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ILENR/RE-WR-92/07

IDENTIFICATION OF
TOXIC SUBSTANCE!
IN THE UPPER .O ^

ILLINOIS RIVER

^i^r--^ 7olo£i:

(sertraline HO)

Illinois Department of
Energy and Natural Resources

Jim Edgar, Governor
John S. Moore, Director

John S. Moore, Director

' — UUfSois Department of Energy and Natural Resources

325 West Adams Street. Room 300

Springfield. IL 62704-1892

217/785-2800

Telefax 217/785-2618

Identification of Toxic Substances in the Upper Illinois River

ILENR/RE-WR-92/07

Dr. Frank S. Dillon was the Project Manager for this research endeavor, and made a
substantial creative contribution to the drafting of the final report. His efforts were
left unacknowledged inadvertently.

A corrected title page and NTIS form are attached. ENR regrets the error.

Printfid on Bervcied Paoer

ILENR/RE-WR-92/07
Printed: October 1992
Reprinted: November 1992
Contract: WR 3691
Project: 89/215

Identification of Toxic Substances in the Upper Illinois River

Final Report

Prepared by:

River Research Laboratory

Forbes Biological Station

Illinois Natural History Survey

P.O. Box 599

Havana, IL. 62644

Principal Investigators:

Richard E. Sparks
Philippe E. Ross

Project Manager:
Frank S. Dillon

Prepared for:

Illinois Department of

Energy and Natural Resources

Office of Research and Planning

325 W. Adams, Room 300

Springfield, IL 62704-1892

Jim Edgar, Governor John S. Moore, Director

State of Illinois Illinois Department of

Energy and Natural Resources

REPORT DOCUMENTATION -"T^ ^ ....g./O;

, Tttl» »nd SuMHU *• •'•oort 0*t«

Identification of Toxic Substances in the Upper Illinois River October 1992

Philippe E. Ross and Frank S. Dillon

ng Organization Rapt. No

River Research Laboratory
Forbes Biological Station
Illinois Natural History Survey
P.O. Box 599
Havana, IL 62644

la Proiact/Taak/Worti Un)t No

89/215

11. Cantract(C) or GranKC)
(C)

,n, WR3691

12. Sponaonna Orsanxation Nama and Addr««t

Illinois Department of Energy and Natural Resources

Office of Research and Planning

325 West Adams Street

Springfield, IL 62704-1892

11. Typa of Rapon & Pariod Covarao

It. AbMTKt (Umtt .aOO worm)

Between 1955 and 1958, several abundant species of acquatic
nail clams practically disappeared from the upper Illinois
equally drastic repercussions on the populations of ducks a
invertebrates. The situation changed very little into the
in water quality. This research found that porewater from
contains a toxic factor that inhibits the filtering ability
also negatively effects the water flea, while stimulating a
All evidence points to ammonia as the toxic agent. Also, a
contained toxic oil products, including polycyclic aromatic
naphthalene. A three phase Toxicity Identification and Eva
utilized in reaching these conclusions.

insects, snails and finger-
River. These declines had
nd fish that fed upon tnese
1980s, despite improvements
relevant river sediments

of fingernail clams and
Iga and bacteria growth.
t two sites, the porewater

hydrocarbons , such as
luation methodology was

Water Pollution, Toxicity, Water Pollution Effects (Animals)
Sedimants, Sedimentation, Suspended Sediments

Sedimentation - Illinois River

Toxicity - Illinois River

Water Pollution - Illinois River

c. cosATi naid/GcuB Bioloqical and Medical Sciences;

Environmental Bioloqv

IS. AvaiiaMiity stMamant jjq restriction on distribution.

1 19. Sacurtty Claat (Thli Raport)

21. No. o« Pagai

Available at IL Depository Libraries or from
National Technical Information Service,
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OPTIONAL FORM 272 (

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Dapartmant of Con>mar

ILENR/RE-WR-92/07
Printed: October 1992
Reprinted: November 1 992
Contract: WR 3691
Project: 89/215

Identification of Toxic Substances in the Upper Illinois River
Final Report

Prepared by:

River Research Laboratory

Forbes Biological Station

linois Natural History Survey

P.O. Box 599

Havana, IL. 62644

Principal Investigators:

Richard E. Sparks
Philippe E. Ross

Prepared for:

Illinois Department of

Energy and Natural Resources

Office of Research and Planning

325 W. Adams, Room 300

Springfield, IL 62704-1892

Jim Edgar, Governor John S. Moore, Director

State of Illinois Illinois Department of

Energy and Natural Resources

This report has been reviewed by the Illinois Department of Energy and Natural Resources (ENR) and
approved for publication. Statements made by the author may or may not represent the views of
the Department. Additional copies of this report are available through the ENR Clearinghouse at
800/252-8955 (within Illinois) or 217/785-2800 (outside Illinois).

Printed by the Authority of the State of Illinois.
Date Printed:

Quantity Printed:

Referenced Printing Order:

October 1992

Reprinted: November 1992

250

200 reprints

One of a series of research publications published since 1975. This series includes the following
categories and are color coded as follows:

Energy Resources
Water Resources
Air Quality
Environmental Health
Insect Pests
Information Services
Economic Analysis

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Illinois Department of Energy and Natural Resources
Office of Research and Planning

325 West Adams, Room 300

Springfield, Illinois 62704-1892

217/785-2800

Printed on Recycled Paper

ACKNOWLEDGEMENTS

This research was administered and partially funded by the Illi-
nois Department of Energy and Natural Resources (IDENR), with the bal-
ance of funding from the Illinois Environmental Protection Trust Fund.
The development of the fingernail clam bioassay, one of five bioassays
used in this project, was funded separately by grant F-94-R from the
Federal Aid in Fish Restoration Program (Wallop-Breaux Act), adminis-
tered by the Illinois Department of Conservation.

Many people contributed to this project. Dr. Anthony A. Paparo,
Department of Zoology and School of Medicine, Southern Illinois Univer-
sity at Carbondale, did much of the early development of a variety of
methods for assessing responses of clams and mussels to contaminants.
Ms. Diane Dillon and Mr. Jeffrey Arnold provided valuable technical
assistance at Western Illinois University, as did Ms. Louann Burnett at
the Natural History Survey in Champaign, and Mr. K. Douglas Blodgett at
the Natural History Survey's Long Term Resource Monitoring Station in
Havana. Ms. Camilla Smith provided secretarial assistance at the River
Research Laboratory of the Stephen A. Forbes Biological Station in
Havana. The research could not have been done without the laboratory
facilities, office space, and equipment provided by Dr. Richard V.
Anderson and the Department of Biological Sciences at Western Illinois
University--to both we express our great appreciation. This project
would not have come into being without the sustained interest of Dr.
John Marl in, Director of the Illinois Pollution Control Board, in the
mysterious die-off of fingernail clams in the Illinois River and the
widespread ecological repercussions of their failure to recolonize.
Finally, we thank our project officers at DENR, Ms. Linda Vogt and Mr.
Will iam Denham.

TABLE OF CONTENTS

Executive Summary xi

1.0. Introduction 1

2.0. Project Goals and General Approach 9

3.0. Methods 11

3.1. Site Description 11

3.2. Sampling Design 12

3.2. Sample Collection Procedures 12

3.3. Chemical analysis 15

3.4. Bioassays 16

3.5. Toxicity identification and evaluation procedures 20

3.5.1. Phase 1 20

3.5.2. Phase II 21

3.5.3. Phase III 24

4.0. Results 25

4.1. Relative toxicity 25

4.2. Toxicity Identification Evaluation - Phase 1 29

4.2.1. 1990 29

4.2.2. 1991 32

4.3. Toxicity Identification Evaluation - Phase II 32

4.4. Toxicity Identification Evaluation - Phase III 40

4.5. Sensitivity of Recolonizing Clams 42

5.0. Discussion. 45

6.0. Li terature C i ted 53

LIST OF FIGURES

Figure 1.1. Sampling stations on the Illinois Waterway 2

Figure 1.2. Sampling stations in the Chicago area 3

Figure 1.3. TIE procedures 7

Figure 1.4. Phase I procedures 8

Figure 4.1. Results of toxicity tests with five test species 26

Figure 4.2. Fingernail clam response to porewaters 28

Figure 4.3. Correlation of toxicity with ammonia 41

LIST OF TABLES

Table 3.1. Location of sampling stations 13

Table 4.1. Toxicity of porewater to Ceriodaphnia dubi'a 30

Table 4.2. Results of treating sediment porewater 1990 31

Table 4.3. Toxicity and ammonia concentrations 33

Table 4.4. Results of treating sediment porewater 1991 34

Table 4.5. Results of treating sediment porewater with zeolite... 35

Table 4.6. Toxicity of sediment porewater 37

Table 4.7. Toxicity of extracts from sediment porewater 38

Table 4.8. Constituents of elutriates 39

EXECUTIVE SUMMARY

In a brief span of 3 years, 1955-1958, several abundant species of aquatic
insects, snails, and fingernail clams practically disappeared from a 170-km reach
of the Illinois River, from Hennepin on the north to the mouth of the Sangamon
River on the south. The declines of the fingernail clam, Musculium transversum,
were particularly spectacular: from average densities of 21,000 animals per square
meter to zero in Peoria Lake and backwater lakes near Havana. The declines had
drastic repercussions on the ducks and fish that fed upon the invertebrates. The
lesser scaup duck, or bluebill, virtually stopped using the Illinois River as a major
migration route, and there was a decline in the condition and growth of bottom-
feeding fish, including sport fish, -such as channel catfish, and commercially
important species, such as common carp.

The situation changed very little into the 1980s, despite improvements in
water quality (e.g., higher disolved oxygen levels attributable to improved waste
treatment in the Chicago-Joliet area and Peoria). This lack of recovery was espe-
cially puzzling because the- invertebrates are capable of rapidly recolonizing
barren areas; seed populations are available in spring-fed areas of Peoria Lake and
in tributaries and these organisms have short, rapid life cycles.

We found that porewater from Illinois River sediments contains a toxic
factor that inhibits the filtering ability of the clam, and the toxicity increases
upstream, peaking near Lockport. We observed the same pattern of sediment
toxicity with a different test organism, also representing a class of important food
organisms for fish and waterfowl: the water flea, Ceriodaphnia dubia. In contrast,
the porewater actually stimulates an alga and bacteria, but this is not surprising
because of the great physiological differences among plants, bacteria and animals.

Toxicity greatly decreased when the porewater was made slightly more acid
and porewater became nontoxic when filtered through a resin that removed
ammonia. Removal of heavy metals with a chelating agent had no effect on toxic-
ity. All the evidence points to ammonia as the culprit, especially since toxicity in
all tests correlated highly with the concentration of ammonia, which is known to
be toxic to aquatic animals. Since ammonia is a nutrient for plants and certain
types of bacteria, the presence of ammonia likewise could explain the stimulation
of these organisms.

Although ammonia appears responsible for the major upstream-downstream
pattern in toxicity, there were two sites where the porewater contained visible
signs of oil and the toxicity was associated with petroleum hydrocarbons, includ-
ing PAHs (polycyclic aromatic hydrocarbons) such as naphthalene.

During the course of this study, several species of fingernail clams, includ-
ing M. transversum, reappeared in the Chicago area waterways and in the Illinois
River at Peoria and Havana. There are at least four possible explanations for this
surprising reappearance of clams in the same general areas where the porewaters
tested toxic. First, we found that clams recolonizing the upper Illinois are more
resistant to ammonia than the clams from the lower Illinois, where the organisms
were obtained for all of the early bioassays. Second, our previous research demon-
strated that the surface layers of sediment in some areas are less toxic than layers
a few centimeters deeper. Toxicity may have been overestimated in tests where
surface and deep layers of sediment were mixed prior to testing. Third, toxic
episodes may be brief and infrequent, allowing organisms to colonize in between
episodes. Fourth, the distribution of toxicity in sediments may be extremely
patchy, so that healthy organisms are found adjacent to barren areas. If the latter
two hypotheses prove to be true, acute toxicity in the Illinois River has changed
recently from a widespread problem to a more localized or episodic problem.
Reduction of toxicity in surface sediments may reflect recent reductions in
ammonia loading from sewage treatment plants in the Chicago area, although it is
not clear whether the sources of ammonia in the porewaters are effluents, the
deeper layers of sediments, or both.

1.0 INTRODUCTION

The quality of sediments is critical to the ecological health of
aquatic ecosystems. Benthic organisms that live in sediments are key
links in food chains that lead from nutrients in water and sediment to
higher level consumers, such as fish and ducks. Sediments in aquatic
systems can be both sinks and sources for inorganic and organic contami-
nants. At present,. the extent of the sediment contamination problem is
largely unknown. Comprehensive assessments of the accumulation of
contaminants from agricultural, municipal, and industrial sources in
sediments of our rivers, lakes and estuaries have not been completed.
Currently, the U.S. Environmental Protection Agency has identified 134
sites with serious sediment contamination problems (USEPA 1988). In
addition, 41 areas in the Great Lakes (IJC 1988), 50 coastal sites, and
85 wildlife refuges have been identified where contaminated sediments
pose a problem (USEPA 1988).

In Illinois, contaminants have been identified in sediments
throughout much of the Illinois River and its associated tributaries and
waterways (Figures 1.1 and 1.2; Cahill and Steele 1986; Cahill and
Autrey 1987; Blodgett et al . 1984; Mathis et al . 1973; Polls et al .
1985; Harrison et al . 1981; Coleman and Sanzolone 1991; Bhowmik and
Demissie 1989; Sparks and Blodgett 1984; and Fitzpatrick and Bhowmik
1990). Two-thirds of the population of the state lives in the Illinois
River basin which drains approximately half the state (Talkington 1991).
The river historically has been one of the most productive rivers in
North America in terms of fish and wildlife populations. In 1908,

Des Plaines
River

•^•-) Detailed Map

Ohio River

Figure 1.1. Location of sediment sampling stations on the Illinois Waterway.
Stations are identified according to river miles: Illinois River miles (IR)
start at Grafton at mile 0.0 and proceed upstream to Chicago. A reference
station was established on the Mississippi River (MR), 377 miles above the
confluence with the Ohio River. ^

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a 320-km (200-mile) reach from the great bend at Hennepin to the conflu-
ence with the Mississippi River at Grafton (Figure 1.1) produced 10% of
the total U.S. harvest of freshwater fish and two thousand commercial
fisherman made a living from the river (U.S. Department of Commerce and
Labor 1911). The commercial yield was 24 million pounds annually, or
about 178 pounds per acre of permanent water (Lubinski et al . 1981). By
the 1950s the yield had dropped to 38 pounds per acre; since the 1970s
the yield has been a low 4 pounds per acre, totaling only 0.32% of the
total U.S. catch (Sparks 1984). Similar downward trends were recorded
over the same period for two other indicators of biological
productivity: waterfowl and sport fish populations (Bellrose et al .
1979; Sparks 1977; Sparks 1992). Major commercial fish species and the
diving ducks feed on bottom-dwelling invertebrates such as clams,
snails, aquatic worms, and aquatic insects. In the early 1900's a
healthy benthic community contributed to the tremendous production of
fish and waterfowl. A major component of that benthic community was a
small clam, Musculium transversum (Family Sphaeriidae) . Now, this clam
as well as other small mollusks, mayflies, midges, and other burrowing
aquatic insects has been virtually eliminated from certain reaches of
the Illinois River (Starrett 1972, Anderson 1977, Sparks et al . 1986).

Declines in the benthic invertebrates of the Illinois River system
have been linked to sediment-associated toxicity (Sparks et al . 1981;
Blodgett et al . 1983; Sparks 1984). Aquatic sediments can accumulate
both inorganic and organic chemicals that are absorbed to particulate
matter or are in solution in sediment porewater (Salomons et al . 1987,
Tessier and Campbell 1987). Porewater (also called interstitial water)
is the water occupying the spaces between the sediment particles. These

contaminants can have acute toxic effects on benthic organisms, or
accumulate slowly in the organisms until some toxic threshold is
reached.

The toxicity to aquatic organisms is known for only a fraction
(<1%) of the approximately 50,000 compounds manufactured in the U.S.
(Martell et al . 1990). This situation is further complicated by the fact
that organisms usually are simultaneously exposed to a number of chemi-
cals (Giesy et al . 1990). The toxic responses associated with these
mixtures of compounds depends on their bioavailability--some contami-
nants are bound to sediment particles or otherwise unavailable to organ-
isms. For instance, the bioavailability of non-ionic organic compounds
depends on the total organic carbon content (TOC) of the sediment
(Nebeker et al . 1989, Swartz et al . 1990, Di Toro et al . 1991) and the
bioavailability of certain cationic metals depends on the acid-volatile
sulfide (AVS) content of the sediment (Di Toro et al . 1990, Ankley et
al. 1991, Carlson et al . 1991). Due to the complex mixtures of contami-
nants present in most toxic sediments, as well as the effects that
sediment matrices may have on the bioavailability of compounds, it has
been difficult to link specific compounds with toxicity. The tradition-
al approach to identifying toxic agents has been to correlate toxicity
with the concentrations of chemicals in the bulk sediment sample (Carr
et al . 1989). This approach does not work well with complex mixtures
and does not address the question of bioavailability. The dose response
curve for biological effects from certain chemicals is not correlated to
the bulk sediment concentration but rather to the porewater concentra-
tion (Di Toro et al . 1991). The recent development of Toxicity Identi-
fication and Evaluation (TIE) methodology has made it possible to iden-

tify specific toxic compounds in complex mixtures (Figure 1.3; Mount and
Anderson-Carnahan 1988; Mount and Anderson-Carnahan 1989; Mount 1988).

TIE procedures use toxicity-based fractionation schemes to charac-
terize and identify compounds in aqueous samples that exhibit toxicity
to aquatic organisms. Although TIE cannot be used on bulk sediments, it
can be applied to the aqueous fraction (porewater). Previous studies
(Adams et al . 1985; Swartz et al . 1985; Knezovich and Harrison 1988;
Connell et al . 1988; Swartz et al . 1988, Di Toro et al . 1992) have shown
a correlation between toxicity or bioaccumulation of a number of contam-
inants by benthic macroinvertebrates, on the one hand, and porewater
concentrations on the other. The TIE procedures are designed to address
multiple toxicant interactions as well as matrix effects on bioavail-
ability. The major strength of TIE is that it allows direct relation-
ships to be established between toxicity and chemical analyses. TIE is
a phased approach that is designed to isolate, identify and confirm the
presence of acutely toxic compounds. TIE methodology for identification
of chronically toxic compounds is currently under development (USEPA
1992). Phase I of TIE consists of a series of chemical and physical
manipulations designed to remove or render biologically unavailable
generic classes of compounds (Figure 1.4). Phase II uses information
from Phase I to focus appropriate analytical methods on toxic fractions.
Phase III consists of methods designed to verify that the suspected
toxicant is the actual toxicant. TIE methodology has been applied to
sediments from the Great Lakes (Ankley et al . 1990) and the Calumet Sag
Channel of the Illinois River system (Schubauer-Berigan and Ankley
1991). We applied these techniques to sediments from the Illinois River
System in an effort to identify the substance or substances responsible
for the declines of the benthic invertebrates.

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2.0 PROJECT GOALS AND GENERAL APPROACH

The primary goal was to identify the toxicants in the sediments of
the Illinois Waterway. In addition, we hoped to identify upstream-
downstream patterns in toxicity and follow toxicity gradients upstream
to sources.

We tested the toxicity of sediments taken widely along the entire
length of the Illinois Waterway, and from one reference site on the
Upper Mississippi River (Figures 1.1 and 1.2). Next, sediments that
exhibited toxicity were subjected to Phase I Toxicity Identification and
Evaluation (TIE) procedures (pH adjustment, addition of a chelating
agent, etc.), to characterize the toxicants (see Figures 1.3 and 1.4).
If the toxicity at all sites varied the same way in response to the
Phase I treatments, we would know we were dealing with one class of
toxicants, or perhaps even one major toxicant, and we could focus
additional sampling on the most toxic reach in the hope of identifying
the major source. At the same time. Phase II and Phase III TIE
procedures would confirm the identity of the toxicant or at least narrow
the range of suspect chemicals. However, if Phase I testing indicated
that different classes of toxicants occurred in different reaches of
the waterway, then we would have a much more complex task of identifying
multiple toxicants and multiple sources--a task that might extend well
beyond the budget and time limitations of this research.

3.0 METHODS

3.1 Site Description

Today's Illinois Waterway is approximately 327 miles (526 km) long
connecting Lake Michigan and the Chicago-Joliet metropolitan area with
the Mississippi River and the agricultural heartland, near Grafton,
Illinois (Figure 1.1). The headwaters are in the highly industrialized
Chicago area where the flow of the Chicago River was reversed to carry
wastes away from Lake Michigan into the Illinois River via the Chicago
Sanitary and Ship Canal and the downstream portion of the Des Plaines
River (Figure 1.2). The Calumet Sag Channel enters the Sanitary and
Ship Canal near Lemont. The Illinois River proper begins with the
confluence of the Des Plaines and Kankakee rivers, and flows through a
predominantly agricultural drainage, although the industrial city of
Peoria is situated approximately mid-way along the waterway.

Locations on the waterways are designated by river mile as record-
ed in river charts prepared by the U.S. Army Corps of Engineers (1987)
and by markers along the waterways, starting with mile 0.0 at the con-
fluence with the Mississippi and proceeding upstream to Chicago. The
following abbreviations are used in the text, figures, and tables to
identify reaches of the waterway, and stations are identified by reach
abbreviation and river mile:

IR Illinois River proper

DP Des Plaines River

CS Calumet Sag Channel

SS Chicago Sanitary and Ship Canal

CR Chicago River

The one reference station on the Upper Mississippi River is located
377.0 miles above the confluence with the Ohio River and is designated
MR 377.0. The locations of the sample stations are given in Table 3.1
and Figures 1.1 and 1.2. In accordance with Corps of Engineers termi-
nology, the designation "left bank" or "right bank" assumes the observer
is facing downstream.

3.2 Sampling Design

Nineteen sampling stations were established throughout the
Illinois Waterway (Figures 1.1 and 1.2). Samples were collected from 15
stations from November 1989 to June 1990, and from all 19 stations from
November 1990 to June 1991 (Table 3.1).

3.2 Sample Collection Procedures

It is important to limit the disruption of the sediment so that
toxicity evaluations are conducted under conditions that closely match
the in situ conditions (ASTM 1991). The most appropriate sediment
sampling device is study specific. Sediment corers generally disrupt
the sediment little but collect a limited sample volume (ASTM 1991).
This study employed a battery of bioassays as well as the TIE proce-
dures, all of which used sediment porewater. The volume of porewater
needed for this work made the use of sediment corers impractical. We
used a 25.4 cm (10-inch) Ekman dredge that works well in the soft to
semi-soft sediments that characterize the Illinois Waterway and collects
a relatively large sample volume (ASTM 1991).

Table 3.1. Location of sampling stations.

River Mile Description

North Branch of Chicago River at Michigan Avenue Bridge
South Branch of the Chicago River at Harrison Street Bridge

Upstream of Division Street Bridge on Calumet Sag Channel
Upstream of 104th Street Bridge on Calumet Sag Channel

-5 m from left bank

-25 m from left bank

-2 m from right bank downstream of Route 171 Bridge

-10 m from left bank upstream from Justice Navigation Light

10 m upstream of sunken barge and 30 m from right bank

Left bank -300 m upstream of Brandon Road Lock and Dam

-30 m from left bank across from 01 in Chemical

Upstream of Du Page River Daymark -500 m from right bank

-100 m upstream of Bal lards Island

Center of Turner Lake

Upper Peoria Lake, south of Chill icothe

SE Corner of Lake Chautauqua

Center of Meredosia Lake

Entrance to Swan Lake

MR 377.0 Montrose Flats, Pool 19, Mississippi River

Note: The Illinois Waterway includes the Illinois River (IR), Des
Plaines River (DP), Chicago Sanitary and Ship Canal (SS), Chicago river
(CR), and Calumet Sag Channel (CS). The mileages start at IR 0.0 at the
confluence with the Mississippi and proceed upstream to Chicago. Mile-
ages on the Upper Mississippi River (MR) start at the confluence with
the Ohio. "Right" and "left" assume the observer is facing downstream.
m = meters.

326

324

318

307

317

315

313

310

292

286

281

277

248

215

180

125

The sampler was rinsed with river water at the site prior to
sediment collection. The sample was placed in prewashed (Biosoap wash,
ultrapure water rinse) high density polyethylene containers. High
density polyethylene containers are relatively inert and are optimal for
samples contaminated with a variety of chemicals (ASTM 1991). The con-
tainers were filled completely to achieve zero sample head space.
Sample containers were placed on ice as soon as possible following
collection (never exceeding 2 hours). Samples were transported to the
laboratory and stored at 4°C for no more than two weeks as recommended
by Anderson et al . (1984).

We used sediment porewater in our toxicity tests. Numerous stud-
ies (Adams et al . 1985; Swartz et al . 1985; Knezovich and Harrison 1988;
Connell et al . 1988; Swartz et al . 1988, Di Toro et al 1992) have shown
that porewater is an appropriate surrogate for bulk sediment. Porewater
can be collected from sediment samples by several methods:
centrifugation, squeezing, suction, and equilibrium dialysis (ASTM
1991). Centrifugation is generally used if large volumes of porewater
are required (Edmunds and Bath 1976). Constituents such as salinity,
dissolved inorganic carbon, ammonia, sulfide, and sulfate are generally
not affected as long as oxidation is prevented; however, dissolved
organic carbon (DOC) and dimethyl sufide may be significantly reduced
using this method (Howes et al . 1985). Sediment porewater was extracted
by centrifugation at 4000 g (g = the acceleration due to gravity) at 4°C
for 45 minutes. Sample porewater was stored with zero head space at 4°C
in a decontaminated cubitainer for a maximum of 1 week. The time from
collection to testing ranged from 1 to 6 days, and averaged 2.6 days for

all sediments.

Surface water samples were collected just prior to collection of
sediment. Surface water was collected from approximately mid-depth in
the water column using a Van Dorn sampler. Samples were placed in pre-
cleaned cubitainers and immediately placed on ice. Surface water sam-
ples were stored at 4°C for a maximum of one week.

3.3 Chemical Analysis

Routine chemical measurements were taken on both surface water and
porewater samples. Samples were brought to ambient temperature (20-
24°C) prior to making the following measurements:

SURFACE WATER PORE WATER

Dissolved Oxygen Dissolved Oxygen

pH (negative log of the hydronium pH

ion concentration (minus {H''"))) Conductivity

Alkalinity

Conductivity Hardness

Alkalinity Total Ammonia-N

Hardness Total CI (chlorine)

Total Ammonia-N (ammonia measured H2S (hydrogen sulfide)

as nitrogen, N) Sulfide

Dissolved oxygen was measured using a standard Y.S.I. Model 57
oxygen meter with a Y.S.I. Model 5739 probe. Temperature and pH were
measured using a Jenco Microcomputer pH-Vision 6071 pH meter with a
temperature-compensating Ross electrode. Specific conductance was
measured using a Y.S.I. Model 35 Conductance Meter with a Y.S.I. Model
3401 probe. Total alkalinity was measured using the ASTM (1982) stand-
ard titration method. Total hardness was measured using the EDTA titra-
metric method (APHA 1989). Total ammonia nitrogen was determined using
the Nesslerization method (APHA 1989), total residual chlorine by the
DPD colorimetric method (APHA 1989), sulfide by the methylene blue

method (APHA 1989) and hydrogen sulfide by the lead sufide method. All
instrumentation was calibrated prior to testing.

We intended to calculate the fraction of the total ammonia that
existed in the un-ionized state during the toxicity tests (see below).
In aqueous ammonia solutions an equilibrium exists between ammonia in
the highly toxic un-ionized form (NH3) and ammonia in the relatively
nontoxic ionized form {NH4"''). The dominant factor regulating the
equilibrium between the two forms is pH, with the temperature having a
lesser effect. We were not able to calculate un-ionized ammonia
concentrations in the toxicity tests because the pH of the porewater
drifted slightly during the tests. Temperature was held constant.
However, our subsequent analysis of the correlation between toxicity and
total ammonia is justified because the initial pHs of the samples were
similar (6.5-7.25) and all drifted in a similar manner, so the un-ion-
ized ammonia concentrations were some consistent fraction of the total
ammonia concentrations in all the test chambers.

Measurements of total organic carbon (TOC) were performed on bulk
sediment samples. The results are expressed in percent organic
carbon.

3.4 Bioassays

Burton (1991) described several components that should be consid-
ered in selecting a bioassay for toxicity assessment:

Components of an Optimal Toxicity Assay

1. Verification components

Ecosystem relevance

Species sensitivity patterns

Appropriate test phase

Short or long exposure period

Definitive response dynamics

2. Resource components

Organism availability
Laboratory availability
Expertise required
Expense and time required

3. Standardization components

Approved standard methods

Reference data base

Interlaboratory validation

Quality assurance and control criteria

Verification components such as ecosystem relevance, sensitivity, and
discriminatory ability are so critical that multiple species and end-
points should be incorporated in testing programs for sediment toxicity
assessments, according to Burton (1991). Therefore we measured the
relative toxicity of the sediment porewater with a battery of bioassays
that included the following test organisms: the marine bacterium, Photo-
bacterium phosphoreum (Microtox''^), the freshwater alga, Selenastrum
capricornutum, the rotifer, Branchionus calyciflorus, the daphnid,
Ceriodaphm'a dubia, and the sphaerid clam, Musculium transversum. The
Microtox''^ assay measures the luminescence of P. phosphoreum (Bulich et
al . 1981). Inhibition of this luminescence is considered a toxic re-
sponse. The 5. capricornutum assay measures the inhibition of photosyn-
thetic activity of an algal culture as a measure of toxicity (Ross et
al . 1988). The rotifer assay is a mortality test (Snell and Personne
1989). The C. dubia assay was the standard USEPA (1985) acute assay
(48-hour mortality). The sphaerid or fingernail clam assay is based on
measuring changes in filtering rates. The dilution water used in the
toxicity tests and for maintaining the organisms was Perrier'^ bottled
water.

The fingernail clam filtering assay used in this study is based on
observations by Aldridge et al . (1987), Sparks and Sandusky (1983),
Sparks et al . (1981), and Anderson et al . (1978) that stresses, includ-
ing toxicants, impair the ability of bivalves to filter particles from
water (including the food particles on which the clams feed). The assay
is outlined below and a detailed description is given in Sparks et al .
(1992). Filtering rates are determined by measuring the fingernail
clams' -ability to filter yeast from a suspension of known concentration.
Fingernail clams are first exposed to the porewater sample for one hour.
They are then placed in a yeast suspension and allowed to filter for one
hour. Two controls are used: the first consists of the yeast suspen-
sion alone and is used to determine the change in concentration due to
settling of the yeast. The second control determines the baseline
filtering rate of clams exposed for 1 hour in clean, uncontaminated
water. The yeast concentrations are measured at the beginning and end
of the filtering period. The filtering rates of the exposure and
control tests are then determined by taking the initial yeast
concentration minus the final concentration minus the amount settled
divided by the weight of the test organisms. Filtering rates are
expressed as the concentration of yeast filtered per unit weight of
organism per unit time.

C^ = initial concentration of yeast

Cf = final concentration of yeast

W = wet weight of clams, in g (grams)

C3 = change in yeast concentration due to settling

- Cr - C-

filtering rate in mg (milligrams) yeast/g clam/hour

The exposure filtering rate is then compared to the control. The test
result is a sublethal response (percent reduction in filtering rate,
relative to the control) as opposed to an "all or none" (death or sur-
vival, toxic or nontoxic) type of response. The inhibition of the
filtering performance of the clams is proportional to the severity of
the stress (Sparks et al . 1992). For purposes of evaluating sediments
for toxicity, it is useful to be able to rank sites based on relative
toxicity. Only the 1990-1991 porewater samples were evaluated using
this assay because it was not fully developed until late 1990.

The results of the various assays were standardized for easier
comparisons. The treatment results were divided by the control results
and then 1 was subtracted from the quotient. A negative value indicates
inhibition (toxicity), a positive value indicates stimulation, and 0
indicates no response (no difference with respect to the control). If
we use the fingernail clam filtering bioassay as an example:

T = test response to sample of sediment porewater

C = control response to uncontaminated dilution water

T = 3.4 mg yeast/g clam/hour

C = 6.5 mg yeast/g clam/hour

T/C = 3.4/5.5 = .52

.52 - 1.00 = -.48 A decline of 48% from the

control value, a marked inhibition of the filtering

ability of the clams.
Results of the C. dubi'a bioassay are expressed in toxicity units,
as well as 48-hour LC50s, where toxicity units = 100/(48-hour LC50).
The 48-hour LC50 is the percent dilution of porewater (or treated pore-
water) that kills 50% of the test organisms in 48 hours. For example,

if a 7% solution (by volume) of porewater in dilution water is the LC50
(see site CS307.4, Table 4.1),

7% = 48-hour LC50

100/LC50 = 100/7 = 14.3 toxic units
meaning that the toxicity in the porewater is more than 14 times the
lethal dose.

3.5 Toxicity Identification and Evaluation Procedures

Samples exhibiting acute toxicity to C. dubia were subjected to
Toxicity Identification and Evaluation (TIE) procedures developed at the
USEPA's National Effluent Toxicity Assessment Center (NETAC). The goal
is to separate toxicants from nontoxic compounds, using sample fraction-
ation techniques in combination with bioassays to determine which frac-
tions contain most of the toxicity. We used C. dubia as the TIE test
organism, because it is a widely-accepted reference species. The TIE
approach consists of three phases outlined in Figure 1.3.

3.5.1 Phase I characterizes the physical and chemical properties of the
sample toxicants by altering or rendering biologically unavailable
generic classes of compounds (Mount and Anderson-Carnahan 1988). After
Phase I the toxicants are classified as having characteristics of ca-
tionic metals, non-polar organics, volatiles, oxidants, or substances
not affected by Phase I methods. Phase I manipulations are outlined in
Figure 1.4. The primary tool of Phase I is manipulation of sample pH.
The questions asked are: (1) Is toxicity different at different pHs?
(2) Does sample manipulation at different pHs affect toxicity? (3) Is
toxicity attributable to cationic metals, such as copper or lead? (4)
Is toxicity associated with oxidizing agents, such as chlorine or

chloramines? The graduated pH test answers the first question and is
designed to indicate a pH-dependent toxicant such as un-ionized ammonia.
The second question is answered by performing the following tests at
different pHs: aeration, filtration and reverse-phase solid phase ex-
traction (SPE). Aeration tests determine whether toxicity is attributa-
ble to volatile or oxidizable compounds. The filtration tests indicate
whether toxicity is associated with filterable components. Reverse-
phase SPE indicates whether toxicity is attributable to non-polar com-
pounds. Presence of toxic cationic metals is indicated if addition of a
chelating agent, ethylenediaminetetraacetic acid (EDTA), diminishes
toxicity. Presence of chlorine or other oxidizing agents is indicated
by a reduction in toxicity following addition of the reducing agent,
sodium thiosulfate.

3.5.2 Phase II uses chemical fractionation techniques in parallel with
toxicity tests to isolate suspected toxicants (Mount and Anderson-Carna-
han 1989). Our Phase I results strongly implicated ammonia as a toxi-
cant, so we retested the samples after selectively removing ammonia
using a zeolite ion exchange resin, following the methods of Mount and
Anderson-Canaragan (1989), Ankley et al . (1990), and Schubauer-Berigan
and Ankley (1991). Zeolites are naturally-occurring or synthetically-
created crystalline hydrated alkali -aluminum silicates. A column was
prepared by packing a glass tube with a commercially available zeolite
product. The sample was passed over the zeolite column using a metering
pump, at a flow rate of approximately 10 ml/min (milliliters per minute).
Post column samples were analyzed for total ammonia and screened for
acute toxicity.

In addition to implicating ammonia, Phase I testing also indicated
that toxicity in some samples was associated with non-polar organic
materials and with material that was retained on the filters. We
applied the following Phase II isolation techniques that were used in a
similar situation by Schubauer-Berigan and Ankley (1991). To verify
that toxicity was due in part to material retained on the filters, the
filters were extracted with methylene chloride. Filters used in Phase I
for samples from the Des Plaines River site DP277.0 and the Calumet Sag
Channel site CS307.4 were soaked in 10 ml of methylene chloride for 1
hour. The solvent was evaporated from the beakers and dilution water
was added to the same volume as the original filtered sample. The
extracts then were screened for acute toxicity.

Having checked the toxicity of the material on the filters, we
next investigated the nonpolar organics using solid phase Cjg absorption
columns and subsequent chromatography. To maximize the extraction of
possible toxicants, filtration was omitted and porewater was centrifuged
at 10,000 g for 30 minutes to remove particles that would clog the Cjg
column. The supernatant from the centrifugation step was checked for
toxicity. If toxicity was present, a 200-ml sample of the supernatant
was passed over a 6-ml Cjg column that had been conditioned with 25 ml
of methanol followed by 25 ml of Millipore^^ ultrapure water. Post
column aliquots were collected after passage of 25 ml and 100 ml of
methanol and tested for toxicity.

Toxicity was not recovered from the DP277.0 sample using 100%
methanol elutions of the C^g columns as suggested by Mount and Anderson-
Carnahan (1989), so we eluted the columns with increasingly nonpolar
mixtures of methylene chloride in methanol (1, 5, 10, 15, 20, 25, 30,
35, 40, 45, 50, and 100%) as suggested by Schubaurer-Berigan and Ankley

(1991). The elutriates were collected in hexane-rinsed scintillation
vials. Toxicity may not have been recovered by methanol extractions
because either the toxic compounds or the oil and grease they are sorbed
to are highly nonpolar. The methylene chloride was evaporated and the
sample restored to volume using methanol. The fractions were tested for
toxicity using 150 u^ (microliters), 75 ul , and 37.5 ul of the fraction
in 10 ml of dilution water. The methanol concentrations were below the
48-hour LC50 for C. dubia, so toxicity was attributed to the nonpolar
organic solutes, rather than to the methanol solvent.

The toxic fractions were sent to Daily Analytical Laboratories in
Peoria, Illinois for analysis on a Hewlett-Packard 5890A gas chromato-
graph with a 5970A Series mass selective detector along with a 7673A
autosampler. The methanol concentrate was injected into a 30-m (meter)
x 0.25-mm (millimeter)-i .d. DB-5 J&W capillary column. The temperature
program was 40°C for 4 minutes followed by an increase at a rate of 10°
C per minute to a peak of 300° C for 10 minutes. Run time was 40
minutes with a scan start time at 3 minutes. The peak detection
threshold was 10,000 counts, with a threshold at 100 counts. A
splitless injection mode was used along with a linear scanning method
from 40-450 mhz (megahertz). The samples had 40 ug (micrograms)/ml of
internal standards of the following compounds; l,4-Dichlorobenzene-d4,
Naphthalene-dS, Acenaphthene-dlO, Phenanthrene-dlO, Chrysene-dl2 and
Perylene-dl2. After the sample was analyzed by the mass selective
detector, they were compared to library searches using the NIH (National
Institutes of Health) EPA (U.S. Environmental Protection Agency) Mass
Spectral Database. Identifications were based on the best fit with a
minimum search fit of 70%.

3.5.3 Phase III confirms the identity of toxicants that are provision-
ally identified in Phases I and II. We employed two methods from the
suite of Phase III techniques suggested by Mount (1988): (1) We corre-
lated toxicity with measured concentrations of suspect chemicals in our
test solutions, and (2) we compared the relative sensitivity of our test
species to known toxicants and to our samples. The correlation analy-
sis was performed on the toxicity tests which used the standard refer-
ence animal, C. dubia. As mentioned earlier, the correlation analyses
used total ammonia concentrations, rather than un-ionized ammonia con-
centrations. The drift in pH during the toxicity tests made it impossi-
ble to calculate un-ionized ammonia concentrations based on measurements
of total ammonia and the pH of the test solutions.

4.0 RESULTS

4.1 Relative Toxicity

There were marked differences in the responses of the five test
organisms to sediment porewater from the same sites (Figure 4.1).
Luminescence of the marine bacterium, Photobacterium phosphoreum,
(Microtox test) was inhibited by 34% at SS313.0 on the Sanitary and Ship
Canal and 32% at CS307.4 on the Calumet Sag Channel. Maximum stimula-
tion of approximately 50% occurred at the next site upstream on the
Calumet Sag Channel, CS318.5. Responses to porewaters from other sites
were slight and variable, sometimes mildly inhibitory and sometimes
mildly stimulatory.

Photosynthesis by the freshwater alga, Selenastrum capricornutum,
was markedly stimulated, by a factor of nearly 2, by sediment porewaters
from the mouth of Swan Lake, IR6.0, and the Sanitary and Ship Canal,
SS310.0. Stimulation is an indication of nutrient enrichment; e.g., by
nitrogen and phosphorus (Ross et al . 1988). The greatest inhibition,
-86%, was caused by sediment porewater from Lake Chautauqua, IR125.5,
although inhibition also occurred at IR72.0, IR281.1, SS313.0, SS315.3,
and CS307.4.

A large percentage of the rotifers, Branchionus caTciflorus, died
in porewaters from Meredosia Lake (IR72) and Lake Chautauqua (IR125.5),
but the rotifers exhibited no significant responses to samples taken
anywhere else (Figure 4.1).

In contrast to the microorganisms (bacterium, alga, and rotifer),
the macroinvertebrates C. dubi'a and M. transversum were remarkably

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consistent in their responses to the sediment porewaters. Both organ-
isms exhibited no inhibitory response to porewaters from the lower
Illinois River or from the reference site in the Upper Mississippi River
(Figure 4.1). The stimulation of filtering performance in the finger-
nail clam, M. transversum, may have been caused by favorable ratios of
dissolved sodium, potassium, calcium, and magnesium salts in porewaters
from the lower river. Anderson, Sparks and Paparo (1978) demonstrated
the importance of these salts in regulating the beating of the cilia on
the gills of the clams. Salts that affect the cilia are likely to
affect filtering performance because the lateral cilia produce the water
currents that bring food into the clam and the latero-frontal cilia act
as filters. Also, the presence of organic matter in the sediment pore-
waters may have stimulated a feeding response in the clams, which are
deposit feeders, as well as water column filterers. The clam and the
water flea likewise are consistent in indicating toxicity in the upper
waterway. Filtering performance in the clam was inhibited starting with
sediment porewaters from IR248.2 near Marseilles and water flea mortali-
ty started at DP277.0, just above the mouth of the Du Page River near
the Interstate 55 bridge. Sediment porewaters from 7 of the 13 upstream
sites were toxic to C. dubia, and 12 of 13 inhibited the fingernail clam
(Figure 4.1).

Since the fingernail clam is the organism of main interest in this
study, the response of the clams is extracted from Figure 4.1 and
presented separately in Figure 4.2. With the exception of a stimulatory
response to porewater from one station in the Sanitary and Ship Canal,
SS315.3, all the upper waterway stations exhibited some degree of

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toxicity to the fingernail clam, with the most toxic stations located in
the reach between the mouth of the Du Page River, DP277.0, and the
Summit-Stickney area, SS313.0.

4.2 Toxicity Identification Evaluation - Phase I

Standard toxicity identification evaluations (TIE) use C. dubia to
determine whether various treatments reduce the toxicity of porewater.
We felt it was unnecessary to use the nonstandard clam bioassay in TIE
because the fingernail clams and C. Dubia responded similarly to the
sediment porewaters, and C. dubia appeared to be an adequate surrogate
for the clam.

Seven sites were acutely toxic to C. dubia in 1990 and 1991, with
six sites in common between the two years (Table 4.1). One site on the
upper Calumet Sag Channel, CS318.5, was acutely toxic in 1990 but not in
1991. The mouth of the Du Page River, DP277.0, was not sampled in 1990.
The level of toxicity ranged from 1.1 to 14.3 times the lethal dose,
with the greatest toxicity observed in the Calumet Sag Channel (CS307.4)
in 1991. The second greatest toxicity (7 times the lethal dose) was
also observed in 1991 near the mouth of the Du Page River (DP277.0).
The following discussion of the TIE Phase I results is summarized by
year, 1990 and 1991.

4.2.1 1990. The only sample manipulation that consistently
reduced toxicity in the 1990 samples of sediment porewater was the
graduated pH test (Table 4.2). Toxicity at pH 8.5 was greater than at
pH 7.5 and pH 6.5 indicating a pH-dependent toxicant. Some ionic
compounds, e.g., cationic metals, can be more toxic at a higher pH;
however, EDTA chelation tests did not remove toxicity. Another common

Table 4.1. Toxicity of porewater to Ceriodaphnia dubia (48-hour LC50,
reported as a % of porewater in test solution, and as toxic units,
100/LC50).

]

1990

1991

Site

LC50

Toxic Units
1.5

LC50

xic Units

CR 325.4

(48-79)

1.5

CR 324.8

1.4

1.1

CS 318.5

(44.91)

. 1.5

>100

CS 307.4

(47-80)

1.6

(5-12)

14.3

SS 317.0

>100

SS 315.3

>100

SS 313.0

(25-80)

2.0

1.1

SS 310.0

>100

SS 292.2

1.4

2.8

DP 285.3

1.4

DP 281.1

>100

DP 277.0

(10-23)

7.1

IR 248.2

>100

IR 215.0

>100

IR 180.0

>100

IR 125.5

>100

IR 72.0

>100

IR 6.0

>100

....

>100

Notes:

a. >100 indicates that 100% porewater did not kill at least half the
test organisms in 48 hours.

b. Dashes ( ) indicate that toxic units could not be calculated

because porewater was not lethal within the 48-hour exposure period.

c. Numbers in parentheses indicate range of dilutions that caused 50%
mortality in 48 hours (48-hour LC50s).

NS=not sampled

Table 4.2. Results of treating sediment porewater to reduce
toxicity and characterize the toxicant. Porewater was
obtained from acutely toxic Illinois Waterway sediments in
1990.

Sample

ite

Phase I Treatments

286.3

SS 292.

313.0

CS 307.4

pH adjustment

Aeration

Filtration

Reverse-phase SPE

Oxidation reduction

EDTA chelation

Graduated pH

Sample

"si

ite

Phase I Treatments

318.5

CR 324.

326.4

pH adjustment

Aeration

Filtration

Reverse-phase SPE

Oxidation reduction

EDTA chelation

Graduated pH

NR= No reduction in toxicity
R= Reduction in toxicity
SPE= solid phase extraction

aquatic toxicant that is strongly pH-dependent is ammonia. Total ammo-
nia concentrations in the acutely toxic samples ranged from 32.7 mg/1
(milligrams per liter) to 59.8 mg/1 (Table 4.3).

4.2.2 1991. Five of the seven sites evaluated in 1991 had the same
characterization pattern as in 1990 (Table 4.4). The only manipulation
to consistently reduce toxicity was the graduated pH test, again
indicating a pH-dependent toxicant such as ammonia (Table 4.3). Total
ammonia concentrations in the 1991 samples ranged from 28.6 mg/1 to 51.2
mg/1 (Table 4.3). The characterization pattern differed for porewaters
from DP277.0 on the Des Plaines River and CS307.4 on the Calumet Sag
Channel (Table 4.4). Toxicity in these porewaters was reduced by
filtration and solid phase extraction with a Cjg column, indicating that
toxicity is due to non-polar organic compounds associated with
filterable particles. These samples contained visible quantities of
oil.

In summary, Phase I results from 1990 and 1991 indicate that acute
toxicity in most sediment porewaters from the Upper Illinois Waterway is
attributable to a pH-dependent toxicant, most likely ammonia. Porewater
from one location in the lower Des Plaines River and one location in the
lower Calumet Sag Channel contained toxicity attributable to non-polar
organics associated with oil or grease.

4.3 Toxicity Identification Evaluation - Phase II

Phase II techniques were used to isolate toxicants in porewaters
from the seven sites where ammonia was suspect and the two sites where
non-polar organics were suspect. The zeolite columns completely removed
acute toxicity from porewaters where ammonia was suspect (Table 4.5).

Table 4.3. Toxicity and ammonia concentrations in sediment
porewater in 1990 and 1991. Toxic units = 100/LC50, where
LC50 is the % dilution that kills 50% of the exposed Cerio-
daphnia dubia in 48 hours. Ammonia concentrations in the
porewater are expressed as total ammonia nitrogen, N, in
mg/1.

1990

Site

Toxic Units

CR 324.8
CR 326.4

1.4
1.5

CS 307.4
CS 318.5

1.6
1.5

SS 292.2
SS 313.0

1.4
2.0

Total Ammonia-N

37.8
25.6

35.4
42.7

41.5
59.8

DP 286.3 1.4 23.8

1991

CR 324.8 1.1 34.2

CR 326.4 1.5 51.2

SS 292.2 2.8 33.5

SS 313.0 1.1 28.6

DP 286.3 1.4 30.5

Table 4.4. Results of treating sediment porewater to reduce
toxicity and characterize the toxicant. Porewater was
obtained from acutely toxic Illinois Waterway sediments in
1991.

Sample S-

ite

Phase I Treatments

DP 277.0

DP 286.

SS 292.2

CS 307.4

pH adjustment

Aeration

Filtration

Reverse-phase SPE

Oxidation reduction

EDTA chelation

Graduated pH

Sample S-

ite

Phase I Treatments

CS 313.0

CR 324,

CR 326.4

pH adjustment

Aeration

Filtration

Reverse-phase SPE

Oxidation reduction

EDTA chelation

Graduated pH

NR= No reduction in toxicity
R= Reduction in toxicity
SPE= solid phase extraction

Table 4.5. Results of treating sediment porewater with zeolite to
remove ammonia. Porewater was obtained from acutely toxic Illinois
Waterway sediments.

1990

Pre-

Zeolite

Post

Zeolite

Site

Ammonia-N

(mg/i)

Toxicity

Ammonia-N
(mg/1)

Toxicity

CR 326.4
CR 324.8

25.62
37.82

T
T

1.70
1.46

NT
NT

CS 318.5
CS 307.4

42.70
35.38

T
T

1.98
2.24

NT
NT

SS 313.0
SS 292.2

- 59.78
41.48

T
T

2.58
1.22

NT
NT

DP 286.3

23.67

1991

4.88

CR 326.4
CR 324.8

51.29
34.16

T
T

1.86
1.70

NT
NT

SS 313.0
SS 292.2

28.60
33.55

3.05
1.95

DP 286.3

30.50

1.70

T = Acute toxicity was present, as determined by toxicity tests with

Ceriodaphnia dubia.
NT = No acute toxicity

Since the zeolite selectively removes ammonia, these results support
identification of ammonia as the toxicant.

The suspect nonpolar organics at sites CS307.4 and DP281.1 seemed
to have different chemical and physical properties because no toxicity
could be obtained from DP281.1 by column absorption and elution with
methanol alone, whereas CS307.4 did yield toxicity with the methanol
extraction (Table 4.6). Moreover, toxic materials were eluted by a wider
range of methylene chloride/methanol mixtures (20-50%) from the CS307.4
sample than from the DP277.0 sample (25-40%, Table 4.7). Also, the
greatest toxicity in porewater sample DP281.1 was associated with resi-
due left on the filters after passage of porewater, whereas the greatest
toxicity in sample CS307.4 was in supernatant left after centrifuging
out most of the particles (Table 4.6). The DP277.0 elutriate contained
no organics detectable by gc-mass spectrography, whereas 34 organic
compounds were detected in the CS307.4 elutriate (Table 4.8). This was
surprising because the DP277.0 elutriates contained toxicity (Table
4.7), but perhaps there were undetectable quantities of nonpolar organ-
ics that were highly toxic.

The elutriates from sample CS307.4 contained different combina-
tions of nonpolar organics (Table 4.8). No compounds were found above
the detection limits in the 20% fraction. The 25% fraction contained
the polycyclic aromatic hydrocarbon (PAH) naphthalene. The 30% fraction
contained primarily cyclic and branched hydrocarbons (cyclohexane,
octane) and PAHs. The 35-50% fractions contained numerous long chain
hydrocarbons such as heptadecane, undecane and dodecane. The 40 and 45%
fractions also contained the alkenes, eicosene and dotriacontanol . In
general, toxicity in these samples appears to be primarily due to
petroleum hydrocarbons and PAHs. Scubauer-Berigan and Ankley (1991)

Table 4.6. Toxicity of sediment porewater following fractionation by
filtration, centrifugation, and column absorption and extraction.
Fractions were tested for toxicity using Ceriodaphnia dubia. Toxicity
is expressed as the 48-hour LC50, reported as % of sample fraction in
test solution, and as toxic units, (100/LC50). Numbers in parentheses
indicate a range of dilutions that caused 50% mortality in 48 hours.

277,

307,

LC50

Toxic
Units

LC50

Toxic
Units

^Filter
extraction

27
(18-40)

3.7

62
(48-71)

1.6

^Centrifugation
whole sample

1.4

11.1

cpost Ci8
(25ml)'"

>100

5.6

^POSt Cio

(100 ml)

>100

5.6

^Porewater was filtered, then the filters were extracted with methylene
chloride. Extracts tested for toxicity.

'^Centrifugation at 10,000 g for 30 minutes to settle the particles in
the porewater. Supernatant tested for toxicity.

^200-ml samples of supernatants from b were passed through absorption
columns, then aliquots were taken after passage of 25 ml and 100 ml of
methanol through the columns. Aliquots tested for toxicity.

>100 indicates that the undiluted fraction did not kill at least half
the test organisms in 48 hours.

indicates that toxic units could not be calculated because the

undiluted fraction was not lethal within the 48-hour exposure period.

Table 4.7 Toxicity of extracts from sediment porewater. The porewater
was obtained from sediments at site CS 307.4 and site DP 277.0 where
nonpolar organic chemicals were suspected of contributing to toxicity.
The porewater was passed through Cjo absorption columns and then the
columns were eluted with increasingly nonpolar mixtures of methylene
chloride in methanol (1%-100% methylene chloride). The elutriates were
tested for toxicity on Ceriodaphm'a dubia.

Fraction of Site Site

Methylene Chloride DP 277.0 CS 307.4

in Methanol

100% NT NT

50% NT T

45% - NT T

40% T T

35% T T

30% T T

25% T T

20% NT T

15% NT NT

10% NT NT

5% NT NT

1% NT NT

T = Toxic
NT = Not Toxic

Table 4.8 Constituents of elutriates from solid phase Cjg absorption
columns. Toxic sediment porewaters from sites CS 307.4 and DP 277.0
were passed through the columns, which then were elutriated with mix-
tures of methylene chloride in methanol. The elutriates were analyzed
with a gas chromatograph. Values in table are concentrations, in mg/1 ,
calculated from areas under the peaks in the chromatographs.

Chemical

cyclohexadecane
cyclohexane, dimethyl
cyclohexane, trimethyl
cyclohexane, methyl propyl
cyclopentane, 1-methyl-

3-4-(l-methylethyl)
cyclopentane, methyl propenyl
decane, trimethyl
dodecane, trimethyl
1-dotriacontanol
3-eicosene
5-eicosene

heneicosan, ethylpropyl
heptacosane
heptadecane
heptadecane, trimethyl
heptadecane, tetramethyl
heptane, 3-ethyl -5-methyl
1-heptanol, 2-propyl
nonadecane

nonahexacontanoic acid
nonane, dimethyl
octane, trimethyl
octane, dimethyl
octadecane, chloro
3-octadecanol

naphthalene, decahydro-2-methyl
tetratetracontane
tritetracontane
tridecane, methyl
tritetracontane
undecane, dimethyl
undecane, 2,5-dimethyl
undecane, 3,6-dimethyl
undecane, 6-methyl

307.4

DP 277.0

Methylene

chloride/methanol %

30 35

No peaks

48
32

above

detection

limits

32
13
18

44
16

17
13

10 28

13
16
19

-- = not detected

identifed non-polar organics associated with oil and grease as a source
of toxicity in sediments in the upstream portions of the Calumet Sag
Channel and Lake Calumet.

4.4 Toxicity Identification Evaluation - Phase III

Toxicity in sediment porewaters from the upper Illinois Waterway
is correlated with total ammonia concentrations (r = 0.85, Figure 4.3).
Jones and Lee (1988) found that of more than 30 contaminants measured in
sediments from New York Harbor, only ammonia concentrations correlated to
observed toxicity in grass shrimp. Ankley et al . (1990) identified
ammonia as a major toxicant in sediments from the lower Fox River and
Green Bay, Wisconsin.

The fingernail clam, Musculium transversum, is sensitive to ammo-
nia. Anderson, Sparks and Paparo (1978) found that un-ionized ammonia
concentrations of 0.08-0.09 mg/1 inhibited the cilia on the gills of the
clams, and the growth of the clams in the laboratory was reduced at
concentrations between 0.20 and 0.34 mg/1 NH3-N. Un-ionized ammonia
concentrations greater than these are likely to occur in sediment pore-
waters of the upper Illinois Waterway, based on total ammonia concentra-
tions we measured (23.8-59.8 mg/1 NH3-N) and pH ranges known to occur in
the water column. The clams must draw oxygenated water from the water
column down their burrows to survive, and in doing so, they might shift
the pH from the low levels characteristic of anaerobic sediments to
higher levels characteristic of the water column, thereby increasing the
fraction of the total ammonia that exists in the toxic un-ionized form.

In summary, several lines of evidence lead to the conclusion that
ammonia in sediment porewater was limiting macroinvertebrate populations
in the Illinois Waterway at the time this study was conducted. First,

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the porewater contains a pH-sensitive toxicant that is not affected by
chelation with EDTA, as heavy metals would be (Tables 4.2 and 4.4).
Second, toxicity to both C. dubia and the fingernail clam M. transversum
is associated with total ammonia concentrations in sediment porewater
(Table 4.3, and Figures 4.1, 4.2, and 4.3). Third, removal of ammonia
by treatment with zeolite removes the toxicity (Table 4.5). Finally, H.
transversum is known to be sensitive to un-ionized ammonia at levels
that are likely to occur in the porewaters.

Not all the toxicity found in the upper Illinois Waterway was
associated with ammonia. Porewaters from the Calumet Sag Channel
(CS307.4) and the lower Des Plaines River (DP277.0) contained visible
signs of oil and grease and toxicity associated with PAHs (including
naphthalene) and other compounds found in petroleum.

4.5 Sensitivity of Recolonizing Clams

Much to our surprise, we found several species of fingernail
clams, including Muscuh'um transversum, at several sites in the upper
Illinois Waterway: the Chicago Sanitary and Ship Canal (SS317.0), the
North Branch of the Chicago River (CR325.4), and the Calumet Sag Channel
(CS318.5). Biologists from the Long Term Resource Monitoring (LTRM)
Station at Havana also reported finding fingernail clams in mud they
happened to bring up on their sampling nets and boat anchors.

We wondered if these clams had acquired some resistance to the
toxicants in the sediments, so we tested their responses to a sediment
sample from the Chicago Sanitary and Ship Canal (SS317.0) and another
from the Calumet Sag Channel (CS318.5). At the same time, we tested
clams from Swan Lake on the lower Illinois River, where we had obtained

all the clams used in our previous bioassays. This work was done
outside the scope of our original research proposal and is of a very
prel iminary nature.

The preliminary results support the hypothesis of differential
resistance. The clams from the Sanitary and Ship Canal showed virtually
no impairment of filtering performance in response to porewater from the
place where they where taken or to the Calumet Sag porewater (see
below). The clams from Swan Lake were sensitive (36% decline in filter-
ing performance) to porewater from the Sanitary and Ship Canal, while
they had only a slight negative response, comparable to the Cal-Sag
clams, to the Cal-Sag porewater. The Ship Canal porewater was not
tested on the Cal-Sag clams.

SOURCE OF CLAMS
CS318.5
SS317.0
Swan Lake

SOURCE OF POREWATER

SS317.0 CS318.5

-0.17

-0.08 0.00

-0.36 -0.13

5.0 DISCUSSION

Two different patterns of toxicity occur in the sediment
porewaters of the Illinois Waterway. There is a gradient of increasing
toxicity in the upstream direction, associated with increasing
concentrations of total ammonia in the sediments. The second pattern is
characterized by patches of toxicity associated with polycyclic aromatic
hydrocarbons (PAHs), such as naphthalene, and long-chain hydrocarbons,
both evidently derived from petroleum. One of the latter sites was
located on the lower Pes Plaines River section of the waterway, near
several refineries. Previous studies have measured elevated levels of
metals, pesticides, PAHs, and PCBs in the sediments of the upper Illi-
nois Waterway (lEPA 1990) and demonstrated that sediments are toxic
(Sparks et al . 1981; Blodgett et al . 1984; Schubauer-Berigan and Ankley
1991). The two toxicity problems might even be related: Ankley et al .
(1991) suggested that natural microbial processes in aquatic ecosystems
may be compromised by organic loading or selective toxicity. The alter-
ation of microbial processes could play a role in the incidence of
ammonia accumulation and subsequent toxicity in sediments in the Upper
Illinois Waterway.

It is well established that certain sediments can contain high
concentrations of ammonia (Keeney 1973, Berner 1980). Nitrogen-contain-
ing organic matter is decomposed in sediments by heterotrophic bacteria.
The amount of ammonifi cation that takes place depends in part on oxygen
availability (Kleerekoper 1953). Ammonia can accumulate to toxic levels
under anaerobic conditions (Berner 1980). Serruya (1974) found that
ammonia formation is greatest about 10 cm (centimeters) below the
sediment-water interface. In this situation, ammonia probably diffuses

from the deeper sediments to surficial sediments, and perhaps even to
the overlying water, especially if sediments are resuspended by currents
or boat- or wind-driven waves. The fingernail clam, Musculium
transversum, the organism of primary interest in this study, makes
shallow burrows in the sediment and may be exposed to much higher levels
of ammonia than organisms living in the water column, at the mud-water
interface, or on plants, rocks, and woody debris.

Ammonia toxicity is due to the un-ionized (NH3) form (USEPA 1985).
The proportion of total ammonia existing in the un-ionized form is
controlled primarily by pH and temperature (Emerson et al . 1975). The
pH of sediments can fluctuate dramatically on a seasonal basis, and the
pH of the overlying water can fluctuate daily, so that episodes of
toxicity may occur even if the total ammonia concentration remains
relatively constant. Ammonia loading of rivers tends to increase during
winter because the microorganism-mediated conversion of ammonia to
nitrate stops at cold temperatures. Also, aquatic vegetation does not
remove ammonia (a plant nutrient) during winter dormancy. Water quality
standards frequently allow higher levels of ammonia in the winter be-
cause the proportion of total ammonia existing in the toxic, un-ionized
form is less at cold temperatures. However, the sensitivity of fish to
ammonia increases at cold temperatures, so even though there may be less
un-ionized ammonia, acute toxicity may still occur (Reinbold and Pesci-
telli 1990). Research is needed to determine the effect of cold temper-
atures on the sensitivity of invertebrates, as well as fish, to ammonia.

Musculium transversum is sensitive to ammonia. Anderson, Sparks
and Paparo (1978) found that un-ionized ammonia concentrations of 0.08-
0.09 mg/1 (expressed as un-ionized ammonia nitrogen, NH3-N, in mg/1)

inhibited the cilia on the gills of the clams, and the growth of the
clams in the laboratory was reduced at concentrations between 0.20 and
0.34 mg/1 NH3-N. The C. dubia acute LC50 for ammonia is 1.04 mg/1 NH3-N
(Ankley et al . 1990). Arthur et al . (1987) reported un-ionized ammonia
toxicity to 5 invertebrates ranged from 1.95 to 18.3 mg/1 NH3-N and
mollusks (snails) were most sensitive.

iPECIES

Snail

Physa gyn'na - adult
Hell soma trivolvis - adult

LC50 (mg/1)

1.95
2.17

Amphipod
Crangonyx pseudogracih's - adult

3.12

Mayfly
Callibaetis skokianus - nymph

3.12

Isopod
Asellus racovitzai - adult

5.02

Caddisfly
Phi Tardus giaeris - larvae

10.1

Crayfish
Orconectes immunis - adult

18.3

Concentrations of this magnitude (1.0-8.0 mg/1 NH3-N) are commonly found
in the sediments in the Upper Illinois Waterway, based on total ammonia
concentrations (23.8-59.8 mg/1) and naturally occurring pHs. Ammonia
places organisms in double jeopardy because it exerts an oxygen demand
in the process of nitrification (conversion to nitrites and then ni-
trates) and low oxygen levels place organisms under additional stress
(USEPA 1985). Ammonification may be occurring in the deep, anaerobic
zones of the sediments and nitrification in the shallower, aerobic

zones, or in the boundary water at the sediment surface, so benthic
invertebrates are exposed to the worst of both worlds. They are exposed
to ammonia and to low oxygen at the same time.

The highest ammonia concentrations in sediments are associated
with nitrogen-enriched sediments or high organic loading, as from sewage
treatment plants (Brezonik 1973; Ankley et al . 1990; and Schubauer-
Berigan and Ankley 1991). Although most sewage treatment plants remove
a substantial portion of carbon in municipal waste, most do not remove
nitrogen, but convert it from ammonia into nitrate. It is possible that
nitrate is carried down into the sediments where it is converted back
into ammonia in the anaerobic zones. If this is the case, ammonia
toxicity in the sediments might be reduced by reducing the nitrogen
loading of the river.

During the course of this study, several species of fingernail
clams, including M. transversum, reappeared in the Chicago area water-
ways and in the Illinois River at Peoria and Havana. There are at least
four possible explanations for this surprising reappearance of clams in
the same general areas where the porewaters tested toxic. First, we
found that clams recolonizing the upper Illinois were more resistant to
ammonia than the clams from the lower Illinois, where the organisms were
obtained for all of the early bioassays. Second, our previous research
demonstrated that the surface layers of sediment in some areas were less
toxic than layers a few centimeters deeper (Sparks, Sandusky and Paparo
1981; Blodgett et al . 1984). Toxicity may have been overestimated in
tests where surface and deep layers of sediment were mixed prior to
testing. Third, toxic episodes may be brief and infrequent, allowing
organisms to colonize in between episodes. Fourth, the distribution of

toxicity in sediments may be extremely patchy, so that healthy organisms
are found adjacent to barren areas. If the latter two hypotheses prove
to be true, toxicity in the Illinois River has changed recently from a
widespread, chronic problem to a more localized or episodic problem.
Reduction of toxicity in surface sediments may reflect recent reductions
in ammonia loading from sewage treatment plants in the Chicago area,
although it is not clear whether the sources of ammonia in the pore-
waters are effluents, the deeper layers of sediments (as described
above) , or both.

We remind the reader that all the toxicity tests we conducted were
short-term, acute tests. The fingernail clams, MuscuTium transversum,
were exposed to sediment porewater for only 1 hour and then their fil-
tering performance was tested in clean dilution water. The water flea,
Ceriodaphm'a dubia, was exposed to porewater for just 48 hours. The
organisms in the waterways are exposed to contaminants for their entire
life spans. In the past, more sensitive tests with fingernail clams
have demonstrated toxicity even in downriver sediments, including Peoria
Lake and Quiver Lake (Sparks, Sandusky and Paparo 1981).

In addition to being a problem for the benthic invertebrates that
fish feed upon, ammonia in the Illinois Waterway may be a problem for
the fish themselves. In 1987, the U.S. Fish and Wildlife Service simu-
lated resuspension of bottom sediments by boat- or wind-driven waves by
stirring sediments in clean water, allowing the sediment to settle for
24 to 48 hours, then exposing larval fathead minnows, Piwephales prome-
las, to the water. Water mixed with surface sediments from the Chicago
River and the Des Plaines River killed all the fish within 24 hours.
Surface sediments from Lake Chautauqua, a bottomland lake and federal

wildlife refuge along the Illinois River at Havana, killed 15% of the
test fish in 96 hours; deeper sediments, taken at the 30,5-45.7 cm (12-
to 18-inch) depth, killed 25%. Fish mortality correlated (R = 0.71, P <
0.01) with the concentration of un-ionized ammonia released from the
sediment and both ammonia and fish mortality increased upstream toward
Chicago. The Long Term Research Monitoring Station (LTRM) at Havana
started measuring ammonia concentrations in Anderson Lake, a floodplain
lake of the Illinois River and a state fish and wildlife area, on 1 May
1990, 2 days after a fish kill. The total ammonia nitrogen
concentration was 0.90 mg/1 and the un-ionized ammonia nitrogen was
calculated to be 0.36 mg/1 at the temperature of 16.6° C and pH of 9.34.
NH3-N concentrations of 0.32 mg/1 at 3-5° C and 1.35 mg/1 at 24-25° C
were acutely lethal to bluegill sunfish, Lepomis macrochirus (Reinbold
and Pescitelli 1990). The fish kill might have been caused by ammonia,
if the un-ionized ammonia had peaked at higher concentrations before our
samples were taken.

Elevated un-ionized ammonia concentrations might be triggered by
resuspension of sediments or episodes of elevated pH resulting from
phytoplankton blooms. Plants remove carbon dioxide from the water, in
the form of carbonic acid and bicarbonate, and thereby elevate the pH of
the water, which in turn increases the proportion of ammonia existing in
the toxic, un-ionized form. The Havana LTRM station (unpublished data)
measured pHs as high as 10.12 in backwater lakes of the Illinois River
in July 1990 and values between 9.0 and 10.0 occur fairly often.
Episodes of acute ammonia toxicity thus may be occurring sporadically in
places other than just the upper Illinois River, and it takes only one
brief episode per year to kill or reduce populations of invertebrates or

fish that take many months or years to build up. Potential sources of
ammonia or nitrogen, besides sewage plants and anaerobic sediments,
include industrial plants (especially refineries and munitions plants),
feedlots, and agricultural fields.

Although a general recovery does seem to be beginning in the
Illinois River, with the return of fingernail clams in some areas where
they have been absent at least 30 years and appearance of largemouth
bass throughout the Illinos River proper, the pace and permanence of
recovery still appears to be threatened by ammonia, even if the problem
now turns out to be episodic instead of chronic. Reports of fingernail
clam and mussel die-offs in the Upper Mississippi River and other rivers
(Wilson et al . submitted; Blodgett and Sparks 1987; Neves 1987) indicate
that drastic population declines in macroinvertebrates that burrow in
sediments are not unique to the Illinois River.

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REPORT O0CUMENT.T,ON | .. «- ~ .„„.gj^„,

1. RcclpMiTf 1 Accasuon No

4. TItto and SuMltIa

Identification of Toxic Substances in the Upper niinois River

5. Report Oat*

nrtnher 1992

7. Apthortt) , . ^

Richard E. Sparks and Philippe E.

Ross

Parformmt Or«anliation Rapt.

I. Parforminc Orsanlxatton Nama and Addraa*

River Research Laboratory
Forbes Biological Station
Illinois Natural History Survey
P.O. Box 599
Havana, IL 62644

la Pro|act/Ta»k/Wor* Unit

89/215

II. ContracKC) or Grant(C) No

WR3691

12. Sponaofinc Organization Nama and Addraas

Illinois Department of Energy and Natural Resources

Office of Research and Planning

325 West Adams Street

Springfield. IL 62704-1892

13. Typa of Raport & Pariod Covarad

insects, snails and finger-
River. These declines had
nd fish that fed upon these
1980s, despite improvements
relevant river sediments

of fingernail clams and

ga and bacteria growth,
t two sites, the porewater

hydrocarbons, such as
luation methodology was

Water Pollution, Toxicity, Water Pollution Effects (Animals]
Sedimants, Sedimentation, Suspended Sediments

b. M«ntlfi«rs/09«n-End«d Tarmt

Sedimentation - Illinois River

Toxicity - Illinois River

Water Pollution - Illinois River

c. COSATI Flald/Group

Biological and Medical Sciences; Environmental Biology

i«. Avaiiabiirty statamant fjQ restriction on distribution.
Available at IL Depository Libraries or from
National Technical Information Service,

■^pr-i nr|f-io1H yA

19. Sacurtty Clait (This Raport)

llnclassi fipu

21. No. of Pagai

3. Sacuhty Clait (This Paga)

Unclassi f led

(Saa ANSI-

39.18)

Saa Inttructiont on Ktimn*

OPTIONAL FORM 272 (1-77)
(Formerly NTIS-35)
Dapartmant of Commarca

Printed by the Authority of the State of Illinois
Printed on Recycled Paper

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