toxic chemical contaminants that are difficult to analyze
in biological tissues.
PROCEDURES
Fish are collected by non-chemical means (i.e., by electroshocking, net-
ting or hook and line) according to specified instructions. Sometimes fish
must be purchased from local commercial fishermen known to Fish in the vi-
cinity of the collection site. All specimens are adult fish, preferably of
uniform size, and weighing no more than 22.7 kg (5 lb) each.
99
100
TABLE 2. FRESHWATER FISHES RECOMMENDED FOR COLLECTION FOR TISSUE
CONTAMINANT RESIDUE DETERMINATIONS (NPMP), LISTED BY CATEGORY,
HABITAT AND (IN THE ORDER OF PREFERENCE) SPECIES
Category of fish, habitat, and species I
Predator
Cold water
Rainbow trout, Salmo gairdneri
Brown trout, S. trutta
Brook trout, Salvelinus fontinalis
Lake trout, S^. namaycush
Cool water
Walleye, Stizostedion vitreum
Yellow perch, Perca flavescens
Sauger, S^, canadense
Northern pike, Esox lucius
White perch, Roccus americanus
Other percid (Percidae) or temperate bass (Perichthyidae)
Warm water
Largemouth bass, Micropterus salmoides
Other sunfish (Centrarchidae)
Bottom Dwelling
All habitats
Carp (Cyprinius carpio)
Channel catfish (Ictalurus punctatus)
White sucker (Catostomus commersoni )
Other locally abundant sucker (Catostomidae) or catfish
(Ictaluridae)
Ipredator species are listed in order of preference for each habitat; order
of preference for bottom dwelling species is the same for all habitats.
101
TABLE 3. CONTAMINANT RESIDUES MEASURED AND DETECTED IN NPMP
FRESHWATER FISH SAMPLES, 1967 THROUGH 1976-77
Year
Contaminant
1967
1968
1969
1970
1971
1972
1973
1974
1976-1977
p,p' - DDE
.1
+
+
+
+
+
+
+
+
p,p' - DDD
+
+
+
+
+
+
+
+
+
p,p' - DDT
+
NA^
+
+
+
+
+
+
+
+
o,p' - DDE
NA
NA
+
+
+
+
+
+
o,p' - DDD
NA
NA
NA
+
+
+
+
+
+
o,p' - DDT
NA
NA
NA
+
+
:i
+
+
+
Aroclor 1242
NA
NA
NA
NA
NA
-
+
+
Aroclor 1248
NA
NA
NA
NA
NA
NA
NA
+
+
Aroclor 1254 2
NA
NA
+
+
+
+
+
+
+
Aroclor 1260
NA
NA
NA
NA
NA
+
+
+
+
Aldrin & dieldrin
+
+
+
+
+
+
+
+
+
Endrin
+
+
-
+
+
+
+
+
+
Lindane-^
+
+
NA
NA
NA
NA
NA
NA
+
a-benzene hexa-
chloride (a-BHC)^
NA
NA
+
+
+
+
+
+
+
Heptachlor & hepta-
chlor epoxide
+
+
+
+
+
+
+
+
+
Chlordane
+
+
+
+
+
+
+
+
+
Toxaphene
NA
NA
NA
NA
+
+
+
+
+
Hexachlorobenzene
(HCB)
NA
NA
NA
NA
+
+
+
+
+
Arsenic
+
Selenium
+
Mercury
+
Lead
+
Zinc
+
^On the body of the table, + indicates that the contaminant was detected in
at least one sample and - indicates that none was detected. NA - not
analyzed.
2Total PCB as Aroclor 1254, 1969-1971.
^Lindane (y- benzene hexachloride) separated beginning 1976.
^BHC as technical, 1969-74; as a - BHC beginning 1976.
102
Five fish (no fewer than three) are pooled to make up a sample and no
sample may exceed 113.4 kg (25 lb). Fish are rinsed in tap water and care
is taken to insure that they do not come into contact with potential con-
taminating surfaces such as plastics, printed paper, metal, or mud. Each
fish is weighed and measured (total length) and the age of each fish is
determined whenever possible. Fish are then wrapped individually in clean
aluminum foil and labeled, after which the specimens making up each pooled
sample are placed into a heavy bag and frozen immediately in dry ice. The
samples are then transported frozen, by air freight, to CNFRL for analysis,
Fish samples are kept frozen until the time of analysis. The five
specimens are then thawed, homogenized and appropriate subsamples are re-
moved for analysis of metals or chlorinated organic contaminant residues.
Metals are analyzed by atomic absorption spectrometry, and organochlorine
compounds are measured by gas-liquid chromatography; organic residues in
some samples are confirmed using mass spectrometry. Selected samples are
sent to an independent laboratory for analysis as a means of confirming
results .
SELECTED TEMPORAL AND GEOGRAPHIC TRENDS IN CONTAMINANT RESIDUES
Residues of DDT and it metabolites in fishes from the nation's major
rivers and lakes have shown a continuing downward trend. The steady de-
crease in total DDT, as reflected in summed p,p'-homologues (Figure 2) il-
lustrates the effectiveness of the 1972 ban on the use of DDT in the United
States. Although DDT residues remain high in some areas where it was used
extensively in the past, the overall trend has been downward. Even in those
areas where total DDT residues remain high, the p-p'-homologue, DDE, is pre-
sent in much greater proportion than in the past (Table 4), indicating sub-
stantial degradation of DDT and DDD in the environment.
The number of collection sites where DDT has been observed in at least
one samples has also decreased somewhat since 1970 (Table 5). Although the
present occurrence of p,p'-DDT appears to have increased in recent years
(1976-77), this change can probably be attributed to improved analytical
techniques that enable better resolution and higher sensitivity for organo-
chlorine contaminants.
PCBs have become virtually ubiquitous, reflecting the former widespread
use of these persistent industrial compounds as hydraulic fluids and as heat
transfer agents in capacitors and other electrical equipment. Fish contain-
ing residues of 0.5 ijg/g (wet weight, whole fish), the criterion established
for the protection of piscivorous fishes and wildlife, are collected re-
gularly from all NPMP stations near urban and/or industrial areas, and trace
levels are present in fish from the major watershed of all 50 states.
Definite trends in the overall magnitude of PCB residues are more diffi-
cult to discern due to the evolution of analytical methods between 1970 and
1974 (Tables 3 and 4). While there appears to be a slight downward trend
nationwide, especially in Aroclor 1254 residues, more data produced by
103
1.20 r
1.00
Q.
0
D
■D
■(/)
I-
Q
Q
o
0.80
0.60
0.40
0.20
0.00
69
71
I
73
Year
JL
±
J
75 76+77
Fiaure 2. Geometric mean total DDT residues (p,p' - homologues)
in freshwater fish, 1969-1976/77.
104
TABLE 4. GEOMETRIC MEAN RESIDUES OF ORGANOCHLORINE COMPOUNDS
AT 74 SELECTED NPMP STATIONS, 1970-1976/77
Y
ear
Compound
1970
1971
1972
1973
1974
1976-77
p,p'-DDT
0.27
0.19
0.11
0.07
0.05
0.05
p,p'-DDD
0.34
0.25
0J8
0.12
0.14
0.08
p,p'-DDE
0.47
0.35
0.40
0.30
0.37
0.24
Total DDT
0.98
0.73
0.64
0.44
0.52
0.35
Aroclor 1254
1.20
1.03
1.21
0.58
0.82
0.49
Total PCB
1.20^
1.03^
2
i.2r
0.78^
0.95^
0.87^
Toxaphene
NA^
0.01^
0.13
0.17
0.17
0.36
Aldrin + Dieldrin
0.08
0.07
0.07
0.05
0.09
0.06
Endrin
0.01
0.02
0.01
0.01
0.01
0.01
^p,p'-homologues
2As Aroclor 1254
^Aroclor 1242 + 1254 + 1260
^Aroclor 1242 + 1248 + 1254 + 1260
^Not analyzed
^Not analyzed
105
TABLE 5. PERCENTAGE OF 74 NPMP STATIONS WHERE DETECTABLE RESIDUES OF
IMPORTANT 0R6AN0CHL0RINE COMPOUNDS WERE FOUND, 1970-1976/77
Year
Compound
1970
1971
1972
1973
1974
1976-77
p,p'-DDT
100
98.6
74.3
41.9
48.6
87.8
p,p'-DDD
100
98.6
97.3
71.6
78.4
TOO
p,p'-DDE
100
98.6
97.3
95.9
95.9
100
Total DDT^
100
98.6
100
100
97.3
100
Total PCB
98.62
98.62
83.82
70.33
93. 2^
91. 9^
Toxaphene
NA^
13.5
9.5
12.2
14.9
60.8
Aldrin + Dieldrin
100
100
81.1
70.3
52.7
95.9
Endrin
31.1
82.4
10.8
20.3
2.7
48.6
^p,p'-homologues
2as Aroclor 1254
3Aroclor 1242 + 1254 + 1260
^Aroclor 1242 + 1248 + 1254 + 1260
^Not analyzed
106
today's methods are needed to substantiate this trend. However, residues at
the most heavily contaminated sites appear to be declining more noticeably.
PCBs occur in fish tissues most frequently and at the highest concentra-
tions in the industrial northeastern and midwestern sections of the United
States (Figure 3). Though no longer manufactured in the United States, PCBs
are still used and continue to contaminate the environment as a result of
spills and improper disposal of waste hydraulic fluids and discarded elec-
trical components.
Mean toxaphene residues are increasing in freshwater fishes of the
United States (Table 4). The national geometric average has increased from
0.13 yg/g in 1972 to 0.36 yg/g in 1976-77, and residues exceeding 1.0 yg/g
are not uncommon. Studies by CNFRL have shown that toxaphene residues of
1 yg/g may be associated with impaired growth and developmental abnormal-
ities in young fish.
Toxaphene also occurs much more widely now than it did in past years
(Table 5). Formerly found only in fish from the cotton growing regions of
the Southeast and Southwest, it now occurs in fish throughout the United
States (Figure 4). Its growing ubiquity may be explained by the increased
use of toxaphene in agriculture, largely as a substitute for DDT and other
compounds that have been banned. However, this interpretation is compli-
cated by findings indicating the possible occurrence of chlorinated cam-
phenes that behave like certain toxaphene components during gas chromato-
graphic analysis. Particularly high residues of this compound have been
found in fishes from the Upper Great Lakes. Despite extensive investiga-
tion by gas-liquid chromatography and mass spectrometry, neither the iden-
tity nor the source of this compound has yet been satisfactorily determined.
Nationally, average residue of dieldrin and endrin in fish tissues have
remained essentially unchanged from 1970 through 1977 (Table 4). Dieldrin
residues remained widespread (Table 5), reflecting the extensive use of this
compound (and aldrin) before 1974. The apparent variation in endrin occur-
rence (Table 5), however, may merely indicate changing analytical resolu-
tion; endrin residues have remained generally low (Table 4).
Using newly developed capabilities to measure trace metals, we at CNFRL
analyzed the fish samples collected in 1977 (representing 54 stations) for
residues of Cd, Pb, Hg, As, and Se. 'Background' levels for the five metals
in whole fish samples was determined, as well as geographic areas where
these levels are exceeded. As examples, we found As levels ^.5 yg/g in fish
from Texas, Oklahoma, and the Upper Great Lakes; Se of >] .0 yg/g at many
stations in the Upper Missouri River system, and at both stations in
Pennsylvania; Pb >] .0 yg/g at a group of stations in the central Missouri
River system; Hg 20-25 yg/g in the Great Lakes and in some Gulf Coast
rivers; and Cd ^0.15 yg/g at two Upper Missouri stations.
Discerning geographic and temporal trends in contaminant residues is not
the only result of NPMP monitoring activities. More importantly, the re-
sults of these efforts are reflected in the planning of research at CNFRL.
For example, unknown gas chromatograph peaks are resolved using mass spec-
107
I-;.-.-. ■.•.•.;■:■ ■-■:-.
1
-1
1
o
o
o
in
c
CO
HI
9
en
LU
OC
CO
o
Q.
LU
O
o
o
o
E
Q.
Q- in Q.
'- . o
O to
' TO ■
Q- ro o
O I en
30N3HdnOOO°
30N3HHnOOO°o
30N3yano30°o
o
o
in
o
33N3aanooo°o
Figure 3. Percent occurrence of polychlorinated biphenyl _{PCB) residues
in freshwater fish, by U.S. Fish and Wildlife Service Region, 1976/77.
108
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''\
''<-.
\l
r<.
'-y^-0\
A
II
M
ai
III
7"
Ol
Ol
111
I
b
je
je
(0
Ol
en
0.
E
E
l~^
I
en
w
s.
i
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01
s-
£
o
C7)
(U
0)
u
X
(U
10
(U
(U
s-
(U
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a.
IB
X
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(U
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109
tral analysis, which in turn may generate a list of candidate compounds for
toxicity testing. Or, the consistent occurrence of a given compound from
one location may stimulate a cooperative effort with Fish and Wildlife Ser-
vice Regional personnel, as in the cases of recent investigations of DDT in
the lower Rio Grande and toxaphene in the Great Lakes, to determine the
source and magnitude of the regional problem. And finally, questions aris-
ing from the analysis of NPMP samples continue to stimulate the development
of new analytical approaches.
no
SECTION 9
ACCUMULATION AND METABOLISM OF PERSISTENT PESTICIDES
IN FRESHWATER FISH
F.Ya. Komarovskiy and A.Ya. Malyarevskaya^
Long term utilization of persistent organochlorine pesticides, espe-
cially DDT, and BHC on a world wide basis has led to their distribution and
accumulation in a wide variety of media including soil, water, sediments,
and aquatic organisms. The accumulation of persistent pesticide residues in
organs of aquatic species and their tendency to be transformed in trophic
food chains are additional factors aggravating the danger of water pollution
by pesticides, both for regeneration of the biological resources of aquatic
ecosystems and for the health of man using fish for food.
Studies of the last decade indicated the possibility of understanding
the fundamental principles of DDT distribution in the biosphere, including
the world ocean; its accumulation in the biota; the role of DDT in eco-
systems of different types; demonstrated the biological danger of DDT resi-
dues for animals and man; and established the mechanism of its metabolism
in abiotic media and in aquatic organisms. While our knowledge has in-
creased and information on the subsequent biological effects of wide-scale
DDT utilization has increased, a great number of unsolved problems requiring
further research remain. For example, comparatively little data are avail-
able on DDT accumulation in brain tissue of warm-blooded animals and fish,
even though the neurophilicity of the compound suggests that it should have
received the greatest attention. There are \/ery few studies available which
show that the development of clinical symptoms of intoxication in warm-
blooded animals correlates with an increase of DDT accumulation in brain
tissue (Hayden 1960).
One of the most important principles of the biotic circulation of or-
ganochlorine pesticides, especially DDT, is their accumulation and trans-
formation in trophic chains, and their tendency to concentrate in the
highest links of these chains. This phenomenon is well demonstrated in
studies of terrestrial and marine ecosystems (Mayer-Bode 1966;
Andryuschchenko and Pishcholka 1975), but has received little attention in
freshwater ecosystems.
1 Institute of Hydrobiology of the Ukrainian Academy of Sciences, 44,
Vladimirskaya St., Kiev, 252003, USSR.
Ill
Recently, the processes of organochlorine pesticide accumulation in
trophic chains have been experimentally modeled to obtain more detailed in-
formation on the transformation mechanism in ecosystems. Metcalf, et^ al_.
(1971) used this experimental approach to select organochlorine pesticides
with lowest accumulation factors, i.e., those which were poorly accumulated,
and which were not transformed in trophic chains.
The question of metabolic pathways in tissues of animals, and metabolic
transition through final products is of considerable ecological importance.
Though the DDT metabolic processes have been well described by Kelvin, et_
a1 . (1959), additional detail for varying aquatic organisms are required.
The intent of this communication is to demonstrate the peculiarities of ac-
cumulation and distribution of residues of DDT and its metabolites in organs
and tissues of freshwater fish. Further, the factors characterizing the
development of intoxication will be considered.
Experimental efforts directed toward three major areas: 1) a demonstra-
tion of the level of persistent pesticides in the aquatic ecosystems and the
organisms under examination; 2) perform experiments in vitro to demonstrate
the accumulation of residues of DDT and its metabolites in selected organs
and tissues of fish, and to describe the developmental characteristics of
the intoxification process in time; and 3) conduct studies in experimental
basins to establish accumulation and transformation of persistent pesticides
at different trophic levels. In these studies, the following fish species
were used: bream (Abramis brama), pike perch (Lucioperca lucioperca), pike
(Esox esox), perch (Perca f luviatilis), carp (Cyprinus carpio), crucian carp
(Carassius carassius), silver carp (Hypophthalmychthys mol itrix) . The food
organisms tested included tubificids (Tubifex tubifex), and water fleas
(Daphnia magna) .
The residue level of DDT and its metabolites in water, silt, and tissues
of fish was determined by the gas chromatography technique.
Systematic examination for DDT and its metabolites (DDE and DDD) in the
water and sediments of the investigated water-bodies showed that this pesti-
cide was not always found. Their concentration in water were found to be in
the parts per trillion (ppt) and (ppb) parts per billion range. Sediment
values were in the range of parts per billion (ppb) and parts per million
(ppm). Since DDT solubility in water is expressed by a range of 1-5 ppb,
the availability of DDT and its metabolites in freshwater ecosystems is not
a function of physio-chemical transformations, but rather of biological
transformation of this substance, and its accumulation in trophic levels on
the basis of biological increases of 1 order of magnitude per trophic level.
As a result, it is possible to find rather high concentrations accumulated
in the second, third, and subsequent links of trophic chain. In both bio-
logic tissues and in the abiotic environment, DDT alone is not isolated.
Rather, the sum of its metabolites, DDD and DDE together with DDT proper is
usually expressed as the sum of DDT (DDE + DDD + DDT).
In freshwater fish (pike perch, bream, pike, carp, perch, etc.) from the
water-bodies investigated, the distributions of accumulated DDT and its
metabolites in organs and tissues is rather clearly observed, although the
112
content of DDT and its metabolites in tissues is comparatively low. The
greatest accumulation of DDT residues is found in the inner fat and brain
tissue of fish. Internal organs (liver, stomach and intestine) contain a
considerable quantity of the metabolites (DDE and DDD), but comparatively
little DDT. An even lesser amount of residual DDT is found in gonads and
spawn, while the lowest levels of residues of this pesticide are found in
muscular tissue (Komarovskiy, et al_. 1975).
Thus, residual quantities of DDT and its metabolites are mainly accumu-
lated in fatty and brain tissues. Having been taken into the fish, DDT
undergoes substantial metabolic changes. This fact is indicated by pre-
dominance of the metabolites DDE and DDD in storage organs.
It should also be noted that the results did not demonstrate the pre-
sence of polychlorinated biphenyls in organs and tissues of fish from the
study sites. However, corresponding analysis of fish specimens from the
Black Sea and the Barents Sea were positive for the presence of PCB (chro-
matograms showed saw-tooth peaks, analogous to those of the Baltic fish
that were convincingly shown by Swedish scientists to be associated with
PCB's). Chromatograms of the freshwater fish associated with the present
investigation showed only peaks typical for DDT and its metabolites.
The experimental research associated with this study provided the op-
portunity to confirm data on specific differences in accumulation and dis-
tribution of DDT residues in fish tissue, and to demonstrate differences
conditioned by the functional role of tissues, and the metabolic rate of
DDT during intoxication. Pesticide accumulation depends upon metabolic ac-
tivity. For example, DDT accumulation is much greater in tissues of preda-
tory fish, notable for their elevated level of metabolism. Total DDT con-
tent in the liver of fish from the experimental water-bodies was as follows:
pike - 1.400 ppm, zander - 0.220 ppm, silver carp - 0.115 ppm, and carp -
0.047 ppm. Crucian carp, subjected to the effect of high concentrations
(40 ppm) of this pesticide had DDT accumulation in intestine 0.850 ppm by
the end of the exposure, while pike perch had 1.430 ppm.
Pesticides accumulation was conditioned by the functional role of
tissues. It was the greatest in the tissues playing an important role in
the detoxification of pesticides (liver), and those having a comparatively
high content of lipid (liver, inner fat, and intestine). For example,
total DDT uptake under experimental intoxication for pike perch was as fol-
lows: liver - 0.220 ppm, intestine - 3.175 ppm, inner fat - 5.635 ppm,
muscles - 0.057 ppm.
The clinical picture of fish intoxication as a result of acute DDT expo-
sure was characterized by a marked behavioral change. Intensive locomotor
activity gave way to deceleration and a disturbance of coordinative move-
ments, loss of balance, adynamia and death. Dissection of the fish re-
vealed marked hemorrhaging of the brain and other vital organs (gills,
liver, heart, kidneys, etc.), as well as necrotic changes, especially in the
liver.
113
Chromatographic analysis showed comparatively rapid (within hours) ac-
cumulation of DDT and its metabolites {o,p' - DDE, o,p' - DDD, o,p' - DDT,
p,p' - DDD and p,p' - DDT) in fish tissues. Estimation of the DDT residue
content at different phases of intoxication enabled an understanding of the
dynamics of this process during fish convulsions (Phase 1), and at adynamia,
preceding death (Phase 2).
The quality of DDT and its metabolites increased in the tissues during
the processes of the development of intoxication, within a few hours. Total
DDT content in the muscles of silver carp increased from 0.103 ppm during
the first phase to 0.501 ppm at the second phase of intoxication. In liver
this increase was from 1.99 ppm during the first phase to 3.38 ppm during
the second phase. Similarly, in the intestine the range was from 2.83 ppm
at the first phase up to 0.79 ppm during the second phase.
Accumulation of DDT and its metabolites in fish was also accompanied by
a phase change of a number of biochemical indices, the group B vitamins in
particular. For example, vitamin Bi content increased in carp liver by
131 percent when locomotor activity was increased (Phase 1), and decreased
by 14 percent at the time of adynamia (Phase 2) when compared with control
values. These data are indicative that vitamins are of considerable impor-
tance in the process of intoxication.
During the first phase of intoxication, the vitamin B] content, which
is of considerable importance in metabolic processes, increases. During the
second phase when metabolism processes are disturbed, the organism's vital
resources are exhausted and the vitamin B] quantity is greatly reduced.
Changes in the levels of nicotine-amide enzymes in the fish tissues were
also indicative of alterations in the oxidation-reduction processes. The
total quantity of oxidized and reduced forms of nicotine-amide enzymes de-
creased in fish liver as a result of the action of lethal quantities of DDT,
from 554 ppm in the control group to 307 ppm in test animals. Similarly,
the ratio of oxidized and reduced forms also decreased in the liver tissue
from 2.26 ppm in control fish to 0.96 ppm in test species. Since nicotine-
amide enzymes are of great importance in the regulation of cellular respira-
tion, the alterations observed were indicative of considerable metabolic
disturbances in fish tissues under the influence of DDT.
Coupled with these observations was an extensive formation of metabo-
lites of DDT in organs and tissues rich in lipids. The formation of p,p' -
DDE; o,p' - DDT; p,p' - DDD; p,p' - DDT metabolites in intestine and inner
fat were of analogous character. DDT accumulation in fatty tissue during
the first phase of intoxication is accompanied by the formation of the
metabolite n,n' - DDD, while levels of o,p' - DDT and p,p' - DDE increase.
During the second phase this ratio changed to domination by p,p' - DDD and
o,p' - DDT. In the intestine, p,p' - DDT, and o,p' - DDT predominated
during the first phase, and by the second phase p,p' - DDD was dominant. In
the muscles of silver carp during the first phase of intoxication, p,p' -
DDT content was the greatest, while o,p' - DDT and p,p' - DDD were pro-
nounced in the second phase.
114
The liver, unlike other organs, was notable for greater stability in
content of DDT metabolites. This was conditioner^ by rapid transformation of
DDT in this organ. During the second phase of intoxication, o,p' - DDT,
p,p' - DDT, and p,p' - DDD were predominant.
Thus, the accumulation of DDT and its metabolites in organs and tissues
of fish is conditioned by their specific peculiarities, functional purpose,
and time of development of intoxication.
With the intent of studying accumulation of persistent pesticides, the
level of transformation in aquatic organisms, and their distribution and
transmission in trophic chains, experiments in aerated aquaria and pools
were carried out. In the process of studying the transformation of DDT and
its metabolites in the food chain, forage organism (Tubifex tubifex and
Daphnia magna), consumer fish (Cyprinus carpio), and predatory fish (Perca
f luviatilis and Esox lucius) were modeled.
Food organisms poisoned by chemically pure p,p' - DDT (1.1 to 3 ppm)
Con-
ac-
. ..v.,, — ,„^ >,..„„, ,^„, ^^ ^^...^.^^ ....v.,^ w. ...^ ^.^f....^ ■ were con-
trolled, and the complex of morphlogical and functional indices charac-
terizing the development of intoxication were studied (Braginskiy, et al .
1976).
Food organisms poisoned by chemically pure p,p' - DDT (1.1 to 3 pf
were fed to yearling carp, which in turn were eaten by predatory fish,
trol fishes were given food without DDT. During the experiments, DDT
cumulation and metabolism at selected levels of the trophic chain were
Investigations have shown that the DDT residue from water was taken into
the tissues of the Daphnia and tubificids in a very short time period, prac-
tically within the first day. When these organisms were fed to fish, con-
siderable concentrations of DDT residues were found in organs and tissues,
especially in fatty layers and in brain tissues, as early as the first 3
days, with a constant increase throughout the experiments. In the forage
species, (Daphnia and the tubificids), DDT metabolizes primarily to DDD,
while DDE is formed Mery slowly. In carp, the general accumulation of
pesticides with high specific weights of the DDE metabolite greatly in-
crease. An analogous picture is characteristic of perch and pike. When
these species are fed for an extended time with food containing DDT, the ac-
cumulation of this substance in their lipid containing tissues increases,
with a prevalence of the metabolites DDE and DDD.
Tubificids metabolize DDT only to DDD; Daphnia to DDD and DDE, carp to
DDD and DDE, and perch and pike to DDD and DDE, but with different percent-
age ratio.
Experimental research has shown that in parallel with fatty tissue, DDT
accumulates extensively in fish brain tissue, reaching critical values
(Braginskiy, et a_l_. 1979). It was found that using poisoned natural food,
the developing of intoxication in fish was, in fact, connected with accumu-
lation of DDT and its metabolites. It was stated that the fish died from
toxicosis at critical levels of DDT accumulation in the brain (3 ppm and
greater). These findings correspond to the results obtained during the in-
vestigation of analogous phenomena in warm-blooded animals (Dale, et al .
1963).
115
The modeling experiments show that DDT accumulates in trophic chain
quickly and effectively. Accumulation of DDT and its metabolites in fish
organs of vital importance was observed. These findings were distinctly
manifested in the fish brain tissue.
Toxicological symptoms appear in parallel with increasing levels of DDT
in target organs, especially in the brain. Clinical and pathological-ana-
tomical intoxication may be reproduced by experimental modeling rather
quickly and synonymously, and the fish behavior and clinical symptoms are
similar to those of acute intoxication.
When DDT and its metabolites (DDE and DDD) accumulate up to 3 ppm in
the brain tissue of fish (perch, pike), the fish die with obvious symptoms
of cumulative toxicosis. It should be noted that mammals and birds present
an analogous picture, i.e., convulsive phenomena as DDT accumulation ap-
proaches the lethal level, and death at definite accumulation levels (Hayden
1960; Dale, et al_ 1963; Ludwig and Ludwig 1969).
Thus, these investigations enabled the development of principles of the
actions of DDT and its metabolites. Their distribution in organs and
tissues of freshwater fish, the development of a model of cumulative toxi-
cosis in fish under experimental conditions, and an understanding of the
basis of accumulation of DDT, along with its metabolism, depending upon the
functional role of the tissue and species of aquatic organism.
REFERENCES
Andryushchyenko, V.V. and Yu.K. Peshcholka. 1975. DDT in certain elemen-
tary biocoenosis of the Black Sea and the Delta of the Danube. In:
Studies of the Biological Production and Protection of Waters of the
Ukraine. Scientific Thought, Kiev, pp. 100-101.
Braginskiy, L.P., F.Yah. Komarovsky, and Yu.K. Petsolka. 1976. Experi-
mental modeling of the mechanism of DDT intoxication in predatory fish.
In: Experimental Aquatic Toxicology. Zenatnyeh, Riga, pp. 204-215.
Braginskiy, L.P., F.Yah. Komarovsky, and A.I, Myehryehzuko. 1979. Per-
sistent pesticides in the ecology of freshwater. Scientific Thought,
Kiev, 143 p.
Calvin, M.M. 1969. Metabolism of pesticides. Special Scientific Report
Wildlife. Washington, No. 127, 293 p.
Dale, W.E., T.B. Daines, and W.J. Hayer. 1963. Poisoning of DDT relation
between clinical signs and concentration in rat brain. Science, 142,
No. 3598, pp. 1474-1479.
Hayden, R.E. 1960. Effects of DDT on birds. N.Z. Gardiner, 17, No. 1, pp.
66-73.
116
Komarovsky, F.Yah., V.V. Matyelyev, and Yu.K, Peshcholka. 1975. DDT and
its metabolism in organs and tissues of fish. In: The Formation and
Control of the Quality of Surface Waters. Scientific Thought, Kiev,
Volume 1, pp. 79-84.
Ludwig, J. P. and C.E. Ludwig. 1969. The effect of starvation on insecti-
cide, contaminated nerring gulls removed from a Lake Michigan colony.
Proc, 12th Conf. on Gr. Lakes Res., Ann Arbor, Michigan, pp. 185-192.
Mayer-Bodyeh, G. 1966. Residue of pesticides. Peace, Moscow, 350 p.
Metcalf, R.L., G.K. Sangua, and I. P. Kapoor. 1971. Model ecosystem for the
evaluation of pesticide biodegradability and ecological magnification.
Environ. Sci. Techn., No. 5383, pp. 709-719.
117
SECTION 10
SOME FACTORS AFFECTING THE TOXICITY OF AMMONIA TO FISHES
Robert V. Thurston^
INTRODUCTION
Ammonia can be a serious toxicant to fishes and other aquatic life. It
can enter natural water systems from several sources, including industrial
wastes, sewage effluents, coal gasification and liquefaction conversion pro-
cess plants, and agricultural discharges including feedlot runoff. It is
also a metabolic waste product of fishes, and as such presents a major pro-
blem in fish culture.
In aqueous solutions, ammonia assumes two chemical species, illustrated
by the following equation.
NH
3(g) ' ^^2^{l) ^ N"3-"^20(aq) ^ ''^" ' 0^" ' ^'-'^""z^U)
These species are the gaseous or un-ionized form (NH3), bound to at least
three water molecules, and the ionized form (NH4"'"). In this presentation,
the term NH3 will refer to un-ionized ammonia, NH4"'" will refer to ionized
ammonia, and total ammonia will refer to the sum of these. The aqueous am-
monia equilibrium is strongly dependent upon the pH of the solution, and to
a lesser extent upon temperature and ionic strength. As the pH increases,
increasing the hydroxide ion concentration, the equilibrium shift of ammonia
is toward the un-ionized (NH3) species. Within the pH range acceptable to
most freshwater fishes, an increase of one pH unit will increase the NH3
concentration approximately tenfold (Thurston et al_. 1974). Temperature in-
crease also favors the NH3 species, but to a lesser extent; ionic strength
increase, at low concentrations, favors the NH4"'" species (ibid).
Early reported research on the toxic effect of ammonia (Chipman 1934;
Wuhrmann et a^. 1947; Wuhrmann and Woker 1948) implicated NH3 as being the
toxic form of ammonia, and NH4+ was considered non-toxic or appreciably less
toxic. Because of the recognized toxicity of NH3, and the belief that NH4"'"
is not significantly toxic, most toxicity values reported in the literature
are as NH3. Sometimes total ammonia values have also been reported, but too
'Fisheries Bioassay Laboratory, Montana State University, Bozeman, Montana
59717.
118
frequently pH, temperature, and other water quality parameters have been
omitted, making it difficult to reconstruct reported test conditions.
Much of the literature on ammonia toxicity to fishes has recently been
reviewed in the EPA "Red Book" (U.S. EPA 1977) and the American Fisheries
Society "Red Book Review" (Thurston et al_. 1979). Reported acute toxicity
values in tests from 1 to 4 days duration on salmonids range from 0.25 to
0.85 mg/liter NH3; values for comparable tests on non-salmonids range be-
tween 0.4 and 4 mg/liter NH3.
Published reports on chronic toxicity of ammonia do not include any
life-cycle mortality data, but effects of ammonia on both warm- and cold-
water fishes at sublethal concentrations of ammonia for periods of time
ranging from 1 week to 3 months have been reported by several researchers.
Within the concentration range of 0.06 to 0.4 mg/liter NH3, these reported
effects include swelling and diminishing of number of red blood cells, ir-
reversible blood damage, inflammation and degeneration of gills and other
tissues, and lessening of resistance to disease (Reichenbach-Klinke 1967;
Flis 1968; Smart 1976). Within the range 0.05 to 0.15 mg/liter NH3, re-
duced food uptake and assimilation and growth inhibition have been reported
(Ministry of Technology 1972; Robinette 1976; Schulze-Wiehenbrauck 1976;
Burkhalter and Kaya 1977). In a test of 6 months duration on rainbow trout
(Salmo gairdneri ) it has been reported that concentrations as low as 0.01
mg/liter NH3 caused not only reduced growth rates, but pathological changes
to gills and livers (Smith and Piper 1975). Ball (1967) indicated that al-
though it may appear that different species of fishes exhibit dissimilar
susceptibilities to ammonia toxicity under acute exposure conditions, such
is not the case under long-term exposures. He theorized that trout and
carp, given time to react, may be equally susceptible to ammonia, and that
although acute responses are different, the ultimate response by both
fishes to a given concentration of ammonia may be the same.
In sunmary, reported acute toxicity ammonia values for a variety of
species of fishes range between 0.25 and 4 mg/liter NH3, and other mani-
festations of the effects of ammonia have been reported at concentrations
as low as 0.01 mg/liter NH3. There is some evidence that differences in am-
monia tolerance among fish species may be less under chronic conditions than
under acute conditions. Based on the published literature, the European
Inland Fisheries Advisory Commission (EIFAC 1970) has recommended a crite-
rion of 0.025 mg/liter NH3 as being the maximum which can be tolerated by
fishes for an extended period of time, and the United States Environmental
Protection Agency (1977) has published a criterion of 0.02 mg/liter NH3,
just slightly more restrictive than that recommended by EIFAC.
Mery possibly these criteria are "safe" for most water bodies which sup-
port aquatic life, but some questions remain unanswered as to whether they
are reasonable for all waters at all times under all conditions. Tabata
(1962) has attributed some toxicity to NH4"'", concluding that it may be
l/50th as toxic as NH3 to Daphnia pulex. Robinson-Wilson and Seim (1975),
testing coho salmon (Oncorhynchus kisutch), have demonstrated correlation
between pH and the acute toxicity of ammonia expressed as NH3. More re-
cently Armstrong et aj_. (1978), in tests on larvae of the prawn Macrobra-
119
chium rosenbergii , concluded that NH4''" is toxic. The work of these re-
searchers raises questions about an ammonia criterion based solely on NH3.
In addition, it is also known that prior acclimation, temperature, and dis-
solved oxygen may also affect the toxicity of ammonia to fishes. Consider-
ing the large number of industrial and agricultural discharges which contain
ammonia, and the tremendous expenditure of energy and resultant cost to
treat these discharges for ammonia reduction to meet statutory requirements,
it is reasonable to ask whether a single water quality standard for ammonia
can be justified. Certainly some of the factors that increase or decrease
the toxicity of ammonia should be considered further.
EFFECT OF ACCLIMATION
The question of whether fishes can acquire an increased tolerance to am-
monia by acclimation to low ammonia concentrations is an important one. In
certain real-world environmental situations, such as a stream receiving ef-
fluent from a sewage treatment plant, fishes may be subjected to high am-
monia concentrations for short and/or intermittent periods of time. If a
fish had an increased ammonia tolerance, developed due to acclimation or
conditioning to low ammonia levels, it would perhaps be able to survive what
might otherwise be acutely lethal ammonia concentrations.
There is some information in the literature reporting that the effect of
previous exposure of fishes to low ammonia concentrations reduces or does
not affect their tolerance to lethal ammonia levels. Steinmann (1928) re-
ported that the minnow Alburnus bipunctatus was more susceptible to ammonium
hydroxide if previously exposed. Observations by McCay and Vars (1931) in-
dicated that bullheads (Ameiurus nebulosus) subjected to several successive
exposures to ammonia, alternated with recovery in fresh water, acquired no
immunity from the earlier exposures to the later ones. Fromm (1970) accli-
mated goldfish (Carassius carassius) to low (0.5 mg/liter) or high (5.0 or
25.0 mg/liter) ambient NH3 for periods of 20 to 56 days and found that urea
excretion rate in subsequent 24-hour exposures to concentrations ranging
from 0.08 to 2.37 mg/liter was independent of the previous acclimation con-
centration or duration.
There is a larger body of information, however, which indicates that
prior exposure of fishes to low concentrations of ammonia increases their
resistance to lethal concentrations. Vamos (1963) conducted an experiment
in which carp (species not specified) were exposed to 0.67 and 0.52 mg/liter
NH3 for 75 minutes, revived in fresh water for 12 hours, and then subjected
to ammonia at a concentration of 0.7 mg/liter NH3. Control fish, exposed
only to the latter ammonia concentration, developed ammonia-poisoning symp-
toms within 20 minutes, but the previously exposed fish did not exhibit
these symptoms until 60-85 minutes. M'^lci'cea (1968) subjected carp (Rhodeus
sericeus amarus Bloch) for 4 days, and minnows (Phoxinus phoxinus L. ) for 3
days to "acclimation" solutions of ammonium sulfate (0.26 mg/liter NH3).
The "adapted" carp and "unadapted" control group were then exposed to lethal
concentrations of ammonium sulfate (5.1 mg/liter NH3). The mean survival
time of the adapted carp was 88 minutes and that of the unadapted carp was
78 minutes. The minnows were subjected to lethal toxic concentrations of
120
2.4 mg/liter NH3 in ammonium sulfate solution. Mean survival time of
adapted minnows was 65 minutes, and of the unadapted control group was 45
minutes.
Fromm (1970) has measured urea excretion rates of rainbow trout initial-
ly subjected to either 5 or 0.5 mg/liter NH3, and then subjected to 3
mg/liter NH3. The trout previously exposed to 5 mg/liter NH3 excreted
slightly less urea than those previously exposed to the lower concentration.
Lloyd and Orr (1969) measured urine flow rates of rainbow trout exposed for
24 hours to 0.27 mg/liter NH3, and then exposed for another 15 hours to 0.53
mg/liter NH3. Pretest urine flow rates of 2.8 ml/kg/hr increased first to
6.4 and then to 8.0. One fish died during the lower ammonia level exposure
and none during the higher exposure. A control batch of fish with a pretest
urine flow rate of 0.75 ml/kg/hr was subjected directly to the higher (0.53
mg/liter NH3) armionia concentration. The urine flow rate jumped to 11
ml/kg/hr, and all fish died within 3 hours.
In a second experiment by Lloyd and Orr (1969), rainbow trout were sub-
jected to 0.32 mg/liter NH3 for successive 22-hour time periods, separated
by a 24-hour non-exposure period. Although urine flow rates were higher
during exposure periods than during pre-exposure, they were less during the
second exposure period than during the first. This suggests that some ac-
climation was developed and subsequently retained, at least for a 1-day rest
period. A third experiment indicated that this acclimation was not retained
during a 3-day rest period between two similar ammonia exposures.
Schulze-Wiehenbrauck (1976) conducted a study on the effect of sublethal
ammonia exposures on young rainbow trout growth, food consumption, and food
conversion. In one experiment, trout were acclimated for 3 weeks at 0.007
(control), 0.131, and 0.167 mg/liter NH3; the fish from these three tanks
were then subjected to concentrations of approximately 0.45 mg/liter NH3 for
8.5 hr. Fish from the two ammonia acclimation concentrations had 100 per-
cent survival, whereas only 50 percent of the control group survived the
test period. In the second experiment, the acclimation concentrations were
0.004 (control) and 0.16 mg/liter NH3; these fish were placed in NH3 concen-
trations of approximately 0.5 mg/liter for 10 hours. There was 100 percent
survival of the ammonia acclimated fish, and 85 percent survival of the con-
trol fish. The results of these experiments thus showed an increase in re-
sistance of rainbow trout to acutely toxic concentrations of ammonia after
prior exposure to sublethal ammonia concentrations.
At Fisheries Bioassay Laboratory we have conducted experiments to inves-
tigate the effect of acclimation of rainbow trout to sublethal ammonia con-
centrations on the fish's response to acutely lethal ammonia concentrations.
Seven 96-hour flow-through bioassays (using NH4CI) were conducted, six of
these on fish that had been acclimated for 29 days to concentrations ranging
from 0.018 to 0.078 mg/liter NH3, and the seventh on a control group accli-
mated at 0.001 mg/liter NH3. For each bioassay there were 5 test tanks and
1 control tank containing 10 fish each; mean fish sizes for the tests were
12 to 15 g. Additional details of these tests and data treatment will be
reported elsewhere (Thurston and Russo, in preparation).
121
Figure 1 shows the toxicity curves for these tests (LC50 in mg/liter NH3
vs. time). There was a statistically significant correlation between the
NH3 concentration at which the fish were acclimated and their subsequent re-
sistance to acutely toxic NH3 concentrations. The higher the NH3 concentra-
tion at which the fish were acclimated, the more tolerant the fish were to
acutely lethal levels during the 96-hour test period. The shapes of the
curves also show that there is a general trend for fish acclimated at higher
ammonia concentrations to take longer to arrive at an eventual asymptotic
LC50 value.
We also performed some experiments to determine whether the length of
time of acclimation to low ammonia concentrations affected the fish's re-
sponse in subsequent exposure to lethal NH3 levels. Duration of acclimation
to ammonia in these experiments ranged from 29 to 154 days; the subsequent
lethal tests were all 96-hour bioassays as described above. Results showed
that there was a significant relationship between 96-hour LC50 and length of
time of prior acclimation; the longer the acclimation period, the more tol-
erant the fish were to high ammonia levels. Our calculations took into con-
sideration the fact that fish weight also increased as acclimation duration
increased. We also investigated whether there was an effect on fish's tol-
erance to ammonia if they were placed in fresh (ammonia-free) water for
periods of 2, 14, and 28 days after acclimation and before exposure to
lethal concentrations. From limited data, our experiments indicated that
fish rapidly (less than 2 days) started to lose the tolerance to ammonia
built up by acclimation once they were placed in ammonia-free water.
In summary, there is reasonable evidence that fishes with a history of
prior acclimation to some sublethal concentration of ammonia are better able
to withstand an acutely lethal concentration, at least for some period of
hours and possibly days. The concentration limits for both acclimation and
subsequent acute response need definition and explanation.
EFFECT OF TEMPERATURE
There is limited information in the literature on the effects of temper-
ature on ammonia toxicity to fishes. Generally, the toxicity of total am-
monia decreases with lower temperatures, attributable mainly to a decrease
in the concentration of NH3. Woker (1949), testing chub (Squalius cephalus)
within the range of 10-25 C, concluded that water temperature had practi-
cally no effect on the manifestation time of toxic symptoms resulting from
ammonia. On the other hand, Colt and Tchobanoglous (1976) observed that the
tolerance of channel catfish (Ictalurus punctatus) to ammonia increased as
the experimental temperatures were increased up to the fish's reported opti-
mum temperature for growth (29-30 C). It is reasonable to expect that at
temperature conditions which are marginal for any given fish species, the
species will not be able to function optimally to resist toxic effects of
ammon i a .
We have conducted eight 96-hr ammonia bioassays on 2- to 12-g rainbow
trout at elevated temperatures within the range 12-19 C. Test conditions
were similar to those employed in the acclimation experiments reported
122
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I-
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00 n m t^ CO O) T-
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Figure 1. Effect of prior ammonia acclimation on the acute toxicity
of ammonia to rainbow trout.
123
above. Fish were acclimated to test temperature for 1.5 to 2 days prior to
introduction of ammonia toxicant. Ninety-six hour LC50 values ranged be-
tween 0.6-1.2 mg/liter NH3, but there was no correlation between ammonia
toxicity and temperature. Statistical treatment showed that size was not a
factor. We also conducted nine similar tests on 1-g cutthroat trout (S.
clarki ), within the range 13-19 C. Ninety-six hour LC50 values ranged be-
tween 1.0-1.5 mg/liter NH3, but again there was no temperature/ammonia toxi-
city relationship. In 15 tests on fathead minnows (Pimephales promelas),
however, within the range 13-22 C, we did find a definite correlation be-
tween temperature and susceptibility to ammonia toxicity. The toxicity
curves for these tests are shown in Figure 2. As temperature decreased,
toxicity increased. A plot of 96-hour LC50 values (mg/liter NH3) vs. tem-
perature, and a statistically computed correlation curve are illustrated in
Figure 3. It should be noted that in the case of the two trout species
tested, the temperature range studied was above their normal environmental
temperature; in the case of the fathead minnows, the range tested reached
several degrees below that for their optimum growth. We have not tested
fathead minnows at temperatures above, nor have we tested trout below, their
optimum growth temperature ranges.
Our results for trouts agree with those reported by other researchers
within the temperature range 10-20 C (Herbert 1962; Lloyd and Orr 1969).
The British Ministry of Technology (1968), however, has reported that the
toxicity of ammonia to both adult and juvenile rainbow trout was much
greater at 5 C than at 18 C. Based on our analysis of their data as re-
ported, their case for juvenile trout appears stronger than that for adults.
The European Inland Fisheries Advisory Commission (1970) has cautioned that
acceptable concentrations of ammonia may be less at temperatures below 5 C.
Although this temperature value may be arbitrary, we conclude that there is
some merit to the argument that a drop in temperature below some optimum
range for a given species of fish may increase its susceptibility to ammonia
toxicity. It is important that this relationship be further studied. The
available evidence that temperature, independent of its role in the aqueous
ammonia equilibrium, affects the toxicity of ammonia to fishes argues for
further consideration of the temperature/ammonia toxicity relationship.
EFFECT OF DISSOLVED OXYGEN
The discharge of ammonia is frequently associated with a reduction in
oxygen levels in the receiving water. This is brought about by any of
several causes, including the oxygen demand of the ammonia itself as it is
converted by natural microbial oxidation to nitrite and nitrate; the chemi-
cal and biological oxygen demand of other chemicals which may be, and fre-
quently are, discharged along with ammonia; and ihe reduction in oxygen-
carrying capacity of the receiving water if the discharge causes a rise in
its temperature. If the receiving water body is rich in nutrients and
highly productive, as is frequently the case downstream from a sewage treat-
ment plant, there is the effect of diurnal and seasonal fluctuations in dis-
solved oxygen caused by plant growth.
124
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10 12
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TEMPERATURE, °C
Figure 3. Actue toxicity of ammonia vs. temperature for fathead minnows,
[LC50 = 1.086 + 0.002203 (temperature)2] .
126
Several researchers, working with a variety of warm-water fishes, have
reported that the acute response to ammonia was not affected when dissolved
oxygen levels dropped from saturation to approximately one-half or one-third
saturation, but below that resistance decreased (Wuhrmann 1952: Wuhrmann and
Woker 1953; Merkens and Downing 1957; Danecker lb»64; Vamos and Tasnadi
1967). Reports on rainbow trout generally agree that this species is more
sensitive than warm-water fishes to the combined effects of low dissolved
oxygen and ammonia, and that any reduction in dissolved oxygen or any reduc-
tion below two-thirds saturation will decrease rainbow trout tolerance to
amnonia (Allan 1955; Downing and Merkens 1955; Merkens and Downing 1957;
Danecker 1964), One of the findings reported by Downing and Merkens (1955),
who tested young rainbow trout in experiments lasting up to 17 hours, was
that a decrease in dissolved oxygen from 8.5 to 1.5 mg/liter shortened the
periods of survival at all ammonia concentrations tested; this decrease was
proportionally greatest at the lowest concentrations of ammonia. In longer
tests, lasting up to 13 days, these same researchers reported similar re-
sults (Merkens and Downing 1957).
To explain the accelerated action of ammonia toxicity under reduced
oxygen conditions, Lloyd (1961) presented the argument that a given toxic
effect is produced by a specified concentration of toxicant passing across
the fish gill surface at a rate governed by the fish gill movement. At re-
duced oxygen concentrations the rate of movement increases, resulting in an
increased rate of gill exposure to the toxicant. He hypothesized that a re-
duction in CO2 excretion at the gill surface, resulting from reduced O2 in-
take, will raise the pH at the gill surface. Such an increase in pH will
favor the more toxic ammonia species (NH3) resulting in an even more accele-
rated toxic effect of ammonia than might be expected solely by an increased
rate of gill movement. However, CO2 loss at the gill surface is also con-
nected with the fish's ammonia excretion mechanism, and recent research on
the possible toxicity of NH4"'" suggests that a complete explanation may be
more complex.
To examine the effect of dissolved oxygen on ammonia toxicity we con-
ducted two series of 96-hour flow-through bioassays, one of these (15 bio-
assays) on rainbow trout, and the other (10 bioassays) on fathead minnows.
Test conditions were similar to those described earlier, and test fish were
acclimated to the test oxygen level for at least 2 days prior to introduc-
tion of ammonia toxicant. The rainbow trout for all tests were from the
same stock, and the stock fish grew in size over the several weeks that the
tests were conducted so the average test fish size gradually increased from
2 to 10 g. The tests were not run in any particular sequence of dissolved
oxygen level, however, and subsequent statistical treatment showed that
there was no correlation between test result and fish size. Figure 4 shows
a plot of the 96-hour LC50 value (mg/liter NH3) for each test vs. the dis-
solved oxygen level at which the test was conducted. The correlation for
rainbow trout between LC50 and dissolved oxygen was striking (correlation
coefficient 0.9346, P = 0,00001); the lower the dissolved oxygen concentra-
tion, the greater the toxicity of ammonia. Although a regression line for
the fathead minnow tests was obtained, the slope of this line is not statis-
tically different from zero (P = 0,365), We conclude that there is most de-
127
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DISSOLVED OXYGEN, mg/l
Figure 4. Effect of dissolved oxygen on the acute toxicity of ammonia to
fathead minnows and rainbow trout.
128
finitely a correlation for the rainbow trout tests, but we cannot draw the
same conclusion for the fathead minnow tests.
In an attempt to study the reduced dissolved oxygen effect on ammonia
toxicity in relation to time, we analyzed our data for the rainbow trout
tests, comparing the dissolved oxygen vs. LC50 correlations for the tests at
12, 24, 48, 72, and 96 hours. This showed a very clear and statistically
defensible trend (Figure 5); the shorter the time period, the more pro-
nounced the correlation. This trend suggests at least two possibilities:
either individual fish which require higher oxygen concentrations succumb
early in the tests, and/or those fish which do survive become increasingly
acclimated to the ammonia and oxygen test conditions as time progresses.
The EPA Red Book (U.S. EPA 1977) has recommended a minimum concentration
of 5.0 mg/liter dissolved oxygen to maintain good freshwater fish popula-
tions. At that dissolved oxygen concentration the regression line for the
rainbow trout tests reported above indicates a 96-hour LC50 of 0.5 mg/liter
NH3 (Figure 4). At dissolved oxygen concentrations of 8.0 mg/liter and
above, more common to natural cold-water fish habitats, the test results re-
gression line indicates 96-hour LC50's in excess of 0.7 mg/liter NH3. For
this particular stock of test fish, tested under the given bioassay condi-
tions, there was a 30 percent decrease in the medium lethal concentration of
ammonia when the dissolved oxygen concentration dropped from 8 to 5
mg/liter. If this ammonia LC50/dissolved oxygen correlation bears up under
further testing using this and other species, the need for reconsideration
of both ammonia and dissolved oxygen criteria is clear.
EFFECT OF pH
A premise of both the EIFAC (1970) and the U.S. EPA (1977) criteria for
ammonia is that NH4'^ is not appreciably toxic to aquatic life. The empiri-
cal basis for this was mentioned earlier, and has been explained by the
ability of NH3 to diffuse across the gill membrane whereas NH4+ requires
active transport. The research by Tabata (1962), Robinson-Wilson and Seim
(1975) and Armstrong et al_. (1978), however, raises questions about the
criteria premise.
We have conducted two series of bioassays to investigate the toxicity of
ammonia under different pH conditions. The fishes tested were rainbow
trout and fathead minnows, and the pH range was 6.5 to 9.0. We chose this
pH range because its limits are those recommended by the U.S. EPA (1977) as
being the limits acceptable to freshwater fishes. We treated the data from
each test by the trimmed Spearman-Karber method (Hamilton et aj_. 1977) to
determine both the total ammonia and the un-ionized ammonia 96-hour LC50
values. Again, for each bioassay there were five test tanks at different
ammonia concentrations and one control tank; eac*" tank contained 10 test
fish. The pH of the water in all tanks for any one test was uniform; this
was achieved by adjusting the normal pH (7.8) of the test water either up
by means of a metered sodium hydroxide solution, or down using a solution of
hydrochloric acid. During any given test, the ammonia concentration, pH,
and temperature in each test tank were monitored between 5 and 8 times, and
129
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minute adjustments in pH were made as appropriate. Mixing of the test water
and additives was virtually instantaneous, ensuring uniform water chemistry
conditions throughout any one tank. This was confirmed by repeated sampling
studies.
The average size of rainbow trout was 9-11 g, all fish from the same
stock, and that for the fathead minnows was 1.8-2.0 g, again from a single
stock. Tests were conducted on successive weeks, three at a time: one
acid, one base, and one at the normal pH of the test water. The normal pH
test was repeated each time the acid and base tests were conducted; com-
parable results from the normal pH tests verified that test conditions from
week to week were comparable, and that the test fish stock had not changed
appreciably over time.
The results of the tests on rainbow trout are illustrated in Figure 6.
Ninety-six hour LC50 values and their confidence limits in terms of both
total ammonia-nitrogen and un-ionized ammonia-nitrogen are plotted for each
pH test. A log scale for the LC50 values has been used so that visual com-
parison of total ammonia and NH3 values can easily be made. The excellent
reproducibility of the tests run at normal test water pH is apparent. If
the un-ionized form of ammonia (NH3) were solely responsible for the toxic
action on the test fish, then one would expect that the LC50 values, in
terms of NH3, would be reasonably constant for all tests regardless of the
solution pH and total ammonia present. This did not turn out to be the
case. Figure 7 illustrates the results of the tests on fathead minnows.
The LC50 values, in terms of both total ammonia-nitrogen and un-ionized
ammonia-nitrogen, are higher than those for rainbow trout because the fat-
head minnow is a more ammonia- tolerant fish, but the LC50 vs. pH trend is
the same.
Our findings provide support for the conclusions of Tabata (1962) and
Armstrong et ^. (1978), and are in conflict with the more widely accepted
notion that the toxicity of NH3 is independent of pH. The LC50 values in
terms of NH3 for our 96-hour acute toxicity tests on rainbow trout are
strikingly similar to those reported by Robinson-Wilson and Seim (1975) for
coho salmon within the pH range 7.0 to 8.5. These authors explain the cor-
relation of solution pH with NH3 LC50 values to be related to changes in the
CO2 concentration, hence pH, at the surface of the fish gill tissue. Our
conclusion at this time is that the NH4''' ion exerts a heretofore not fully
recognized toxic effect on fishes, and/or that the toxicity of NH3 increases
as the H+ ion concentration increases.
Regardless of the explanation for it, the correlation between LC50 in
terms of NH3 and pH has been demonstrated, and the rationale for water
quality criteria for ammonia needs to address this.
CONCLUSION
I have discussed briefly just four factors affecting the toxicity of am-
monia. I have used these as examples of how the many chemical and physical
parameters involved in aqueous systems are interrelated in affecting the
131
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UN-ION/ZED AMMONIA
1 1 1 1
6.5 7.0
7.5 8.0 8.5 9.0 9.5
pH
Figure 6. Acute toxicity of ammonia to rainbow trout:
96-hour LC50 vs. pH.
132
lUUU
—
—
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-z.
LU
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TOTAL AMMONIA
§100
-z.
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6.5 7.0 7.5 8.0 8.5 9.0 9.5
PH
Figure 7. Acute toxicity of ammonia to fathead minnows;
96-hour LC50 vs. pH.
133
toxicity of a pollutant. Time limitations have necessitated a cursory
treatment of both the published literature and the new research reported
here. More complete information on this and other ammonia toxicity research
being conducted both at Fisheries Bioassay Laboratory and at the Sunoga
Laboratory here in Borok is now in preparation for journal publication in
the Soviet Union and in the United States. The information I have presented
illustrates some of the complexities involved in establishing water quality
criteria and setting standards. It also underscores the necessity for con-
tinued collaborative effort between fish physiologists and water chemists,
from laboratories such as ours and Sunoga, in conducting and interpreting
the results of aquatic toxicity tests.
REFERENCES
Allan, R.H. 1955. Effects of pollution on fisheries. Verh. Int. Ver.
Limnol. 12: 804-810.
Armstrong, D.A., D. Chippendale, A.W. Knight and J.E. Colt. 1978. Inter-
action of ionized and un-ionized ammonia on short term survival and
growth of prawn larvae, Macrobrachium rosenbergii. Biol. Bull. 154:
15-31.
Ball, I.R. 1967. The relative susceptibilities of some species of fresh-
water fish to poisons - I. Ammonia. Water Res. 1: 767-775.
Burkhalter, D.E. and CM. Kaya. 1977. Effects of prolonged exposure to
ammonia on fertilized eggs and sac fry of rainbow trout (Salmo gaird-
neri). Trans. Am. Fish. Soc. 106(5): 470-475.
Chipman, W.A., Jr. 1934. The role of pH in determining the toxicity of am-
monium compounds. Ph.D. Thesis, University of Missouri, Columbia, MO.
153 p.
Colt, J. and G. Tchobanoglous. 1976. Evaluation of the short-term toxicity
of nitrogenous compounds to channel catfish, Ictalurus punctatus.
Aquaculture 8: 209-224.
Danecker, E. 1964. Die Jauchevergiftung von Fischen -- eine Ammoniak-
vergiftung. (The jauche poisoning of fish — an ammonia poisoning).
Osterreichs Fischerei. 3/4: 55-68. (In English translation).
Downing, K.M. and J.C. Merkens. 1955. The influence of dissolved oxygen
concentration on the toxicity of un-ionized ammonia to rainbow trout
(Salmo gairdnerii Richardson). Ann. Appl . Biol. 43: 243-246.
European Inland Fisheries Advisory Commission. 1970. Water quality cri-
teria for European freshwater fish. Report on ammonia and inland
fisheries. EIFAC Tech. Paper No. 11: 12 p. (Also in Water Res. 7:
1011-1022 (1973)).
134
Flis, J. 1968. Anatomicohistopathological changes induced in carp (Cyp-
rinus carpio L.) by ammonia water. Part II. Effects of subtoxic
concentrations. Acta Hydrobiol. 10: 225-238.
Fromm, P.O. 1970. Toxic action of water soluble pollutants on freshwater
fish. Water Pollution Control Research Series 18050 DST 12/70, U.S.
Environmental Protection Agency, Washington, D.C. 56 p.
Hamilton, M.A., R.C. Russo, and R.V. Thurston. 1977. Trimmed Spearman-
Karber method for estimating median lethal concentrations in toxicity
bioassays. Environ. Sci. Technol. 11(7): 714-719. Correction 12(4):
417 (1978).
Herbert, D.W.M. 1962. The toxicity to rainbow trout of spent still liquors
from the distillation of coal. Ann. Appl . Biol. 50: 755-777.
Lloyd, R. 1961. Effects of dissolved oxygen concentrations on the toxicity
of several poisons to rainbow trout (Sa1mo gairdnerii Richardson). J.
Exp. Biol. 38: 447-455.
Lloyd, R. and L.D. Orr. 1969. The diuretic response by rainbow trout to
sub-lethal concentrations of ammonia. Water Res. 3: 335-344.
Malacea, I. 1968. Untersuchungen uber die Gewohnung der Fische an hohe
Konzentrationen toxischer Substanzen. (Studies on the acclimation of
fish to high concentrations of toxic substances). Arch. Hydrobiol.
65(1): 74-95. (In English translation).
McCay, CM. and H.M. Vars. 1931. Studies upon fish blood and its relation
to water pollution. Pages 230-233 Ln A biological survey of the St.
Lawrence Watershed. Supplement to 20th annual report. New York Conser-
vation Dept.
Merkens, J.C. and K.M. Downing. 1957. The effect of tension of dissolved
oxygen on the toxicity of un-ionized ammonia to several species of fish.
Ann. Appl. Biol. 45(3): 521-527.
Ministry of Technology. 1968. Water Pollution Research 1967. H.M. Sta-
tionery Office, London. 213 p.
Ministry of Technology. 1972. Water Pollution Research 1971. H.M. Sta-
tionery Office, London. 129 p.
Reichenbach-Klinke, H.-H. 1967. Untersuchungen uber die Einwirkung des
Ammoniakgehalts auf den Fischorganismus. (Investigations on the in-
fluence of the ammonia content on the fish organism). Arch. Fisch-
ereiwiss. 17(2): 122-132. (In English translation).
Robinette, H.R. 1976. Effect of selected sublethal levels of ammonia on
the growth of channel catfish (Ictalurus punctatus). Prog. Fish-Cult.
38(1): 26-29.
135
Robinson-Wilson, E.F. and W.K. Seim. 1975. The lethal and sublethal ef-
fects of a zirconium process effluent on juvenile salmonids. Water
Resour. Bull. 11(5): 975-986.
Schulze-Wiehenbrauck, H. 1976. Effects of sublethal ammonia concentrations
on metabolism in juvenile rainbow trout (Salmo gairdneri Richardson).
Ber. dt. wiss. Kommn. Meeresforsch. 24: 234-250.
Smart, G. 1976. The effect of ammonia exposure on gill structure of the
rainbow trout (Salmo gairdneri ). J. Fish Biol. 8: 471-475.
Smith, C.E. and R.G. Piper. 1975. Lesions associated with chronic exposure
to ammonia. Pages 497-514 ln_ The pathology of fishes. W.E. Ribelin and
6. Migaki (Eds.), University of Wisconsin Press, Madison, WI.
Steinmann, P. 1928. Toxikologie der Fische. Handbuch der Binnenf ischerei
Mitteleuropas. 6: 289-342. (Cited in Chipman 1934).
Tabata, K. 1962. Suisan dobutsu ni oyobosu amonia no dokusei to pH, tansan
to no kankei. (Toxicity of ammonia to aquatic animals with reference to
the effect of pH and carbonic acid). Bull. Tokai Reg. Fish. Res. Lab.
34: 67-74. (In English translation).
Thurston, R.V., R.C. Russo, and K. Emerson. 1974. Aqueous ammonia equili-
brium calculations. Tech. Rep. No. 74-1, Fisheries Bioassay Laboratory,
Montana State University, Bozeman, MT. 18 p.
Thurston, R.V., R.C. Russo, CM. Fetterolf, Jr., T.A. Edsall, and Y.M.
Barber, Jr. (Eds). 1979. A review of the EPA Red Book: Quality cri-
teria for water. Water Quality Section, American Fisheries Society,
Bethesda, MD. 313 p.
U.S. Environmental Protection Agency. 1977. Quality criteria for water.
Office of Water and Hazardous Materials, U.S. Environmental Protection
Agency, Washington, D.C. 256 p.
Vamos, R. 1963. Ammonia poisoning in carp. Acta Biol. Szeged 9(1-4):
291-297.
Vamos, R. and R. Tasnadi, 1967. Ammonia poisoning in carp. 3. The oxygen
content as a factor influencing the toxic limit of ammonia. Acta
Biol. Szeged 13(3-4): 99-105.
Woker, H. 1949. Die Temperaturabhangigkeit der Giftwirkung von Ammoniak
auf Fische. (The temperature dependence of the toxic effect of ammonia
on fish). Int. Assoc. Theor. Appl. Limnol. 10: 575-579. (In English
translation) .
136
Wuhrmann, K. and H. Woker. 1948. Beitrage zur Toxikologie der Fische. II,
Experimentelle Untersuchungen iiber die Ammoniak- und BTausaurever-
giftung, (Contributions to the toxicology of fishes. II. Experimental
investigations on ammonia and hydrocyanic acid poisoning). Schweiz. Z.
Hydrol. 11: 210-244. (In English translation).
Wuhrmann, K., F. Zehender, and H. Woker. 1947. Uber die f ischereibiolo-
gische Bedeutung des Ammonium- und Ammoniakgehaltes fliessender
Gewasser. (Biological significance for fisheries of ammonium ion and
ammonia content of flowing bodies of water). Vierteljahrsschr. Natur-
forsch. Ges. Zurich 92: 198-204. (In English translation).
Wuhrmann, K. 1952. Surquelques principes de la toxicologie du poisson.
(Concerning some principles of the toxicology of fish). Bull. Cent.
Beige Etude Doc. Eaux 15: 49-60. (In English translation).
Wuhrmann, K. and H. Woker. 1953. Uber die Giftwirkungen von Ammoniak- und
Zyanidlbsungen mit verschiedener Sauerstoffspannung und Temperatur auf
Fische. (On the toxic effects of ammonia and cyanide solutions on fish
at different oxygen tensions and temperatures). Schweiz. Z. Hydrol.
15: 235-260. (In English translation).
137
SECTION 11
THE PREDICTION OF THE EFFECTS OF POLLUTANTS ON AQUATIC ORGANISMS
BASED ON THE DATA OF ACUTE TOXICITY EXPERIMENTS
O.F. Filenko and E.F. Isakoval
The increasing number of pollutants requires acceleration of the ability
to assess their toxicity, and to determine acceptable levels in the environ-
ment. These needs, coupled with a reduction of analytic costs, require a
reduction in the length of experimental effort, and, at the same time, an
increase in the reliability of the response.
To accelerate capabilities of assessment of toxicity, attempts were made
to connect the biological activity of compounds with their physico-chemical
properties. The correlation of toxicity of individual compounds with ap-
proximately 40 different physico-chemical properties were investigated
(Filov and Liublina 1965). Naturally, a high correlation of these data for
one organism is not sufficiently reliable for a group of species. It is
known that reactions of different organisms, and occasionally even one or-
ganism, to the same toxin are different under altered conditions. In such
cases, toxicity can differ by many orders of magnitude.
Another direction in the search has been an attempt to find the specific
and especially sensitive reactions of organisms to the action of a given
pollutant. These attempts have mostly failed. The sensitive and specific
index for poisoning by land, an increasing level of 8-amino levulic acid in
blood and urea, proved to be less sensitive than in the case of poisoning by
mercury (Jackim 1973).
Usually such biophysical, biochemical, and physiological indices assist
in identifying harmful effects after they have produced irreversible changes
in the organism. The natural fluctuations of many of these indices in or-
ganisms are so wide that changes produced by chronic toxic action are usual-
ly unrecognizable. The picture is further complicated by the varying re-
actions of the organisms under the influence of toxic substances in varying
environmental conditions.
^Moscow State University, Biological Faculty, Lenin Hills, Moscow, USSR.
138
Thus, to be reliable, the index applicable to the rapid determination of
biological effects of pollutants must take into account the peculiarities of
both compounds and organisms. An example of one such approach to the prob-
lem can be found in the relationship of toxicity of organic in compounds in
fish to values of their concentration gradients on the blood-brain barrier
(Filenko and Parina, In press). It may be assumed that compounds of a homo-
logous series have equally effective toxic potentials, but varying tissue
accumulation capabilities, and that this is the principal reason for differ-
ent resulting toxicity.
However, such general biological indices as survival and fecundity are
still the most reliable. To decrease the time required for assessment of
toxicity of a compound, instead of using the more reliable chronic experi-
ments, acute toxicity tests of the compounds over a period of 24-96 hours
usually are used. Application of such data for other conditions, concentra-
tions, and species specific coefficients and factors can be used (Steinberg
1974). This approach is primarily useful as a quick screening methodology.
When experiments are shortened, a portion of the reliability of response can
be retained by increasing the number of experimental tests. Therefore, it
becomes a question of the acceptability of the degree of simplification of
conditions, and the reduction of the length of the experiment to that which
is essential, and which involves a sufficient number of tests to make a
reasonably reliable estimation of the probable effect of the material on
the specific index in question for a period which exceeds the length of the
time of observation.
An attempt to investigate aspects of this problem and some associated
difficulties, are described in this paper. It should be noted, however,
even the most carefully made predictions cannot equal the reliability of re-
sults from experimental verification.
METHODS
The experimental design utilized the water flea, Daphnia magna (Straus)
in densities of 10 animals per 500 ml. The toxicity of individual compounds
that are potential industrial and agricultural pollutants of water was
assessed. The calculation of regression equations was made by the least
squares method.
RESULTS AND DISCUSSION
The toxic effect of compounds on Daphnia was assessed by organism sur-
vival. The typical mortality curve for varying concentrations of compounds
are shown in Figure 1. To demonstrate the regularity of this phenomenon,
the coefficients for different equations that could describe the mortality
of Daphnia in time were calculated. The results of such calculations for
trimethyl tin chloride (TMTCh) are given in Table 1. The exponential,
power, logarithmic, and parabolic functions were calculated. Tne fit of
theoretical and experimental points was examined using correlation coeffi-
cients. The larger the coefficients, the greater the correspondence to a
139
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TIME, days
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Figure 1. The relationship of the number of dead Daphnia magna (N)
with time (T) under the influence of various concentrations
of trimethyl tin chloride.
140
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141
high degree of fit. Equations for varying numbers of time observations from
the start of experiment were calculated. A comparison of correlation
values, calculated for different functions, shows that they are largest for
power and parabolic functions, suggesting that these equations describe the
regularity more accurately.
This conclusion was correct for varying concentrations of TMTCh. The
power function IgN = a + b IgT, where N is number of dead Daphnia, in per-
centage, and T is the time in days, in logarithmic coordinates becomes a
straight line (Figure 1), and it is possible to construct the curve based
upon two points. Examples of the transformation of regularity of Daphnia
mortality with time in logarithmic coordinates for organic tin and other
compounds are given in Figure 2. It should be noted that the experimental
and calculated values are not close enough. This fact is reflected by low
values of correlation coefficients. It is possible that fluctuations de-
pend on factors that are difficult to take into account in calculations,
e.g., varying development of adaptive processes in organisms, and their al-
tered reactions to environmental influences when exposed to different con-
centrations of compounds.
An attempt to analyze the dynamics of mortality in toxic solutions was
made in order to understand the relationship of observed regularities to
time. It is obvious for both groups of organisms, and for individuals, that
they are influenced by the solution of toxic compounds, and that the toxic
reaction increases through time, either as a function of continuous accumu-
lations of the toxic materials, or as a result of the volume of alterations
in the organism. The outcome for individual Daphnia will be the increasing
of probability of death, and for a test group, there will be an increasing
ration and rate of mortality. Thus, the slope of the curve increases dra-
matically in acute lethal experiments with organic tin compounds. In
chronic studies, the curve progresses in a step-wise form. This reflects a
sudden reduction in the rate of mortality with continuous exposure to toxic
influences.
The explanation for this phenomenon lies in a combination or sum of two
processes, (1) mortality under the influence of toxic substances, and (2)
acceleration and enhancement of adaptive process^^s within the organism that
inhibit mortality (Figure 3). The increase in toxicity proceeds more or
less regularly with time, forming the basis for the adaptive processes that
occur after the development of harmful effects in response to the toxins.
It is not yet clear what activates these adaptive processes, the level of
compound, the results of the deleterious effects in tissues, or the rate of
increase of accumulation. It is possible to determine the rate of decrease
or absence of mortality in toxic concentrations. Both of these two compo-
nents, harmful effects and adaptation, can be described by adequate equa-
tions that can be used for further elementary analysis of the dynamics of
the curve of mortality. However, the unique reactivity of living systems
under the influence of toxic substances complicates the regularities that
could describe the results of toxic effects. However, after calculating the
coefficients a and b for the equation of power function, it is possible,
with high degree of probability, to calculate the mortality of any percent-
age of Daphnia for a given period of time.
142
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Stroganov (1975) recommends as acceptable the use of toxins that produce
not more than 25 percent mortality. The equations present here calculate
the data of death of 25 percent of Daphnia (T25). As a rule, interpolations
have been made, but extrapolation is also possible.
Practically, it is important to determine the minimum time period of ob-
servation that is sufficient for reliable calculations. To this end, re-
gression equations for various time/mortality points were calculated. This
enables a determination of the number of time points that would be suffi-
cient for calculation of the T25, value that does not differ significantly
from the experimental value for 30 days. In Table 2 the dependence of cor-
relation coefficients and the T25 value from the length of experiment is
shown. The value of T25 does not significantly change for different periods
of observation. This information makes it possible to limit the duration of
experiments. For the calculation of coefficients for the equation, two dots
are enough, but the reliability of calculated values will be low. For reli-
able results, it is advisable to have 3 to 4 time/mortality points for every
concentration. In relatively high concentrations and with frequent record-
ing of results, the time period can be \jery short. Thus, experimental re-
sults can be completed and quickly specify a preliminary assessment of
acceptable concentrations.
As a result of these calculations a set of data is available that
characterize the time of death of test organisms in varying concentrations
(Table 3). The graphical relationship of concentration to time of death of
25 percent of Daphnia can be given as shown in the Figure 4A. This rela-
tionship can also be described by regression equations. From examined re-
gularities (exponential, power, logarithmic and hyperbolic) the power func-
tion was found to be most suitable (Table 4). The correlation coefficients
for the power function are highest, and it can be simply calculated by usual
methods. This function is also suitable from a logical standpoint. Indeed,
the curve of this function can never cross the axes, because time cannot be
negative function, and there are enough small concentrations that do not in-
fluence the life-span of Daphnia. The concentration that does not effect
Daphnia corresponds to the vertical asimptote.
There is certain diversity in the relationship of concentrations of pol-
lutants to their effects (Warren 1971). However, these relationships can be
described with a high degree of approximation by power or other simple func-
tions. Using logarithmic axes, the power function becomes a straight line
(Figure 48), and approximate equation coefficients can be calculated from
two concentrations.
By using these equations for certain compounds, we can evaluate the time
of death for other concentrations, and estimate the concentration that
causes the death of 25 percent of Daphnia in a given time period. The
period of life-span can be limited to 30 days, and mortality to 25 percent.
The concentration, that corresponds to these data, will be an acceptable
concentration in terms of survival (Table 5). In this table the acceptable
concentrations were calculated from data of concentrations, and a comparison
with values that were accepted from experimental evidence is made. It is
natural that there are some differences between experimental data and
145
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1.5
Figure 4. The relationship of time of death of 25 percent of Daphnia magna
with the concentration of trimethyl tin chloride.
150
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151
TABLE 5. ACCEPABLE CONCENTRATIONS OF COMPOUNDS FOR SURVIVAL OF
DAPHNIA MAGNA CALCULATED WITH EQUATIONS OF POWER FUNCTION
Compound
Maximal acceptable concentrations (mg/1)
Determined in chronic
experiments
Calculated on the data of
acute toxicity
TMTCh
0.01
TETCh
0.01
TPTCh
0.001
TATCh
0.0005
THTCh
0.002
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0.01
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0.001
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0.0001
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0.0002
0.0003
0.001
0.003
0.005
0.00026
1
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0.01
0.005
0.85
152
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Graphical determination of acceptable concentrations of
trimethyl tin chloride for Daphnia magna.
153
theoretical data. However, even in experimental determinations there may be
a diversity in repetitions, as well as deviations caused by the toxicologi-
cal experiment itself, especially when the test concentrations utilized
differ by orders of magnitude.
The reported results were derived for in experiments on Daphnia, but
this approach is applicable for other aquatic organisms as well. This ap-
proach has been shown to be particularly effective in experiments with long-
lived species (Parina et al_. 1979). It is assumed that these regularities
are applicable for other indices of the effects of toxic substances on
aquatic organisms.
In summary, it may be concluded that of all newly developing methods of
quick screening of toxic effects of pollutants on the aquatic organisms
using forecasting techniques, the most effective method, is still the use of
mathematical extrapolation of data from acute experiments. The dynamics of
the results of toxic influence for aquatic organisms (mortality) can be
shown as a combination of simpler processes. The connection of mortality
with time, and the onset of given effects with concentration can best be
described by power function equations. For evaluation of regression equa-
tions describing these statistically reliable relationships, 3 to 4 experi-
mental points are necessary. These equations can be used for determination
of the effects of the pollutant on the organisms for a period that exceeds
the duration of observation, and for concentrations, that have not been
experimentally investigated, including an approximation acceptable concen-
trations of the pollutant in the aquatic environment. It is particularly
advisable to use this approach for work under field conditions, or with
long-lived organisms, when the possibility of long-term observations does
not exist. This approach may also be used for investigations into a wide
spectrum of concentrations of a given pollutant.
REFERENCES
Filenko, O.F. and O.V. Parina. 1979. The distribution in organ systems,
as factor in determining the toxicity of tri-alkyl tin chlorides for
carp. In press.
Filov, V.A. and E.I. Liublina. 1965. The connection of toxic activity of
volatile organic compounds with their physico-chemical properties.
Biophysica, 10: N 4, pp. 602-608.
Jackim, E. 1973. Influence of lead and other metals on fish 8-aminolevuli-
nate dehydrase activity. J. Fish. Res. Board Can. 30: 560.
Parina, O.V., O.F. Filenko and O.P. Siutkina. 1979. The connection of
toxicity with some physico-chemical properties of organic tin compounds
in carp. D^ The Reaction of Aquatic Organic Tin Compounds. Ed. by N.S.
Stroganov. Moscow State University, pp. 147-155.
154
Steinberg, M.A. 1974. A review of some effects of contaminants on marine
organisms. Indo-Pacific Fish. Counc. Proc. 15th Session. Wellington,
Bangkok, pp. 8-23.
Stroganov, N.S. and L.V. Kolosova. 1971. The keeping of laboratory culture
and determination of fecundity of Daphnia in range of generations. ln_
Methods of Biological Investigations of Water Toxicology. Ed. by N.S.
Stroganov, Nauka, Moscow, pp. 210-216.
Stroganov, N.S. 1979. Some general problems of analysis of influence of
organic tin compounds on aquatic organisms. 2ll The Reaction of Aquatic
Organisms to Organic Tin Compounds. Ed. by N.S. Stroganov. Moscow
State University, pp. 241-259.
Warren, C.E. 1971. Biology and water pollution control. W.B. Sounders
Co., Philadelphia.
155
SECTION 12
AGE SPECIFICS OF SENSITIVITY AND RESISTANCE OF FISH TO ORGANIC
AND INORGANIC POISONS
V.I. Lukyanenko^
The continually increasing interest of researchers in the age aspects of
toxicoresistance of fish (Mironov 1972; Kuhnbold 1972; Eisler 1972; Mitrovic
1972; Shmalgauzen 1973; Samylin 1974; Waldiehuk 1974; Danilchenko 1975;
Dethlefsen 1975; Patin 1977, etc.) results from many factors, two of which
are particularly interesting.
1. The first is the need to understand the paths of direct toxic
influence of various substances entering the water on ichthyo-
fauna and, in the final analysis, on the productivity of the
reservoir. As we know, toxic substances affect all stages of
the life cycle of fish: from fertilization of eggs to sexually
mature individuals. However, from the ecologic standpoint, the
early stages of ontogenesis of fish (embryonal and immediate
postembryonal) are most vulnerable from the standpoint of the
toxic factor, since they cannot actively migrate and avoid pol-
luted water. It follows from this that the reaction of a popu-
lation of fish to chemical pollution will be determined by the
effect of the toxic factor on these early stages of ontogenesis
if they are less resistant than mature fish.
2. The second factor determining the activation of research in the
area of the age factor in ichthyotoxicology is the search for
the most vulnerable stage in individual development of various
species of fish, which should be used as the test object in the
determination of the basic parameters of toxicity of various
groups of substances and subsequent determination of maximum per-
missible concentrations (MPC) for these substances. It is quite
understandable that the least resistant stages of ontogenesis
development of fish are of primary interest for those involved
in development of the problem of biologic testing of the quality
of natural and waste waters.
The possible influence of pollution on larvae and fry was first men-
tioned in the last century. For example, the great Russian ichthyologist.
1 Institute of Biology of Inland Waters, Academy of Sciences, Borok, Nekouz,
Jaroslavl, 152742, USSR.
156
O.A. Grimm (1896), in his now classical monograph, "Kaspiysko Volzhskoye
Rybolovstvo" (Fishing the the Caspian and Volga), in analyzing the paths of
influence of petroleum on the "fish content" of this basin, wrote, "It is
quite probable that petroleum kills the fry of the Clupeidae family of fish
and others, which float on the top or accumulate near the bank in shoals".
Somewhat later, H. Clark and G. Adams (1913) concluded that one of the lead-
ing causes of the decrease in the population of whitefish was pollution of
the spawning grounds in the Great Lakes with industrial wastewater. How-
ever, experimental study of the age specifics of toxicoresistance of fish
began only comparatively recently.
One of the first reports in this area is that of N.S. Stroganov and A.M.
Pazhitkov (1941). In experiments with eggs, larvae, fry and mature individ-
uals of perch, it was shown that the early stages of development are less
resistant to the ions of copper and ammonia than mature fish. Given equal
exposure, mature perch survived in solutions of copper 100 times more con-
centrated than the lethal concentration for fry. In experiments with am-
monia, the differ^ence was less striking, but still clearly indicated the
lower stability of embryos and perch larvae than that of mature fish.
The high resistance of mature fish in comparison to larvae and fry for
heavy metal salts was noted by other authors as well (Sollman and Schweiger
1957; Cairns and Seheir 1957). However, in later works, materials have been
presented indicating that the stability of fish in the early stages of on-
togenesis is higher than that of mature individuals, or at least equal
(Mosevich, et al_. 1952; Wurtz-Arle 1959; Katz and Chadwick 1961; Vernidub
1962). For example, N.A. Mosevich, et^ al. (1952), in experiments with eggs,
larvae and first-year perch, establisheT"that the first-year fish were less
resistant to phenol than the eggs and larvae. Developing eggs and recently
hatched larvae were found to be more resistant than mature fish to the
pesticide andrin (Katz and Chadwick 1961). These data agree with the mate-
rials of Ye. A. Veselova, et al_. (1965), who studied the toxicity of still
another pesticide - hexachlorane - and concluded that developing eggs and
larvae of many species of fish (salmon, roach, bleak, perch, rock perch,
pike) are somewhat more stable than mature individuals. Finally, in a work
of D. Wurtz-Arle (1959) performed on developing eggs and fry of trout, it
was shown that their resistance to two detergents (sodium alkylsulfates) de-
creases with age.
Thus, in the mid-1960's there were two mutually opposite points of view.
The proponents of one believed that "the most vulnerable stage of ontogene-
sis in fish for the effects of toxic substances is the stage of the larvae
and fry" (Stroganov and Pazhitkov 1941, p. 68), i.e., the toxicoresistance
increases with age. The other group of authors held the opposite point of
view, assuming that the resistance of fish to poisons decreases with age and
that it is highest in the early stages of ontogenesis.
Analysis of the available literature data has allowed us (Lukyanenko
1967) to find the reasons for this contradiction. It was found that the
proponents of the idea of increased stability of fish in early stages of in-
dividual development based their ideas on data obtained in experiments with
organic poisons (phenols, synthetic detergents, pesticides). Researchers
157
holding the opposite point of view, that of reduced resistance of fish to
poisons in the early stages of ontogenesis, had performed experiments with
inorganic poisons, primarily heavy metal salts. This indicated to us that
the seeming disagreement, concerning the level of toxicoresistance of
various stages of ontogenesis of fish, resulted in fact from the different
nature of the toxic substances studied and, consequently, the differences in
mechanism of action of the poisons, organic and inorganic in nature, on the
developing eggs, larvae and fry.
Considering the importance of this problem, both in the theoretical and
in the practical aspects, we undertook an experimental test of this assump-
tion, concentrating our emphasis on organic poisons. Since in most works on
the age variation of ichthyotoxicology, authors have used some single
"point" of embryonal, larval or fry development, we decided to study the
dynamics of toxicoresistance in each of the three periods of early ontogene-
sis. In our report, we summarize the results of many years of studies per-
formed on bony fishes (rainbow trout, bream, zope, carp) and cartilagenous
ganoids (Russian sturgeon, Caspian sturgeon, sterlet and giant sturgeon).
The toxins used represented a broad range of concentrations of phenol, cer-
tain pesticides (metaphos, yalan and propanid), as well as chlorides of
cadmium and cobalt, in order to determine the age specifics of toxicoresis-
tance of the fish to inorganic poisons.
In our initial experiments, performed jointly with V.M. Volodin and B.A.
Flerov on the eggs, larvae, fry and mature individuals of two systematically
similar species of the genus Abramis; the bream (A. brama) and zope (A,
ballerus), exposed to the toxic effects of 12 different concentrations of
phenol (from 1 to 5000 mg/liter), we found that the toxicoresistance of
mature fish was significantly lower than that of the eggs, embryos and lar-
vae (Volodin, et^ aj_. 1965, 1966). This was reflected both in the lethal
concentrations for fish of the various age groups, and in the time of sur-
vival of each of the age groups studied with identical or similar concentra-
tions of toxic substance.
The decrease in resistance of fish to phenol from young age groups to
older age groups agrees with the available data from the literature; how-
ever, in these same experiments we found that, within each of the three main
stages of early ontogenesis; embryonal, larval and fry, toxicoresistance
undergoes significant changes. For example, the least stable period of em-
bryongenesis was found to be the earliest - from the beginning of division
to the formation of the embryo, particularly the stage of gastrulation. Be-
ginning with the early formation of the embryo, resistance to phenol greatly
increases. Suffice it to say that, with a phenol concentration of 100
mg/liter, zope eggs in the early stages of development die 8 times more
rapidly than in the stage of formation of the embryo. After emergence from
the shell, resistance of the embryos decreases greatly and embryos without
shells die in half the time as those still in the shells. The significant
decrease in the resistance of embryos after hatching from the shell indi-
cates the great significance of the shell, preventing penetration of the
poison and its accumulation in the organisms during the embryonal period of
development.
158
During subsequent ontogenetic development, resistance of fish to phenol
continues to drop. The survival time of zope larvae in the stage of mixed
feeding in phenol solutions of 100 and 150 mg/liter was found to be 48 and
30 hours, respectively. This is 1/5 the time of survival of the embryos in
the stage of beginning of pulsation of the heart (240 hours) and 1/2 the
time of survival of hatched embryos. Whereas, during the embryonal period
of development, the toxicoresistance of the zope undergoes significant
changes throughout the entire larval period of development; i.e., at the
beginning, middle and end, it remains more or less at the same level. Then,
in the early fry period of development, the stability of the zope to phenol
drops greatly (by a factor of more than 10) and the mean survival time in
phenol solutions of 150 and 100 mg/liter becomes 2-3 hours. However, the
least resistance was noted for mature zope, which survived only 6-8 hours in
a phenol solution of 25 mg/liter, i.e., 1/4-1/6 the concentration used in
the experiments with the fry. Let us recall that the eggs, embryos and lar-
vae survive and develop without any significant deviations from the norm in
a solution of this concentration. In order to cause death of eggs in this
same time interval, the concentration of phenol must be increased to 1000
mg/liter, i.e., by a factor of 40.
Thus, the resistance of the zope in the early stages of ontogenesis to
one of the most widespread organic poisons, phenol, undergoes significant
changes. The least resistance is that of the eggs in the stage of gastrula-
tion; the greatest, that of the eggs in the stage of pulsation of the heart.
Subsequently, the level of toxicoresistance decreases continually from
hatching embryo to larvae, from larvae to fry and fry to adults. An analo-
gous variation was observed in experiments with eggs, hatched embryos, lar-
vae, fry and mature individuals of another species of the genus Abramus, the
common bream.
In experiments with still another species of carp (Carassius carassius),
we succeeded in comparing the toxicoresistance of four age groups: current
year's brood, 1-, 2- and 3-year fish (Lukyanenko and Flerov 1963). The
criterion of resistance was the time of survival of experimental fish in
toxic solutions of phenol (17-800 mg/liter). As was to be expected, the
most resistant carp was the current year's brood, which survived many times
longer than older fish. For example, in a phenol solution of 50 mg/liter,
the mean survival time of the current year's brood was 137.4 hours, of fish
1-2 years old - 34.9 hours, of fish which had completed 2 years of life -
12.4 hours, of fish over 3 years old - 5.7 hours. Analysis of these mate-
rials indicates that the survival time of the current year's brood in com-
parison to carp 1+ years old is 3.9 times greater, than that of carp 1+
years old in comparison to carp 2+ years old 2.8 times greater. The dif-
ference between the next two age groups (2+ and 3+ years) is still less, a
factor of 2. The impression is gained that, as age increases, the resist-
ance of the fish, after reaching a certain level, undergoes only moderate
changes. However, there is no doubt that fish in the younger age groups are
more resistant to phenol than fish in the older age groups.
This is also indicated by the results of a comparative study of the
level of toxicoresistance of the current year's brood and two-year-old rain-
bow trout (Salmo irideus Gibb) which we performed (Lukyanenko and Flerov
159
1956) using the phenol intoxication model. The elevated resistance of the
current year's trout brood in comparison to 2+ year old individuals was re-
flected both in the absolute values of CLM (minimal lethal concentration),
CMT (maximum tolerant concentration) and LC50 (concentration causing death
of 50% of experimental fish), as well as the mean time of survival at all
concentrations of phenol tested (5, 7,5, 12.5, 15, 20, and 25 mg/liter). In
experiments with the current year's brood, the CLM was 15 mg/liter, LC50 -
11 mg/liter, CMT - 7.5 mg/liter, while in experiments with 2+ year old fish
the figures were 10 mg/liter, 7.5 mg/liter and 5 mg/liter, respectively.
The differences between two age groups of trout in terms of time of survival
at a given concentration of phenol were still more sharply expressed. The
mean time of survival of two-year-old trout in a phenol solution at 12.5
mg/liter was only 95 minutes, i.e., less than 1/6 the survival time of the
current year's brood - 601 minutes. No less demonstrative were the differ-
ences found in comparison of times of survival of the current year's brood
(272 minutes) and two-year-old fish (40 minutes) in a solution of 15
mg/liter phenol, survival being almost 7 times longer for the current year's
brood.
The increased resistance of younger age groups, which we found in our
experiments with phenol in highly resistant carp and more susceptible trout,
indicates that what we have here is a general regularity of reactions of
fish of different levels of organization to organic poisons. In order to
test this assumption, we performed experiments (Kokoza 1970) on fry, 35-70
days of age, of three species of sturgeon: the Russian sturgeon, Caspian
sturgeon and sterlet, representing the evolutionarily more ancient group of
cartilagenous fish. The experiments involved phenol at 50 mg/liter. We
will not take the time to present the results of this series of experiments
in detail, but rather shall note only the clearly expressed specific differ-
ences in the level of toxicoresistance, manifested in the fry period of
development. The mean survival time of 40-45 day old fry of Russian stur-
geon (12 hours 24 minutes) was 4 times greater than that of sterlet fry of
the same age (3 hours 05 minutes), and 2.6 times greater than that of
Caspian sturgeon of the same age (4 hours 40 minutes). Sexually mature
Russian sturgeon, which survived in a phenol solution of 40 mg/liter for 5
hours 30 minutes, were also characterized by higher toxicoresistance in com-
parison to the Caspian sturgeon (1 hour 20 minutes) and sterlet (1 hour 35
minutes) (Lukyanenko 1967).
However, in this case, we would like to concentrate our primary atten-
tion, not on the specific differences of toxicoresistance of the sturgeons
during their fry period of life, but rather on age differences, i.e., to
compare the time of survival of mature individuals of each of the three
species and 1-2 month fry of the same species. This comparison showed
clearly that the resistance of mature fish, as indicated by survival time in
phenol solutions of similar concentrations (40 and 50 mg/liter), is only 1/2
to 1/3 the resistance of fry. In other words, the conclusion which we have
reached, that of decreasing level of resistance of fish with increasing age
in terms of organic poisons, is true not only for the evolutionarily young
and highly organized bony fish, but also for the cartilagenous fish, lower
on the evolutionary scale.
160
The materials, which we have accumulated in our laboratory In the past
10 years, indicate clearly that the resistance of various groups of fish to
poisons differs in different stages of ontogenesis. Periods of high stabil-
ity (eggs in the stage of pulsating heart, larvae in stage C2 and current
year fish) alternate with periods of low resistance (eggs in the stage of
gastrulation, fry in their early period, sexually immature individuals).
Particular attention should be given to the end of the larval period of
development and the beginning of the fry period, when toxicoresistance drops
sharply, approaching that of mature individuals, or falling somewhat below
it. On the whole, however, the resistance of various species to organic
poisons decreases with continuing ontogenetic development and reaches its
minimum in mature individuals. We relate this fact, observed repeatedly in
our laboratory, to the formation of various functional systems in the or-
ganism in ontogenesis and their neurohormonal control, which determines the
level of reactivity of the entire organism to various physical and chemical
irritants. An important role should be paid by the central nervous system
and its synaptic structures, since the toxic effects of many poisons in the
organic series are manifested by disruption of this activity, and conse-
quently dysfunction of the basic physiologic systems (Lukyanenko 1967).
This point of view is held by a number of domestic researchers. In the
opinion of O.I. Shmalgauzen (1973), the younger stages of sturgeons (Caspian
sturgeon and Russian sturgeon) are more resistant to phenol than larvae as
they go over to active feeding. Whereas, a phenol concentration of 40
mg/liter is sublethal for eggs and only the teratogenic effect of phenol is
manifested, for larvae which have begun active feeding this concentration of
phenol is lethal. Larvae die with symptoms of acute phenol poisoning, des-
cribed for mature fish by O.I. Shmalgauzen (1973), indicating the "phenol
acts" on the larvae as a poison specifically damaging the nervous system
(page 7).
An objective study of the resistance of various species of fish in early
ontogenesis to certain toxins was undertaken by A.F. Samylin (1974). Com-
paring the resistance of Salmo salar to ammonium carbonate during various
periods of ontogenesis, he came to the conclusion that as the eggs of the
fish increased, the survival time in the same concentrations of the sub-
stance decreased. A similar picture was observed in experiments with urea
(carbamide): Fry were less resistant to this toxin than eggs and larvae.
The decrease in resistance of salmon with increasing age observed in this
experiment was also seen in experiments with three pesticides; hexachlorane,
pentachlorophenol and copper naphthenate. We must note that the toxins used
in this work differ significantly in their mechanism of action and a number
of other properties, particularly their cumulative properties. Whereas am-
monium carbonate is a physiologically cumulative poison, hexachlorane is a
materially cumulative poison. Nevertheless, a decrease in toxicoresistance
with increasing age was observed in experiments with all of the substances.
Summarizing the results of the experiments, performed with five different
toxins, differing greatly in their degree of toxicity, the author emphasizes
that as ontogenetic development continues, the resistance of the salmon de-
creases. In complete agreement with our earlier published data on the age
dynamics of the resistance of fish to phenol (Lukyanenko 1967), A.F. Samylin
(1974), concludes that there is a significant change in the level of toxi-
161
coresistance during various periods of ontogenesis. The least resistance
was noted in salmon in the stage of gastrulation in the embryonal period of
development, during transition of larvae to active feeding in the larval
period of development and during transformation of larvae to fry, i.e., in
the early fry period of development. These "points" of decreased resistance
of each of the three stages of early ontogenesis, found in experiments on
salmon with various toxins, are identical to those which we found in our ex-
periments (Volodin, et al_. 1966) with phenol using the eggs and larvae of
the zope and bream.
Thus, at the present time there is sufficient proof of increased resist-
ance of the early stages of ontogenesis, primarily the embryonal period of
development, to organic poisons. The materials of a number of authors, in-
dicating that the resistance of fish in the early stages of ontogenesis to
organic poisons is significantly less than that of mature fish, are not in
agreement. For example, according to S.A. Patin (1977), developing eggs and
particularly larvae of the Stauridae are hundreds of times less resistant to
the effects of polychlorinated biphenols than are mature fish of related
species. He also noted higher resistance of embryonal and larval periods of
development in comparison to mature individuals in experiments with other
organic poisons, petroleum and surfactants. Recalling that these data do
not agree with many reports in the literature on the elevated resistance of
the embryonal period of life of fish obtained, true primarily with fresh-
water forms or transient forms, S.A. Patin assumes that one reason for the
disagreement is the salinity of the medium, which may change the toxic pro-
perties of detergents. We can add to this the fact noted earlier
(Lukyanenko 1967) of decreased resistance of sea fish in comparison to
fresh-water fish which, apparently, is true for all stages of individual
development.
Still, it is difficult to understand the reasons for the reduced resist-
ance of the embryonal period of life in comparison to later stages of onto-
genesis to organic poisons. However, increased toxicoresistance in the
early stages of ontogenesis, in our opinion, is quite easily explained. As
we know, fish embryos in the early stages of development are protected by
the egg shell, which is an effective barrier for foreign substances, includ-
ing toxic substances (Skadovskiy 1955). This factor causes the unique con-
ditions of influence of organic toxins on the embryonal stage of development
of fish. No matter how toxic a substance dissolved in water may be, in
order to manifest its toxicity it must penetrate the egg shell and reach the
perivitelline fluid. The toxic effect is a function of concentration of the
substance and time of action. Therefore, it can manifest its action only if
a quantity of the substance accumulates in the egg sufficient to influence
the metabolic processes of the embryo and, in the final analysis, the course
of morphogenesis. It follows from this that the more difficult it is for a
substance to penetrate the egg shell, the less toxic it is for the embryo
still in the egg. Therefore, we must realize that in those cases when we
record increased resistance for the embryonal period of development of fish
to organic poisons, it is determined not only by the fact that the substance
has little influence on the metabolism of the developing organism, but also
the fact that the concentration of the substance penetrating through the
egg shell into the perivitelline fluid is significantly lower than that dis-
162
solved in the water. Quite understandably, we can determine the true causes
for increased resistance of the embryonal period of life of fish to toxins
only if we have information on the concentration of the substance not only
in the water, but also within the egg. Of course, it is difficult to pro-
duce this information, but the first studies in this area (Rosenthal and
Sperling 1974; Dethlefsen, et al. 1975; Rosenthal, et al_. 1975; Westernhagen
and Dethlefsen 1975; Patin 1977]" confirm the existence of a relationship
between manifestation of the toxic effect and the degree of permeability of
the egg shell. True, most works have been performed with inorganic poisons,
with heavy metals, and particulary with cadmium. It has been found that the
egg shell can form strong complex bonds with the metal, thus preventing its
penetration to the embryo (Rosenthal and Sperling 1974; Westernhagen and
Dethlefsen 1975). The thicker the shell, the greater the supply of active
centers bonding the metal and the greater the quantity of metal it can ac-
cumulate. However, the coefficients of accumulation of metal by the larva
are determined not only by the morphophysiologic properties of the shell,
but also by the physical-chemical status of the metal in the water. Ionic
and molecular forms of zinc and copper, which easily form strong complexes
with biologic substrates, have a higher coefficient of accumulation than
cadmium and particularly lead, which are more frequently present in
hydro lyzed and suspended form in the marine medium. We can agree with the
opinion of those authors, who believe (Patin 1977) that adsorption of a
metal onto the egg shell does not mean that it has penetrated to the em-
bryo. Such metals as lead or cadmium, bonding firmly with the active cen-
ters in the shell, apparently find it considerably more difficult to pene-
trate into the shell than the easily soluble ionic forms of zinc or copper.
These two latter metals can penetrate into the perivitelline fluid and ac-
cumulate in the embryo. Based on the concept of increased vulnerability of
the early stages of ontogenesis for toxic substances as a whole, and heavy
metals in particular, the resistance of the eggs to zinc and copper should
be lower than the resistance of the larvae. However, according to the in-
formation of Skidmore (1974), the eggs of fish are 20 times more resistant
to the toxic effects of zinc than are the larvae, while the toxic effect of
copper, which also easily penetrates the shell barrier, is approximately the
same for eggs and larvae (Patin 1977). It follows from this that even with
respect to inorganic poisons (metals), the idea of decreased toxicoresist-
ance of the embryonal period of life requires some significant adjustment,
for two reasons:
First of all, the available factual data indicate that eggs are not less
resistant to all inorganic poisons than, say, the emerging prelarvae and
larvae (Skidmore 1974; Patin 1977, Bengtson 1974; Blexter 1977). Secondly,
and this is particularly important, the high specific surface of embryonal
and postembryonal stages of development of fish, which are small in this
period, should lead to accumulation of higher concentrations of the toxic
substance (if they penetrate the biologic membranes) than, e.g., in larger
individuals of the same species in later stages of development. In any
case, the radioecology of fish provides us with data indicating the presence
of some feedback between the specific surface of hydrobionts, including fish
eggs, and the intensity of accumulation of radioactive substances. The
smaller the dimensions of the hydrobiont and, consequently, the greater the
surface of contact with the surrounding medium, the higher the concentration
163
of the toxic substance in the organism. In order to conclude reduced re-
sistance of eggs in comparison to larvae or, say, fry, we must compare their
survival time at various concentrations of toxin actually penetrating into
the organism. Therefore, any author stating that fish eggs have reduced re-
sistance of so-called increased sensitivity must present data on the concen-
tration of the toxic substance in the developing organism. Unfortunately,
such data have not yet been presented.
As concerns the statement, sometimes seen, of increased vulnerability or
reduced resistance of eggs to organic poisons, they simply do not agree with
the multitude of factual data accumulated at the present time in both the
domestic and foreign literature (Bandt 1948; Mosevich, et al_. 1952;
Wurtz-Arle 1959; Katz and Chadwick 1961; Veselov 1965; Volodin, et al. 1966;
Lukyanenko 1967; Samylin 1974; Danilchenko 1975; Hakkila and Nilmi T973;
Wilson 1976; Wienberg 1977; Paflitscher 1976).
The increased toxicoresistance of developing eggs to organic poisons can
be easily understood if we keep in mind that most of these substances cannot
penetrate the shell or penetrate very slowly, so that is is difficult for
them to reach effective concentrations inside the shell. Thus, according to
S.A. Patin (1977), the lethal concentration (LC50) of polychlorinated bi-
phenyls are 8 times less for developing fish eggs than for larvae, which the
author correctly relates to the inability of these substances to penetrate
to the embryo through the egg shell. In earlier observations, H. Bandt
(1949) noted increased resistance of larvae to hexachlorane, which was pre-
sent at 2.5 mg/liter, many times greater than the lethal concentration for
mature roach, his test species. Studying the toxicity of organic compounds
of tin or eggs and larvae of several bony fish and cartilagenous fish (stur-
geons), P.O. Danilchenko (1975), on the example of triethyl tin chloride,
showed that embryonal development occurs in bony fish in solutions of this
substance 10 times greater; in sturgeons, 100 times greater than the concen-
tration in which prelarvae survive.
The decreased penetration of the shell for most organic poisons does not
of course mean that they do not penetrate into the perivitelline fluid at
all and do not reach the embryo. Organic chlorine pesticides, for example,
have been found in the eggs (Dethlefsen 1975), but they are apparently ad-
sorbed on the surface of the egg and only cases of high concentration and
permeability disorders of the shell have a toxic effect on the embryo.
The increased sensitivity of eggs to toxins of various natures, as well
as the difficulties arising in interpretation of experimental data obtained
in experiments on eggs, lead to the need to use other substrates as test
data in ichthyotoxicologic studies in evaluating the level of resistance of
fish in the early stages of ontogenesis. Prelarvae, larvae and, parti-
cularly, fish fry which, like mature individuals (after the transition to
gill breathing), have direct contact with the toxic agents, i.e., are under
conditions comparable to those in which experiments are performed on mature
fish, have doubtless advantages. Therefore, from the practical standpoint,
our primary emphasis must be on data characterizing the dynamics of toxi-
coresistance of fish in the larval and fry periods of life, both to organic
and inorganic poisons.
164
In the first part of our report, we analyzed the age specifics of the
resistance of bony fish and cartilagenous fish in the larval and fry stage
of life, using the model of phenol intoxication of fish performed in our
laboratory. The fact of gradually decreasing resistance from larvae to fry
and from fry to immature individual, we found has been repeated by many re-
searchers in experiments with other organic poisons, including pesticides
and detergents.
In contrast to organic poisons, toxic substances of inorganic nature
and, in particular, heavy metal salts, are most toxic for fish "in the lar-
val and fry stages" (Stroganov and Pazhitkov 1941). However, what are the
dynamics of toxicoresistance of fish in the larval and fry periods of life,
i.e., in the early stages of ontogenesis, we do not know due to the sparse
nature of studies of this problem. D. Blaxter (1975) considers, for
example, that the "sensitivity" of plaice larvae (meaning decreased resist-
ance) increases with age. If "young" larvae survive in 1000 pg Cu/liter,
32-42 day larvae died at a concentration as low as 300 pg Cu/liter. G.
Larson, et aV. (1977) studies the acute toxicity of inorganic chloramino
compounds for larvae with the yellow sac, fry and juvenile American brook
trout (Salvelinus fontinalis). The fry were less resistance than the larvae
and the lethal concentration (LC50) of inorganic chloramines at 96 hours ex-
posure for them was 82 yg/liter, for larvae with the yellow sac - 90-105 yg/
liter. In the larvae, a decrease was noted in the resistance with increase
in body weight.
In our laboratory in the last three years, we have performed a cycle of
studies involving students from the ARE - Abbas Said Abu El-Ess, and from
Iraq - Talyal Al Kubeysi and Adnan Musa Edzhad - on the age dynamics of
toxicoresistance of larvae and fry of sturgeons with respect to common
metals, cadmium and cobalt.
The experiments were performed on 1, 5, 10, 20 and 30-day-old larvae, as
well as 40, 60, 90 and 120-day-old fry of the giant sturgeon, Russian stur-
geon and Caspian sturgeon. We used the following concentrations of salts:
cadmium chloride - 0.01, 0.1, 0.5, 1, 2, 4, 5, 8 and 10 mg/liter; cobalt
chloride - 0.1, 1, 4, 5, 8, 10, 16, 32 and 64 mg/liter. The indication of
resistance of the larvae and fry was the percentage of deaths and the time
of survival in a solution of a given concentration of toxic substance. The
duration of the experiments was 48 hours; observations were performed around
the clock.
Summarizing the results of many series of experiments in this cycle, we
conclude that the level of toxicoresistance of larvae and fry of these stur-
geons differs significantly and that the larvae are significantly less re-
sistant in comparison to the fry. However, within each of these two age
groups of early ontogenesis, there is a significant change in toicoresist-
ance, as indicated by the percentage and time of death of fish at the same
concentration, as well as the threshold lethal concentration. For example,
the toxicoresistance of the Russian sturgeon gradually decreases from the
early stages of larval development to later stages, becoming minimal in the
transition period (from larval to fry), then increases once more from the
early age group to the later age groups, reaching a rather high level by the
165
60th day of age. Whereas, in the fry period of life in all three species we
see the same direction of change of toxicoresistance (an increase from
younger age to older age), in the larval period of li1^e we see species
specificity of the dynamics of toxicoresistance. In the giant sturgeon, the
10-day-old larvae were least resistant; in the Caspian sturgeon, the 20-day-
old larvae; in the Russian sturgeon, the 30-day-old larvae.
Among the three species of sturgeons studied, the larvae of the giant
sturgeon were least resistant to the salts of heavy metals, the larvae of
the Caspian sturgeon were most resistant. The larvae of the Russian stur-
geon occupied an intermediate position. The species specificities of toxi-
coresistance, which we observed, were manifested for each of the three in-
dexes, lethal concentration, percent death and time of survival of experi-
mental larvae in toxic solutions. For example, the lethal concentrations of
cadmium chloride for larvae of the giant sturgeon of various ages were 0.1-1
mg/liter (LC50 = 0.5 mg/liter); cobalt chloride, 0.1-10 mg/liter (LC50 10
mg/liter). A change in concentration of cadmium chloride by a factor of 100
had practically no influence on the level of toxicoresistance of the giant
sturgeon in early ontogenesis, and the mean time of survival did not undergo
significant changes in any of the three age groups of larvae. This is also
fully true of the level of resistance of various age groups of larvae of the
giant sturgeon in relationship to cobalt, although its toxicity is about
1/10 the toxicity of cadmium chloride.
The lethal concentration of cadmium chloride (LCioo) for Russian stur-
geon (4 mg/liter) was 1/2 that for the giant sturgeon (8 mg/liter). The
elevated resistance of Caspian sturgeon larvae, in comparison to Russian
sturgeon, was also found in experiments with cobalt chloride, lethal concen-
trations of which were 64 and 32 mg/liter, respectively.
Age variability and the level of toxicoresistance in the early stages of
ontogenesis are determined primarily by the degree of formation of various
functional systems, to a lesser extent by changes in size (mass) of the
body. A change in body mass by a factor of 4 for 10-120 day old fry (from 3
to 12 g) does not lead to any significant increase in the survival time of
the fry of Russian sturgeon in toxic solutions of the metals studies.
As we know, cadmium is a highly toxic metal. Suffice it to say that the
lethal concentrations of this metal for many species of fresh-water and
marine fish fall in the range of 0.01-2 mg/liter (Lukyanenko 1976; Patin
1977). However, according to our data, a concentration of cadmium chloride
of 4 mg/liter leads to the death of 10-day-old Russian sturgeon larvae in
14.6 hours; of 20-day-old larvae in 29.7 hours; 30-day-old larvae in 8.5
hours; while 60-day-old fry survive for 48 hours. Furthermore, 4-month-old
fry survive in a solution of cadmium chloride of 8 mg/liter for 48 hours
(only 105 of the experimental animals die). All of these data indicate that
the cartilagenous fish, in this case Russian sturgeon, are significantly
more resistant to the toxic effect of cadmium in comparison to marine and
fresh-water species of bony fish in the early stages of ontogenesis.
166
Summing up our report on the age specifics of the sensitity and resist-
ance of fish to poisons, I would like to draw the attention of participants
in the symposium to still another "jery important, in my opinion, question.
I am speaking of the great need for a clear delineation between the concepts
of "sensitivity" and "resistance" of fish to poisons, which are quite dif-
ferent in their physiologic and toxicologic significance (Lukyanenko 1967).
Unfortunately, quite frequently in both domestic and foreign literature, the
concept of sensitivity and that of resistance of hydrobionts to various fac-
tors in the aquatic environment, as well as toxins, are either identified or
sensitivity is considered to be the reverse of resistance. The use of these
concepts as synonyms can lead and does lead to negative results, including
difficulty in understanding the degree of scientific foundation of the con-
clusion of various authors who have estimated the age differences of toxi-
coresistance of fish.
There is a generally agreed idea, concerning the meaning of the concept
of resistance of an organism to abiotic factors in the environment, concern-
ing toxins of various natures. An estimate of the degree of resistance is
based either on the concentration of the substance causing death of a cer-
tain percentage of experimental animals (LC50 or LC]on) in a certain period
of time (24-48-96 hours or more), or the time of survival in a toxic solu-
tion of a predetermined concentration. Resistance is the capacity to sur-
vive low concentrations of a toxic substance for longer periods of time, or
to survive higher concentrations of the same substance for a fixed short
period of time by the operation of various regulatory mechanisms. Quite
understandably, the earlier these regulatory mechanisms are brought into
play (detoxication, excretion of the substance, etc.), supporting short-term
or long-term adaptation of the organism to the toxic agent, the longer will
be the time of survival of the organism and the more probable that, in the
case of interruption of the toxic effect on the organism, it will survive.
However, it is also obvious that regulatory mechanisms will be brought into
play earlier, the more sensitive the organism is to the toxin at the given
stage of individual development.
In terms of their physiologic content, the concept of "sensitivity" is
close to or coincides with the concept of "excitability", the level of which
determines the threshold of excitability. In turn, a measure of excit-
ability is the minimum force of an irritant; in this case a chemical factor,
which exceeds the threshold of irritation. The greater the minimum force
of the chemical irritant necessary to call forth a reaction, the higher the
threshold of irritation, the lower the excitability, the lower the sensi-
tivity of the organism to the substance in question. Quite understandably,
the lower the threshold of irritation, the higher the excitablity, and the
higher the sensitivity. This is a generally known physiologic truth, in
light of which we must analyze the question of sensitivity of the organism
or cell to a toxic irritant. It follows from all of this that, in order to
estimate the level of sensitivity of the organism to a given toxin, the
question of the primary reaction of the organism to this irritant is of pri-
mary significance. I propose that there is no need to prove that neither
the concentration of the substance causing the death of a certain percentage
of experimental fish, nor the time of survival of fish at a fixed concentra-
tion, can be used in any way as an indication of the primary reaction to a
167
chemical irritant. It becomes obvious from this that the widespread concept
of sensitivity of fish to a poison as the "inverse of resistance" is without
foundation.
We turned our attention to this inconsistency more than 10 years ago
(Lukyanenko 1967) in our study of specific peculiarities of the toxicore-
sistance of mature fish to poisons on the model of phenol intoxication.
Using rapid motor activity as an indication of the primary reaction of
mature fish to the phenol irritant, its latent period, and the time of sur-
vival of the experimental fish as an indication of stability, we proved
(Lukyanenko and Flerov 1965) that high sensitivity of a species is not al-
ways accompanied by low resistance and vice-versa. Of course, our concept
of the degree of sensitivity of fish to various toxins will change depending
on which functional system is selected as the indication of primary reac-
tion. Everything is determined by the understanding of the mechanism of ac-
tion of the toxic substance being studied, and the precise knowledge of the
"functional target", since only using this function can we adequately deter-
mine the level of sensitivity. It is difficult to determine the target
function, even in mature fish, to say nothing of the early stages of onto-
genetic development and especially embryonal development. In the embryonal
period, a toxic substance which penetrates the shell in many cases has its
harmful influence not on organs and functions as such, but rather on pro-
cesses determining the development of organs or the genesis of functions.
If we agree with the current opinion (Bocharov 1975) that the sensitivity of
the developing organism varies in various portions of the embryo, the task
of evaluating the sensitivity of the embryo as a whole becomes still more
difficult and responsible.
However, in many works dedicated to the toxicology of embryonal or lar-
val stages of development of fish, the concept of "sensitivity" is used
quite broadly and most frequently as the reverse of resistance. Therefore,
the decreasing stability of developing larvae to a toxin is taken as evi-
dence of increased sensitivity in comparison to mature, fully formed indivi-
duals of the same species. If we agree with this point of view, we must say
that the organism of the fish as it develops, accompanied by formation of
organs and development of functions, including the receptor function of the
peripheral nervous system, somehow loses its sensitvity to chemical irri-
tants (in this case toxins) in comparison to the developing embryo. From
the physiologic standpoint, this interpretation of the change in sensitivity
of the organism in onotogenesis is hardly acceptable. The developing egg
contacts the surrounding medium and, consequently, receives external irri-
tants with its entire surface. If a chemical substance which has toxic pro-
perties penetrates through the shell, its reception may be performed by the
plasmatic membrane of the cells of the developing embryo, the ancient func-
tion of which is the reception of stimuli. However, it is hardly possible
that the sensitivity, i.e., excitability of these cells, which are simple
acceptor-receptor systems, could be higher than that of the specialized ner-
vous system of a complex multicell organism such as a mature fish, respon-
sible for the function of reception, conduct and acceptance of stimuli of
physical or chemical nature.
168
We propose that in describing the reactions of fish to toxic irritants
in the embryonal and immediate postembryonal periods of development (prelar-
val and larval), the concept of resistance be universally used. Sensitivity
or susceptibility can be spoken of only if it is specially studied using
adequate methods of investigation.
Returning to the primary point of the present report, I would like to
emphasize that over the past decade, new data have been obtained, indicating
the presence of clear age specifics in the sensitivity of fish to poisons.
However, the level of toxicoresistance is determined not only by the direc-
tion and intensity of metabolic processes of fish in various stages of onto-
genesis, but also by the nature of the toxic agent used. The resistance of
various species of fish to many organic poisons decreases with ontogenetic
development and reaches a minimum in sexually mature fish. However, this
process is not uniform and periods of high resistance (egg in stage of pul-
sating heart, larva in C2 stage and current year's brood) alternate with
periods of low resistance (egg in stage of gastrulation, larva at end of
larval period, immature individuals). Particular attention should be given
to the end of the larval and the beginning of the fry period of development,
when the resistance of fish to organic poisons drops sharply. As concerns
the resistance of fish to inorganic poisons and, in particular, to heavy
metal salts, it is minimal in the larval and fry period of individual devel-
opment. The resistance of the fry (embryonal period of development), both
to organic and to inorganic poisons, is significantly higher in comparison
to the larval and fry periods. The nature of the increased toxicoresistance
of the egg remains unclear. This factor makes the use of eggs as test ob-
jects (reference objects) undesirable in studies of the degree of toxicity
of various substances for various stages of the ontogenesis of fish and
biologic testing of natural and waste waters (larvae and fry are prefer-
able).
REFERENCES
Danilchenko, O.P. 1975. Effects of toxic substances on certain fresh-water
bony and cartilagenous fishes in the embryonal period of development.
Cand. Diss., Moscow State University, 150 pp.
Grimm, O.A. 1896. Kaspiisko-volzhskoe rybolovstvo, St. Peterburg, 153 pp.
Lukyanenko, V.I. and B.A. Flerov. 1963. Toxicoresistance of current year's
brood of carp. Materialy po biologii i gidrologii volzhskikh vodokhran-
ilishch. Izd. AN SSSR, Moscow-Leningrad.
Lukyanenko, V.I. and B.A. Flerov. 1963. Materials on the age toxicology of
fish. Farmakol. i Toksikol., No. 5.
Lukyanenko, V.I. and B.A. Flerov. 1965. Species peculiarities in the sen-
tivity and resistance of fish to phenol. Gidrobiologicheskii Zhurnal,
No. 2.
169
Lukyanenko, V.I. and B.A. Flerov. 1966. Comparactive study of the resist-
ance of two age groups of rainbow trout to the toxic effects of phenol.
Biologiya ryb bolzhskikh.
Lukyanenko, V.I. 1967. Toksikologiya ryb (Toxicology of fish), Moscow,
Pishchevaya promyshlennost' Press, 216 pp.
Lukyanenko, V.I. 1973. Physiologic criterion and methods of determination
of toxicity in ichthyology, Eksperimental 'naya vodnaya toksikologiya.
No. 4, pp. 10-30
Vernidub, M.F. 1962. Experimental analysis of processes caused by poison-
ing with nonvolatile (resinous) phenols in the Baltic salmon during the
larval period of life. Uchenyye zapiski LGU Seriya biologicheskikh
nauk.. No. 48.
Veselov, Ye. A., I.V. Pomazovskaya, Ye. I. Remezova, and S.Ye. Cherepanov.
1965. Toxic effect of hexachlorane on fish and aquatic invertebrates.
Voprosy gidrobiologii , Moscow, Nauka Press, p. 65.
Volodin, V.M., V.I. Lukyanenko, and B.A. Flerov. 1965. Dynamics of changes
in the resistance of fish to phenol in early stages of ontogenesis.
Voprosy gidrobiologii, Moscow, Nauka Press, p. 82.
Volodin, V.M., V.I. Lukyanenko, and B.A. Flerov. 1966. Comparative des-
cription of the resistance of fish to phenol in early stages of onto-
genesis. Biologiya ryb volzhskikh vodokhranilishch. Moscow-Leningrad,
Nauka Press, pp. 300-310.
170
SECTION 13
SYNERGISTIC EFFECTS OF PHOSPHORUS AND HEAVY METAL LOADINGS ON
GREAT LAKES PHYTOPLANKTON
E.F. Stoermer, L Sicko-Goad and D. Lazinskyl
INTRODUCTION
The Laurentian Great Lakes are one of the major physiographic features
of North America. They represent a tremendous resource to the people of
Canada and the United States. They provided European colonizers a route of
access to the interior of the continent and continue to provide an important
transportation artery, particularly for the raw materials of heavy industry.
In the early decades of the present century the Great Lakes supported an im-
portant fishing industry and their waters furnished a seemingly inexhaust-
ible supply of high quality potable water and industrial process and cool-
ing water. As a result of these favorable circumstances the shores of the
Great Lakes were a favored site for early settlement and have supported the
growth of several major population and industrial centers.
Unfortunately, the byproducts of these populations and industrial con-
centrations have had effects on the Great Lakes ecosystem which damage the
yery resource potential which allowed their growth and development. During
the past several decades important fish stocks have been severely damaged
or, in some cases, entirely lost. Some of the stocks remaining have been
contaminated by heavy metals or organics to the point that there are serious
questions regarding their suitability for human consumption. Eutrophication
has also caused modifications in the composition and abundance of primary
producer communities which have had direct effects on the utility of Great
Lakes waters. Overproduction and changes in composition of the phytoplank-
ton assemblages of the Great Lakes have led to taste and odor problems in
municipal water supplies and additional treatment costs for removal of
biological materials from the water. Extreme overproductivity of benthic
communities has resulted in nuisance growths of attached algae such as
Cladophora.
These problems have been recognized and considerable effort has been
directed towards defining the causes of water quality and associated re-
^Great Lakes Research Division, University of Michigan, Ann Arbor, Michigan
48109.
171
source deterioration and implementing management strategies which will con-
trol or eliminate the particular problems. In many cases management strate-
gies are clearly evident and considerable success has been obtained by their
implementation. Perhaps the clearest case of success is the restriction of
use of certain chlorinated hydrocarbon pesticides which has reduced the con-
tamination levels of Great Lakes fish. In the Great Lakes system primary
productivity is clearly controlled by phosphorus availability and efforts
are underway to limit inputs of this material to the system. This limita-
tion has proven more difficult to implement and positive effects, to this
point, have not been dramatic.
As we become more familiar with the characteristics of the Great Lakes
ecosystem it becomes more and more apparent that effective management will
demand a detailed understanding of ecosystem characteristics and functional
relationships in order to develop management strategies which can control
subtle and multiplicative causes of ecosystem deterioration. Consideration
of the unique characteristics of the Laurentian Great Lakes leads to the
conclusion that these bodies of water may present the most demanding chal-
lenge to effective water quality management found in any freshwater system.
Several considerations are involved in this conclusion:
1. In their pristine state the Laurentian Great Lakes were an
almost perfectly exploitable system. They were a source of
water which could be utilized without extensive treatment
and supported a fishery for very highly valuable species.
They were also a source of aesthetic enjoyment and recrea-
tional activities for a significant portion of the popula-
tion. Minimal levels of perturbation led to disproportion-
ately large damage to the resource potential compared to
other systems.
2. The Great Lakes are a geologically ^^ery young ecosystem, com-
pared to most large lakes of the world. The fauna and flora
are unique but have not had time to develop stable adaptations
to their environment. Such communities might be expected to
be particularly susceptible to environmental perturbation and
this expectation has been realized in the history of biological
changes observed.
3. The Great Lakes are ^^ery long residence-time systems compared
to most other freshwater biotopes. This means that introduced
contaminants may have very prolonged effects.
4. Because of the great dilution volume of the Great Lakes con-
taminants may be present in quantities so low that they are
difficult to measure by conventional chemical methods although
their effects may be crucial to the biota.
5. It is quite clear that the classification and perception of
water quality developed for other freshwater systems is not ap-
propriate for the Great Lakes. Paradoxically, drastic and
possibly irreversible modifications of the Great Lakes eco-
172
system have occurred in regions that would be classified as
"oligotrophic" according to the normal criteria.
In the following report we will attempt to address some of the interac-
tive effects of two types of contaminant loadings, phosphorus and heavy
metals, which might not be discerned by conventional limnological methods.
The research was originally initiated in an attempt to explain the apparent
differential influence of phosphorus enrichment on particular species of
phytoplankton advected through zones of phosphorus pollution. Loadings,
biological availability, and biological pathways of this nutrient in the
Great Lakes system are of particular interest because it is the primary
nutrient controlling eutrophication. Most undesirable anthropogenic modi-
fications of the Great Lakes ecosystem are directly related to increased
phosphorus loadings resulting from increased population densities, intro-
duction and widespread usage of phosphorus containing detergents, and poor
land management practices. In the course of this investigation we found
that the mechanism allowing differential sequestering of phosphorus was
intimately assotiated with heavy metal concentration in the water and that
the same mechanism could permit excessive uptake of certain toxic metals.
Since this bioaccumulation mechanism could have both effects on the aquatic
ecosystem and potential effects on human health we have attempted to deter-
mine some of the factors involved.
Since the problem we are dealing with has not, to our knowledge, been
previously investigated in the context of large lake limnology and since
some of the methods we have adopted have not been widely employed in water
quality investigations it would perhaps be helpful to give a brief chronolo-
gical outline of the development of this investination before discussing re-
sults.
During an investigation of Saginaw Bay, one of the more grossly polluted
regions within the Great Lakes ecosystem, it became apparent that certain
species of phytoplankton were surviving transport out of the bay into Lake
Huron. This was unexpected because the species involved have high nutrient
requirements which cannot be satisfied in Lake Huron. We hypothesized that
populations within the bay were taking up phosphorus in gross excess of
their immediate physiological requirements and subsequently surviving trans-
port out of the nutrient-rich environment by using these internal stores.
In order to verify this hypothesis we examined the internal cellular con-
stituents of these populations by analytical electron microscopy. This ana-
lysis confirmed the presence of internal stores of phosphorus in the form of
polyphosphate bodies. X-ray analysis further showed that the polyphosphate
bodies also contained appreciable quantities of lead. Subsequent field ob-
servations in areas subjected to combined phosphorus enrichment and heavy
metal contamination indicate that the phenomenon observed in Saginaw Bay is
common in other parts of the Great Lakes system. Laboratory studies were
also carried out to determine if other metals behave in the same manner as
Pb.
173
MATERIALS AND METHODS
The observations reported here come from natural phytoplankton assem-
blages collected and fixed under field conditions, natural assemblages
brought into the laboratory and subjected to experimental nutrient and heavy
metal additions, and populations isolated from the lakes and maintained in
the laboratory.
Culture Conditions
Natural assemblages used for experiments were returned to the laboratory
within 5 hours of collection in 20-£ prerinsed plastic containers. Contain-
ers were placed in an insulated, light-tight box for transport to avoid
temperature and light shock. In the laboratory experimental material was
maintained in a culture chamber at the temperature of collection (+ 1.0°C),
and 200 y Ein m"2 sec-^ of illumination on an alternating 16-hr day, 8-hr
night cycle.
Cultured material was grown in FM medium (Lin and Schelske 1978) at 15°C
at the same illumination and daylength conditions used for natural assem-
blages.
Light Microscopy
All observations reported were made with a Leitz Ortholux microscope
with irmiersion objectives furnishing numerical aperature of at least 1.30.
Cells were stained for polyphosphates by the method of Ebel et al^. (1958)
and were observed and photographed either in temporary aqueous mounts or in
permanent mounts embedded in Epon prepared by the same method used for elec-
tron microscopy. Photographs were taken with a Leitz Orthomat photo appara-
tus.
Electron Microscopy
Material was fixed with 3% (vol. /vol.) biological grade glutaraldehyde
in 0.05 M cacodylate buffer (pH 7.2) for one hour at 4°C and post-fixed in
1% OSO4 for 1 hour. Cells were dehydrated in a graded ethanol -propylene
oxide series and embedded in Epon (Luft 1961).
Thin sections were cut with a diamond knife, collected on 300 mesh grids
and stained with uranyl acetate (Stempak and Ward 1964). Sections were exa-
mined on a Zeiss EM 9S-2 electron microscope. Microscope magnification
calibrations were made by use of a grating replica.
X-Ray Analysis
Sections for X-ray analysis approximately 60 nm thick were cut with a
diamond knife and collected on 75X300 mesh titanium grids. Sections were
examined at 100 KV in STEM mode in a JEM lOOC electron microscope equipped
with a KEVEX series 7000 energy dispersive X-ray analysis system. The
specimen was tilted 30° toward the detector. Specimen to detector distance
was 18 mm. Spot analysis of inclusions was made with a spot size of 50 A.
174
Stereology
Quantitative estimates of cellular components were developed by techni-
ques described by Sicko-Goad et al_. (1977). Fifty micrographs were examined
for each experimental treatment analyzed. A transparent 12.5 mm square sam-
pling lattice was superimposed over the micrographs for point count measure-
ments. Although several sections were collected on one grid, only one sec-
tion per grid was used in the analysis. Blocks were retrimmed after each
series of sections had been cut in order to avoid repeated sampling of adja-
cent material within the same organism. For species where cells are con-
nected in a colony, only one cell per colony was included in the statistical
sample.
RESULTS
Figure 1 shows the distribution of Fragilaria capucina Desm. in southern
Lake Huron in June of 1974. This distribution is atypical in that this
species generally becomes abundant in areas of the Laurentian Great Lakes
which are severly eutrophied (Hohn 1969) but does not survive in the less
nutrient rich offshore waters. Electron micrographs of cells of this
species taken within Saginaw Bay (Figure 2) show that they contain numerous
small vacuolar inclusions having the general form and appearance of poly-
phosphate bodies. Although the formation of polyphosphate bodies has not
been widely reported in eukaryotic phytoplankton organisms. X-ray analyses
of the inclusions (Figure 3) confirm that their elemental composition is es-
sentially similar to that of polyphosphate bodies reported from prokaryotic
organisms (Sicko-Goad et aj_. 1975). The primary difference is that the
bodies found in Fragilaria capucina are much smaller than those found in
most prokaryotic organisms and that they are found within the vacuole of the
eukaryotic cells.
X-ray spectra of the polyphosphate bodies found in Fragilaria capucina
in this locality also indicate the presence of appreciable quantities of Pb
as a constituent of the bodies (Figure 3).
Observations of other eutrophication tolerant phytoplankton species in
Saginaw Bay indicated the widespread occurrence of polyphosphate bodies,
even in areas where chemical analyses of the water showed low levels of dis-
solved phosphorus in the water. Polyphosphate bodies were particularly ap-
parent in cells of some of the potentially nuisance producing blue-green al-
gae in the assemblages. These observations also show that the distribution
of populations containing polyphosphate bodies within the bay is restricted
primarily to stations along the southern and southwestern shore of the bay
(Figure 4) .
Subsequent observations utilizing staining techniques which permit
visualization of polyphosphate bodies at the light microscope level (Ebel
et al_. 1958) show that polyphosphate bodies are developed in phytoplankton
populations present in several areas of the Great Lakes system which receive
relatively high loadings of phosphorus and other contaminants.
175
EAST TAWAS •
•GODERICH
4-8 June 1974
• PORT HURON
Figure 1 Outline map of the southern Lake Huron showing the distribution
of the eutrophication tolerant diatom FragilaHa ca£U£ina Desm. in the
waters of Lake Huron outside Saginaw Bay in early June 1974.
176
Figure 2. Transmission electron micrograph of a cross section of Fragilana
capucina. Numerous small polyphosphate bodies (PP) are present in the
vacuole (V). Other cytoplasmic organelles are normal. Large chloro-
plasts (c) are positioned under the valve face of the frustule (F).
Golgi apparatus (G) appears somewhat disorganized because the inter-
calary bands (B) are being formed prior to next cell division, (flagni-
fication X29,000).
Figure 3. X-ray spectrum of a polyphosphate body contained in the vacuole of
Fragilaria capucina. The labelled peaks are P (Ka) and Pb (Ma, La). A
minor calcium peak (Ka 3.69 Kev) is also present. Unlabel! ed peaks are
CI (Ka 2.62 Kev), a component of the epoxy embedding medium, and Cu (Ka
8.04, 8.02 Kev; KB 8.90 Kev), which originates from the grid.
177
Rings proportional to number of
polyphosphate body occurrences
during period of sanripling.
Figure 4. Outline map of Saginaw Bay, Lake Huron showing the abundance of
algal populations containing polyphosphate bodies in different segments
of the bay (Smith et al- 1977). Average circulation is
counterclockwise and polyphosphate bodies are most
common downstream of the Saginaw River pollution source.
178
The form and position of these inclusions is somewhat different in the
various major physiological groups of phytoplankton. Polyphosphate bodies
in the blue-green algae may become large compared to the volume of the cell
within which they are contained and their position within the cell is highly
variable (Figure 5). In the green algae, as in most other eukaryotic cells,
polyphosphate bodies are restricted mainly to the vacuole. In the species
we have examined so far, there is considerable variation in the relative
size and position of the bodies present (Figures 6 and 7).
In diatoms polyphosphate bodies are usually \/ery small (< 0.5 ym) (Fi-
gure 2) and are usually positioned near the vacuolar membrane inside the
vacuole, although they may become dispersed in the vacuole (Figures 2 and
8).
Among the flagellate groups, polyphosphate bodies similar to those found
in diatoms have been noted in various members of the Chrysophyceae (sens.
str. ) and the Prymnesiophyceae. Interestingly, they seem not to be present
in the Cryptophyceae and we have not found them in Euglenoids, although our
samples of these organisms are small, since they are very rare in the Great
Lakes.
Since we had observed accumulation of Pb, but not other metals in field
samples, we decided to test for possible differential uptake of different
metals under controlled conditions. The metals tested were Pb and copper,
which is known to be rather acutely toxic to many species of algae
(Fitzgerald and Faust 1963). A unialgal culture of Diatoma tenue var.
elongatum Lyngb., originally isolated from Lake Michigan was grown in FM
medium. Since phosphorus limitation followed by phosphorus excess is one of
the conditions known to initiate polyphosphate body formation (Jensen and
Sicko 1974) phosphorus starvation and phosphorus excess were simulated in
the following manner. Four-day-old cultures which were in logarithmic
growth (controls) were packed by gentle centrifugation, washed twice with
sterile distilled water, then inoculated into a medium of the same composi-
tion of FM medium except that it lacked phosphate salts. Cells were incu-
bated in this medium for 3 days to induce phosphorus starvation. At the end
of the starvation period, during the fourth hour of the culture light cycle,
cells were again packed by centrifugation and resuspended in one of the 3
following media as treatments:
1. Medium containing twice the phosphorus concentration of FM
medium with no other additions.
2. Medium containing twice the phosphorus concentration of FM
medium + 0.05 yg-at/K. Pb.
3. Medium containing twice the phosphorus concentration of FM
medium + 0.08 yg-at/2, Cu.
Cells were incubated under normal culture conditions in these treatments
for 2 hours then fixed and prepared for electron microscopy along with con-
trol samples. Splits of the samples were also stained for polyphosphates
and prepared for observation under the light microscope.
179
Figure 5. Transmission electron micrograph of Anacystis sp. containing large
polyphosphates bodies (PP). (X53,000).
Figure 6. Transmission electron micrograph of Scenedesmus sp. showing large
polyphosphate bodies (PP) in the vacuole. (X23,000).
Figure 7. Light micrograph of Scenedesmus sp. stained for polyphosphates by
the technique of Ebel et al . (1958) . Material is from a natural phyto-
plantkon assemblage enriched with phosphorus and heavy metals.
(XI, 700).
Figure 8. Light micrograph of Fragilaria crotonensis Kitton stained for poly-
phosphates by the technique of Ebel et al . (1958) . Material is from a
natural phytoplankton assemblage enriched with phosphorus and heavy
metals. (X800).
180
Electron micrographs of sectioned material from control cultures and all
treatments were analyzed by stereology to quantify polyphosphate body abun-
dance under the conditions tested and to determine other changes in cellular
structure which might be induced by the treatments. Sectioned material was
also subjected to X-ray analysis to verify polyphosphate body composition
and metal accumulation. The results of this analysis is given in Table 1.
Preliminary results from work currently in progress indicates that heavy
metal stress results in increased polyphosphate body formation in Plectonema
boryanum Gom. (Figures 9-11). These results further indicate a differential
effect depending on the degree of direct toxicity of the metal to the alga
subjected to the stress. In Plectonema Pb and zinc cause an approximately
10-fold increase in polyphosphate bodies per cell after 3 days exposure.
Copper and cadmium treatments result in a ca. 5-fold increase, but increased
apparent cellular damage at the ultrastructural level.
DISCUSSION
Our results are indicative of the complex and poorly understood cellular
level interactions which may occur in algal populations of large lakes sub-
jected to nutrient and toxicant contamination. Previous reports in the
literature suggest polyphosphate accumulation may be triggered by several
types of nutrient imbalance (see Sicko 1974 for review). It is important to
note that the mechanism may be triggered either by deficiency in some criti-
cal nutrient in the presence of excess exogenous phosphorus (Lawry and
Jensen 1979), stress invoked by excess levels of micronutrients, or simply
by the restoration of excess exogenous phosphorus to cells previously
stressed by deficiency of this nutrient.
Any or all of these conditions are apt to be present in mixing zones
where contaminated stream flows enter the Laurentian Great Lakes. It is
thus highly probable that rapid uptake of phosphorus in these areas is not
directly related to the immediate growth potential of the algal populations
affected. This is illustrated by our results from Saginaw Bay (Figure 4).
The normal water circulation of the bay is counterclockwise with water exit-
ing the bay along the southern shore (segments 3 and 5 in Figure 4) being
replaced by Lake Huron water entering the bay along the northern coast
(Danek and Saylor 1977). The primary source of nutrient enrichment and
heavy metal contamination is the Saginaw River (Smith et ^. 1977) which en-
ters the far southwestern tip of the bay. In this case polyphosphate bodies
are much more abundant in phytoplankton populations taken at stations down-
stream, in the sense of the average current vector, of the source than in
other segments of the bay. It further appears that phosphorus bound in this
form is transported out of the bay since polyphosphate bodies are found at
stations near the mouth of the bay. The eventual fate of this material in
the Lake Huron system cannot be determined on the basis of our observations.
We would speculate, however, that at least two effects may occur. The first
is that phosphorus bound in this form may eventually be reutilized allowing
the survival of phytoplankton populations which are usually restricted to
eutrophic areas in the open waters of Lake Huron. Other investigations
(Stoermer and Kreis, in press) have shown that populations which appear to
181
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Figure 9. Transmission electron micrograph of cytologically normal PI ectonema
boryanum. Note regular cell septae (arrow) and polyhedral bodies (PB).
(X28,000).
Figure 10. Transmission electron micrograph of PI ectonema boryanum treated
with 0.1 yg-at/£ Pb. Note increased vacuolization of the cell, ap-
parent reduction in number of polyhedral bodies, and the presence of
numerous polyphosphate bodies (PP). (X28,000).
Figure 11. Transmission electron micrograph of PI ectonema boryanum treated
with 0.1 yg-at/£ Zn. Note increased vacuolization of the cell, ap-
parent reduction in the number of polyhedral bodies, numerous poly-
phosphate bodies (PP), and lack of cell division. (X45,000).
183
originate in Saginaw Bay under certain conditions can survive transport into
the extreme southern part of Lake Huron, It is difficult to imagine this
occurring unless these populations were growing fast enough to replace
grazing and sinking losses. The other plausible effect is that death of
these populations, through grazing or other process, will release additional
phosphorus and thus stimulate eutrophication of the offshore waters of Lake
Huron. To our knowledge this type of biological loading has not been con-
sidered in the limnological literature, but it may be an important mechanism
of pollutant dispersal in the Laurentian Great Lakes.
Our data also suggest that incorporation of Pb in polyphosphate bodies
may be an important mechanism for dispersal of the toxicant in aquatic
systems. To our knowledge, our report of the polyphosphate-lead association
is the first demonstration of this mechanism in naturally occurring popula-
tions. The fact that this type of uptake can be produced in the laboratory
conditions and Crang and Jensen's (1975) demonstration of titanium incor-
poration in polyphosphate bodies in Anacystis nidulans Dr. and Daily sug-
gests that binding of heavy metals in osmotically inert inclusions such as
polyphosphate bodies could be a general mechanism for protecting phytoplank-
ton cells (at least temporarily) against heavy metal toxicity. Our results
to date suggest that this is probably not the case. Our experiments with
metals more directly toxic to algae, such as Cu and Cd, as well as Zn show
that although stress induced by the presence of these elements at relatively
low levels may induce polyphosphate body formation, these elements are not
sequestered in the polyphosphate bodies to any measurable extent. This
situation should be further investigated as it is possible that organisms
other than those so far investigated may be able to affect heavy metal in-
corporation in polyphosphate bodies or that incorporation may take place at
concentrations other than those tested.
Our results are also interesting in respect to previous reports of heavy
metal accumulation in algae. Silverberg (1975) demonstrated that Pb accumu-
lated in the cell wall and in the peripheral vacuole of Stigeoclonium tenue
(Ag.) Kutz. Silverberg (1976) also found that exposure of 3 species of
green algae to relatively high levels of Cd resulted in degenerative changes
in the mitochondria of the cells and the formation of granules within the
mitochondria which apparently contained Cd. Although we have observed some
changes in cellular organelle structure in our experiments, we have not ob-
served measurable accumulation of Cd or Zn associated with any organelle of
specific cellular site. It should be noted that the concentrations used in
Silverberg's experiments were 3 to 10 times higher than the concentrations
tested in our experiments. It is probable that the cellular modifications
he noted are symptomatic of acute toxicity.
At this stage of our investigations many questions remain to be an-
swered. We are, none the less, encouraged in that the application of modern
instrumentation and techniques has provided some insight to the complex
interactions of nutrient and heavy metal contamination in large aquatic
systems. It is clear that an understanding of cellular level processes is
essential to understanding system level processes and the development of
effective management strategies. In the particular case of the Saginaw Bay
pollution problem application of these techniques has elucidated a mechanism
184
which would be exceedingly difficult to discover by conventional limnologi-
cal methods.
REFERENCES
Crang, R.E. and T.E. Jensen. 1975. Incorporation of Titanium in polyphos-
phate bodies of Anacystis nidulans. J. Cell Biol. 67: 80a.
Danek, L.J. and J.H. Saylor. 1977. Measurements of the summer currents in
Saginaw Bay, Michigan. J. Great Lakes Res, 3: 65-71.
Ebel, J. P., J. Colas and S. Muller. 1958. Recherches cytochimiques sur les
polyphosphates inorganiques contenus dans les organismes vivants. II.
Mise au point de methods de detection cytochimiques specifiques des
polyphosphates. Exptl. Cell. Res. 15: 28-36.
Fitzgerald, G.P. and S.L. Faust. 1963. Factors affecting the algicidal and
algistatic properties of copper. Appl. Microbio. 11: 345-351.
Hohn, M.H. 1969. Qualitative and quantitative analyses of plankton diatoms
in the Bass Islands area. Lake Erie, 1938-1965, including synoptic sur-
veys of 1960-1963. Ohio Biol. Surv., N.S., Vol. 3. 211 p.
Jensen, T.E. and L.M. Sicko. 1974. Phosphate metabolism in blue-green al-
gae. I. Fine structure of the "polyphosphate overplus" phenomenon in
Plectonema boryanum. Can. J. Microbiol. 20: 1235-1239,
Lawry, N.H. and T.E. Jensen. 1979. Deposition of condensed phosphate as an
effect of varying sulfur deficiency in the Cyanobacterium Synechococcus
sp, (Anacystis nidulans). Arch. Microbiol. 120: 1-7.
Lin, C.K. and C.L. Schelske. 1978. Effects of nutrient enrichments, light
intensity and temperature on growth of phytoplankton from Lake Huron.
Univ. Michigan, Great Lakes Res. Div,, Spec. Rep. No. 63. 61 p.
Luft, J.H. 1961. Improvements in epoxy resin embedding methods. J.
Biophys. Biochem. Cytol. 9: 409-414.
Sicko, L.M. 1974. Physiological and cytological aspects of phosphate meta-
bolism in Plectonema boryanum. Ph.D. dissertation. The City Univ. of
New York, N.Y.
Sicko-Goad, L.M., R.E. Crang and T.E. Jensen. 1975. Phosphate metabolism
in blue-green algae. IV. ln_ situ analysis of polyphosphate bodies by
X-ray energy dispersive analysis. Cytobiologie. 11: 430-437.
Sicko-Goad, L., E.F. Stoermer and B.G. Ladewski . 1977. A morphometric
method for correcting phytoplankton cell volume estimates. Protoplasma.
93: 147-163.
185
Silverberg, B.A. 1975. UUrastructural localization of lead in Stigeo-
clonium tenue (Chlorophyceae, Ulotrichales) as demonstrated by cyto-
chemical and x-ray microanalysis. Phycologia. 14: 265-274.
Silverberg, B.A. 1976. Cadmium-induced ultrastructural changes in mito-
chondria of freshwater green algae. Phycologia. 15: 155-159.
Smith, V.E., K.W. Lee, J.C. Filkins, K.W. Hartwell, K.R. Rygwelski and J.M.
Townsend. 1977. Survey of chemical factors in Saginaw Bay (Lake
Huron). Ecol. Res. Series, U.S. Environmental Protection Agency,
Duluth, MN, Rep. No. EPA-600/3-77-125. 143 p.
Stempak, J.F. and R.T. Ward. 1964. An improved staining method for elec-
tron microscopy. J. Cell Biol. 22: 697-701.
Stoermer, E.F. and R.G. Kreis, Jr. In press. Phytoplankton composition and
abundance in southern Lake Huron. Univ. Michigan, Great Lakes Res.
Div., Spec. Rep. No. 65. 382 p.
186
SECTION 14
REVERSIBILITY OF INTOXICATION AND FACTORS GOVERNING IT
I.V. Pomozovskaya^
Criteria characterizing the poor state of the aquatic environment and
its inhabitants, their degradation and pathology when affected by various
kinds of pollutants have been developed intensively during recent years.
One of the industries with the largest water requirement is the pulp and
paper industry. Wastes coming from this type of enterprise are among the
most complicated and multi-factorial toxic complexes. In this connection,
the attention given to the study of the effects exerted by wastes from these
enterprises on bodies of water and aquatic organisms is quite natural.
Aquatic toxicological experimentation conducted in the zone of action
of such mills have provided valuable data on the real danger of waste
waters, the effects of their separate components, and their complexes upon
aquatic organisms of varying organisation and taxonomic ranking. These
studies have enabled a comparison of biological effects, related to the
functioning of various waste treatment plants, and have provided recommenda-
tions for their most economic and rational reconstruction and exploitation.
In this type of work carried out for a few years in Karelia, the main
criteria of toxicity chosen were the survival time of organisms, symptoms of
intoxication, changes in growth development and reproduction (fecundity,
quality of progeny, rate of maturation and spawning, etc), and alterations
in indices of the functional state; such as gas exchange, hematology, and
the degree and pattern of reversibility of intoxication.
The problem of reversibility of intoxication of organisms occupies a
special position in the whole complex of methodical approaches. Intoxica-
tion of fish and other organisms is highly probable, even in the presence
of a space limited point-sources pollution, since such sources may be on the
direct route of migration of the organism.
An inquiry into the problem of the possible reversibility of intoxica-
tion may assist in predicting results for organisms that undergo short
duration exposure in the polluted zone during crises, and in the case of
salvo discharges. This index should be considered when the remote conse-
quences of prolonged low-dose intoxication are in question, in assessing
^Karelian Branch of the Academy of Sciences of the USSR, Division of Water
Problems, Prospect Uritskogo, 68, KASSR, Petrozavodsk, USSR.
187
the degree of toxicity of one chemical reagent or another, and in deter-
mining the resistance of organisms to toxicants.
Reversibility of intoxication implies the recovery of organisms to their
normal physiological state after some pathological shifts brought about by a
toxic agent. The reversibility of pathological processes is possible only
at a definite concentration, and at a given duration of exposure to a toxic
substance. It may be said that pharmacological practice is based on this
phenomenon, since all pharmaceuticals employed are also toxins; but in a de-
finite combination they are of use for the organisms. Such combinations, at
which changes occurring under influence of poisons demonstrate reversi-
bility, should also be understood in the area of aquatic toxicology.
Data from literature on this problem are fairly scanty and, in some in-
stances, contradictory. Evidence of these facts can be found in the works
by Jones (1947, 1951, and 1957), Schweiger (1957), Wuhrmann and Woker
(1950), and Stroganov and Pozhitkov (1941), in which reversibility of in-
toxication in fish as affected by cyanides, sulphides, chloromercury, ethyl
alcohol, salts of heavy metals, and phenols, has been investigated.
The dynamics of phenol intoxication reversibility have been described
in a study by Lukyanenko and Fluorov (1963). Studies by Mann (1958),
Ludemann (1962), Chernysheva (1968) and others have been concerned with
reversibility of intoxication in fish as affected by insecticides. In
these reports, the possibility of restoring the vital activity of fish
which have been intoxicated with organophosphates is shown. Similarly, the
irreversible phenomena arising from contact with organochlorine compounds
is also demonstrated. A high degree of reversibility has been demonstrated
under the influence of detergents (Libmann 1960), but the resistance of
fish to various diseases decreases drastically.
This study has employed unpurified multi-component wastes from sulphate
pulp production as toxicants in various modifications and dilutions.
Further, sewage from sewage treatment plants has also been used. Waste
waters utilized contained methyl mercaptans, sulphides, hydrosulphides,
sulphates, acids and alkalis, methyl alcohol, furfurol, acetone, ammonia
and other organic and mineral compounds. The water in the natural effluent
receiver is similar in chemical composition to the average composition of
wastes resulting directly from production. It is nearly oxygen-free and has
a high carbon dioxide content (25.1 mg/0- Different quantities of sulphur-
containing compounds have been found in wastes from boiling and evaporating
shops. They possess a strong hydrogen sulphide smell. These wastes contain
alkali and some fairly toxic organic substances, including terpentine,
methanol, acetic and other acids. Wastes from the heat-and-power stations
are distinguished by a considerable amount of mechanical suspensions, the
result of burning slurry lignin, bark, and fuel oil, and by their sulphur
trioxide and sulphur dioxide content.
Atlantic salmon (Salmo salar), Cisco (Coregonus albula), roach (Rutilus
ruti lus), perch (Perca f luviatilis) and pike (Esox~lucius) were test
species. Fish of the first year of life (from the moment of hatching until
188
the transition to the fingerling stage) were used in contrast to fish of
older age groups.
The species of the fish, its age, average weight, and state (motor ac-
tivity, respiratory rhythm, pattern of food uptake, response to external
stimuli, etc.) were determined before the experiment.
Fish were pre-adapted to laboratory conditions and were placed for a de-
finite time in both concentrated and diluted waste water. When characteris-
tic signs of intoxication appeared, these organisms were transferred to pure
lake water where changes in their state, and the time and sequence of re-
storation of the functions lost were subsequently recorded.
The main sign of intoxication, which served as a signal for transferring
fish to pure water, was most often the loss of the equilibrium reflex, and a
transition to the inverted state. In some cases, the fish were subjected to
a sequence of two to four exposures in the waste waters. The degree and
dynamics of intoxication reversibility depended upon the temperature, the
concentration of toxicants, the duration of exposure, the test species, and
the age of the fish.
The maximum duration of the experiments was 30-35 days. Observations
have shown that the resistance of organisms to toxicants depends on all of
the factors noted above, but primarily upon the concentration of the agent,
its chemical structure, and duration of exposure.
Symptoms of intoxication of similar types can be traced in the behavior
of fish in test medium. The first phase of this phenomena involves in-
creased excitability (violent movements, sometimes whirling, with increased
respiratory activity). This phase is followed by a passive state (loss of
the equilibrium reflex, lateral or inverted position, respiration depressed,
refusal of food, loss of the shoaling effect, and changes in color. The
degree, time, and pattern of manifestation of intoxication symptoms are also
dependent on quite a number of factors. The most distinct, although brief,
symptoms of intoxication are observed in concentrated media. In some cases
these effects are obscure, especially in juveniles. In some phases they are
entirely absent.
In this paper, attention was focused mainly on juvenile fish, since they
inhabit the littoral part of a body of water which is most subject to con-
tamination. Furthermore, special investigations have indicated that wastes
issuing from sulphate pulp mills do not possess repellent properties for
fish. Numerous experiments have demonstrated that brief contact with con-
centrated or weakly diluted wastes results in an irreversible intoxication
of fish.
Thus, in 7-day-old larvae of Atlantic salmon (average weight 98 mg) kept
in both undiluted and diluted (1:1, 1:1) waste, vigorous excitation was in-
stantly recorded, coupled with serpentine movements and whirling activity.
After six minutes, the larvae descended to the bottom in lateral position,
failed to respond to stimuli, and their rate of respiration was diminished.
After the larvae were transferred to pure water, restoration of normal
189
breathing activity was observed after 15-20 minutes. They began to respond
to external stimuli, and by the end of the first day of detoxification, the
test larvae could not be distinguished from the controls by appearance
alone. During the first day no deaths were observed. By the 7th day the
larvae transferred from the undiluted wastes died. The dynamics of the sur-
vival rates for fish in pure water after intoxication are shown in Figure 1.
An approximately similar situation was observed when 37-day-old salmon
larvae (mixed feeding stage) were exposed. The characteristic symptoms of
intoxication were recorded after an exposure duration of four minutes. The
whole complex of symptoms (strong excitation, persistent loss of equili-
brium, and inverted position) was clearly seen in concentration wastes. In
pure water, the fish died within the first day after exposure.
In dilutions 1:1 and 1:2, test organisms were very excited. When trans-
ferred to pure water, they retained this increased motor and respiratory ac-
tivity for 30 minutes, subsequently sinking to the bottom of the tank and
reacting to stimuli with only weak movements of the caudal fin. Food was
refused and by the end of the third day of detoxification, the survival
rate was only 10% (Table 1).
TABLE 1. REVERSIBILITY OF INTOXICATION CAUSED BY EFFLUENTS
IN JUVENILE SALMON
(Age - 37 days. Mean weight - 144 mg. Temperature - 24°C,
Exposure 4 minutes)
Dilution
Condition
of fish
Survival
(%) in c'
lean water
of
toxicant
after exposure
1 Day
2 Days
3 Days
Control
Active
100
100
80
1:2
^ery active
20
20
10
1:1
^ery excited
20
10
10
Undiluted
Equilibrium
waste
reflex disturbed
0
-
-
The temperature factor significantly influences the rate of development
of the intoxication process and its results. A comparison of the data in
Table 1 and 2 shows that at 24°C, the death of the bulk of organisms ensues
within 72 hours. At an initial temperature of 13.5°C with an increase to
17.5°C , the first signs of intoxication appeared considerably later. Only
a repeated exposure to wastes (four exposures, 15 hours cumulatively) at in-
tervals with detoxification periods of 10-15 days (total 36 cumulative days)
lead to irreversible consequences for fish.
A short (6 minutes) exposure of roach (mean weight 16.7 g) to wastes
caused a persistent loss of the equilibrium reflex in fish. In diluted
wastes, this symptom appeared only in selected species.
190
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By the end of the first day of detoxication in pure water, the state of
the majority of fish did not differ from that of the controls. They ac-
tively swam, obviously reacted to external stimuli, and consumed food. The
fish exposed to the point of equilibrium loss during intoxication, restored
horizontal positioning in pure water only at intervals, ultimately sinking
to the bottom and dying on the second day. The survivors did not differ
from controls after 25 days of detoxification. They were again subjected
to the action of the toxicants. During a repeated 5-minute exposure with
concentrated sewage, the inverted position was observed. In diluted
sewage, unstable reactions were noted, but an equilibrium state was re-
corded. In pure water, the roach exposed to the concentrates died on the
third day, 20 percent of fish exposed to weak dilutions survived (Figure 2).
The dynamics of perch survival rate in pure water after 7 minutes expo-
sure is illustrated in Table 3. A situation similar to that described above
was observed when fish were exposed to concentrated and weakly diluted in-
dustrial wastes (boiling shop, evaporating and hydrolysis shops), and to the
waters of a natural waste water receiver, the isolated bay of a reservoir.
Experiments determining reversibility of intoxication in fish after a
brief exposure to effluents from a heat-and-power station were also re-
vealing, since they are considered to be relatively pure by industry. After
fish were exposed to effluents from a heat-and-power station diluted in
ratios of 1:5, 1:10, and 1:25 for 6, 10, and 24 minutes, respectively, only
a minor suppression of activity was observed. At the dilution 1:5 there was
a thin coating of coal observed on the fins. Mortality during the 10 day
period of detoxification was only 20 percent. However, additional exposure
of fish at the same dilutions of wastes for 7, 16, and 24 minutes led to the
death of the fish after 20 minutes in the first case, after a day in the se-
cond case, and only at a dilution of 1:25 did 40 percent of the experimental
fish survive (Table 4). These examples convincingly demonstrate the high
toxicity of treated wastes of sulphate pulp manufacturing.
The results of the experiment given in Table 5 are good evidence for the
dependence of the result on the duration of exposure.
The data show that only a four minute difference in exposure marked ef-
fects in the outcome of intoxication.
The dependence of the reversibility rate on concentration in roach lar-
vae is shown in Table 6.
As was demonstrated earlier, the main factors determining the resistance
and degree of restoration of activity, are the duration of exposure and the
concentration of the toxicant. This is also demonstrated in Table 7, which
shows that the purified wastes from treatment plants loose their toxic pro-
perties to a considerable degree, and although there are some symptoms of
intoxication, life activity is restored in pure water. Table 8 gives an in-
dication of the reaction of juvenile fish of various species to toxicants.
Thus, an extensive investigation into the pattern of intoxication from
effluents and its possible reversibility demonstrated that even brief expo-
193
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TABLE 4. REVERSIBILITY OF INTOXICATION IN JUVENILE SALMON CAUSED BY
EFFLUENT FROM A HEAT-AND-POWER STATION
(Mean weight - 125 mg)
First exposure
Second exposure
Dilution
Exposure
time
(min)
Condition
of fish
after
exposure
Survival
in clean
water, %
1-10 days
Exposure
time
(min)
Condition
1 of fish
after
exposure
Survival
in clean
water, %
1-10 days
Control
100
100
1:25
24
Insignifi-
cant de-
crease in
activity
100-80
24
Poorly mo-
bile,
"stand" on
the bottom
in vertical
position
80-40
1:10
10
Insignifi-
cant de-
crease in
activity
100-80
16
Poorly ac-
tive, thin
coating on
fins
Death
within
24 hrs
1:5
6
Increased
activity,
thin coat-
ing of car-
bon on fins
100-80
7
Equilibrium
reflexes
disturbed,
coating on
fins
Death
within
20 min
TABLE 5. REVERSIBILITY OF INTOXICATION IN JUVENILE SALMON CAUSED BY
EFFLUENT WATER (Dilution 1:5) FROM A HEAT-AND-POWER STATION
First exposure
Second exposure
Survival
Survival
Exposure
Condition of
in clean
Exposure
Condition of
in clean
time
fish after
water
time
fish after
water
(min)
exposure
1-10 days
(min)
exposure
1-10 days
2
Fins slightly
covered by
coating of
carbon
100-70
8
Carbon coating
on fins, fre-
quently "stand"
on the bottom
60-30
1
6
"Stand" on the
40-30
8
Lie on the bot-
Death
bottom, carbon
tom, carbon
within
coating on fins.
coating on
30 min-
convulsion of
fins
utes
the body
196
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sures to sulphate-cellulose discharges, with subsequent migration to pure
water does not guarantee fish safety. These factors are especially dan-
gerous when combined with high temperature regimes.
While the general symptoms of intoxication can be identified, there are
specific variations, depending upon the composition of the complex ef-
fluents, its concentration, the duration of exposure, and the species of
the test organism.
ACKNOWLEDGMENTS
My thanks are due to the research workers of the Petrozavodsk State
University named after O.W. Kuusinen; A.N. Ryzhkov, G.A. Tikka, and L.G.
Tyutyunnik for their help in performing the experiments.
REFERENCES
Jones, J. 1947. Journ. Exptl. Biol., 23, pp. 298-311.
Liebmann, H. 1960. Handbuch der Frischwasser und Abwasser Biologie. R.
Oldenbourg, Munchen, Bd. II, N 5-6.
Ludemann, D. and H. Neumann. 1962. Auz Schodlings Kunde, 35 (5-9).
Lukyanenko, V.I. and B.A. Flerov. 1963. The dynamics of reversibility of
phenol intoxication of crucian carp. Collected articles. "Materials
on hydrobiology and biology of the Volga reservoirs". Moscow-Leningrad,
Acad. Sc, USSR.
Lukyanenko, V.I. 1967. Toxicology of fish. "Food industry". Moscow, p.
47-52.
Mann, H. 1958. Fischwirt, 8 (217-220).
Schweiger, G. 1957. Arch. Fischereiwissenschaft, 8 (54-78).
Stroganov, N.S. and A.T. Pozhitkov. 1941. The action of industrial wastes
on aquatic organisms. Moscow State University, Moscow, USSR.
Wuhrmann, K. and H. Woker. 1950. Schweiz. Zeits, Hydrol., 12.
200
SECTION 15
ASPECTS OF THE INTERACTION BETWEEN BENTHOS AND SEDIMENTS IN THE
NORTH AMERICAN GREAT LAKES AND EFFECTS OF TOXICANT EXPOSURES
John A. Robbins^
The sediments of the North American Great Lakes are mostly overlain by
well -oxygenated waters and support a diverse and abundant population of ben-
thic (bottom-dwelling) organisms. Principal species include the freshwater
shrimp, Mysis relicta; the amphipod, Pontoporeia hoyi; many species of Oli-
gochaete worms such as Tubifex tubifex and Limnodrilus hoffmeisteri; the
midge larvae Chironomus anthracinus and a variety of freshwater clams such
as Sphaerium and Pisidium spp. Many of these organisms occur in great abun-
dance throughout the Great Lakes. The deposit feeding Oligochaete worms
occur in polluted harbors in numbers exceeding 1,000,000 m-2 (P. McCall,
pers. comm.), and even in the profundal sediments of Lake Erie in densities
approaching 50,000 m-2. Characteristically, Pontoporeia hoyi occurs in
densities on the order of 1,000 m-2 throughout much of the Great Lakes. In
Lake Erie, as well as in the inshore areas of the other Great Lakes,
Chironomid larvae densities are roughly 500 m-2 (P. McCall and D. White,
pers. comm.). These organism densities represent an enormous biomass
dwelling in or interacting with the sediments.
Not only are certain benthos an important link in the food chain, but
many of them significantly affect the stratigraphy of sediments (Robbins et
al . 1977) and the exchange of nutrients between sediments and water through
such activities as burrowing, feeding, respiration, and excretion. As the
fine-grained sediments are both the ultimate sink and a partial source (cf
Remmert et^ aj_. 1977) of nutrients in the Great Lakes, the life activities of
the benthos are likely to be an important factor in the nutrient cycle. If,
in turn, the behavior, physiology, or mortality of benthos are affected by
aquatic pollutants, there can be potentially novel and important effects on
major nutrient cycles. While there has been considerable work done on the
role of benthos in sediment mixing and exchange of substances across the
mud-water interface in other lakes (see Petr, 1976 for a review), \/ery
little has been done in the Great Lakes. The aim of this paper is to illu-
strate the effects of selected benthos on particle and solute transport and
^Great Lakes Research Division, University of Michigan, Ann Arbor.
Michigan 48109.
201
to indicate some preliminary results of exposing benthos-sediment microcosms
to toxic substances.
STRATIGRAPHIC EFFECTS OF NATURAL POPULATIONS
In early attempts to interpret radioactivity profiles in sediments of
the Great Lakes (Robbins and Edgington 1975) it became clear that signifi-
cant mixing of material occurred over the upper 10 cm of sediment. From
later work (Robbins et^ aj_. 1977) it was evident that the sediment mixing was
due to the presence of benthic organisms. At two locations in Lake Huron,
twelve cores of fine-grained sediment were taken for comparison of the
vertical distributions of the naturally occurring radionuclide, lead-210,
and fallout cesium-137 with the distributions of benthic macroinvertebrates.
In the absence of mixing, the activity of lead-210 should decrease exponen-
tially with sediment depth reflecting radioactive decay (T]/2 " 22.26 yr) on
burial. In actuality, the lead-210 activity was constant down to 6 cm in
cores at one location and 95% of the total invertebrates occurred within the
zone of constant activity. At the other location, the zone of constant
activity was only 3 cm deep but more than 90% of the benthos were confined
to it. In each case comparison of published tubificid reworking rates with
sediment accumulation rates showed that the activities of benthos were able
to account for the mixing of sediments. An example of the effect of sedi-
ment mixing on cesium-137 profiles is given in Figure 1 for a core from Lake
Erie where the sedimentation rate is exceptionally high. The observed al-
teration in the radioactivity profile over that expected in the absence of
steady-state mixing is consistent with the measured vertical distribution of
benthos which at this location consists primarily of mature and immature
Oligochaete worms. Studies of the distribution of natural and fallout
radionuclides in cores from Lake Erie (Edgington and Robbins 1979), Lake
Huron (Johansen and Robbins 1977) and Lake Michigan (Edgington and Robbins
1975) show that the mixing of surface sediments occurs widely in the Great
Lakes.
It may thus be expected that altered patterns of sediment mixing result-
ing from exposure of benthos to aquatic pollutants could result in altered
and possibly uninterpretable radioactivity and heavy metal profiles. From
our studies (Robbins 1977) it is apparent that the time resolution with
which lake-wide pollution changes can be reconstructed from sedimentary re-
cords is limited by benthic reworking (bioturbation) . Increased benthos
mortality would be likely to improve the long-term resolution because of the
associated reduction in sediment mixing.
LABORATORY STUDIES USING RADIOTRACERS
To investigate the role of benthos in the transport of sediment parti-
cles in a controlled ans systematic way, experiments were set up in the lab-
oratory using a particle-bound radiotracer, cesium-137. Illite clay parti-
cles with adsorbed cesium-137 were added as a submi llimeter layer to the
surface of fine-grained sediments contained in plastic cells of a rectangu-
lar cross section stored in a temperature-regulated aquarium. A well-col li-
202
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Figure 1. Distribution of benthos and cesium-137 in a core from Lake Erie.
In the absence of sediment mixing the expected distribution of cesium-317
should reflect the history of atmospheric fallout (dashed line). The
measured distribution is shown as the solid line histogram. The
theoretical distribution shown as the continuous curve is based on the
assumption of steady state mixing to a depth of 8.7 cm. The measured
distribution of Oligochaete worms is consistent with this mixing depth.
The results illustate the stratigraphic effects of benthos which are
observed widely in the Great Lakes.
203
mated sodium iodide crystal gaima detector scanned the length of each of
several cells at daily or weekly intervals over a period of several months
in order to determine how several benthic species transported labeled parti-
cles away from the sediment surface. The experimental set up with an ex-
posed aquarium containing several cells, a detector and a counting system
are shown in Figure 2. Details of the construction and operation of the
system are found in Robbins et al_. (1979). The actual and measured distri-
bution of activity from a submi llimeter line source is shown in Figure 3.
The nearly Gaussian profile of measured activity mainly reflects collimator
geometry. The limited broadening of the line source in the control cell
(with no benthos present) is due to molecular diffusion.
When Oligochaete worms are added to surface-labeled sediments, the
radioactivity profile evolves over a six-month period as illustrated in Fi-
gure 4. The shaded areas represent the profile corrected for the effects
of finite detector resolution. The initial effect of the worms on the dis-
tribution is one of burial. This is, of course, consistent with the well-
known behavior of these organisms. They penetrate sediments to about 10 cm
depth to feed while at the same time holding their tails above the sediment
surface to defacate. This behavior has led Rhoads (1974) to describe such
organisms as "conveyor-belt" species. In time, the marked layer is buried
to the point where it encounters the zone of feeding and begins to reappear
at the sediment surface. During the initial burial period, the reworking
rate is essentially constant as can be seen in Figure 5 which shows the lo-
cation of the peak activity versus time. The burial rate is about 0.052 +
0.007 cm/day at 20 degree C. Error bars primarily reflect uncertainty in
locating the sediment-water interface due to irregular pile up of fecal
mounds.
The interaction of the amphipod, Pontoporeia hoyi , with sediments
strongly contrasts with that of Oligochaete worms. As can be seen in Fi-
gure 6, the activity spreads downward from the surface under the action of
Pontoporeia without significant advection. This species burrows randomly
through the upper several centimeters of sediment and thus serves to move
sediment particles in a manner akin to eddy diffusion. Shown in Figure 7
are the corrected peak width versus time plus a theoretical relationship
based on the assumption that particle motion is truly eddy diffusional in
character. Details of the calculation are given in Robbins et aj_. 1979.
The diffusion coefficient implied by the data is 4.4 cm^/yr for an amphipod
density of 16,000 cm-2.
While the two benthic species investigated have Mery different modes of
interaction with sediments, their effect on vertical particle movement can
in each case be quantitatively described and measured with a precision and
rapidity which suggests the radiotracer method as a useful behavioral bio-
assay technique. Mery precise reworking rates, expressed either in terms of
a sediment burial rate or eddy diffusion coefficient, can be determined
under realistic conditions in a matter of a few days. This radiotracer
method of observing a particular organism's behavior offers the special ad-
vantage of being noninteracti ve to a \jery high degree. The gamma radiation
passes readily through the cell walls and, once radionuclides have been
added to the system, no further interaction with the microcosm is required
204
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207
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to make quantitative observations of processes occurring within it. The no-
tion of using radiotracers to measure aspects of behavior in a noninterac-
tive way can be extended to include species other than benthos, other be-
haviors, and other aquatic microcosms.
So far, we have not applied the method for carefully controlled assay of
toxic substances but only for several trivial cases where it was important
to demonstrate that the addition of major ions to water overlying cells
would not affect reworking rates. Still, general features of the experi-
ments are useful to consider. Either sulfate (SO4) or chloride (CI) ions
were added as Na2S04 or NaCl to a series of cells containing Oligochaete
worms and a surface-labeled layer of sediment. Prior to addition of the
ions, the reworking rate in each cell had been measured with the scanning
system. Results are shown in Figure 8 for addition of NaCl. Below 5000
micrograms Cl/ml (ppm) no change in reworking rate was observed while the
rate decreased abruptly following addition of NaCl at a concentration in
overlying water of 10,000 ppm. In this case, the reduction was probably not
a behavioral but rather a mortality effect. The results for sulfate are
given in Figure 9 for two species of Oligochaete worms, Tubifex tubifex de-
rived from Lake Michigan sediments and laboratory culture of Limnodrilus
hoffmeisteri . The ratio of final to initial reworking rate is shown versus
concentration. Again, significant decreases in reworking rates occur only
for the very high concentrations used in the experiment. These levels of
course far exceed any encountered in most lakes. There appear to be signi-
ficant differences in the response of the two Oligochaete populations to SO4
additions. More important experiments will involve additions of metals and
toxic organics to these microcosms. Such work would represent a continua-
tion of studies by others, notably Brkovic-Popovic and Popovic who have in-
vestigated the effect of heavy metals on survival (1977a) and on respiration
rate (1977b) of tubificid worms. Problems will arise in the interpretation
of the effects of nonconservative substances on the system, which were far
less significant in the case of conservative ions like sulfate and chloride.
Nonconservative materials may rapidly adsorb to sediment particles and
little meaning may be attached to the concentration in overlying water. As
sophistication develops in the use of such radiotracer behavioral assay
methods, it will be desirable to take the community approach as there is
considerably evidence for species interaction effects (Petr 1976). In re-
lated studies, it would be desirable to look at the relation between toxic
substance exposure and the ability of benthos to avoid predation (Hall et
al . 1979). A further effect which can be studied with relative ease with
this method is the response of benthos to chronic oxygen depletion. Under
conditions of oxygen depletion, feeding of tubificid worms is minimal. In
sediments of Lake Easthwaite, worms spend most time at the mud-water inter-
face (Stockner and Lund 1970) but resume feeding with the restoration of
aerobic conditions.
The scanning method described above can also be used to investigate the
effects of benthos on interstital transport. By using both a particle-bound
radiotracer such as cesium-137 (Kj '^ 5000, Robbins et al. 1977) and a rela-
tively conservative gamma emitting isotope, sodium-22 Xl<(j ^ 2, Lerman and
Weiler 1970), reworking rates and molecular diffusion rates can be deter-
mined simultaneously. In a prototype experiment, we have investigated the
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effect of Oligochaete worms on soulte transport. In each of two cells con-
taining natural sediment and lake water, we added a submillimeter layer of
cesium-labeled sediment and about 20 microCuries ».f sodium-22 as NaCl. To
one cell we then added worms to achieve a density of about 70,000 m-2. Fol-
lowing this treatment the two cells, control and worm, were scanned about
once a day for over 10 days. The results of the experiment are illustrated
in Figure 10. Profiles of cesium-137 and sodium-22 are shown after an
elapsed time of about 200 hours. In the control cell, there is no signifi-
cant displacement of the marked layer while in the worm cell, the layer has
moved downward by an amount corresponding to a rate of about 0.055 cm/day.
In the worm cell, the Na-22 has penetrated further into the sediments than
in the control. Note that the measurements in the worm cell were made about
40 hours earlier than in the control. Thus the downward movement of the
sodium-22 would be even more pronounced if the profiles could have been
taken at the same time. The solid curve is the expected distribution of
sodium-22 based on a solution to the diffusion equation with values of the
diffusion coefficient chosen to give the best least squares fit to the data.
In the control cell, the effective diffusion coefficient is 3.9 x 10"^
cm^/sec while in the worm cell, the value is 13.1 x 10"^ cm^/sec. Thus, the
presence of tubificid worms at a density of about 70,000 m"^ enhances the
diffusion coefficient by over a factor of 3. In a separate experiment where
the sediments had been conditioned by allowing worms to create an equili-
brium system of burrows, but where there was no active reworking at the time
of adding radiotracers, the diffusion coefficient for Na-22 transport was
still enhanced (x2) over its value in a control cell having no conditioned
sediments. Therefore, it seems that the enhancement of pore water diffusion
by tubificid worms results from their loosening of the sediments through the
creation of a system of burrow channels rather than to their momentary life
activities. Thus, the short-term effect of reducing or terminating the bur-
rowing activity of worms through exposure to aquatic pollutants would seem
to be small but the long-term result would appear to be the collapse of the
burrow structure with an associated reduction in the ability of ions to
migrate through pore fluids.
With proper experimental design, the radiotracer method could be used to
examine the effect of aquatic pollutants on benthos-mediated transport of
solutes. However, a more direct approach is to relate measured sediment-
water fluxes to the density of activities of benthos.
NUTRIENT FLUXES FROM UNDISTURBED SEDIMENT CORES
We have taken this approach in collecting a series of cores from various
locations in the Great Lakes (Remmert et aj^. 1977; Robbins et^ al_. 1976).
Undisturbed 7.5 cm diameter cores of fine-grained sediments from Lakes
Michigan, Huron, and Erie were stored in in situ temperatures ('v^50° C) in
their original plastic liners along with a^out 10 cm of overlying water.
Increases in the concentration of reactive dissolved silica over periods of
hours to days in stirred, oxygenated overlying water provided estimates of
the rate of exchange of dissolved silicon across the sediment water inter-
face. The increases in the concentration of silicon (ppm Si) versus time
is shown for a core from Saginaw Bay, Lake Huron in Figure 11. The release
214
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EPA-SB-78
CRUISE 7
CORE 31' -3
1.0-
^ 0.5
0
0
40
80
TIME (hrs.)
120
160
Figure 11. Concentration of soluble reactive silicon in water overlying
sediments stored without disturbance in a core liner collected from
Saninaw Bay, Lake Huron. Over about the first hundred hours, the
release of Si into overlying v/ater is essentially constant.
216
rates for each of the lakes was similar despite different seasons of coring
each lake and averaged about 2000 (i.e., 2000 yg Si/cm2/yr) micrograms
Si/cm2/yr. If this flux represents an annual average then the amount of Si
regenerated from sediments each year in Lake Erie for
The vertically integrated amount of dissolved silicon
is a maximum of about 200 micrograms/cm^, so the time
the Si removed from the water (through incorporation
example is enormous,
in the water column
required to replenish
into diatoms) by re-
generation from sediments is 0.1 year. Robbins and Edgington (1979) found
that the flux of Si from sediments in Lake Erie is proportional to the con-
centration of amorphous silicon in surface sediments suggesting the flux is
dominated by dissolution of particulate cilica recently deposited on the
sediment surface. This result indicates a particular role for organisms
like the larvae of Chironomids which are shallow water plankton detritus
feeders and whose effect on the release of silicon from sediments was noted
many year ago by Tessenow (1964).
By comparing silicon fluxes with benthos densities in a series of repli-
cate cores taken on two cruises in Saginaw Bay, Lake Huron, last year (fall
1978), we have been able to confirm Tessenow's observations for our particu-
lar Great Lakes environment. Shown in Table 1 is the density of benthos in
TABLE 1. BENTHOS DENSITY AND SILICON FLUX: SAGINAW BAY, LAKE HURON
Dens'
ity (m-^)
Tubi
f icids
Silicon Flux
Cruise
Core
Mature
Immature
Naididae
Chironomids
(yg/cm2/yr)
7I
1
850
40,000
5900
0
1100
3
850
65,000
7900
0
770
T
280
8,200
850
560
1800
2'
1700
29,000
8200
850
2700
3'
2300
18,000
560
560
1680
8^
1
1400
23,000
0
1130
3600
2
280
5,600
0
0
1300
8
280
5,600
0
280
1600
2'
280
29,000
13000
560
2400
"•October, 1978.
^November, 1978,
each of several replicate core
timed sampling of overlying wa
the dominant species in terms
However, densities of these or
as can be seen from Table 2.
most correlations are not sign
Si flux and the density of Chi
for both observation periods (
ents were measured as well and
both cruises are underlined
s along with the silicon flux measured via
ter as described above. It can be seen that
of numbers are the immature tubificid worms,
ganisms correlate poorly with the silicon flux
Because of the limited number of observations
ificant. However, the correlation between the
ronomids is outstandingly high and significant
Figure 12). In this experiment, other nutri-
correlations which are persistently high over
The observed decrease in the concentration of
217
TABLE 2. CORRELATIONS BETWEEN NUTRIENT FLUXES AND ORGANISM DENSITIES
(CRUISE 7)
Organism
Group
Phosphate
(PO4)
Ammonia Nitrate
(NH3) (NO3)
Sulfate
{SO4)
Silicon
(Si)
Tub. Mature
0.93
-0.09 -0.17
-0.75
-0.07
Immature
0.07
-0.82 0.6*^
-0.54
-0.36
Naididae
-0.19
-0.49 0.13
-0.41
-0.29
Chironomids
0.11
0.74 0.04
0.04
0.97
Total
0.06
-0.69 0.55
-0.56
-0.26
(CRUISE 8)
Organism
Group
Phosphate
(PO4)
Ammonia Nitrate
(NH3) (NO3)
Sulfate
{SO4)
Silicon
(Si)
Tub. Mature
0.41
0.92 0.23
0.97
0.88
Immature
0.23
0.14 0.93
0.22
0.49
Naididae
-0.30
0.63 0.78
-0.50
0.13
Chironomids
0.09
0.63 0.25
0.76
0.99
Total
-0.9
-0.05 0.94
0.01
0.62
218
STATION 31/31'
SAGINAW BAY
4000
CRUISE 7: OCT 1978
^^ 2000
•^^^^"^
6
► F= 873+ 1.86 Nc
_l
1.1,1,1
o
o
CRUISE 8 : NOV. 1978
_l
^ 4000
^^^^^^
^^^^^m
2000
i^-^"^^^^
n
F= 1130+ 2.15 Nc
1
0
400 800 1200
CHIRONOMID DENSITY (m^)
1600
Figure 12. Relationship between the flux of Si from sediments and the
density of Chironomid larvae in a series of replicate cores taken from
Saginaw Bay, Lake Huron, on two separate cruises in 1978.
219
phosphate in overlying water is marginally associated with the presence of
mature tubificid worms, the increase in ammonia is persistently associated
with Chironomids, and the reduction in nitrate levels over time appears to
be associated with the population of immature tubificids and/or the total
macrobenthos population.
The results for silicon suggest the relationship:
Flux = 1000 + 2 X Chironomid larvae density,
where the flux is in micrograms Si/cm^/yr and the density is in numbers m"^.
As the mean density of Chironomid larvae at this location is about 500 m"^,
roughly half the flux of silicon from the sediments is attributable to the
presence of these organisms. This circumstantial evidence for the effect of
Chironomids is strengthened by considering Tessenow's experiments with sedi-
ments from Lake Heiden, Germany (Tessenow 1964) in which he demonstrated a
casual relationship. Addition of Chironomids (Pulmosus group) to his sedi-
ments resulted in enhanced silicon release. Converting Tessenow's results
to the above form, we find that for his experiments:
Flux = 1000 + 4 X Chironomid larvae density.
Graneli (1977) has also observed that Chironomus Pulmosus larvae increase
the release of silica as well as phosphorus from sediments of several lakes
in Sweden. It would therefore seem likely that at least in shallow waters
of the Great Lakes where fine-grained sediments can be found, such as lower
Saginaw Bay, and in most of Lake Erie, Chironomid larvae may play a major
role in the regeneration of silicon from sediments. In Lake Erie, average
Chironomid densities may be as high as 1000 m-2 (p. McCall, pers. comm.).
That these organisms may enhance silicon fluxes does not necessarily mean
that their removal or inhibition through exposure to aquatic pollutants will
result in a long-term reduction in the capacity of the sediments to return
silicon to overlying waters. It is always possible that the ecological
niche represented by diatom detritus processing can be filled by another
biotic or abiotic component. In other words, the role of Chironomid larvae
may be mainly a kinetic one.
Several preliminary experiments have been undertaken to determine the
effect of removing the influence of macrobenthos on release of silicon. A
method must be chosen which results in minimal alteration of the structure
or composition of sediments. In one experiment, a core incubated at in situ
temperatures was exposed to 5 megaRads of cobalt-60 gamma radiation, enough
exposure to completely sterilize the sediment core and overlying water. The
results of this experiment are shown in Figure 13. Prior to irradiation,
the silicon flux was 2000 micrograms Si/cmVyr. After irradiation, the flux
dropped to 900 micrograms Si/cm^/yr. It is interesting to note that the
factor of two reduction in flux is consistent with the relation given above
for the flux as a function of Chironomid larvae density. In this particular
core, the density of benthos was not measured. A major reduction in the
silicon flux also resulted from addition of Chlordane in amount sufficient
to destroy the macrobenthos population (about 1 ml of Chlordane in a disper-
sant). Results of this and other treatments are given in Table 3. No
220
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221
TABLE 3. EFFECTS OF SELECTED TREATMENTS OF SILICA RELEASE FROM SEDIMENTS
Treatment
Release
Ratel
Core
Before
Treatment
After Treatment
NLH 2-11
Rotenone
.126
+
.021
.091 +
.020
NLH 2-14
Chlordane
.116
+
.016
.036 +
.010
NLH 4-4
Control
.136
+
.012
.086 +
.009
NLH 44-1
Gatimia Radiation
.234
+
.010
.116 +
.013
LM 5-1
Tubif icids
.241
+
.022
.177 +
.012
LM 5-3
Pontoporeia
.217
+
.015
.202 +
.016
LM 5-2
Control
.166
+
.017
.120 +
.014
NLH 44-3
Sediment Sti
at Coring
rred
.0132 +
.016
—
1
vg/cm /hr,
222
significant reduction in flux occurred following addition of rotenone (about
1 ml of saturated solution in ethyl alcohol). Note that reduction of the
flux in the control cell reflects the progressive approach toward an equili-
brium concentration of silicon in overlying water. In another set of ex-
periments, tubificid worms and Pontoporeia were added to cores so as to in-
crease the natural population densities by about a factor of two. As can be
seen in Table 3, these additions did not result in a significant increase in
the silicon flux. In retrospect, it appears likely that the addition of
Chironomid larvae would have produced the increase in the flux.
Our results suggests an important role for benthos in the cycling of
silica (and possibly other nutrients) in the Great Lakes. As silica is a
major and probably limiting nutrient for the diatom productivity, it is im-
portant to understand the role of benthos and Chironomid larvae in particu-
lar in nutrient regeneration and the possible effect of aquatic pollutants
on their interaction with sediments.
ACKNOWLEDGEMENTS
The author wishes to acknowledge the help of Cheryl Hoyt, Karen Husby,
Kjell Johansen, John Krezoski, and Maranda Willoughby in various aspects of
field and laboratory work. Great Lakes Research Division Contribution Num-
ber 254.
REFERENCES
Brkovic-Popovic, I., and M. Popovic. 1977a. Effects of heavy metals on
survival and respiration rate of tubificid worms: Part I - Effects on
survival. Environ. Pollut. 13: 65-72.
Brkovic-Popovic, I., and M. Popovic. 1977b. Effects of heavy metals on
survival and respiration rate of tubificid worms: Part II - Effects on
respiration rate. Environ. Pollut. 13: 93-98.
Edgington, D.N., and J. A. Robbins. 1975. The behavioral of plutonium and
other long-lived radionuclides in Lake Michigan: II. Patterns of depo-
sition in the sediments. IAEA Symposium on the Environmental Effects of
Nuclear Power Generation, (IAEA-SM/198/40), l^elsinki, Finland, June,
1975.
Edgington, D.N., and J. A. Robbins. 1979. History of plutonium deposition
in Lake Erie sediments. Twenty Second Annual Conference on Great Lakes
Research of the International Association for Great Lakes Research,
Rochester, New York, April 30 - May 3, 1979. Abstacts p. 49.
Graneli, W. 1977. Sediment respiration and mineralization in temperate
lakes. Ph.D. Dissertation, Institute of Limnology, University of Lund,
Sweden, Summary 9 pp.
223
Hall, R.J., A.M. Forbes, J.J. Magnuson, P. A. Helmke, and J. P. Keillor.
1979. Effects of mercury and zinc on the behavior of Pontoporeia hoyi
(Amphipoda). (Submitted to J. Great Lakes Res., 1979).
Johansen, K.A., and J. A. Robbins. 1977. Fallout cesium-137 in sediments of
Lake Huron. Twentieth Annual Conference on Great Lakes Research of the
International Association for Great Lakes Research, Ann Arbor, Michigan,
May 10-12, 1977.
Lerman, A., and R.R. Weiler. 1970. Diffusion and accumulation of chloride
and sodium in Lake Ontario sediment. Earth Planet. Sci. Lett. 10:
150-156.
Petr, T. 1976. Bioturbation and exchange of chemicals in the mud-water
interface, Ijn Interactions between sediments and freshwater (Ed. H.L.
Golterman). Proceedings of the International Symposium held at
Amsterdam, The Netherlands, September 6-10, 1976.
Remmert, K.M., J. A. Robbins, and D.N. Edgington. 1977. Release of dis-
solved silica from sediments of the Great Lakes. Twentieth Annual Con-
ference on Great Lakes Research of the International Association for
Great Lakes Research, Ann Arbor, Michigan, May 10-12, 1977.
Rhoads, D.C. 1974. Organism-sediment relations on the muddy sea floor.
Oceanogr. Mar. Biol. Ann. Rev. 12: 263-300.
Robbins, J. A. 1977. Recent sedimentation rates in southern Lake Huron.
Fortieth Annual Meeting of the American Society for Limnology and Ocean-
ography, East Lansing, Michigan, June 20-23, 1977.
Robbins, J. A., and D.N. Edgington. 1975. Deter-^i nation of recent sedimen-
tation rates in the Great Lakes using lead-210 and cesium-137. Geochim.
Cosmochim. Acta 39: 285-304.
Robbins, J. A., and D.N. Edgington. 1979. Release of dissolved silica from
sediments of Lake Erie. Twenty Second Conference on Great Lakes Re-
search of the International Association for Great Lakes Research,
Rochester, New York, April 30 - May 3, 1979. Abstracts, p. 9.
Robbins, J. A., J.R. Krezoski, and S.C. Mozley. 1977. Radioactivity in
sediments of the Great Lakes: postdepositional redistribution by de-
posit feeding organisms. Earth Planet. Sci. Lett. 36: 325-333.
Robbins, J. A., K. Remmert, and D.N. Edgington. 1976. Regeneration of sili-
con from sediments of the Great Lakes. Radiological and Environmental
Research Division Annual Report, Ecology, Argonne National Laboratory,
Argonne, Illinois, Jan. through Dec, 1976, pp. 82-86.
Robbins, J. A., P.L. McCall, J.B. Fisher, and J.R. Krezoski. 1979. Effect
of deposit feeders on migration of cesium-137 in lake sediments. Earth
Planet. Sci. Lett. 42: 277-287.
224
Stockner, J.G., and J. Lund. 1970. Live algae in postglacial lake depo-
sits. Limnol. Oceanogr. 15: 41-58.
Tessenow, U. 1964. Experimental investigations concerning the recovery of
silica from lake mud by Chironomid larvae (Pulmosus group). Archiv f.
Hydrobiol. 60: 497-504.
225
SECTION 16
RECENT ADVANCES IN THE STUDY OF NITRITE TOXICITY TO FISHES
Rosemarie C. Russo^
Nitrite has not until recently received much attention as a toxicant to
aquatic organisms. However, it has been established that nitrite is very
toxic to fishes and aquatic invertebrates. Furthermore, nitrite has been
implicated in the formation of N^-nitroso compounds (Archer et a^. 1971;
Wolff and Wasserman 1972; Mirvish 1975), and nitrosamines have been shown
to be carcinogenic to zebra fish (Brachydanio rerio), rainbow trout (Salmo
gairdneri), and guppy (Lebistes reticulatus) (Stanton 1965; Ashley and
Halver 1968; Sato et al . 1973). Recently nitrite has been reported to in-
duce cancer in rats directly, rather than through formation of nitrosamines
(Newberne 1979).
In the past few years much research has been done to investigate the
toxicity of nitrite to aquatic organisms. This includes the study of
nitrite toxicity to additional fish species, the effects of water chemistry
conditions on nitrite toxicity, and some work on the mode of toxic action
of nitrite.
Nitrite is produced as an intermediate product in the nitrification pro-
cess. In this process, the biological oxidation of ammonia to nitrate,
Nitrosomonas bacteria convert ammonia to nitrite, and Nitrobacter converts
nitrite to nitrate. The effectiveness of the conversion process is affected
by several factors, including pH, temperature, dissolved oxygen concentra-
tion, numbers of nitrifying bacteria, and presence of inhibiting compounds.
Under normal circumstances the first conversion, ammonia to nitrite, is the
rate-limiting step in the process; the second conversion, nitrite to ni-
trate, is relatively rapid. For this reason, nitrite is generally present
in only trace amounts in most natural freshwater systems. In sewerage
treatment plants utilizing the nitrification process, the process may be im-
peded, causing discharge of nitrite at elevated concentrations into the re-
ceiving water. Also, water reuse systems using the nitrification process
may malfunction, resulting in increased nitrite levels in the treated water.
^Fisheries Bioassay Laboratory, Montana State University, Bozeman,
Montana 59717.
226
It has been demonstrated (Anthonisen et aj_, 1976) that the nitrification
process can be inhibited in the presence of nitrous acid {HNO2) and un-
ionized ammonia (NH3). The total ammonia in a wastewater treatment system
is present as ammonium ion (NH4"*') and un-ionized ammonia {NH3). If the pH
of the solution increases, either naturally or by addition of a base, the
concentration of un-ionized ammonia will increase. Un-ionized ammonia in-
hibits nitrobacters at concentrations (0.1-1.0 mg/1 NH3) appreciably lower
than those (10-150 mg/1) at which it inhibits nitrosomonads. This impedes
the conversion of nitrite to nitrate, causing nitrite to accumulate. When
the pH decreases, as ammonium and nitrite are oxidized, an increase in ni-
trous acid (HNO2) concentration occurs. Nitrous acid inhibits both nitro-
bacters and nitrosomonads at concentrations between 0.22 and 2.8 mg/liter.
This inhibition of the process can also result in an increase in nitrite.
Several organic compounds likely to be found in significant concentra-
tions in industrial wastes have been shown to inhibit the nitrification pro-
cess (Hockenbury and Brady 1977). Dodecyl amine, aniline, and r[-methyl ani-
line at concentrations less than 1 mg/liter caused 50 percent inhibition of
ammonia oxidation by Nitrosomonas; £-nitrobenzaldehyde, £-nitroaniline, and
r[-methylaniline at concentrations of 100 mg/liter inhibited nitrite oxida-
tion by Nitrobacter.
The loss of nitrification flora, especially resulting from the use of
antibiotics, has also been indicted (Patrick et al_. 1979) as a potential
cause of large amounts of nitrite accumulating in natural waters.
In view of these considerations, nitrite may be present under some cir-
cumstances in natural waters at concentrations high enough to be deleterious
to freshwater aquatic life. Some field data have been reported documenting
this. Klingler (1957) has reported nitrite concentrations of 30 mg/liter
nitrite-nitrogen (NO2-N) and higher in waters receiving effluents from
metal, dye, and celluloid industries. McCoy (1972) has reported concentra-
tions up to 73 mg/liter NO2-N in Wisconsin lakes and streams. We have ob-
served levels of 0,1 mg/liter NO2-N in a reasonably clean cold water trout
stream in Montana (Russo and Thurston 1974).
The literature through 1977 on nitrite toxicity to fishes has been sum-
marized elsewhere (Russo and Thurston 1977, 1978; U.S. EPA 1977). Most of
the data available do not include 96-hour LC50 values, but some comparisons
can be made. From this and more recent literature there appear to be some
differences, at least on a short term (less than four days) basis, in the
relative susceptibilities to nitrite of different fish species. Concentra-
tions as low as 0.2 mg/liter NO2-N are acutely lethal to several species,
with trout and salmon being the most susceptible. Concentrations in the
range of 2 to 15 mg/liter NO2-N have been reported to be lethal to some
warmwater species, such as fathead minnows (Pimephales promelas) and channel
catfish (Ictalurus punctatus). Some fish species, such as creek chub
(Semotilus a. atromaculatus) and carp (Cyprinus carpio), succumb only at
higher concentrations, up to 100 mg/liter NO2-N. Of the fish species
studied, those most tolerant to nitrite were: common white sucker
(Catostomus commersoni), quillback (Carpiodes cyprinus), and mottled sculpin
(Cottus bairdi). These species incurred no mortalities during short expo-
227
sures to NO2-N concentrations of 67 to 100 mg/liter. Manifestations of the
acutely toxic effects of nitrite can thus vary widely, depending on fish
species.
Little information has been reported on the effects of nitrite exposure
for periods of time longer than 1-4 days. We have conducted 36-day expo-
sures on cutthroat trout (S. darki) fry (Thurston e^ aj_. 1978) and found
LC50 values at 36 days to be only slightly lower than 96-hour values.
Wedemeyer and Yasutake (1978) exposed steelhead trout (S. gairdneri ) to low
NO2-N concentrations (0.015-0.060 mg/liter) over a 6-month period and found
no serious deleterious effects. Growth and ability of the fish to adapt to
seawater were not impaired. Varying degrees of gill hyperplasia and lamel-
lar separation were observed early in the test but the fish seemed to re-
cover and after 28 weeks these abnormalities were no longer observed.
Fish size has also been thought to be a factor influencing fishes' sus-
ceptibility to nitrite. Rainbow trout sac fry, and 2-g fry, were found to
be less susceptible than were larger (12-, 14-, and 235-g) rainbow trout
(Russo et ^. 1974); 4.5-g fingerling rainbow trout were reported to be more
tolerant than were 100-g yearlings (Smith and Williams 1974). Coho salmon
(Oncorhynchus kisutch) fry (0.65 g) were less susceptible than were
yearlings (22 g) (Perrone and Meade 1977). We have now conducted 20 96-hour
nitrite bioassays on rainbow trout over the size range 2 to 387 g. These
experiments were all conducted under similar water chemistry conditions
(Table 1). The results are given in Table 2; over this larger range of fish
size than that reported previously, there does not appear to be any rela-
tionship between fish size and susceptibility to nitrite. This is illu-
strated in the graphs of LC50 vs. fish weight and length, shown in Figures
1 and 2.
We have also studied the effect of chloride ion (CI") on nitrite toxi-
city to rainbow trout (Russo and Thurston 1977). We conducted a series of
nitrite toxicity tests in which we added CI" (as NaCl) in concentrations
ranging from 1 to 41 mg/liter. A significant reduction in nitrite toxicity
resulted from increased levels of CI" (Figure 3), and this effect was
linearly correlated (Figure 4). The 96-hour LC50 was raised from 0.46
mg/liter NO2-N in the presence of 1 mg/liter CI" to 12.4 mg/liter NO2-N at
41 mg/liter CI". Similar conclusions have been reported for coho salmon
(Perrone and Meade 1977) and for steelhead trout (Wedemeyer and Yasutake
1978). We have conducted some nitrite bioassays with addition of bromide
(Br"), sulfate (SO42-), phosphate (PO43-), and nitrate (NO3"); the results
of these tests indicate that these other anions also exhibit, in different
degrees, an inhibitory effect on nitrite toxicity. It is apparent that the
toxicity of nitrite is highly dependent on the chemical composition of the
water.
Crawford and Allen (1977) studied the effect of calcium (Ca^"^) and of
seawater on nitrite toxicity to chinook salmon (0. tshawytscha) . The acute
toxicity of nitrite in seawater was markedly less than that in freshwater,
logically so because of the chloride effect discussed above. Crawford and
Allen also found that increasing the calcium concentration both in fresh-
water and in seawater decreased the toxicity of nitrite.
228
TABLE 1. CHEMICAL CHARACTERISTICS OF THE DILUTION WATER USED IN
BIOASSAYS. (ALL VALUES ARE MG/LITER UNLESS OTHERWISE NOTED)
Alkalinity,
as
CaC03
171
Hardness, as
CaC03
200
PH
7.70
Temperature,
C
I
9.8
S.E.C., ymho/cm
25 C
339
TOG
3.3
Turbidity, NTU
1.6
NH3-N
0.00
NO2-N
0.00
NO3-N
0.14
ci-
0.16
F"
0.35
PO43-
0.05
SO/,2-
17.2
Al
<1
As
0.0012
Ca
52.1
Cd
<0.005
Cr
<0.005
Cu
0.007
Fe
0.004
Hg
0.00030
K
0.82
Mg
16.7
Mn
0.002
Na
2.5
Ni
<0.005
Pb
<0.015
Se
0.00085
Zn
0.01
229
TABLE 2. ACUTE TOXICITY OF NITRITE TO RAINBOW TROUT (SALMO GAIRDNERI)
UNDER UNIFORM WATER CHEMISTRY CONDITIONS
Test
Number
Average
Wt.{q)
Fish Size
Length(cm)
96-hour LCE
(mg/1
iO (95% C.I.)
N02-N)
117
2.3
—
0.38 (
0.34-0.43)
579
3.1
6.3
0.25 (
0.21-0.30)
585
7.0
9.1
0.40
N-C.""
587
8.0
8.6
0.36 (
0.33-0.39)
590
8.2
8.7
0.30 (
0.26-0.34)
182
8.8
8.8
0.14
0.12-0.16)
597
10.0
9.3
0.21 (
0.19-0.24)
600
10.4
9.2
0.17
[0.15-0.20)
120
11.9
—
0.21
0.18-0.24)
121
12.1
—
0.21
;0. 19-0.23)
605
12.8
10.3
0.21
[0.19-0.24)
610
13.2
10.3
0.22
(0.18-0.27)
102
14.0
—
0.26
(0.21-0.32)
323
20.6
11.8
0.27
N.C.
326
24.3
12.3
0.28
(0.24-0.32)
243
53.1
15.7
0.27
(0.22-0.32)
244
60.5
16.6
0.27
(0.23-0.32)
423
188
23.6
0.19
(0.15-0.24)
138
235
—
0.20
(0.16-0.24)
505
387
29.7
0.24
(0.17-0.33)
^N.C. = Confidence interval not calculable.
230
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232
TIME, days
Figure 3. Toxicity curves showing effect of chloride on nitrite
toxicity to rainbow trout (Salmo gairdneri) .
233
50
- 40
E
<
cc
H
LU
O
o
o
LU
9
o
-J
X
CJ
30
20
10
— ▲■
^m •
ly
0
4NB
1
1
*36-hr data.
1
0 10 20 30 40
LC50, mg/l NITRITE NITROGEN
Finure 4. Effect of chloride on nitrite toxicity to rainbow
trout (Salmo gairdneri) .
234
An additional factor that should be considered in regard to nitrite
toxicity is the pH of the solution. Nitrite ion establishes the following
aqueous equilibrium.
NO2" + H+ t HNO2
The concentration of nitrous acid (HNO?) is 4-5 orders of magnitude less
than the concentration of nitrite ion (N02~) within the pH range 7.5 to
8.5; in going from pH 7.5 to 8.5, the N02~ concentration stays essentially
constant, whereas the HNO2 concentration decreases tenfold. Because this
equilibrium is pH-dependent, we studied the toxicity of nitrite to rainbow
trout over the pH range 6.4 to 9.0, to examine the effect of pH on nitrite
toxicity and to see whether toxicity could be attributed to one or the other
of the chemical species.
The results for a series of these experiments are shown in Figures 5
and 6. The first figure is a plot of 96-hour LC50 vs. pH for total NO2-N.
It shows that the toxicity of nitrite decreases with increasing pH. If the
toxicity of nitrite were solely due to the N02~ ion, this plot would be a
horizontal line. The second figure shows a plot of LC50 vs. pH for nitrous
acid (as N). If all the toxicity were attributable to this nitrite species,
this plot would be horizontal. Neither plot is horizontal, suggesting that
neither chemical species alone is responsible for the entire toxicity. Over
the pH range studied, both species are significantly, although not neces-
sarily equally, toxic. It is not possible to separate the toxicity into its
components without additional data, but in order to obtain these data by the
design we chose, experiments would have to be carried out beyond the pH
range acceptable for fishes.
The question of mode of toxic action of nitrite on fishes has also been
studied. Oxygen is transported in fish blood by the respiratory blood pig-
ment hemoglobin. The iron in hemoglobin is present in the ferrous, Fe(II),
state. Hemoglobin combines loosely with oxygen to form the easily disso-
ciated compound oxyhemoglobin, in which iron is still in the Fe(II) state.
The transport of oxygen by blood is dependent on the ease with which hemo-
globin unites with oxygen and with which oxyhemoglobin gives up oxygen. If
the iron in hemoglobin is oxidized to the ferric, Fe(III), state, methemo-
globin is formed. Methemoglobin is not capable of combining reversibly with
oxygen, and thus sufficiently high concentrations can cause hypoxia and
death. Nitrite in the blood oxidizes hemoglobin to methemoglobin, thereby
increasing the amount of methemoglobin present and impairing the ability of
the blood to transport oxygen.
It has been established that increased nitrite concentrations produce
increased methemoglobin levels in fish blood (Smith and Williams 1974; Smith
and Russo 1975; Brown and McLeay 1975; Crawford and Allen 1977; Perrone and
Meade 1977; Bortz 1977). The presence of high levels of methemoglobin in
fish blood is visually apparent in that the blood becomes brown-colored.
Different levels of methemoglobin have been reported as the concentrations
causing mortality in fishes. Species differences and differences in overall
physical condition may influence fishes' tolerance to different methemoglo-
bin levels.
235
■* 1
H I •— »
I •— I
CN
00
00
00
o X
00 Q-
h«H
CD
CN
iir«-
I I I I I I L
CO
CM
00
d
00
,co
N3D0dllN 31iailN l/6ui '°^3n
Figure 5. LC5(i (as NO2-N) vs. pH.
236
to
>
I
o
to
c
ID
O
U3
(U
s-
01
N3D0diiN aiovsnoaiiN i/Bu
'09
on
237
Some work has been done on treatment of methemoglobinemia. Ascorbic
acid administered intravenously reduced methemoglobin in rainbow trout blood
(Cameron 1971). Methylene blue administered either by injections (Bortz
1977) or by addition to test water (Wedemeyer and Yasutake 1978) also re-
duced methemoglobin levels. Removal of fish to nitrite-free water results
in a reduction of methemoglobin levels, although to a smaller extent than
found for methylene blue treatment (Wedemeyer and Yasutake 1978). Methylene
blue reduces methemoglobin levels rapidly, within a few hours. The treat-
ment appears to be temporary, in that methemoglobin levels gradually rise
again (Bortz 1977).
Methemoglobinemia, then, is one mechanism by which nitrite is toxic to
fishes. It is probably not the only mode of toxic action. Observations by
Smith and Williams (1974) that mortality occurred for some rainbow trout
with blood methemoglobin levels lower than other rainbow trout which sur-
vived led them to suggest that those fish died from a toxic reaction to ni-
trite itself rather than from methemoglobinemia. Crawford and Allen (1977)
observed that in seawater with added nitrite, chinook salmon had high (74%)
methemoglobin levels but very low (10%) mortality; in freshwater with added
nitrite, lower (44%) methemoglobin levels were found in the salmon, but 70%
mortality occurred. They further observed that fish dying in freshwater
often had red gill lamellae, not the brown color typically caused by methe-
moglobinemia. This indicates that the toxicity of nitrite in freshwater may
be attributable to something else besides or in addition to methemoglo-
binemia. More research is needed to determine what this mechanism is.
The effect of chloride and calcium also needs more study to elucidate
the mechanism by which these ions reduce nitrite toxicity. It has been sug-
gested (Perrone and Meade 1977) that chloride may compete with nitrite for
uptake through gills, or for entry into the red blood cell, thus suppressing
methemoglobin formation. Calcium does not appear to be affecting methemo-
globin formation, because raising the calcium level of freshwater did not
reduce methemoglobin levels in chinook salmon (Crawford and Allen 1977).
These are important areas for further research.
In conclusion, it is apparent that the toxicity of nitrite to fishes is
highly dependent on the chemical composition of the test water, and that
more research is needed to define the mechanism(s) of nitrite toxicity and
to learn more about ways to protect fish from nitrite poisoning.
REFERENCES
Anthonisen, A.C., R.C. Loehr, T.B.S. Prakasam, and E.G. Srinath. 1976. In-
hibition of nitrification by ammonia and nitrous acid. J. Water Pollut.
Control Fed. 48(5): 835-852.
Archer, M.C., S.D. Clark, J.E. Thilly, and S.R. Tannenbaum. 1971. Environ-
mental nitroso compounds: Reaction of nitrite with creatine and creati-
nine. Science 174: 1341-1343.
238
Ashley, L.M. and J.E. Halver. 1968. Dimethylnitrosamine-induced hepatic
cell carcinoma in rainbow trout. J. Nat. Cancer Inst. 41(2): 531-552.
Bortz, B.M. 1977. The administration of tetramethylthionine chloride as a
treatment for nitrite-induced methemoglobinemia in rainbow trout (Salmo
gairdneri ) . M.S. Thesis, American University, Washington, D.C. 54 p.
Brown, D.A. and D.J, McLeay. 1975. Effect of nitrite on methemoglobin and
total hemoglobin of juvenile rainbow trout. Prog. Fish-Cult. 37(1):
36-38.
Cameron, J.N. 1971. Methemoglobin in erythrocytes of rainbow trout. Comp.
Biochem. Physiol. 40A: 743-749.
Crawford, R.E. and 6.H. Allen, 1977. Seawater inhibition of nitrite toxi-
city to Chinook salmon. Trans. Am. Fish. Soc. 106(1): 105-109.
Hockenbury, M,R, and C.P,L. Grady, Jr. 1977. Inhibition of nitrif ication--
effects of selected organic compounds. J. Water Pollut. Control Fed,
49(5): 768-777.
Klingler, K. 1957. Natriumnitrit, ein langsamwirkendes Fischgift. (Sodium
nitrite, a slow-acting fish poison.) Schweiz. Z. Hydrol. 19(2): 565-
578. (In English translation).
McCoy, E,F, 1972. Role of bacteria in the nitrogen cycle in lakes. Water
Pollut, Control Res, Ser, 16010 EHR 03/72. Office of Research and Moni-
toring, U.S. Environmental Protection Agency, Washington, D.C. 23 p.
Mirvish, S.S, 1975, N^-Nitroso compounds, nitrite, and nitrate: possible
implications for the causation of human cancer, 15 pp. ln_ Proc. Con-
ference on nitrogen as a water pollutant. Vol, I. Analysis, sources,
public health, August 18-20, 1975, Copenhagen. International Associa-
tion on Water Pollution Research, London.
Newberne, P.M. 1979. Nitrite promotes lymphoma incidence in rats.
Science 204: 1079-1081,
Patrick, R., J,E, Colt, R,E. Crawford, B.A. Manny, R.C, Russo, R,V.
Thurston, and G.A. Wedemeyer. 1979. Nitrates, nitrites. Pages 158-
162. _l£ A review of the EPA Red Book: Quality criteria for water.
R.V. Thurston, R.C. Russo, CM, Fetterolf, Jr., T.A. Edsall, and Y.M.
Barber, Jr. (Eds.). Water Quality Section, American Fisheries Society,
Bethesda, MD,
Perrone, S.J. and T.L. Meade. 1977. Protective effect of chloride on ni-
trite toxicity to coho salmon (Oncorhynchus kisutch). J. Fish. Res.
Board Can, 34(4): 486-492,
Russo, R.C, CE. Smith, and R.V. Thurston. 1974. Acute toxicity of ni-
trite to rainbow trout (Salmo gairdneri ) . J. Fish. Res. Board Can.
31(10): 1653-1655.
239
Russo, R.C. and R.V. Thurston. 1974. Water analysis of the East Gallatin
River (Gallatin County) Montana 1973. Tech. Rep. No. 74-2, Fisheries
Bioassay Laboratory, Montana State University, Bozeman, MT. 27 p.
Russo, R.C. and R.V. Thurston. 1977. The acute toxicity of nitrite to
fishes. Pages 118-131. ^Recent advances in fish toxicology. R.A.
Tubb (Ed.). EPA Ecol. Res. Ser. EPA-600/3-77-085, U.S. Environmental
Protection Agency, Corvallis, OR.
Russo, R.C. and R.V. Thurston. 1978. Ammonia and nitrite toxicity to
fishes. Pages 75-82. ln_ Proc. of the Second USA-USSR Symposium on the
Effects of Pollutants upon Aquatic Ecosystems, June 22-26, 1976, Borok,
Jaroslavl Oblast, USSR. W.R. Swain and N.K. Ivanikiw (Eds.). EPA Ecol,
Res. Ser. EPA-600/3-78-076, U.S. Environmental Protection Agency,
Duluth, MN.
Sato, S., T. Matsushima, N. Tanaka, T. Sugimura, and F, Takashima. 1973.
Hepatic tumors in the guppy (Lebistes reticulatus) induced by aflatoxin
B , dimethylnitrosamine, and 2-acetylaminof luorene. J. Nat. Cancer
Inst. 50(3): 767-778.
Smith, C.E. and W.G. Williams. 1974. Experimental nitrite toxicity in
rainbow trout and Chinook salmon. Trans. Am, Fish, Soc, 103(2): 389-
390.
Smith, C.E. and R.C. Russo. 1975, Nitrite-induced methemoglobinemia in
rainbow trout. Prog. Fish-Cult, 37(3): 150-152.
Stanton, M.F. 1965. Diethylnitrosamine-induced hepatic degeneration and
neoplasia in the aquarium fish, Brachydanio rerio. J. Nat, Cancer Inst.
34(1): 117-130,
Thurston, R,V,, R,C, Russo, and C,E. Smith. 1978. Acute toxicity of ammo-
nia and nitrite to cutthroat trout fry. Trans. Am. Fish. Soc. 107(2):
361-368.
U.S. Environmental Protection Agency. 1977. Quality criteria for water.
Office of Water and Hazardous Materials, U.S, Environmental Protection
Agency, Washington, D,C. 256 p.
Wedemeyer, G.A. and W.T. Yasutake. 1978. Prevention and treatment of ni-
trite toxicity in juvenile steelhead trout (Salmo gairdneri ), J, Fish,
Res. Board Can. 35(6): 822-827.
Wolff, I. A. and A.E. Wasserman. 1972. Nitrates, nitrites, and nitrosa-
mines. Science 177(4043): 15-19.
240
TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
EPA-6QQ/9-8n-n^4
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
Proceedings of the Third USA-USSR Symposium on the
Effects of Pollutants Upon Aquatic Ecosystems :
Tfieoretical Aspects of Aquatic Toxicology
5. REPORT DATE
July 1980 Issuing Date
6. PERFORMING ORGANIZATION CODE
AUTHOR(S)
Environmental Protection Agency-USA
Soviet Academy of Sciences -USSR
8. PERFORMING ORGANIZATION REPORT NO.
, PERFORMING ORGANIZATION NAME AND ADDRESS
Large Lakes Research Station
Environmental Research Laboratory -
Grosse He, Michigan 48138
10. PROGRAM ELEMENT NO.
Duluth
A30B1A
11. CONTRACT/GRANT NO.
Joint US-USSR Project
02.02-13
12. SPONSORING AGENCY NAME AND ADDRESS
Environmental Research Laboratory - Duluth MN
Office of Research and Development
U.S. Environmental Protection Agency
Duluth, Minnesota 55804
13. TYPE OF REPORT AND PERIOD COVERED
Inhouse
14. SPONSORING AGENCY CODE
EPA/600/03
15. SUPPLEMENTARY NOTES
Prepared in cooperation with the Institute for the Biology of Inland Waters,
Soviet Academy of Sciences, Borok, Jaroslavl Oblast, USSR
16. ABSTRACT
The Joint US-USSR Agreement on Cooperation in the Field of Environmental Pro-,
tection was established in May of 1972. These proceedings result from one of the
projects. Project 02.02-13, Effects of Pollutants Upon Aquatic Ecosystems and Per-
missible Levels of Pollution.
As knowledge related to fate and transport of pollutants has grown, it has be-
come increasingly apparent that local and even national approaches to solving pollu-
tion problems are insufficient. Not only are the problems themselves frequently inter
national, but an understanding of alternate methodological approaches to the problem
can avoid needless duplication of efforts. This expansion of interest from local and
national represents a logical and natural maturation from the provincial to a global
concern for the environment.
In general, mankind is faced with very similar environmental problems regardless
of the national of political boundaries which we have erected. While the problems may
vary slightly in type or degree, the fundamental and underlying factors are remarkably
similar. It is not surprising, therefore, that the interests and concerns of
environmental scientists the world over are also quite similar. In this larger sense,
we are our brother's brother, and have the ability to understand our fellowman and his
dilemma, if we but take the trouble to do so. It is this singular idea of concerned
scientists exchanging views with colleagues that provides the basic strength for this
iproject,
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b. IDENTIFIERS/OPEN ENDED TERMS
c. COSATI Field/Group
Freshwater
Phosphorus
Ni trogen
Pesticides
Fishes
Stream Flow
Toxici ty
Bioassay
Communi ties
Phytoplankton
Nutrients
Waste Treatment
Water Quali ty
Toxic Substances
Macrobenthos
Microbiota
Water Quality Criteria
Great Lakes
Maximum
Permissible
Concentrations
57H
68D
18. DISTRIBUTION STATEMENT
Release to Public
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unclassified
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unclassified
22. PRICE
231.
EPA Form 2220-1 (Rev. 4-77)
PREVIOUS EDITION iS OBSOLE
241
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