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-M W

PARTITIONING, TOXICITY AND MUTAGENICITY OF IN-PLACE
CONTAMINANTS OF SEDIMENTS FROM THE GRAND CALUMET RIVER
AND INDIANA HARBOR, INDIANA

BY

Robert Alan Hoke

A DISSERTATION

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Department of Fisheries and Wildlife
Institute of Environmental Toxicology

1992

Copyright by:

Robert A. Hoke

1992

ABSTRACT

PARTITIONING, TOXICITY AND MUTAGENICITY OF IN-PLACE CONTAMINANTS OF
SEDIMENTS FROM THE GRAND CALUMET RIVER, INDIANA

BY

Robert A. Hoke

The presence of a wide variety of in—place chemical contaminants in
sediments from the Grand Calumet River-Indiana Harbor Canal, Indiana Area
of Concern, their chemical partitioning between environmental
compartments, and their potential toxicological effects were the primary
focus of this study. The partitioning behavior of non-polar organic
chemicals (NPOCs) present in both sediment and sediment pore waters was
investigated to determine the importance of NPOC binding to dissolved
organic carbon (DOC) in sediment pore water. Better concordance was
observed between estimated and actual field partition coefficients with a
three-phase partitioning model. These results highlight the potential
importance of the DOC-binding phenomenon in determinations of the
bioavailability and effects of NPOCs present in sediment pore waters.

The Microtox® assay, 48 h Daphnia maqng and Ceriodaphnia dubia
tests and a 10-d Chironomug tentm test were used in a toxic units
approach to assess the toxicity of sediments and sediment pore waters and
to conduct a preliminary identification of the potential toxicants. Based
on the results of these analyses, ammonia, polycyclic aromatic
hydrocarbons, metals, petroleum hydrocarbons and bicarbonate ion were the

major contaminants of concern to benthic invertebrates within the study

area. Separate experiments to determine the mechanism of bicarbonate ion
toxicity to D. magna suggested that toxicity was due to the inhibition of
the active uptake of Cl" from water. Therefore, pore water alkalinity
should be considered when interpreting the results of aqueous phase
toxicity tests with cladocerans and, perhaps, other species of
invertebrates and fish.

Evaluation of the comparative mutagenicity of solvent extracts of
sediments from the study area was conducted with the Ames and MutatoxQ
assays. Extracts were mutagenic with metabolic activation in both assays,
however, few samples contained direct acting mutagens. The lack of
mutagenicity in Mutatox® assays of pore waters indicates that short-term
human exposure to mutagens in pore waters and sediments is likely to be
non-problematicu Greater concern is required for the potential ecological
and human health effects due to food chain transfer of mutagenic compounds

in sediments from the study area.

This dissertation is dedicated to the memory of my father and to my
son, Ian, whose insatiable curiosity about the world around him is a

constant source of wonder.

ACKNOWLEDGEMENTS

I would like to thank the members of my dissertation committee for
sharing their knowledge and insight, as well as for being patient with my
meanderings. Thanks are also due to my fellow graduate students in the
Aquatic Toxicology Program at Michigan State University. Don Tillitt was
responsible for helping me maintain sight of the goal while still "having
a life" and.Will Gala was responsible for providing unusual insight at the
most incongruous times, as exhibited by his role in the development of
"weekend science". I am jparticularly grateful for having had the
opportunity to study and work with Dr. John Giesy. Both in and out of the
laboratory, John provides a multi-faceted learning environment which
encourages students to explore new ideas and develop their full potential.
I wish to thank him for providing me with the opportunity to expand my
horizons, both personally and professionally. I have benefited both from
his professional knowledge and advice and from his friendship.

My son, Ian, has been a constant source of inspiration and curiosity
about the world at large, as well as an occasional unpaid lab assistant.
Finally, I would like to thank Kathee for her patience with me and the
vagaries of graduate school. Her constant love, support and understanding

made things easier along the way.

ii

TABLE OF CONTENTS

LIST OF TABLES O O O O O O O O O O O O O O O O O O O O O C O O 0
LIST OF FIGURES . . . . . . . . . . . . . . . . . . . . . . . .

GENERAL INTRODUCTION 0 O O O O 0 O O O O O O O O O O O O O O O 0
Literature Cited . . . . . . . . . . . . . . . . . . . . .

CHAPTER 1

A Field Evaluation of Equilibrium Partitioning in Sediments
from the Grand Calumet River and Indiana Harbor, Indiana .
Introduction . . . . . . . . . . . . . . . . . . . .
Material and Methods . . . . . . . . . . . . . . . .
Chemical Analysis . . . . . . . . . . . . . .

Equilibrium Partitioning Calculations . . . . .

Results . . . . . . . . . . . . . . . . . . . . . .
Discussion . . . . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . .
Literature Cited . . . . . . . . . . . . . . . . . .

CHAPTER 2

Toxicity of Sediments and Sediment Pore Waters from the

Grand Calumet River and Indiana Harbor, Indiana . . . . .
Introduction . . . . . . . . . . . . . . . . . . . .
Materials and Methods . . . . . . . . . . . . . . .

Sample Collection . . . . . . . . . . . . . .
Pore Water Extraction . . . . . . . . .
Chemical Analysis . . . . . . . . . . . . . .

Organics . . . . . . . . . . . . . . . .
Metals . . . . . . . . . . . . . . . . .
Miscellaneous Parameters . . . . . . . .
Toxicity Tests . . . . . . . . . . . . . . . .
Photobacterium phosphorgum . . . . . . .
Daphnia magna and Ceriodaphnia dubia . .
Chironomus tentans . . . . . . . . . . .
Statistical Analysis . . . . . . . . . . . . .
Results . . . . . . . . . . . . . . . . . . . . . .
Chemical Analysis . . . . . . . . . . . . . .
Organics . . . . . . . . . . . . . . . .
Metals . . . . . . . . . . . . . . . . .
Miscellaneous Parameters . . . . . . . .

iii

vi

ix

11
12
16
16
19
21
45
50
51

55
S6
61
63
63
63
63
65
65
66
66
66
68
69
69
69
69
83
93

TABLE OF CONTENTS - continued

Toxicity Tests . . . . . . . . . . . . . . . . . . 93

Photobacterium phosphoreum . . . . . . . . . 93

Daphnia magng and Ceriodaphnia dubia . . . 99
Chironomus tentans . . . . . . . . . . . . 103

Discussion . . . . . . . . . . . . . . . . . . . . . . 104
Acknowledgements . . . . . . . . . . . . . . . . . . . 113
Literature Cited . . . . . . . . . . . . . . . . . . . 114

CHAPTER 3

Bicarbonate as a Potential Confounding Factor in Cladoceran
Toxicity Assessments of Pore Waters from Contaminated

Sediments . . . . . . . . . . . . . . . . . . . . . . . . . 127

Introduction . . . . . . . . . . . . . . . . . . . . . 128

Materials and Methods . . . . . . . . . . . . . . . . 133

Toxicity of Na+ and HCO3' . . . . . . . . . . . 133

Calculation of Free C02 and MCO3' . . . . . . . 134

X-ray Dispersive Microanalysis . . . . . . . . . 135

Data Analysis . . . . . . . . . . . . . . . . . 138

Results . . . . . . . . . . . . . . . . . . . . . . . 138
Toxicity of Na+ and HCO3' . . . . . . . . . . . 138
Calculation of Free C02 and HCO3' . . . . . . . 139
X-ray Dispersive Microanalysis . . . . . . . . . 139

Discussion . . . . . . . . . . . . . . . . . . . . . . 143

Acknowledgements . . . . . . . . . . . . . . . . . . . 150
Literature Cited . . . . . . . . . . . . . . . . . . . 151

CHAPTER 4

Mutagenicity and 2,3,7,8-Tetrachlorodibenzo-p-dioxin

Equivalents in Organic Solvent Extracts of Sediments from

the Grand Calumet River, Indiana . . . . . . . . . . . . . . 158
Introduction . . . . . . . . . . . . . . . . . . . . . 159

Materials and Methods . . . . . . . . . . . . . . . . 163
Sample Collection . . . . . . . . . . . . . . . 163
Pore Water Extraction . . . . . . . . . . . . . 165

Mutatox Assay . . . . . . . . . . . . . . . . . 165
Ames Assay . . . . . . . . . . . . . . . . . . . 167
H4IIE Assay . . . . . . . . . . . . . . . . . . 168
Chemical Analysis . . . . . . . . . . . . . . . 169
Statistical Analysis . . . . . . . . . . . . . . 170
Results . . . . . . . . . . . . . . . . . . . . . . . 170

iv

TABLE OF CONTENTS - continued

Mutatox® . . . . . . . . . . . . . . . . . . . . 17o
Ames Assay . . . . . . . . . . . . . . . . . . . 171
H4IIE Assay . . . . . . . . . . . . . . . . . . 174
Discussion . . . . . . . . . . . . . . . . . . . . . . 174
Acknowledgements . . . . . . . . . . . . . . . . . . . 182
Literature Cited . . . . . . . . . . . . . . . . . . . 183

SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191

Table

Table

Table

Table

Table

Table

1.

2.

LIST OF TABLES

Non-polar organic compounds

waters.

water partition

AQUIRE database were used to calculate estimated
organic carbon partition coefficients (Koc) with

analyzed for and
present in Grand Calumet River sediments and pore
Measured (M) or estimated (E) octanol-
from the

coefficients

the following equation:
0.983 10910 Kow (27)

Measured
concentrations

uQ/L-

non-polar

in
sediment (BS), mg/kg, and pore waters ( PW, Cp),
"dissolved"

Calculated

loglo Koc = 0.00028 4-
organic chemical (NPOC)
Grand Calumet River bulk

(K

pore water NPOC
concentrations (PW, Cd) are reported as ug/L and

ow)

sediment TOC and pore water DOC are reported as %
respectively

dry wt. and mg/L

Loglo apparent field partition coefficients,
(Csoc/Cp); actual field partition coefficie
theoretical

KP! (Cece/Cd)

sediments and pore waters

I

and

EB;

partition
coefficients Kt (foc x Kow) for non-polar organic
chemicals (NPOCs) measured in Grand Calumet River

Organic compounds analyzed for but not detected
in sediments or sediment pore waters from the
The limit of detection
is given for the matrix in which the compound was

Grand Calumet River,

not detected .

IN.

0

0

Concentrations of organic chemicals in bulk or

whole sediments from the Grand Calumet River,

IN.

Results are reported as mg/kg dry wt., except for
(% dry wt.) and

organic carbon,

oil and grease

2,3,7,8-dibenzo—p-dioxin (pg/kg)

Concentrations of
waters of sediments
River, IN.

organic

chemicals
from the Grand Calumet

in pore

Results are reported as pg/L, except

for total inorganic and organic carbon, which are

reported as mg/L

vi

23

25

31

70

73

79

LIST OF TABLES - continued

Table 7. Concentrations of metals and acid volatile
sulphide (AVS) in sediments from the Grand
Calumet River, IN. Results are reported as gm/kg
dry wt., except AVS, which is reported as pM S/gm
dry wt . . . . . . . . . . . . . . . . . . . . . . . . . 84

Table 8. Total concentrations of metals (mg/L) in pore
waters of sediments from the Grand Calumet River,

IN 0 O O O O O O O O O O C O O O O O O O O O O O O O O O 87

Table 9. Concentrations of miscellaneous chemical
compounds in pore waters of sediments from the
Grand calumet River, IN 0 O O O O O O O I O O O O O O O 91

Table 10. Results of toxicity tests conducted with bulk
sediments or pore waters of sediments from the
Grand Calumet River, IN. All data are expressed
as either % pore water, or % response (e.g. Q;
tentans inhibition of dry wt gain relative to
control) . . . . . . . . . . . . . . . . . . . . . . . . 94

Table 11. Calculated and neasured pore water toxic units
(TU) for selected parameters based on pore water
chemical concentrations and 15-min ECSO values
from Microtox'ID tests of pore waters and pure
chemicals. Microtoxo tests were osmotically
adjusted with NaCl . . . . . . . . . . . . . . . . . . . 98

Table 12. Calculated and measured pore water toxic units
(TU) for selected parameters based on pore water
chemical concentrations and 48 h L050 values from
D. magna and C. dubia acute toxicity tests of
pore waters and pure chemicals . . . . . . . . . . . . 100

Table 13. Results of routine chemical analyses of sediment
pore waters from the Grand Calumet River-Indiana
Harbor Canal IJC AOC tested in 48-h acute assays
with Q; magna and g; dubia. Maximum measured pH,
conductivity, hardness and alkalinity are
reported from the 100% pore water exposure
treatment for each location, as well as
calculated free 002 and HCO3' concentrations in

vii

Table

Table

Table

Table

Table

Table

14.

15.

16.

17.

18.

19.

pore water based on measured pH,
temperature (25°C) from the cladoceran assays

Toxicity to Q; magna and g; dubia of Na+ as NaCl
or NaHCO3 and of HCO3- as NaHCO3

Results of X—ray microanalysis experiments with
Values reported are mean
each

2;

= 0.05)

Variance
analysis

data

Results of Ames and Mutatoxo assays of organic
solvent extracts of sediments from the Grand
Calumet River and Indiana Harbor,
without

were

Pearson product moment correlation coefficients
from analyses of Ames assays with S9 activation
and organic chemical analyses of sediment from
Correlations

tested

ma na.
ratios

for

component
of

with
activation in both assays

variance

and

experimental
Significant treatment effects

the Grand Calumet River,

reported were statistically significant at p s

0.05

TCDD-EQ from the H4IIE assay in comparison to
TCDD-EQ based on measured concentrations of total
PCBs (as Aroclor 1248) and TCDD in bulk sediments
from the Grand Calumet River and Indiana Harbor,

IN

C 0

viii

analyses
of

O

IN.

of

LIST OF TABLES - continued

IN.
S9

results
X-ray microanalysis

alkalinity,

(SD)
treatment.
(indicated with
asterisk) for elemental P/B ratios were based on
a significant ANOVA for the individual element (a
followed by Bonferroni
indicated a significant difference from either
the control or lowest experimental treatment

and

P/B

t-test which
from

Extracts
metabolic

130

136

140

142

172

175

176

Figure 1.

Figure 2.

Figure 3.

Figure 4.

Figure 5.

LIST OF FIGURES

Sampling locations in the Grand Calumet River and
Indiana Harbor, Indiana . . . . . . . . . . .

Loglo Kt, (f0C x Row), versus loglo apparent
field . (CSochC/C ), for all non- polar organic
compoun s analyzed for and present in Grand
Calumet River sediments and pore waters. The
relationship between Kt and KP. data is plotted
(dashed line) and shown in the equation while the
solid line represents a theoretical relationship
with r = 1.0 . . . . . . . . . . . . . . . . .

Loglo Kt (foc x K ow)' versus loglo apparent field
Kp. (Csoc/Cplh for all non- polar organic
compounds Wit loglox ow values <3. 0 analyzed for
and present in Grand Calumet River sediments and
pore waters. The relationship between Kt and
data is plotted (dashed line) and shown in t e
equation while the solid line represents a
theoretical relationship with r = 1.0 . . . .

Loglo Kt (foc x Kow), versus loglo apparent field
. (Csoc/C) for all non-polar organic
compounds wit tfi.loglox Kow values >3. 0 analyzed for
and present in Grand Calumet River sediments and
pore waters. The relationship between Kt and
data is plotted (dashed line) and shown in t e
equation while the solid line represents a
theoretical relationship with r = 1.0 . . . .

Loglo Kt, (foc x Kow), versus loglo apparent
field KP. (Cece/Op)! open squares) and actual
(Cece/Cd)' closed squares) for all non-polar
organic compounds ‘with loglo Kow *values >3.0
analyzed for and present in Grand Calumet River
sediments and pore waters. The equation in the
upper left corner presents the relationship
between Kt and K for these compounds while the
equation in the Tower right corner presents the
relationship between Kt and Kp. . . . . . . .

ix

0

17

39

4O

42

Figure

Figure

Figure

Figure

Figure

7.

10.

LIST OF FIGURES - continued

Log10 KtI (fo

(Cage/Cd)!

Comparative

triangles),

C

apparent field K .
, P

(closed Circles),

Socha

(CSOC

and

squares) and Oliver (29,

(foc

x Kow),
for all non-polar organic compoun s
analyzed for and present in Grand Calumet River
sediments and pore waters

versus

X

Kow).
/Cp) data from this study

Kadeg and Pavlou open

Carpenter
open diamonds)

10910

I

versus

(23,

open

Sampling locations in the Grand Calumet River and

Indiana Harbor,

Comparative
osmotic

. e
Microtox

Indiana

effects of
adjustment on
assays of study site pore waters

NaCl
lS-m ECSO

versus
values

sucrose

for

Sampling locations in the Grand Calumet River and

Indiana Harbor,

Indiana

0

43

44

62

96

164

GENERAL INTRODUCTION

Many rivers, harbors and connecting channels of the Great Lakes have
sediments which contain a wide array of chemical compounds, including
organic xenobiotics (Pranckevicius 1986, Fallon and Horvath 1985) and
metals (Pranckevicius 1986, Fallon and Horvath 1985, Hamdy and Post 1985).
Due to municipal, industrial and non-point source waste discharges and the
tendency of many chemicals to become associated with sediments (Knight
1984, Cairns et al. 1984), contaminated sediments are particularly
problematic near densely populated, industrialized urban areas, such as
the area surrounding the Grand Calumet River-Indiana Harbor (GCR-IH)
International Joint Commission (IJC) Area of Concern (AOC) in northwest
Indiana. Rodgers et a1. (1985) have summarized the recent environmental
status, relative to anthropogenically-introduced chemical contamination,
in these area of the Great Lakes.

The drainage basin of the Grand Calumet River comprises 43,000 acres
within the Calumet Lake plain in northwest Indiana which have been greatly
affected by development activities. The Calumet Lake plain occupies an
area of low relief which historically was formed by the bottom of Lake
Chicago and three subsequent lakes prior to the formation of Lake Michigan
(IDEM 1988). High linear coastal sand dunes and many low beach ridges
characterized the Grand Calumet River basin prior to the beginning of
industrial development in the late 1800's. The wetland areas between the

beach ridges has subsequently been filled with sand from the ridges and

2
dunes and slag from the area's many steel mills (Reshkin et a1. 1975,
Hartke et a1. 1975).

The hydrology of the Grand Calumet River basin is complex as a
result of modifications which were initially begun by Indians living in
the area (i.e. a channel to connect the Grand Calumet River with the river
Draining Lake Calumet). Additional alterations in the hydrology of the
basin were made by various local and federal agencies to facilitate
industrial development and protect the water intake of Chicago (and
southern Lake Michigan) from contaminants in the many municipal and
industrial effluent discharges to the river. Surface water in the system
may flow east toward the Marquette Park lagoons, west to the Mississippi
River or north to Lake Michigan depending on the local weather conditions
and municipal/industrial effluent discharge rates and volumes. Flow in
the east branch of the Grand Calumet River may be comprised of up to 93 %
municipal and industrial effluent whilezmunicipal.waste treatment effluent
may account for 100 % of the flow in the west branch (U.S. EPA 1985,
Crawford and Wangsness 1987). The direction of flow in the west branch
also has been observed to change from 34 cfs to the east to 42 cfs to the
west over a 24 hour period (IDEM 1988).

The severely contaminated condition of the GCR-IH ADC is based on
water quality problems with conventional pollutants, metals, and organic
chemicals, contaminated sediments, impacted aesthetics and biota, and fish
consumption advisories (IDEM 1988). The ADC covers the entire length of
the east branch and the first two miles of the west branch of the Grand
Calumet River, the Indiana Harbor Canal, Indiana Harbor and the Lake
Michigan nearshore zone of the harbor.

The presence of a wide variety of in-place chemical contaminants in

3

sediments from the ADC and their potential effects on benthic
macroinvertebrate populations were the primary focus of this study. A
wasteload allocation study of the Grand Calumet River conducted by
HydroQual, Inc (1984) identified the presence of numerous priority
pollutants in sediments from 10 sampling locations within the GCR-IH
system. Concentrations of 13 metals (As, Cd, Cr, Cu, Pb, Hg, Ni, Ag, Se,
Zn, Sb, Be, T1) and cyanide were analyzed with increased concentrations of
at least one metal (generally more than one) or cyanide observed in each
of the 10 samples. Concentrations of 21 pesticides were generally below
0.05 pg/g and were always below 5.0 pg/g. Various PCB Aroclors (1016,
1221, 1232, 1242, 1260) also were present at concentrations below 1.0 pg/g
although Aroclors 1248 and 1254 were present at concentrations as high as
17.0 and 6.9 pg/g, respectively. Forty-seven base neutral priority
pollutants also were analyzed and concentrations of 24 of these generally
observed to be below 0.025 yg/g. Nine of the remaining base neutral
compounds exhibited concentrations > 100 pg/g in sediment from at least
one location" These nine compounds were polynuclear aromatic hydrocarbons
(PAHs) with maximum observed concentrations in sediments collected from
locations adjacent to steel mill effluent discharges. High concentrations
of other base neutral compounds also were observed near a municipal
wastewater treatment plant effluent discharge. Comparison of the results
of the sediment chemical analyses conducted by HydroQual, Inc (1984) with
results from earlier studies demonstrated little change in concentrations
of chemicals in sediments from the Grand Calumet River (HydroQual, Inc.
1984).

Previous assessments of the effects of chemicals in the water column

and sediments of the GCR-IH AOC on indigenous fauna have relied heavily on

4

surveys of benthic macroinvertebrate and fish community structure (U.S.
EPA 1985, Polls and Dennison 1984, U.S. ACOE 1985, 1986). Effects on
plankton (Cook 1966, Gannon and Beeton 1969) and other wildlife, including
birds and mammals (U.S. ACOE 1985, 1986), also have been conducted on a
limited basis. Elevated contaminant levels in sediment and water have
been implicated in the lack of aquatic and terrestrial organisms in the
GCR-IH system (IDEM 1988). Sediments from Indiana Harbor and Canal have
been demonstrated to be toxic to benthic macroinvertebrates during
laboratory sediment toxicity tests (Gannon and Beeton 1969, U.S. ACOE
1987). Concern also has been raised over the potential for
bioaccumulation of metals and organic chemicals by terrestrial species if
harbor sediments were dredged and moved to an upland disposal site (U.S.
ACOE 1987). To date, however, no such comprehensive sediment toxicity
evaluations have been conducted on sediments collected from the east and
west branches of the Grand Calumet River.

Any meaningful assessment of sediment contamination and toxicity in
the GCR-IH system must include an evaluation of the contaminant-associated
toxicity of sediments to aquatic life including benthic
macroinvertebrates. The data needed (and which is currently lacking) for
such an assessment includes synoptic measures of chemical concentrations
in sediments and interstitial (pore) waters, physical characteristics of
the sediments (particle size, total organic carbon, etc) and measures of
sediment and interstitial water toxicity to relevant benthic test species
(U.S. EPA 1985).

Currently, great interest also exists among regulators, legislators
and the scientific community in the development of sediment quality

criteria (Shea 1989). Numerous approaches have been proposed for the

development of criteria (Anonymous 1985), including' the: use of the
equilibrium partitioning theory (Pavlou 1987). Empirical observations
from various studies (Adams et al. 1985, Ziegenfuss et al. 1986, Adams
1987, Connell et al. 1988, Lake at al. 1990 Swartz et al. 1990) support
the hypothesis that neutral organic chemical bioavailability, toxicity and
bioaccumulation are more closely related to pore water chemical
concentrations of contaminants than to total (bulk) sediment
concentrations. These observations engendered and continue to fuel the
interest in the equilibrium partitioning theory as a mechanism for
predicting potential exposure to, or bioaccumulation of, contaminants by
benthic biota.

The information presented above led to the selection of the GCR-IH
AOC as a study area for the evaluation of the usefulness of several
alternative invertebrate and microbial assays for the comparative
assessment of sediment and/or sediment pore water toxicity and
mutagenicity. The specific objective of chapter one was to conduct a
field evaluation of the equilibrium partitioning theory and the
ramifications of non-polar organic compound binding to dissolved organic
carbon relative to the development of sediment quality criteria using the
equilibrium partitioning theory. Chapter two presents comparative
toxicity data from four toxicity tests commonly used to assess sediments
or sediment pore waters. It also tests the hypothesis that the toxicity
of sediment pore waters can be predicted based on toxic units calculated
from laboratory-derived, chemical-specific dose response relationships.
Chapter three examines the toxicity of bicarbonate ion to the cladocerans
Daphnia magna and Cerioda hnia dubia, evaluates the potential for

bicarbonate toxicity in sediment pore waters and presents a potential

6
mechanism for the observed effects produced by the bicarbonate ion. The
final chapter presents a comparative evaluation of several different
assays for determining effects, other than acute toxicity, which may be

linked to exposure to contaminated sediments.

Literature Cited

Adams, W.J., R.A. Kimerle and. R.G. Mosher. 1985. .Aquatic safety
assessment of chemicals sorbed to sediments. In, my;
Toxicolo and Hazard Assessment, ASTM STP 854, R.D. Cardwell,

R. Purdy and R.C. Bahner, Eds. American Society for Testing
and Materials, Philadelphia, PA. pp. 429-453.

Cairns, M.A., A.V. Nebeker, J. H. Gakstatter and W. Griffis. 1984.
Toxicity of copper-spiked sediments to freshwater
invertebrates. Environ. Toxicol. Chem. 3:435-446.

Connell, D.W., M. Bowman and D.W. Hawker. 1988. Bioconcentration of
chlorinated hydrocarbons from sediment by oligochaetes.
Ecotoxicol. Environ. Safety 16(3):293-302.

Cook, G. 1966. Report on the Calumet area surveillance program for June-
November-l969, Illinois-Indiana. In, Conference in the gatte;

of Pollution of the Interstate Waters of the Grend Calumet

 

Riveg, Little Calumet Riveg. Wolf Lakel Lake Miehigen eng
Thei; Tributaries. Federal Water Pollution Control

Administration, Chicago, IL. pp. 22—105.

Crawford, G.C. and D.J. Wangsness. 1987. Streamflow and Water Quality of

the Grand Calumet RiverJ Leke County, Indiana and Cook County.

Illinoie. October 1984. Water—Resources Investigations Report

 

86-4208, U.S. Geological Survey, Indianapolis, IN. 137 p.

Fallon, M.E. and F.J. Horvath. 1985. Preliminary assessment of
contaminants in soft sediments of the Detroit River. J. Great
Lakes Res. 11:373-387.

Gannon, J.E. and A.M. Beeton. 1969. Stud'es on the f c s

a ls from Selected Great Lakes Harbors on Plan 0 n

Benthos. Center for Great Lakes Studies, University of
Wisconsin-Milwaukee. 82 p.

Hamdy, Y. and L. Post. 1985. Distribution of mercury, trace organics and
other heavy metals in Detroit River sediments. J. Great Lakes
Res. 11:353-365.

Hartke, E.J., J.R. Hill and M. Reshkin. 1975. Environmengel Geelogy of

Lake and Porter Counties. Indiana - en Aid to Planning.

 

Indiana Department of Natural Resources, Geological Survey
Special Report II. Bloomington, IN. 57 p.

HydroQual, Inc. 1984. Grand Calumet River Wasteload Allece§iog §§ggy.
Mahwah, NJ. 188 p.

Indiana Department of Environmental Management. 1988. Northwest Indiana
Environmental Action Plen. Draft Area of Concern Remedial
Action Plan. IDEM, Indianapolis, IN. 183 p.

Knight, A.W. 1984. The Evaluation of Contaminated Sediment o
§elected Benthic Freshwate; Invertebrates. Final.Report, U.S.
Environmental Protection Agency Cooperative Agreement No. CR-
808424. U.S. EPA, Environmental Research Laboratory,
Corvallis, OR.

Lake, J.L., N.I. Rubinstein, H. Lee II, C.A. Lake, J. Heltshe and S.
Pavignano. 1990. Equilibrium partitioning and

bioaccumulation of sediment-associated contaminants by

9
infaunal organisms. Environ. Toxicol. Chem. 9:1095-1106.
Pavlou, S.P. 1987. The use of the equilibrium partitioning approach in

determining safe levels of contaminants in marine sediments.

In, Fate and Effects of Sediment-Bound Chemicals in Ageatic

 

§ysteme, K.L. Dickson, A.W. Maki and W.A. Brungs, Eds.
Pergamon Press, Elmsford, NY. pp. 388-412.

Polls, I and S. Dennison. 1984. Biolo ical and Chemical Wate a1 t
Seggev in Indiana Harbor Canal and Southwestern Lake Michigan
for the U.S. Army Corps of Engineers. Chicago District.
Metropolitan Sanitary District of Greater Chicago, Chicago,
IL. 88 p.

Pranckevicius, P.E. 1986. 1982 Detroit Michigan Area Survey. EPA Report
No. 905-4-86-002, U.S. Environmental Protection .Agency,
Chicago, IL.

Rodgers, P.W., M.S. Kieser and G.W. Peterson. 1985. Su e

 

Exieting Status of theegpper Great Lakes Connecting cgaggele
QEEQ- Limno-Tech, Inc., Ann Arbor, MI.

Shea, D. 1988. Developing national sediment quality criteria. Environ.
Sci. Tech. 22:1256-1261.

Swartz, R.C., D.W. Schults, T.H. DeWitt, G.R. Ditsworth and J.O.
Lamberson. 1990. Toxicity of fluoranthene in sediment to
marine amphipods: a test of the equilibrium partitioning
approach to sediment quality criteria. Environ. Sci. Tech.
9:1071-1080.

U.S. Army Corps of Engineers. 1985. Great Lekes Connecting Channele end
fleebors. Draft Final Feeeibilitv Report end Environmental

‘—

Impact Statement. U.S. ACOE, Detroit, MI. 228 p.

10
Army Corps of Engineers. 1986. Indiana Harbor Confined Dieegeel

Eecilitv end. Maintenance Dredging. Lake County. Indiana.

 

Qraft Environmental Impact Statement. U.S. ACOE, Chicago, IL.
78 p.

Army Corps of Engineers. 1987. Disposel Alternative for PCB-

 

Qontaminated Sediments from Indiana Harbor, Indiana. U.S.

ACOE, Waterways Experiment Station, Vicksburg, MS. 211 p.
Environmental Protection Agency. 1985a. Master Plan for Impgevieg
WeeereQuality in the Grand Calumet RiverjIndiana Harbor Canal.

U.S. EPA, Chicago, IL. 149 p.
Environmental Protection Agency. 1985b. Sediment Quality Critegia
Qeyelopment Workshop. U.S. EPA, Office of Water Regulations

and Standards, Criteria and Standards Division, Washington,

DOC.

Ziegenfuss, P.S., W.J. Renoudette and W.J. Adams. 1986. Methodology for

assessing the acute toxicity of chemicals sorbed to sediments:
testing the equilibrium partitioning theory. In, Ageatic

Toxicology end Environmengel Fate, ASTM STP 921. American

 

Society for Testing and Materials, Philadelphia, PA” pp. 479-

493.

CHAPTER 1

A Field Evaluation of Equilibrium Partitioning in Sediments

from the Grand Calumet River and Indiana Harbor, Indiana

11

Introduction

Currently, great interest exists among regulators, legislators and
the scientific community in the development of sediment quality criteria
(1). Numerous approaches have been proposed for the development of these
criteria (2), including the use of the equilibrium partitioning theory
(EqP) (3,4). Empirical observations from various studies (4-10) support
the hypothesis that non-polar organic chemical (NPOC) bioavailability,
toxicity and bioaccumulation are more closely related to pore water
(interstitial water) or organic carbon-normalized bulk sediment chemical
concentrations of contaminants than to total (bulk) sediment
concentrations. These observations engendered and continue to fuel the
interest in the EqP as a mechanism for predicting potential exposure to,
or bioaccumulation of, NPOCs by benthic biota.

The equilibrium partitioning theory predicts contaminant
bioavailability in pore water by assuming that a thermodynamic equilibrium
is established between the solid phase sediment and pore water
concentrations of NPOCs (4). According to the theory, pore water
concentrations and tissue residues of these compounds should be
predictable based on the bulk sediment concentration of the compound and
the physical and chemical properties of the sediment and compound of
interest (3,4,11).

For NPOCs, the equilibrium concentration in pore water is a result

of partitioning between the solid and dissolved phases. This partitioning

12

13
process is primarily controlled by the concentration of the NPOC of
interest in the bulk sediment, the fractional organic carbon content of
the sediment and the proportionality constant for distribution of the NPOC
between water and organic matrices. This constant is approximated by the

octanol-water partition coefficient (K for the compound of interest

ow)
(4,11). It should theoretically be possible to estimate the pore water

concentration of a neutral organic chemical based on ‘the following

relationship between the solid (bulk) phase and pore water concentrations,

C8 and Cp, respectively (Equation 1),
Cs 2 foc Kow Cp (1)
where foc = fractional organic carbon content of the sediment and Kow =

the octanol-water partition coefficient for the chemical of concern.

If C is defined as the solid phase concentration of the NPOC of

soc
interest normalized to the fractional organic carbon content of the

sediment (Equation 2),

CSOC z foc (2)

then the concentration of NPOC in the pore water can be estimated from the
concentration in the bulk sediment normalized to the fractional organic

carbon content of the sediment and the Kow for the NPOC (Equation 3).

p 0.. (3)

However, determination of chemical concentrations in pore'water, and
thus quantification of potential chemical exposure via one of the major
exposure routes for benthic organisms in contaminated sediments (5,12,13),
may not be this simple. Numerous authors have proposed that a third
phase, dissolved organic carbon (DOC), plays an important role in the
partitioning behavior of NPOCs with a 10910 Kow > 3.0 (14-20). Because
DOC complexation is related to NPOC hydrophobicity, complexation will
increase with an increase in the Kow' and thus Koc and KDOC' of the NPOC.
This phenomenon will result in a greater total concentration of NPOC in
the pore water than would be estimated from equation 3.

An in-depth discussion of the ramifications of DOC-binding of NPOCs
in sediment pore water is presented by DiToro et a1. (4). The essence of
the problem, however, involves the distribution of measured NPOC
concentrations in pore water, i.e., what proportion of the total NPOC
concentration is "dissolved" or "free" and what proportion is bound to
DOC. This distinction is important because NPOC complexed to DOC may be
partially or totally unavailable to cause effects on biota (21-22). Thus
"dissolved" NPOC concentrations in pore water may most accurately reflect
the true chemical exposure for benthic biota.

The relative concentrations of free and bound, soluble NPOC in pore
water should be related to the organic carbon-normalized sediment

concentration (Csoc) (Equations 4-6):

ow p (4)

15

and, Cd = C (6)

 

1 + mDoc KDoc

if, Cd = "dissolved" NPOC concentration in pore water
CDOC = DOC-bound NPOC concentration in pore water
mDOC = DOC concentration in pore water

KDOC = DOC partition coefficient for NPOC

Knowledge of these values would permit an initial examination of the
potential importance of DOC-binding of NPOCs in the field and provide a
better understanding of the potential problems in the application of EqP
for the development of sediment quality criteria.

If EqP can be verified, sediment quality criteria could be developed
by applying water quality criteria to chemical concentrations in pore
water and back-calculating permissible sediment concentrations based on
sediment organic carbon content. This method of developing sediment
quality criteria is frequently referred to as the equilibrium partitioning
theory/water quality criteria (EqP/WQC) approach. To date, attempts to
verify EqP have been restricted to laboratory evaluations of the behavior
of a few NPOCs in sediments containing a limited range of organic carbon
concentrations (8,10) or field evaluations with an equally restrictive
number of compounds and sampling locations (9,11,23).

The objective of this study was to measure bulk sediment and pore

water concentrations of a suite of NPOCs in samples collected from 10

16
locations on the Grand Calumet River near Gary, IN and to evaluate the
potential ramifications of DOC-binding of NPOCs on the EqP by examining
two general types of field data: 1) apparent versus estimated theoretical
partition coefficients for NPOCs as a direct test of the Csoc = Kow CP
relationship, independent of DOC and 2) actual versus estimated
theoretical partition coefficients for NPOCs when DOC-binding is accounted

for by using the relationship Csoc = Kow (Cd + CDOC”

Materials and Methods

Chemical Analysis

Sediment samples were collected with a Ponar grab sampler from 10
locations along the Grand Calumet River in the northwestern corner of
Indiana (Figure 1) between 1 September 1988 and 1989. The sample from
each location was a composite of approximately 80-100 L of wet sediment
from multiple Ponar grabs. Each sample was placed in coolers or plastic
buckets lined with food-grade plastic bags, immediately transported to the
laboratory and placed in 4°C storage.

Sediment samples were dried and percent moisture determined for each
sample. Ten grams of dried sample were mixed with 10 g of anhydrous
sodium sulphate and Soxhlet extracted for 24 h with pesticide-grade
acetone/hexane (1:1, v/v). The extract was passed over a drying column
containing anhydrous sodium sulphate and the column rinsed with
approximately 100 ml of the acetone/hexane 1mixture to complete ‘the
transfer. The extract was transferred to a Kuderna-Danish concentrator
and extract volume reduced to 1 ml. The final extract volume was adjusted

to 10 ml with the acetone/hexane mixture. These procedures and all other

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aspects of sediment sample preparation followed U.S. EPA,Method 3540 (24).

Pore waters were prepared by centrifuging 275-325 g of wet sediment
in acid-rinsed 250 m1 polycarbonate centrifuge tubes for 45 min at 6800
xg. Supernatants were filtered though a Whatman glass fiber filter (GF-F,
0.7 p nominal pore size) in a Millipore stainless steel filtering funnel.
Multiple centrifuge tubes were necessary to provide sufficient sample
volume for all analyses. One liter of filtered pore water was collected,
preserved with 5 ml/L of a 1 g/L solution of HgClz to prevent microbial
degradation, and stored in the dark at 4°C until required for analysis.

One liter of pore water was adjusted to a Ph of >11 with
concentrated NaOH, extracted with three successive 60 ml portions of
pesticide-grade)methylene chloride in a separatory funnel and the extracts
combined for analysis of NPOCs. If a large emulsion was observed,
continuous liquid-liquid extraction was used to complete the sample
extraction. Extracts were passed over an anhydrous sodium sulphate drying
column and reduced to 1 ml in a Kuderna-Danish concentrator. Final
extract volumes were adjusted to 10 ml with methylene chloride. Pore
water extraction followed the protocols outlined in U.S. EPA Method 3510
or, for emulsions, U.S. EPA Method 3520 (24). Cleanup procedures for both
sediment and pore water extracts followed the protocols outlined in U.S.
EPA Methods 3620, 3630 or 3660 (24) and were dictated by the class of
compounds quantified in subsequent analyses.

Identification and quantitation of chemical analytes were performed
with gas chromatography/mass spectroscopy (GC/MS) techniques. U.S. EPA
600 Series methods (25) were used for CC analyses while compound
confirmation was conducted with U.S. EPA GC/MS Methods 8240 and 8250 (24).

GC/MS operating conditions were as follows for all analyses: electron

19
energy 70 eV, mass range 35-550 amu, scan time 1 sec/scan, transfer line
temperature 250°C, source temperature 200-250°C, injector temperature 250-
350°C, injector on column or Grob splitless, sample volume 1 p1 and
carrier gas He at 15 psi. Analytical detection limits for all analysis
have been presented elsewhere (26). Bulk sediment and pore water chemical
concentrations are reported as mg/kg and pg/L, respectively.

Bulk sediment total organic carbon (TOC) was measured with a LECO
carbon analyzer dry combustion technique and the results reported as 8 TOC
on a dry wt. sediment basis. Dissolved organic carbon (DOC) in the pore
water was operationally defined as the material passing a 0.7 pm filter
and was measured with an IO Corporation, Inc. Model 700 TOC analyzer after

sample acidification. Pore water DOC is reported on a mg/L basis.

Equilibrium Partitioning Calculations
Apparent partition coefficients, Kp', for NPOCs were calculated from

measured chemical concentrations (Equation 7):

 

Kp = Csoc (7)
Cp
where: CSOC = C8
foc

while actual partition coefficients K were calculated from Equation 8,

P:

p soc (8)

 

where: Cd = Cp

 

1 + mDoc KDoc

and; Cp = measured total NPOC concentration in pore water
C8 = measured NPOC concentration in bulk sediment
CSOC = organic carbon-normalized NPOC concentration in bulk
sediment
f0c = decimal fraction organic carbon in bulk sediment
Cd = calculated dissolved NPOC concentration in pore water

mDOC = measured DOC concentrations in pore water

KDOC = DOC partition coefficient for NPOC

Theoretical partition coefficients, Kt were calculated*with the fractional
organic carbon content of the sediment and the Kow for the NPOC of

interest (Equation 9).

Kt = foc Kow (9)

For all NPOCs, the loglo Koc (i.e., pKOC) was calculated (Equation 10)

[27].

loglo Koo = 0.00028 + 0.983 loglo K (10)

CW

21
Measured pKow values, if available, were used in all calculations of pKoc.
Estimated pKow values were used if necessary, however, all values, either
measured or estimated (Table l), were obtained from the AQUIRE database
(28). In calculations of potential DOC-binding by NPOCs and dissolved

NPOC concentrations in pore water, KDOC was assumed to be equal to Koc'

Results

A total of 29 NPOCs analyzed for were present in both Grand Calumet
River bulk sediments and pore waters (Table 1). Measured or estimated
pKow values from AQUIRE and calculated pKoc values for each chemical also
are reported in Table 1. As a point of interest, the original organic
chemistry analytical suite for both bulk sediments and pore waters was
composed. of 106 discrete compounds (26). The compounds ‘present in
sediments and pore waters from the study area included phenolic compounds,
industrial solvents and degreasers, pesticides and polycyclic aromatic
hydrocarbons (PAHs).

The concentrations of 29 NPOCs which were present in both bulk
sediment and pore water samples from the study area are presented in Table
2. TOC and DOC concentrations in bulk sediment and pore water,
respectively, are presented in Table 2, as well as calculated ”dissolved"
concentrations of the 29 NPOCs in pore water. Measured concentrations of
the NPOCs in bulk sediment ranged from 10 pg/kg for tetrachloroethylene to
over 100 mg/kg for benzo(a)pyrene. Measured total concentrations of the
NPOCs in pore water ranged from 0.1 pg/L for some of the chlorinated
pesticides (p,p'-DDT, dieldrin, lindane, chlordane) and industrial

chemicals (e.g., hexachlorobenzene) to greater than 450 pg/L for the PAH,

22
naphthalene.
The pKt, which represents the loglo (f0c x Kow) was calculated for

each NPOC and compared. with the ‘pK .,which represents the apparent

P
partition coefficient loglo (Csoc/Cp) between the organic carbon-
normalized sediment and total pore water chemical concentrations and the
pr, which represents the actual partition coefficient loglo (Csoc/Cd)
based on the calculated "dissolved" chemical concentration in the pore
water (Table 3). These results indicate that under the operationally
defined conditions of KDOC = K0C and DOC = <0.7 pm, the binding of NPOCs
to DOC was an important factor determining the distribution of NPOCs in
pore waters from the Grand Calumet River. As the pKow of a compound
increased, so did the importance of DOC-binding in determining the pore
water distribution of a given NPOC. Di-octyl phthalate was the NPOC with
the greatest pKow in the data set. Calculations based on measured total
pore water DOC and di-octyl phthalate concentrations indicated that less
than .01% of the measured total concentration of di-octyl phthalate in
pore water (1.7-27.7 pg/L) was actually present as "dissolved", un-bound
chemical.

To assess the influence of DOC in the pore water on concentrations
of all 29 NPOCs in pore water, pr. was plotted as a function of pKt
(Figure 2). A similar relationship was developed for only those NPOCs
with a measured or estimated pKow value <3.0 (Figure 3) or >3.0 (Figure
4). Each point with confidence limits represents the mean value i one
standard deviation from 5-10 samples (see Table 3) for an individual
chemical.

Little relationship was observed between pr and pKt since pKt

accounts for only approximately 9% (r=0.29, R=0.09) of the variance in

23

 

 

Table 1. Non-polar organic compounds analyzed for and present in Grand
Calumet River sediments and pore waters. Measured (M) or
estimated (E) octanol-water partition coefficients (Kow) from
the AQUIRE database were used to calculate estimated organic
carbon partition coefficients (Koc) with the following
equation: loglo Koc = 0.00028 + 0.983 10910 Kow (27).

Compound CAS No. 10910 Kow (M/E) loglo Koo

Phenol 108-95-2 1.46(M) 1.44

2,4-Dinitrotoluene 121-14-2 1.98(M) 1.95

1-Chloro-2-nitrobenzene 88-73-3 2.24(M) 2.20

1,2,3-Trichloropropene 96-19-5 2.36(E) 2.32

1,1,1-Trichlorethane 71-55-6 2.49(M) 2.45

Chlorobenzene 108-90-7 2.84(M) 2.79

Styrene 100-42-5 2.95(M) 2.90

Ethylbenzene 100-41-4 3.15(M) 3.09

Naphthalene 91-20-3 3.30(M) 3.25

p-Chlorotoluene 106-43-4 3.33(M) 3.28

o-Dichlorobenzene 95-50-1 3.38(M) 3.33

Tetrachloroethylene 127-18-4 3.40(M) 3.35

Lindane 58-89-9 3.61(M) 3.55

Biphenyl 92-52-4 4.09(M) 4.02

Dieldrin 60-57-1 4.32(M) 4.25

Phenanthrene 85-01-8 4.46(M) 4.39

Heptachlor 76-44-8 4.61(E) 4.53

Pentachloronitrobenzene 82-68-8 4.64(M) 4.56

Table 1. (cont.)

24

 

 

Compound CAS No. 10910 Kow (M/E) loglo Koc

Pyrene 129-00-0 4.88(M) 4.80
Fluoranthene 206-44-0 4.95(E) 4.87
Pentachlorophenol 87-86-5 5.24(M) 5.15
Hexachlorobenzene 118-74-1 5.31(M) 5.22
Chlordane 57-74-9 5.54(E) 5.45
Chrysene 218-01-9 5.66(E) 5.57
Benzo(a)anthracene 56-55-3 5.66(E) 5.57
Benzo(a)pyrene 50-32-8 5.97(M) 5.87
p,p'-DDT 50-29—3 6.36(M) 6.25
p,p'—DDE 72-55-9 6.51(M) 6.40
Di-octyl phthalate 117-81—7 8.71(M) 8.56

 

25

 

 

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0.00000 .0 00009

29

 

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31

 

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32

 

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33

 

 

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34

 

 

 

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35

 

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41
pxp'. This is also indicated by the divergence of the dashed line which
best fits the data from the solid line representing unity or perfect
correspondence between pKP' and pKt. When only the data for NPOCs with
pKow values (3.0 were plotted (Figure 3), better agreement was observed
between theoretical predictions and observed results. The line which best
fits the data has an intercept of 0.99 and a slope of 1.35. The

relationship pKt accounts for 48% of the variance in pr. (r=0.69,

R=O.48). The NPOCs with pKow values >3.0 exhibited little relationship

I

between pKt and pr

since pKt explained (3% of the variance in pr.
(r=0.16, R=0.03).

If the data for NPOCs with PKow values >3.0 are corrected for
binding to DOC and plotted, there is much better concordance
between pKt and the actual partition coefficient, pr (Figure 5). After
correction, pKt accounted for 58% of the variance in pr (r=0.76, R=0.58)
as opposed to accounting for only <3.0% of the variance in pr' before
correction for binding to DOC. If the original data for NPOCs with pKow
values <3.0 are added back into the data set, the relationship between pKt

and pK further improves and approximately 62% of the variance in pr can

P

be accounted for by pKt (Figure 6).

I

The original data on pr

versus pKt from this study as well as
similar data from other field studies on PAHs (18, 23) and laboratory
sediment-spiking studies with industrial chemicals and pesticides (29) are
compared in Figure 7. These data fit the general pattern observed for the
Grand Calumet River data presented here, however, it appears that binding
to DOC was potentially more important in determining NPOC distribution in

pore waters from the studies by Socha and Carpenter (18) and Oliver (29)

than in the study by Kadeg and Pavlou (23). This should be the case since

 

42

 

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oogw

 

 

 

45
Kadeg and Pavlou (23) determined "dissolved" concentrations of PAHs in

pore water.

Discussion

The potential for binding of NPOCs to DOC in pore water is an issue
of importance in the development of exposure models for the effects of
NPOCs on aquatic organisms and, in particular, benthic macroinvertebrates.
If DOC-bound NPOCs are unavailable to cause effects on biota as proposed
by McCarthy and Jimenez (21) and Landrum et al (22), then these exposure
models must account for DOC-binding of NPOCs in pore water to accurately
predict exposure concentrations for NPOCs. Currently, equilibrium
partitioning theory (EqP) is being proposed as the theoretical basis for
the development of sediment quality criteria (SQC). Draft SQC criteria
documents have been prepared by the U.S. Environmental Protection Agency
(EPA) for five NPOCs, endrin, dieldrin, phenanthrene, fluoranthene and
acenaphthene. These criteria development efforts have made no effort to
account for DOC-binding of NPOCs in pore water because, to date, there has
been little empirical field evidence conclusively demonstrating the
importance of DOC-binding in determining the distribution, and thus
bioavailability, of NPOCs in pore waters from contaminated sediments.
NPOC binding to DOC may be of the greatest importance in attempts to
extrapolate SQC which are reported on an organic carbon-normalized bulk
sediment concentration basis (pg NPOC/gm organic carbon) to permissible
concentrations of NPOCs in pore water. Binding to DOC could increase NPOC
concentrations in pore water to values greater than permissible based on

extrapolation from SQC while the "dissolved" (i.e, bioavailable)

46

concentration of the NPOC in pore water could actually be acceptable if
binding to DOC was accounted for in the calculation of acceptable NPOC
concentrations in pore water. A major strength of EqP as an exposure
model and potential basis for the development of SQC is that including a
correction for DOC-binding of NPOC requires only a simple extension of the
partitioning theory to DOC and is consistent with the thermodynamic
principals underlying EqP (4).

Current EqP models are best described as two phase because they
account for NPOC partitioning between two sorbent phases, sediment organic
carbon and pore water. The existence of a third sorbent phase, DOC, also
would explain the particle concentration effect which has been reported by
some researchers (26, 30). Experimental limitations make it difficult to
measure empirically the "dissolved" and DOC-bound concentrations of NPOCs
in pore water, however, Voice et al (14) and Gschwend and Wu (15) have
previously proposed that dissolved ligands and colloids, respectively,
were the agents causing the observed particle concentration effects.

The results of this investigation of the DOC-binding of NPOCs in
pore waters from the Grand Calumet River has demonstrated the potential
importance of this phenomenon in determinations of the bioavailable
fraction of NPOCs in pore water. Frequently, critics of the EqP method of
predicting exposure concentrations have decried the lack of relationship
between predicted and measured pore water concentrations of NPOCs. This
conclusion would follow after a cursory examination of the data presented
in Figure 2. This observation is made even more apparent for NPOCs with

values

10910 Row >3.0 if the data are divided into NPOCs with 10910 Kow

<3.0 (Figure 3) and >3.0 (Figure 4).

The importance of NPOC binding to DOC in pore water has previously

47

been reported for two limited groups of NPOCs, PCB congeners (16) and Pass
(18), respectively. Both studies reported that NPOC concentrations in
pore waters from field collected sediments were greater than predicted due
to partitioning to DOC. Correction for DOC-binding resulted in
"dissolved" NPOC concentrations in pore waters which were in better
agreement with predicted values. Chiou et al. (17) also have reported
that the apparent water solubilities of organic chemicals, including
pesticides, increased linearly with an increase in DOC (DOM) concentration
while Brusseau (31) reported the enhanced sorption of three non-ionic
compounds to two aquifer materials of low organic carbon content. The
mechanism for this effect appeared to be increased DOC concentration as a
result of experimental additions of tetrachloroethane.

This study demonstrates the applicability of the EqP predictions to
a much greater range of compounds, if DOC-binding is taken into account.
However, one of the criticisms of our approach to determining the
"dissolved" concentrations of NPOCs in Grand Calumet River pore waters may
be the assumption that KDOC = Koc’ Most studies have reported Koc values

< K values but few studies have evaluated KDOC values in relationship to

OW

K Although several studies have reported KDOC values (Koc' several

oc‘
also have reported that NPOCs bind to organic sorbents in particulate and
dissolved form to a similar extent (32, 34). An important factor in these
evaluations may be determining that the sediment organic matter and the
pore water dissolved organic matter are of the same origin (i.e.
terrestrial, aquatic plant, etc.) and physical/chemical characteristics
(i.e. hydrophobicity, CHNO composition, etc). Numerous examples exist of

altered partitioning of NPOCs to different types of organic matter which

was due to differences in the chemical composition of the organic matter

 

48

Although NPOCs with pK >3.0 are generally of the greatest interest

ow
because of their propensity to resist environmental degradation and to be
bioconcentrated by biota, the behavior of the NPOCs with pKow <3.0 also is
of interest in the Grand Calumet River data set. Based on the data
presented in Table 2 and Figure 3, it appears that these NPOCs either
occurred at greater than expected concentrations in sediments from the
study sites or at concentrations less than expected in the study site pore
waters. The former possibility was reported by Boyd and Sun (36) for
soils and was hypothesized to be due to residual petroleum hydrocarbons
and polychlorobiphenyl oils acting as a sorptive phase with approximately
10 times the partitioning strength of native soil organic matter.
Sediments from the Grand Calumet River do contain significant amounts of
petroleum hydrocarbons (26). However, based on our unpublished
comparisons of measured versus predicted pore water concentrations of the
NPOCs in pore water, it appears that a much more plausible explanation is
that these compounds may be volatilized from the pore water preferentially
by the vacuum filtration step during pore water preparation.

Equilibrium partitioning theory appears to be a viable basis for
predicting the partitioning behavior of NPOCs under field conditions if
consideration is given to the potential for DOC-binding of NPOC. Based on
comparison of actual partition coefficients, pr, versus estimated
partition coefficients, pKt, for the NPOCs measured in Grand Calumet River
sediments and pore waters, binding to DOC in pore water is not important
for NPOCs with pKow values <3.0. NPOCs with pKow values >4.0 can be
expected to demonstrate significant potential for binding to DOC in pore

water. If SQC are to be based on water quality criteria extrapolated to

49
"dissolved" pore water chemical concentrations, binding to DOC in pore
water must be accounted for in determining "dissolved" pore waters
concentrations of the chemicals of interest, otherwise the application of
SQC to determine permissible concentrations of NPOCs in pore water will

overestimate the potential for adverse effects.

Acknowledgements

This research was funded by a grant from the U.S. Environmental
Protection Agency, Great Lakes National ProgramnOffice. JPrevious versions
of the manuscript were typed by Debra Williams and Jane Norlander. Russ
Erickson and Gary Ankley provided insightful review comments which

inproved the final version of the manuscript.

50

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Shea, D. Environ. Sci. Technol. 1988, 22, 1256.

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Environmental Protection. Agency, Office of Water
Regulations and Standards, Criteria and Standards
Division, Washington, D.C. 1985.

Pavlou, S.P. In Fate and Effects of Sediment-Bound Chemicals in
Aquatic Systems, Dickson, K.L.; Maki, A.W.; Brungs, W.A.

Eds; Permagon Press: Elmsford, NY, 1987; pp 388-412.

DiToro, D.M.; Zarba, C.S.; Hansen, D.J.; Berry, W.J.; Swartz, R.C.;
Cowan, C.E.; Pavlou, S.P.; Allen, H.E; Thomas, N.A.;
Paquin, P.R. Environ. Toxicol. Chem. 1991, 10, 1541.

Adams, W.J.; Kimerle, R.A.; Mosher, R.G. In Aquatic Toxicology and
Hazard Assessment. STP 854, Cardwell, R.D.; Purdy R.;
Bahner, R.C. Eds; American Society for Testing and
Materials: Philadelphia, PA, 1985; pp 429-453.

Ziegenfuss, P.S.; Renaudette W.J.; Adams, W.J. In, .Aggatic
Toxicology and Environmental Page. Ninth gymposium, ASTM
STP 921, Poston, T.M.; Purdy, R. Eds., American Society
for Testing and Materials: Philadelphia, PA, 1986; pp

479-493.

51

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7. Adams, W.J. In Fate and Effects of Sediment-Bound Chemicals in
Aquatic Systems, Dickson, R.L.; Maki, A.W.; Brungs, W.A.
Eds., Permagon Press: Elmsford, NY, 1987; pp 219-244.

8. Connell, D.W.; Bowman, M.; Hawker, D.W. Ecotoxicol. Environ. Safety.
1988, 16, 293.

9. Lake, J.L.; Rubinstein, N.I.; Lee II, H.; Lake, C.A.; Heltshe, J.;
Pavignano, S. Environ. Toxicol. Chem. 1990, 9, 1095.

10. Swartz, R.C.; Schults, D.W.; Dewitt, T.H.; Ditsworth, G.R.;
Lamberson. J.O. Environ. Toxicol. Chem. 1990, 9, 1071.

11. DiToro, D.M.; Allen, H.E.; Cowan, C.E.; Hansen, D.J.; Paquin, P.R.;
Pavlou, S.P.; Steen, A.E.; Swartz, R.C.; Thomas, N.A.;
Zarba, C.S. EPA 440/5-89-002. U.S. Environmental
Protection Agency, Criteria and Standards Division,
Washington, D.C. 1989.

12. Knezovich, J.P.; Harrison, F.L. Ecotox. Environ. Safety. 1988, 15,
226.

13. Eadie, B.J.; Landrum, P.F.; Faust, W. Chemosphere. 1982,
11(9), 847.

14. Voice, T.C.; Rice, C.P.; Weber Jr., W.J. Environ. Sci. Technol.
1983. 17, 513.

15. Gschwend, P.M.; Wu, 8. Environ. Sci. Technol. 1985, 19, 90.

16. Brownawell, B.J.; Farrington, J.W. Geochemica et Cosmochemica Acta.
1986, 50, 157.

17. Chiou, C.T.; Malcolm, R.L.; Brinton, T.I.; Kile, D.E. Environ. Sci.
Technol. 1986, 20, 502.

18. Socha, S.B.; Carpenter, R. Geochemica et Cosmochemica Acta. 1987,

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Eadie, B.J.; Morehead, N.R.; Landrum, P.F. Chemosphere. 1990, 20(1-
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Capel, P.D.; Eisenreich, S.J. J. Great Lakes Res. 1990, 16(2), 245.
McCarthy, J.F.; Jimenez, B.D. Environ. Toxicol. Chem. 1985, 4, 511.
Landrum, P.F.; Nihart, S.J.; Eadie, B.J.; Herche, LJR. Environ.
Toxicol. Chem. 1987, 6, 11.
Kadeg, R.D.; Pavlou, S.P. Reconnaissance Field Study for
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Hydrophobic Organic Chemicals. U.S. EPA, Criteria and

 

Standards Division, Washington, D.C. 1987.

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Laboratory Maqqal. Physical/Chemical Methoda. U.S.
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Hoke, R.A.; Giesy, J.P.; Zabik, M.; Unger, 1L Ecotox. Environ.
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AQUIRE. aqqatic Toxicity Information Rgtrieval Database. U.S.
Environmental Protection Agency, ERL-Duluth, Duluth, MN.
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142.

CHAPTER 2

Toxicity of Sediments and Sediment Pore Waters from the Grand Calumet

River-Indiana Harbor, IN Area of Concern.

55

Introduction

The sediments of many rivers, harbors and connecting channels of
the Great Lakes contain a wide array of chemical compounds, including
organic xenobiotics (Pranckevicius 1986, Fallon and Horvath 1985) and
metals (Pranckevicius 1986, Fallon and Horvath 1985, Hamdy and Post 1985).
Due to municipal, industrial and non-point source waste discharges and the
tendency of many chemicals to become associated with sediments,
contaminated sediments are particularly problematic near densely
populated, industrialized urban areas, such as the area surrounding the
Grand Calumet River-Indiana Harbor (GCR-IH) in northwest Indiana. Due to
contaminants and nutrients in the water and sediments from this area, it
has been designated an Area of Concern (AOC) by the International Joint
Commission (IJC 1985). The environmental status, relative to
anthropogenically-introduced chemical contamination, in these areasuof the
Great Lakes has been summarized elsewhere (Rodgers et al. 1985). The
severely contaminated condition of the GCR-IH AOC is based on water
quality problems with conventional pollutants, metals, and organic
chemicals; contaminated sediments, impacted aesthetics and biota, and fish
consumption advisories (IDEM 1988). The presence of a wide variety of in-
place chemical contaminants in sediments from the AOC and their potential
effects on biota were the primary focus of this study.

Assessing the degree of contamination of sediments consists of

basically two aspects: 1) determination of which contaminants are present

56

57
and 2) evaluation of the potential effects of these contaminants on biota
(Bishop 1987). The potential toxicity of sediments to benthic organisms
can. be determined by surveying 'the number and types of indigenous
organisms present in a sediment (Chapman 1986). However, the absence of
a particular macroinvertebrate species in a sediment does not necessarily
indicate that the sediment is toxic. Toxicity of sediment to benthic
invertebrates also can be estimated by quantifying all of the toxic
compounds and elements associated with the sediment by analyzing: 1) bulk
sediment, 2) sediment pore water (Jenne et al. 1980, Batley and Giles
1980), 3) an elutriate of the sediment (Brannon et al. 1980, Laskowski-
Hoke and Prater 1980), or 4) various organic or acid extracts designed to
selectively remove particular classes of toxic substances (Samoiloff et
al. 1983). The potential toxicity of a sediment to benthic organisms then
can be predicted by comparing the observed concentrations of chemicals in
1, 2, 3, or 4 above to dose-response relationships determined under
laboratory conditions for each individual toxicant. However, a number of
parameters, such as organic carbon content and particle size distribution,
can affect the availability of both metals and organic chemicals in
sediments to benthic organisms (Babich and Stotzky 1977, Laxen 1985,
Oliver 1985). In addition, potential interactions due to the presence of
a complex chemical mixture in a sediment are not known. Toxicity tests,
however, can be used to provide a direct assessment and to integrate the
effects of the biologically active fraction of all of the toxic chemicals
present in a sediment, pore water, elutriate or organic solvent extract.
If laboratory toxicity data are available for specific chemicals present
in the samples, the results of toxicity tests can be compared with

measured chemical concentrations in a "toxic units" (TU) approach (Sprague

58
and Ramsay 1965) to help determine potential causes of the observed
toxicity.

This investigation tested the relative sensitivities of
several simple toxicity tests with bacteria and invertebrates in order to
provide a framework for conducting rapid, comprehensive surveys of
potentially toxic sediments and to conduct an assessment of sediment
toxicity in the Grand Calumet River, IN. The tests chosen for use in the
toxicity assessment of sediment pore waters were the Microtoxo test and 48
h acute tests with Daphnia maqna and Ceriodaphnia dubia while a 10-d test
with Chironomus tentans was chosen to assess the toxicity of solid phase
sediments.

The Microtox® test is a bacterial luminescence toxicity test
developed by Beckman, Inc. in 1977 (Bulich 1984) as a rapid screening
alternative to standard acute toxicity testing with fish or invertebrates.
This test is based on the reduction in bioluminescence of the marine
bacterium (Photobacterium phoaphoraqm) (NRRL B-11177) by toxic chemicals.

Comparisons of the Microtox® test and acute toxicity tests with both fish
and invertebrates for a large number of pure compounds and complex
mixtures has demonstrated good general agreement between the fathead
minnow, Pimephales promelas, and. Daphnia. maqna acute ‘tests and 'the
Microtox® test, both within- and among-laboratories (Green et a1. 1985).
The use of both sucrose and NaCl osmotic adjustment techniques in the
MicrotoxO test also has been demonstrated to be of use in determining the
nature of the compound causing the toxicity when testing complex mixtures
(Hinwood and McCormick 1987, Ankley et al. 1990a, Hoke et al. 1992a).

Cladoceran species have numerous advantages for aquatic

S9

toxicity testing, including their sensitivity to a wide variety of
environmental contaminants (Maki 1979, LeBlanc 1980). Daphnia maqna and
D. pulex have been most frequently used in aquatic toxicity tests and a
large data base exists for the effects of pure compounds on D. maqga
(Hunter et al 1990). In addition, the response of D. maqna has been
correlated with the responses of other species to toxicants (Nebeker et
al. 1983, Nebeker et a1. 1986). Acute lethality tests with D. maqna also
seem appropriate for use in sediment pore water toxicity tests because it
is one of the species used to establish surface water quality criteria
(U.S. EPA 1980a).

Ceriodaphnia dubia occurs in natural zooplankton assemblages of

 

the Great Lakes and has recently gained wide—spread acceptance as a test
species for acute and chronic toxicity testing of effluents and various
other types of aqueous extracts (Mount and Norberg 1984, DeGraeve and
Cooney 1987, Mount and Anderson-Carnahan 1988a, b; Mount 1989, Ankley et
al. 1990b, Oris et al. 1991, Kszos and Stewart 1991). Several
investigations have evaluated of the comparative sensitivity of C. dqbia
and other daphnid species to water quality variables and pure chemicals
(Cowgill et al. 1985, Takahashi et al. 1987, Winner 1988, 1989; Mokry and
Hoaglund 1990, Cowgill and Milazzo 1990, 1991a, b). The sensitivity of Q;

dubia also has been compared with the response of the MicrotoxO test to

 

different wastewater fractions (Mazidji et al. 1990), with the responses
of natural zooplankton and benthos communities to chronic copper stress
(Moore and Winner 1989), with the response of fathead minnow tests for
evaluating in-stream toxicity dynamics (Stewart et al. 1990) and with the
responses of the Microtox® and fathead minnow tests to sediment elutriates

(Hoke et al. 1990). Because the comparison of relative species

60

sensitivities to pore water from contaminated sediments was of interest,

acute toxicity tests of sediment pore waters were conducted with both _D_y

maqna and C. dubia.

Chironomus tentans (Diptera:Chironomidae) is a representative of

a group of insects known as the midges, which are widely distributed in

freshwater sediments during their larval stage of development. This

species spends almost all of its life cycle in a tunnel in the upper few

centimeters of sediments (Sadler 1935). Chironomids often comprise a

significant proportion of the benthic biomass and are important in the

cycling of oxygen, nutrients and contaminant residues into and from the
sediments due to bioturbation (Graneli 1979, Hargrave 1975, Matisoff et
al. 1985). Chironomus tentans can be satisfactorily reared in the
laboratory and has previously been used as a sediment toxicity test
organism (Wentzel et al. 1977, 1978, Batac-Catalan and White 1982, Mosher
and Adams 1982, Mosher et al. 1982, Giesy et al. 1988, 1990). In previous

studies of Great Lakes sediments (Giesy et al. 1988), a reduction in Q;

tentans dry weight, relative to control, of approximately 30-40% in the

whole sediment toxicity tests was observed for sediments which did not
support viable communities of benthic invertebrates in the field.
The objectives of this study were as follows:

1) To evaluate the toxicity of sediments and sediment pore waters
from 13 locations in the Grand Calumet River, IN (Figure 8) with
the following toxicity tests and endpoints,

i) Photobacterium pmosphoreum, Microtoxo, inhibition of
bioluminescence in 5, 15, and 30 minute incubations with

sediment pore water in assays using two forms of osmotic
protection (NaCl, sucrose) for the bacteria.

 

ii) Daphnia maqna and gariodaphnia dubia, survival of 24 h old
neonates during 48 h exposures to sediment pore water.

61
iii) Chironomus tentans, survival and inhibition of dry wt. gain,
relative to control, following 10-d exposures to whole
sediments.

2) To test the hypothesis that the toxicity of whole sediments or
pore waters as determined by laboratory toxicity tests can be
predicted from toxic units calculated based on the concentrations
of individual or classes of compounds and known dose-response
relationships.

3) To determine whether sediment toxicity is better predicted from

concentrations of chemical residues in pore‘water than residues in
bulk sediments.

Materials and Methods

Sample Collection

Sediment samples were collected from the Grand Calumet River, IN
on 11 October and 22 November 1988; 10 March, 24 May, 30 October and 13
November 1989; and 12 May 1990. A Ponar grab sampler was used to collect
sediment samples from 10 locations along the Grand Calumet River and three
locations in the Indiana Harbor ship canal (Figure 8). At the time of
sample collection, study sites were located by triangulation of local
landmarks. The sample from each location was a composite of approximately
80-100 L of wet sediment from multiple Ponar grabs. Multiple grab samples
were collected and composited to ensure sufficient sample volume for all
necessary sub-sample collection. Compositing and homogenization of
composite samples were done in a large stainless steel pan with stainless
steel tools. Large debris was removed from the composite samples and two,
1-L aliquants (sub-samples) for quantification of metals and organic
compounds in bulk sediments were placed in l-L glass jars capped with
washed aluminum foil under the lid. Samples for toxicity testing or pore

water extraction were placed in coolers or plastic buckets lined with

62

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63

food-grade plastic bags. Prior to sample collection, sample bottles were
washed and solvent rinsed with hexane. .After collection, sediment samples
were placed on ice in coolers and transported to the laboratory, where
they were maintained in a walk-in cooler at 4° C until processing and
analysis.
Pore Water Extraction

Pore water was extracted from sediments by a combination of
centrifugation and filtration as described by Hoke et al (1992a). The
pore water extracts were placed in glass bottles, the bottles capped with
aluminum foil-lined lids, and the pore water maintained in the dark at 4°
C until used for assays (< seven days) or until extracted for subsequent
chemical analysis. Pore water samples for metals analysis were preserved
with a sufficient amount of concentrated HNO3 to lower the sample pH to S
2.0. Samples for quantification of organic compounds were preserved with
5 ml/L of a 1 g/L solution of HgC12 to prevent microbial degradation. The
cleaning procedure for pore water sample bottles was identical to the
procedure used for sediment sample bottles. Subsamples of sediments and

pore water extracts were archived at 4° C.

Chemical Analysis
Organics

Organic chemical analyses were only conducted.on samples collected
from the 10 locations on the Grand Calumet River (Figure 8). In general,
the analytical methods for the quantification of non-polar organic
compounds in sediment and pore water followed the scheme presented below.
Subsamples of sediment were dried at 105°C for 24 h and percent moisture

determined for each sample. Ten grams of dried sample were mixed with 10

64

g of anhydrous sodium sulphate and Soxhlet extracted for 24 h with
pesticide-grade acetone/hexane (1:1, v/v). The extract was passed over a
drying column containing anhydrous sodium sulphate and the column rinsed
with approximately 100 ml of the acetone/hexane mixture to complete the
transfer. The extract was transferred to a Kuderna-Danish concentrator
and extract volume reduced to 1 ml. The final extract volume was adjusted
to 10 ml with the acetone/hexane mixture. These procedures and all other
aspects of sediment sample preparation followed U.S. EPA.Method 3540 (0.8.
EPA 1986).

One liter of pore water was adjusted to a pH of >11, extracted
with three successiveu60 ml portions of pesticide-gradenmethylene chloride
in a separatory funnel and the extracts combined for analysis of neutral
organic compounds. If a large emulsion was observed, continuous liquid-
liquid extraction was used to complete the sample extraction. Extracts
were passed over an anhydrous sodium sulphate drying column and reduced to
1 ml in a Kuderna-Danish concentrator. Final extract volumes were
adjusted to 10 ml with methylene chloride. Pore water extraction followed
the protocols outlined in U.S. EPA Method 3510 or, for emulsions, U.S. EPA
Method 3520 (U.S. EPA 1986). Cleanup procedures for both sediment and
pore water extracts followed the protocols outlined in U.S. EPA Methods
3620, 3630 or 3660 (U.S. EPA 1986) and were dictated by the class of
compounds quantitated in subsequent analyses.

Identification and quantitation of chemical analytes were
performed with gas chromatography/mass spectroscopy (GC/MS) techniques.
U.S. EPA 600 Series methods (U.S. EPA 1984) were used for CC analyses
while compound confirmation was conducted with U.S. EPA.GC/MS Methods 8240

and 8250 (U.S. EPA 1986). GC/MS operating conditions were as follows for

65

all analyses: electron energy 70 eV, mass range 35-550 amu, scan time 1
sec/scan, transfer line temperature 250 °C, source temperature 200-250 °C,
injector temperature 250-350 °C, injector on column or Grob splitless,
sample volume 1 pl and carrier gas He at 15 psi. Compounds were
identified by comparison to library spectra and confirmed by comparison to
authentic standards.

Metals

Concentrations of metals in whole sediments and pore waters were
determined by a combination of inductively-coupled argon plasma (ICP) and
atomic absorption (AA) spectroscopy. Whole sediments were digested in
concentrated HNO3 (Plumb 1981, APHA 1985). Standard additions were used
to avoid matrix effects.

Miscellaneous parametera

A number of parameters were determined by standard techniques.
Acid volatile sulphides (AVS) were measured in bulk sediments with a
gravimetric technique (Di Toro et al. 1990, 1991). Total oil and grease
was measured by the freon extraction method followed by gravimetric
quantitation (Plumb 1981). Sediment TOC was measured with a LECO carbon
analyzer dry combustion technique.

Selected contaminants (CN, H28, NH3), total organic (TOC) and
total inorganic carbon (TIC) in pore water were measured when samples were
prepared for toxicity tests. Concentrations of cyanide and hydrogen
sulfide were determined by standard methods (APHA 1985). Total ammonia
was measured with an Orion® Model 701A ionanalyzer and an OrionO Model
95-12 ammonia electrode. TOC and TIC were measured with an 10
Corporation, Inc. Model 700 TOC analyzer. Hardness, alkalinity,

conductivity and Ph were measured in the 100% pore water concentration

66
during cladoceran toxicity tests. Reported values are the maximum values

observed during the course of the 48 h toxicity tests.

Toxicity Tests
Photobacterium phosphoreum

The Microtox' bacterial luminescence assay' was performed. on
sediment pore waters with the standard procedure (Bulich et al. 1981) and
the alternate osmotic adjustment procedure developed by Hinwood and
McCormick (1987). Reduction in bioluminescence of the bacterium £_.
phosphoreum was used as a measure of the toxicity of the sediment pore
waters. A Microtox® Model 2055 Toxicity Analyzer (Microbics Co.,
Carlsbad, CA.) was used for all measurements. The calculated ratio of
corrected light emitted to emitted light remaining after 5, 15 or 30
minutes was determined for each sample dilution and all results reported
as the percent pore water causing a 50% inhibition of bioluminescence
(ECSO). ECSO values were calculated with the linearized gamma
distribution.
Daphnia maqna andygariodaphnia dubia

Initial C. dubia cultures were obtained from the Eli-Lilly Co.
(Greenfield, IN) while D. maqna originated from laboratory cultures
maintained by the Michigan Department of Natural Resources (East Lansing,
MI). In the laboratory, C. dubia were either maintained in mass cultures
in 1000 ml beakers containing 900 ml of culture water or in a brood board
as previously described by Hoke et a1 (1990). Daphqia maqqa were
maintained in mass cultures contained in 1000 ml beakers. Mass cultures
of both species were initiated with one gravid female and maintained in a

Scientific Products incubator at 25° C with a 16L:8D photoperiod (light

67

levels - 200 pE/m2 sec). Brood boards were initiated with 60 gravid
females and were maintained in the same incubator as the mass cultures.
Feeding consisted of daily additions of Selenastrum capricornugum (200,000
cells/m1 culture water) and Yeast-Cerophyl-Trout chow (YCT, 3 ml/1000 ml
culture water). Thinning of the mass cultures was performed weekly by
removing approximately 90% of the existing organisms and water with
replacement of an equal volume of fresh culture water. Cultures were
maintained for 14-d (brood boards) and 21-d (mass cultures) prior to
initiation of new cultures for C. dubia and D. ma na, respectively.
Culture, control and diluent waters for all assays was 10% Perrier mineral
water prepared as described by Hoke et a1 (1990). Two pg/L of
cyanocobalamin (vitamin 812) and selenium (as sodium selenate) were added
to the 10 % Perrier water mixture. The aerated 10% Perrier water mixture
had the following chemical characteristics over the entire study period
(mean i SD): DO-7.9 i 0.4 mg/l, Ph-8.1 i 0.4, specific conductance-105 :
26 ymhos/cmz, hardness-60.3 ¢_9.6 mg/L as CaCO3 and alkalinity-59.3 i 11.7
mg/L as CaCO3.

The sediment pore water was prepared 24-48 h prior to assays and
allowed to equilibrate to 25° C immediately before starting a test.
Twenty mL of each test solution were dispensed into 10 replicates for each
pore water dilution concentration or control using a device developed by
Mount and Norberg (1984) for this purpose. Chemical characteristics of
the initial pore water dilutions were measured at test initiation. All
assays were conducted in a Scientific Products incubator at the same
temperature, photoperiod and light level regimes used for culturing the

test organisms.

68

Ceriodaphnia dubia neonates, < 12 hr old and not more than 8 h

 

different in age, from the third brood of the originally isolated neonates
were used to begin all assays. D. maqna neonates <12 h old were collected
by screening all neonates from the mass cultures and then again screening
all neonates from ‘the unass cultures ‘within 12 11. Observations of
mortality (lack of movement or respiration) were made at the termination
of the 48 h tests with both cladoceran species. Acute 48 h LCSO values
were calculated with the TOXCALC program. 'This programnoriginated at U.S.
EPA and was modified for use in the laboratory at Michigan State
University. Unless noted above, all other culture and assay methods
followed those presented by Mount and Norberg (1984) and U.S. EPA (1985a).
Chironomus tentans

Chironomus tentans were originally obtained from Dr.
David White of the University of Michigan and have been
continuously cultured in the laboratory at MSU for over five years.
Cultures were maintained using the methods of Giesy et a1 (1988). Second
instar individuals (12 d post-hatching) were used for all tests. Culture
records indicate second instar C. tentans had a mean dry wt. of 0.503 mg
(n=7, SD=0.136 mg); while recent historical test data from the laboratory
indicate that individual control organisms had a mean dry wt. of 6.98 mg
(n=20, SD=0.94 mg) at test termination (22 d post-hatching).

Tests were conducted with individual second instar C. tentans
using the methods of Giesy et al. (1988) except that a 9:1 mixture of
HPLC-grade deionized water and Perrier mineral water was used in each test
instead of distilled water. Tests were conducted for 10 d at 22 1 1° C
with a 16L:8D photOperiod. At test termination, surviving larvae were

recovered, rinsed with deionized laboratory water and placed in aluminum

69

weigh boats. Larvae were dried at 80° C for 24 h and each larvae weighed
on a Mettler Model H54AR analytical balance. Results of C. tentans tests
are reported as mean percent inhibition in dry wt. gain at each station
relative to a control sediment.
Statistical Analysis

LCSO values from the toxicity tests were used for correlation
analyses with the chemical data. Statistical Analysis System (SAS 1988)
software was used to perform univariate Pearson product moment
correlations among the results of toxicity tests and chemical analyses of
sediments and pore waters. Correlation analysis results are reported only

if they were significant at the P s 0.05 level.

Results

Chemical Analysis
Organics

Analyses were conducted to determine solid phase sediment
concentrations for a total of 104 organic chemicals. Forty-one of these
104 compounds were not present in GCR sediments at concentrations above
the limit of detection (LOD). Analyses were subsequently conducted to
determine pore water concentrations of the 63 compounds observed at
concentrations greater than the LOD in the sediments. Of these 63
compounds, only 44 were observed in pore water samples at concentrations
greater than the LOD (Table 4). Concentrations of organic chemical
analytes detected in either whole sediment or pore water are listed in

Tables 5 and 6, respectively.

70
Table 4. Organic compounds analyzed for but not detected in sediments
or sediment pore waters from the Grand Calumet River, IN.
The limit of detection (LOD) is given for the matrix in

which the compound was not detected.

 

 

Parameter Sediment LOD Pore water LOD
mglkg ug/L

m-Chlorophenol 0.1

m-Cresol 0.01

p-Cresol 0.01

2,3-Dichlorophenol 0.01

2,5-Dichlorophenol 0.01

2,6-Dichlorophenol 0.1

3,4-Dichlorophenol 0.01

3,5-Dichlorophenol 0.01

2,3-Dibromophenol 0.01

2,4-Dibromophenol 0.01

2,5-Dibromophenol 0.01

2,6-Dibromophenol 0.01

3,4-Dibromophenol 0.01

3,5-Dibromophenol 0.01

2,3,4,5-Tetrachlorophenol 0.01

2,3,4,6-Tetrachlorophenol 0.01

2,3,5,6-Tetrachlorophenol 0.1

2,3,4-Trichlorophenol 0.01

2,3,5-Trichlorophenol 0.01

Table 4. (continued).

71

 

 

Parameter Sediment LOD Pore water LOD
mg/kg ug/L

2,3,6-Trichlorophenol 0.01

2,4,S-Trichlorophenol 0.01

2,4,6-Trichlorophenol 0.1

3,4,5-Trichlorophenol 0.01

Mercaptobenzothiazole 0.1

1,2,3,4-Tetrachlorobenzene 0.01

1,2,3,S-Tetrachlorobenzene 0.01

1,2,4,5—Tetrachlorobenzene 0.01

Benzidine 0.01

3,4-Dichloroaniline 0.1

3,3-Dichlorobenzidine 0.1

p,p'-DDD 0.1

Aldrin 0.01

Methoxychlor 0.01

m-Dichlorobenzene 0.1

p-Dichlorobenzene 0.1

1,2-Dichloropropane 0.1

1,3-Dichloropropane 0.01

1,3-Dichloropropene 0.1

Hexachloro-1,3-butadiene

Hexachloroethane 0.01

Pentachloroethane 0.1

1,2,3—Trichlorobenzene 0.1

Table 4. (continued).

72

 

 

Parameter Sediment LOD Pore water LOD
mg/kg ug/L

1,2,4-Trichlorobenzene 0.1

1,3,5-Trichlorobenzene 0.01

1,2,3-Trichloropropane 0.1

Trichloroethene 0.01

Butylbenzylphthalate 0.02

Diethylphthalate 0.02

Dimethylphthalate 0.02

Di-n-butylphthalate 0.5

Di-n-octylphthalate 0.02

1-Chloro-2,4-dinitrobenzene 0.01

1-Chloro-2,6-dinitrobenzene 0.01

1-Chloro-3,4-dinitrobenzene 0.01

1-Chloro-4-nitrobenzene 0.01

2,6-Dinitrotoluene 0.01

Nitrobenzene 0.01

Cresyldiphenyl phosphate 0.01

Trixylene phosphate 0.01

2,3,7,8-Dibenzo-p-dioxin 1 2.0

Dimethyl nitrosamine 0.3

 

1 - units for TCDD LOD = pg/L

73

 

 

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Concentrations of the various compounds present in sediments varied
greatly. Chemicals such as m-chlorophenol; 2,6-dichlorophenol; 2,4,6-
trichlorophenol; 2,3,5,6-tetrachlorophenol; 3,4—dichloroaniline; 3,3-
dichlorobenzidine; p,p'-DDD; tetrachloroethylene; 1,2,3-trichlorobenzene;
1,1,1-trichloroethane; di-Q-butyl phthalate; l-chloro-2-nitrobenzene and
2,4-dinitrotoluene were generally present in the low ug/kg (ppb) range
(Table‘S). Compounds exhibiting the greatest sediment concentrations were
the various polycyclic aromatic hydrocarbons (PAHs), total polychlorinated
biphenyls (PCBs, as Aroclor 1248), p,p’-DDE, toxaphene, p-chlorotoluene,
ethylbenzene and p-dichlorobenzene. These compounds were generally
present in the 2-20 mg/kg range although several of the PAHs were present
at concentrations as great as 100 mg/kg (Table 5). Percent TOC ranged
from 4.4 to 28.1% in the sediments while percent oil and grease ranged
from 1.6 to 13.5% (Table 5).

Most compounds observed in sediment pore waters were present in the
low ug/L range (i.e., 0-10 ug/L, Table 6). Notable exceptions included
phenol and the PAHs, phenanthrene and naphthalene. These compounds were
present at concentrations in pore water as great as 326, 245 and 452 ug/L,
respectively (Table 6). Several other compounds such as p-chlorotoluene,
fluoranthene, pyrene, 2,4-dichlorophenol, 2,4-dinitrophenol,
pentachlorophenol and biphenyl were observed in pore water at
concentrations in the 10—100 ug/L range. Pore water TOC concentrations
ranged from 18.7 to 52.8 mg/L (Table 6).

Metals
Detectable concentrations of most metals analyzed for were present

in all study site sediments (Table 7). Iron, magnesium and manganese were

84

 

 

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generally present in high mg/kg to low gm/kg concentrations in solid phase
sediments (Table 7). Of the metals of toxicological concern in aquatic
systems, zinc, lead and chromium were present at concentrations as great
as 5.23, 3.94 and 1.22 gm/kg, respectively. Copper, nickel and cadmium
concentrations were generally below 500 mg/kg.

Sediment AVS concentrations also are presented in Table 7 along with
the molar metals (Cd + Cu + Pb + Ni + Zn)/Avs ratios for each sediment.
The concentrations of metals in sediments used in the calculation of the
molar metals/AVS ratios were not determined by sequential extraction
during the AVS analysis as recommended by Di Toro et al. (1991). The
molar metal/AVS ratios for the sediments ranged from <0.l (i.e. no free
metal) at site UG-l to 144.2 (potentially a large amount of free metal) at
site UG-Z. Only three sites, UG—2,4 and 5 had molar metal/AVS ratios
>2.0. Although the metals concentrations reported here were not
determined by the sequential extraction (cold, weak acid; Di Toro et al.
1992), the harsher extraction technique used in the analyses reported here
would result in conservative, worst case estimates of metal
bioavailability.

Copper and zinc were the only metals of toxicological concern
typically present in pore waters at concentrations greater than the ICP
LOD for these metals (0.005 mg/L) (Table 8). Aluminum and arsenic were
the only other two metals or metalloids of concern observed in any pore
water sample at concentrations greater than their respective LODs of 0.100

and 0.050 mg/L (Table 8).

87

 

 

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Miscellaneous parameters
Alkalinity and hardness in several of the samples were elevated
(1040-1500 mg/L as CaCO3, Table 9) above the values commonly reported for
surface waters while conductivity and Ph were within the ranges reported
for surface waters (Cole 1988). Cyanide and H28 were present in
detectable, although relatively low, concentrations but unionized N83

concentrations ranged from 0.2-8.1 mg/L (Table 9).

Toxicity Tests
Photobacterium phosphoreum

Five, 15 and 30 minute ECSO values were calculated for sediment
pore waters osmotically adjusted with both NaCl and sucrose (Table 10).
Fifteen minute E050 values for NaCl- and sucrose adjusted pore waters
ranged from 0.3-93.8 % and O.2-39.5 %, respectively; Pore waters
osmotically adjusted with sucrose were generally equitoxic or less toxic
than NaCl adjusted pore waters with the exception of pore waters from
sediments collected at sites UG—7, 8, 9, and 10. Pore waters at these
sites were from 1.3-2.4 times more toxic when osmotically adjusted with
sucrose (Fig 9). Both sucrose and NaCl-adjusted.Microtoxo tests generally
exhibited similar ECSO values at 5, 15 and 30 minutes. However, pore
water from sites UG-2, 4 and 6 became more toxic with time in both
sucrose- and NaCl—adjusted tests while pore water from sites 06-9 and 11
became more toxic with time only in the sucrose-adjusted tests.

Several significant correlations were observed between the results
of the Microtoxo pore water tests osmotically adjusted with NaCl and the
results of the pore water organic chemical analyses. Correlation

coefficients (r) between the Microtox® tests and 9:, and p-chlorophenol,

94

 

 

 

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g-cresol, PCBs (as Aroclor 1248), p,p’-DDE, chloro- and ethylbenzene and
benzo(k)fluoranthene ranged from 0.64-0.81. The results of sucrose-
adjusted tests were also correlated with these same compounds in pore
water. However, the total number of significant correlations was less for
the sucrose-adjusted tests and the observed correlations were generally
with ECSO values calculated after 15 and 30 minute exposures. The
significance of these correlations was questionable due to the fact that
if increasing chemical concentrations were related to greater toxicity
(i.e., smaller ECSO value) the correlations should have been negative
rather than ‘positive. bk: biologically' or statistically significant
correlations were observed between the results of MicrotoxO test results
and the results of the ICP metals or the miscellaneous chemical analyses
of the pore waters.

Toxic units (Sprague and Ramsey 1965) were calculated for pore water
concentrations of unionized ammonia, Q- and p-chlorophenol, Q-cresol,
ethylbenzene, Chlorobenzene, 2,4-dinitrophenol, 1,1,l-trichloroethylene,
pentachlorophenol, phenol and naphthalene by dividing the pore water
concentration by the 15—min ECSO value from NaCl~adjusted Microtoxo tests
of the pure chemical (Kaiser and Ribo, 1988). Ammonia, naphthalene and
phenanthrene were the only compounds present at >0.1 TU in the pore water
samples. The calculated TU units for these three compounds were then
summed and compared to the measured TU units for each sample (Table 11).
Measured TU were calculated by dividing 100% by the lS-min EC50 from NaCl-
adjusted pore water Microtox® tests. The total calculated TU for the pore
water samples ranged from 0.2 — 4.6 while the measured TU ranged from 1.1

- 333.3. Comparison of calculated versus measured TU for sucrose-adjusted

98

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tests was not conducted due to the paucity of data for sucrose-adjusted

tests of pure chemicals.

D. magna and C. dubia

 

Pore water ECSO values for D. magna ranged from 5.5% pore water for
site 06-8 to >100% pore water at sites 06-2, 6 and 7 (Table 10).
_griodaphniaggubia 48 h ECSO values ranged from 3.2% pore water at site
06-8 to >100% at sites UG-2 and 7 (Table 10). With the exception of sites
06-3 and 4, where pore water was approximately 3-4 times more toxic to Q;
maggg than to C. dubia, ECSO values indicated C. dubia were always as
sensitive or more sensitive than D. magna to the toxic effects of the
sediment pore waters.

Daphnia magna 48 h ECSO values were negatively correlated (r = -
0.65) with concentrations of dieldrin in pore water while C. dubia 48 h

ECSO values were negatively correlated with pore water concentrations of

2, 4-dinitrophenol, dieldrin, fluoranthene, phenanthrene and
benzo(k)fluoranthene (r = -0.70, -0.67, -0.72, -0.74 and -O.65,
respectively). No significant correlations were observed between

concentrations of metals measured by ICP in pore water and the results of
the acute daphnid tests. However, significant correlations were observed
between pore water concentrations of unionized ammonia and alkalinity and
the 48 h ECSO values for both D. magna (r = -O.69, -O.6l) and C. dubia (r
= -0.60, -O.74).

In a manner analogous to that used for the MicrotoxO tests, TU were
calculated for pore water concentrations of unionized ammonia, alkalinity
(bicarbonate ion), copper, zinc, naphthalene and phenanthrene based on 48

h LCSO values for tests with the pure compounds (Table 12). Forty-eight

100

 

 

 

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103

hour C. dubia and D. magna LCSO values for unionized ammonia were obtained
from U.S. EPA (1991) and U.S. EPA (1985b), respectively. Ceriodaphnia
dubia LCSO values for unionized ammonia used for TU calculations were pH-
specific based on the greatest pH value measured during the test of each
pore water sample (U.S. EPA 1991) while the D. magna LCSO value was the
mean of five separate L050 values from tests conducted at an approximate
pH of 8.0 (U.S. EPA 1985b). Forty-eight hour LCSO values for alkalinity
(bicarbonate ion), copper and zinc, and the PARS used in TU calculations
were obtained from Hoke et al. (1992b), U.S. EPA (1980b,c); and Millemann
et al. (1984), respectively. LCSO values for the metals were at an
approximate water hardness of 200 mg/L as CaCo3. Based on the pure
chemical data for the two cladoceran species, the total calculated TU for
the pore water samples ranged from 0.4 - 6.9, while the measured TU ranged
from 1.0 - 18.2 (Table 12). Unionized ammonia and bicarbonate ion
accounted for the greatest proportion of the total TU calculated for both
D. magna and C. dubia (Table 12).
Chironomus tentans

Percent inhibition in dry weight gain in C. tentans exposed to
sediments from the Grand Calumet River ranged from 37% at 06-7 to 100% at
sites UG-l, 3 and 4. Only one site, UG-7, exhibited a % inhibition of dry
weight gain of less than 90%. Because the range of response for C_.
tentans was very small if the data from site UG-7 were excluded from the
analysis, no further attempts were made to correlate results with the
results of chemical analyses or to calculate TU based on observed chemical

concentrations.

Discussion

The sediments and sediment pore waters from a number of sites in the
Grand Calumet River system of northwest Indiana contained detectable
concentrations of a wide variety of organic chemicals and metals. Simple
visual inspection of sediments from the system, however, led to the
conclusion that one of the primary contaminants in the system was oil and
grease. The concentrations of this broad category of petroleum
hydrocarbons ranged from 1.6-13.5% on a sediment dry weight basis and
undoubtedly, even in the absence of other contaminants, would exert a
strong influence on the presence and/or distribution of benthic
macroinvertebrates in the system. Previous studies have reported on the
depauperate nature of the benthic’macroinvertebrate community in theeGrand
Calumet River (IDEM 1988). The oil and grease content of the sediment
also was likely at least partially responsible for the high mortality and
inhibition of dry weight gain observed in the C. tentans test.

Petroleum hydrocarbons are a major contaminant of aquatic systems
with as much as 6 million metric tons of these products introduced into
aquatic ecosystems worldwide on a yearly basis (NAS 1975). Petroleum
hydrocarbons are a mixture of hydrocarbons and trace elements which can
effect almost all aquatic macroinvertebrates by direct exposure (Petrakis
and Weiss 1980, Hoehn et al. 1974). Petroleum hydrocarbons may affect
macroinvertebrates by coating gills or other body surfaces responsible for
cutaneous respiration thus limiting oxygen exchange, by direct toxic
action, by bioaccumulation of hydrocarbons or by blanketing the substrate

and preventing colonization. The impact of a massive crude oil spill on

104

105

macroinvertebrate fauna in a stream system*was described by Crunkilton and
Duchrow (1990). A three orders of magnitude decrease in expected numbers
of organisms was observed for a month after the spill. Species diversity
and the number of sensitive species (mayflies, stoneflies) were decreased
for almost a year after the spill. The authors observed that the visible
presence of petroleum hydrocarbons in the stream substratum was an
effective predictor of effects on the benthic macroinvertebrate community
and that total flow volume and occurrence of scouring due to floods were
the major factors controlling recovery. Schloesser et al. (1991) also
observed that the visible presence of petroleum hydrocarbons in sediments
was a good indicator of effects on Hexagenia limbata abundance in the
upper Great Lakes connecting channels. In a study of the effects of
oil/gas field produced.water on the benthic macroinvertebrate community in
a small gradient estuary, Nance (1991) reported that sediment oil content
was a more important determinant of population abundance than salinity.
A sediment hydrocarbon concentration of 2.5 mg/gm dry weight sediment
(0.25%) was observed to cause a minor depression of macroinvertebrate
abundance while 5.0 mg/gm (0.5%) was observed to cause major effects on
abundance. Assuming the presence of no other contaminants, the observed
oil and grease content of sediments in the Grand Calumet River (1.6-13.5%)
alone appears to have been sufficient to prohibit the development of a
viable benthic macroinvertebrate community in the sediments. Based on
personal observations made by the authors during field sampling on the
Grand Calumet River, it also appeared that petroleum hydrocarbon inputs to
the system were a continuing, rather than simply a historical, problem.

Although petroleum hydrocarbons in sediments from the study area

appeared to affect the results of the C. tentans bulk sediment tests, the

106

physical effects of this suite’ of compounds should have» been less
important in determining the results of the pore water tests with
2. phgspgoreum, D. magna and C. dubia. The limited water solubility of
most petroleum hydrocarbons and the filtration step (Whatman GF-F, 0.7 pm
nominal pore size) used in the preparation of sediment pore waters could
have removed all but the most water soluble petroleum hydrocarbons (i.e.
naphthalene, phenanthrene) from the pore waters. Therefore, the results
of the pore water tests provided information on the toxicity of water
soluble compounds which likely was not provided by the C. tentans test
results due to the confounding presence of large amounts of insoluble
petroleum hydrocarbons in the sediments.

The results of separate sucrose and NaCl osmotically-adjusted
MicrotoxO tests on a sample may be helpful in determining the causes of
observed sample toxicity (Hinwood and McCormick 1987, Ankley et al. 1990a,
Hoke et al. 1992a). Changes in sample toxicity over the 30-min course of
the full Microtoxo test also may be helpful in identifying potential
toxicants. Several ionic chemicals (certain metals, ammonia) have been
observed to be more toxic in Microtoxo tests osmotically adjusted with
sucrose (Hinwood and McCormick 1987, Ankley et al. 1990a) while chlorine
was more toxic in NaCl-adjusted tests (Ankley et al. 1990a). Some metals
also exhibit a pattern of increasing toxicity with time while the toxicity
of ammonia and many organic chemicals is constant over time. In general,
little difference was observed in the toxicity of GCR pore waters in
Microtoxm tests osmotically-adjusted with sucrose or NaCl. This may
indicate that ionic toxicants in the pore water were not responsible for
the majority of the observed toxicity. Pore waters from UG-7, 8, 9 and 10

were more toxic in tests osmotically-adjusted with sucrose while pore

107

waters from UG-ll and UG-13 were more toxic in tests osmotically adjusted
with NaCl. Measurable concentrations of Cu and Zn were present and pH and
unionized ammonia also were greater in pore waters from site 06-? through
10 (Tables 8, 9). Zn is known to be more toxic at higher pH (Ankley et
al. 1991) and higher pH also shifts the ammonia equilibrium towards
greater prOportions of unionized ammonia which is the more toxic form of
ammonia to most aquatic species (U.S. EPA 1985b). No increase in toxicity
with time was observed in any Microtox‘ID test conducted on these pore
waters. These observations coupled with the known presence of municipal
wastewater treatment plant discharges on this section of the GCR (IDEM
1988) strongly implicate unionized ammonia as one of the principal
contaminants causing the effects observed in Microtoxo tests of pore*water
from sites UG-7 through 10.

Greater concentrations of metals in pore water (i.e., UG-2,3,4) or
greater concentrations of ammonia (i.e., UG-1,4) were observed but pore
water from these sites also generally had lesser pH values. The increase
in toxicity with time observed over the course of both NaCl-and sucrose-
adjusted Microtoxo tests of pore waters from sites UG-2,4 and 6 suggests,
however, that metals were responsible for at least part of the observed
toxicity. No explanation is evident for the greater toxicity in NaCl-
adjusted tests of pore waters from sites UG-ll and 13 because complete
chemical analyses were not conducted on sediments or pore waters from
these sites.

Based on the calculated versus measured TU from Microtoxo lS-min
E050 values for the study site pore waters, a large proportion of the
observed toxicity was unaccounted for at most sites, except sites UG-7

through 06-10. The contribution of the PAH, phenanthrene, to the total

108

calculated TU for a given site was generally equal to or greater than the
contribution of ammonia. Naphthalene TU were generally equal to or less
than the unionized ammonia TU at each site. Copper and zinc TU from NaCl-
adjusted MicrotoxO tests were generally 5 0.1 TU for each site (data not
shown) and it was not possible to calculate fluoranthene TU because no
Microtoxo toxicity data exist for this compound. Calculated TU exceeded
measured TU at sites UG-8,9 and 10, however, when the contribution of the
PAHs was dropped from the calculated TU there was much better agreement
between total calculated and measured TU. An additional factor (i.e.,
pore water bicarbonate ion concentration) discussed below appeared to
cause toxicity in the cladoceran pore water assays. However, no effect on
the Microton test was observed in a series of experiments designed to
evaluate the effects of water hardness and alkalinity at concentrations up
to 3000 mg/L as CaCO3, respectively (data not shown). At sites UG-l
through UG-6 and UG-ll through UG—13, it appeared that unidentified
compounds caused a large portion of the observed effects in the MicrotoxO
tests.

In 48 h pore water tests, C. dubia was, with the exception of site
UG-4, always equally or more sensitive to the effects of pore waters than
was D. maqna. The difference in response between the two species was
generally small, perhaps indicating that the cladocerans were responding
to the same contaminants with C. dubia being slightly more sensitive. At
sites UG-3,5,6 and 9, however, C. dubia.was approximately 1.3-6 times more
sensitive to the effects of the pore waters than was D. magna. This
variability also could have been the result of the effect of pH on the
relative species sensitivity to chemicals. C. dubia sensitivity to

unionized ammonia exhibits a greater pH dependence than does that of Q;

109

maqna (Ankley et al. 1991, U.S. EPA 1985b). Lack of sensitivity to pH
effects on ammonia toxicity also has been observed for Hyalella agtgca
(Ankley et al. 1991). Sensitivity to some metals may exhibit a similar
pattern since the zinc 48 h LCSO for C. dubia tested in very hard
reconstituted water at a pH of 8.0-8.5 was 95 pg/L (Ankley et al. 1991)
while the 48 h LCSO for D. magna tested under similar conditions was 655
pg/L (U.S. EPA 1980c). This factor alone could explain a large portion of
the variation in the responses of the cladoceran tests, especially in
light of the observations that the pore water pH values at sites UG-3,6
and 9 were 8.4, 8.0 and 8.4, respectively (pH was not measured on site UG-
5 pore water) while pore water zinc concentrations were 28, 490 and 114
pg/L, respectively.

Total calculated TU for D. magna and C. dubia exceeded the measured
TU at 8 and 10 of the 13 sites, respectively. However, total calculated
TU for both species assumed that measured pore water concentrations of
copper and zinc were present as the bioavailable free ion. This is
unlikely due to the pH dependent nature of chemical speciation and the
probable presence of many inorganic and organic ligands which could bind
free metal ions in pore waters. A decrease in TU attributable to copper
and zinc would result in a decrease in the total calculated TU for each
pore *water and bring total calculated and. measured TU into closer
agreement.

Toxic units attributable to phenanthrene and naphthalene in W
pore water tests were less than TU attributable to unionized ammonia and
bicarbonate ion” The TU attributable to unionized ammonia and bicarbonate
ion also accounted for the greatest proportion of the total calculated TU

units in the C. dubia tests. Pore water concentrations of unionized

110
ammonia and alkalinity also were negatively correlated with the 48 h ECSO
values from the cladoceran tests, which indicates that greater
concentrations of these chemicals caused greater toxicity (lesser LCSO
value) of the pore waters to the cladocerans.

Ammonia is known to be toxic to cladocerans (U.S. EPA 1985b) and
several authors have recently reported on the effects of carbonate
alkalinity on cladocerans (Cowgill and Milazzo 1991, Hoke et al. 1992b).
In an investigation of the toxicity and potential mode of action of the
bicarbonate ion, Hoke et al. (1992b) demonstrated that bicarbonate ion was
toxic to both D. magna and C. dubia and that sufficient bicarbonate ion
was present in several GCR-IHC pore waters to cause a large portion of the
observed toxicity of the pore waters to these species. These observations
reinforce the importance of unionized ammonia and alkalinity in the TU
calculations for the cladocerans.

It was not possible to calculate the potential contribution of
naphthalene and phenanthrene to the total measured TU for g&_ggbi§ tests
because no 48 h reference toxicant LCSO data for these chemicals were
available for C. dubia. The total calculated and measured TU for both
cladoceran species at each site were generally observed to be in good
agreement with the exception of site UG-8 where the total calculated TU
for both species were much less than the total measured TU. One or more
chemicals not measured during this study must have been present in the
pore water from this site at concentrations sufficient to affect both
cladoceran species.

The relative sensitivities of the tests used during this study were
very different. The C. tentans test appeared to be affected primarily by

the physical presence of petroleum hydrocarbons in the sediments, which

111

also obscured any potential impacts on this species of contaminants in the
sediment pore waters. The cladoceran tests appeared to be affected
primarily by concentrations of unionized ammonia and bicarbonate ion in
the pore water. Unionized ammonia and the PAHs naphthalene and
phenanthrene also were possibly important determinants of the toxicity of
sediment pore waters in the Microtoxo test. However, at several sites,
observed toxicity in both the Microtoxo and cladoceran tests could not be
accounted for solely on the basis of the concentrations of the chemicals
used in the TU calculations presented here.

Reference toxicant data necessary for calculation of TU for a wide
variety of chemicals are not available for either the MicrotoxO or the Q;
m tests. The chemical analysis of study site sediments and pore
waters included only 104 analytes. Although a large number of compounds
were analyzed for in the samples, the potential for additional compounds
to be present in the samples is great. Both of these factors limit the
application of the TU approach, without a concurrent toxicity
identification evaluation (TIE) effort Ankley et al. 1991), for
determining the compounds causing the observed toxicity in tests of whole
sediments and sediment pore waters.

In addition, the TU approach assumes additivity of chemical effects
and complete bioavailability of all measured analytes, which may or may
not be a valid assumptions in complex environmental matrices, depending on
the chemicals under consideration. Antagonism and synergism are equally
plausible consequences for the expression of the toxicity of chemical
mixtures. If chemicals exhibit their effects via different modes of
action, the assumption of additivity of effects will not be met and

calculated TU could be over-estimates of chemical effects. This

112

phenomenon, in addition to the various factors affecting chemical
bioavailability, the lack of reference chemical toxicity data and the
incomplete nature of the chemical analyses previously discussed,
contribute to the observed discrepancy between calculated and measured TU.

The results of the study presented here highlight the necessity and
importance of the battery of tests approach for toxicity assessments which
has been previously advocated by other investigators (Slooff et al. 1983,
Millemann et al. 1984, Williams et al. 1986, Dutka and Kwan, 1988, Giesy
and Hoke 1989). Each test responded in a unique way which emphasized the
intrinsic sensitivity of the test, the differences in exposure routes
among tests and the fate of individual chemicals or classes of chemicals
in the samples. The study also demonstrated the advantages and
limitations of a toxic units approach by itself for evaluating the
potential causes of the observed toxicity in the various assays. A more
utilitarian approach to assessments of contaminated sediments could
incorporate the battery of assays approach and sediment TIE, including the
evaluation of calculated versus measured TU, to identify and confirm the
causes of observed sediment and pore water toxicity and, ultimately, guide

site remediation efforts (Ankley et a1. 1991).

Acknowledgements

Inductively-coupled argon ;plasma spectroscopy and sediment, TOC
analyses were conducted by the MSU Animal Health Diagnostics Laboratory
and the MSU Soil Testing Laboratory, respectively; G. Ankley, V. Mattson,
T. Burton, R. Merritt and several anonymous reviewers provided helpful
comments which improved the final version of the manuscript. Jane
Norlander, Debra Williams and Jane Thompson typed various drafts of the
manuscript. Funding for this research was provided by the U.S.

Environmental Protection Agency, Great Lakes National Program Office,

Chicago, IL.

113

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126
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CHAPTER 3

Bicarbonate as a Potential Confounding Factor in Cladoceran Toxicity

Assessments of Pore Water from Contaminated Sediments

127

Introduction

Toxicity assessment of contaminated sediments has become de rigueur
in planning for remedial actions at International Joint Commission (IJC)
Great Lakes "Areas of Concern" (AOC) (Hileman 1988) and at many U.S. EPA
Superfund sites (Warren-Hicks et a1. 1989). Sediments contaminated by a
wide array of metals and organic chemicals are an acknowledged legacy of
the explosive post-World War II growth in the chemical industry and the
continued reliance on fossil fuels for energy production. Approaches
traditionally used to evaluate contaminated sediments include: 1)
chemical analyses for total sediment concentrations of compounds or
elements of concern, 2) toxicity tests of bulk sediments or aqueous
extracts of sediments using benthic and upper water column test species,
respectively, and 3) field surveys of in-situ effects such as changes in
benthic macroinvertebrate community structure or the occurrence of
histopathological lesions in benthic fish species (for reviews see Giesy
and Hoke 1989, Chapman et al. 1987). In the present study, we focus on
the potential implications of, and one of the problems involved in, the
use of pore water (interstitial water) toxicity tests with cladocerans as
surrogates for bulk sediment tests with benthic invertebrates.

Pore water from contaminated sediment is used as an exposure medium
for toxicity tests (Giesy and Hoke 1989). Comparisons have been made of

the results of tests which evaluated the toxicity to Hyalella azteca and

128

129
Lumbriculus variegatus of bulk sediment, pore water and elutriate from the
same samples (Ankley et al. 1991). These comparisons and the results of
other studies (Nebeker et a1. 1984, van de Guchte and Maas-Diepeveen 1988,
Knezovich and Harrison 1988) suggest that pore water exposures with
benthic species are good surrogates for bulk sediment exposures.

The underlying assumption in the use of pore water exposures as a
surrogate for bulk sediment exposures is that organisms receive most of
their exposure to toxic substances in sediments through contact with the
pore water (Nebeker et al. 1984, Ankley et al. 1991). The relationship
between pore water and bulk sediment exposures has not been clearly
established (Cairns et al. 1984, Nebeker et a1. 1984) and many physico—
chemical processes can be expected to affect the partitioning of metals
and xenobiotics from the solid phase into the pore water (Sinex et al.
1980, Knezovich et al. 1987, U.S. EPA 1989, Di Toro et al. 1990).
Additional assumptions inherent in the use of pore water as a test medium
are that basic chemical properties of pore water, such as major ion
concentrations and/or ratios, are within the physiologically tolerated
ranges of the test species and that the pore water extraction technique
does not cause artifacts in the pore water chemistry. Recent reviews of
the advantages and disadvantages associated with various methods of pore
water preparation can be found in Adams (1991) and Schults et al. (1992).

During an assessment of the toxicity of sediments at the Grand
Calumet River-Indiana.Harbor Canal (GCR-IHC) IJC4AOC, several observations
caused concern over the effects of chemical properties of the pore waters
on the results of toxicity tests with cladocerans. Elevated pore water
alkalinity values (Table 13), gas bubble formation, white crystalline

residue on the inside of test chambers and increases in pore water pH in

130

 

 

 

 

 

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132

test vessels during cladoceran toxicity tests led to the hypothesis that
the test results might be affected by the altered character of the
carbonate-bicarbonate buffer system in the pore waters. At the time the
investigation was begun, no information existed in the literature
concerning the toxicity of bicarbonate ion to cladocerans although some
information was available on the toxicity of C02 (Mount and Anderson-
Carnahan 1988). It also was known that C02 could be used to anesthetize
zooplankton prior to preservation (Gannon and Gannon 1975). Limited
information also was available on the toxicity of bicarbonate ion to fish
(Beatty 1959, Mitchum 1960).

This investigation was conducted to examine the toxicity of the
bicarbonate ion to D. magna and C. dubia. The specific objectives of the
study were: 1) to determine the effects of bicarbonate ion concentrations
on D. magna and C. dubia using NaHCO3 and NaCl as reference toxicants, 2)
to compare calculated concentrations of HCO3' in sediment pore*waters from
the AOC study area to concentrations shown to cause adverse effects on Q;
magpa and C. dubia in laboratory exposures, and, if HCO3'-induced toxicity
was observed, 3) to propose potential mechanisms for the observed HCO3'-
induced effects.

X—ray dispersivexnicroanalysis previously has been used to study the
role of various invertebrate tissues in metal storage and the effects of
different environmental levels of metals on metal localization in
invertebrates (Krantzberg and Stokes 1990, Ballan-Dufrancaisiet al. 1985).
Investigations also have been conducted.on diffusible ions in invertebrate
osmoregulatory systems (Marshall and Wright 1973, Roomans 1988a), and in
bulk specimens of skeletal muscle (Zierold and Schafer 1978). Because of

its previous use in the examination of physiological levels of diffusible

133
ions, X-ray dispersive microanalysis was chosen as the simplest and most
«direct method for monitoring the potential effects of bicarbonate ion on

‘the relative concentrations of several physiologically important elements

in D. magna.

Materials and Methods

Toxicity of Na+ and HCO3‘

Forty-eight hour, acute tests (U.S. EPA 1985) were conducted with
Baker'reagent-grade NaHCO3 and NaCl to establish the toxicity of HCO3' and
Na+ to D. magna and C. dubia. These exposures were initiated with <24 h
old.neonates from laboratory cultures of each species and also with six or
seven-day old, sub-adult D. magna. Six or seven-day old D. magga were
initially necessary to provide specimens of sufficient size for
manipulation in the preparation procedures for the x-ray microanalysis and
were specifically selected for further x-ray microanalysis because of
their larger size. Test organisms were not fed and the temperature and
photoperiod were maintained at 25°C and 16L:8D, respectively, during all
tests. New NaHCO3 and NaCl stock solutions were prepared for each test
(3000 mg/L and 8000 mg/L nominal concentrations, respectively). A 0.56X
dilution series was used to prepare dilutions of the stock solutions for
testing. A minimum of five toxicant concentrations and a control were
prepared for each test. Five replicates, each containing five organisms,
wwere tested for each toxicant concentration or control. The laboratory

0 mineral

culture/diluent water was prepared by mixing bottled Perrier
water and HPLC-grade laboratory water (1:9 v/v) followed by vigorous

aeration for 24 h to ensure equilibration of gases prior to culture or

134
test use (U.S. EPA 1989). Two pg/L each of cyanocobalamin (Vitamin 812)
and selenium, as sodium selenate, also were added to the culture/diluent
water (Lanno 1989). Control survival in toxicity tests was always greater
than 90%. Forty-eight hour LCSO values were calculated using the binomial

test and are reported as nominal HCO3' or Na+ concentrations in mmoles/L

(Table 14).

Calculation of Free 002 and HCO3_

A series of twenty-six, 48-h acute pore water tests were performed
xwith D. magna and C. dubia as part of the original AOC sediment toxicity
.assessment (Hoke et. al 1992). Data from these exposures subsequently
\were used to calculate theoretical equilibrium concentrations of free 002
{and HCO3' in the pore waters for comparison with effects concentrations
:from the literature or the reference toxicant tests described above. Free
<302 and HCO3' concentrations at equilibrium were calculated using
‘temperature (25°C), pH and alkalinity values measured in the 100% pore
‘water exposure treatment from the cladoceran tests (Table 13). Alkalinity
‘was measured using a Hach kit (accuracy = t 0.4 meq/L) and pH was measured
tasing an Orion Model 701A ionanalyzer equipped with. a glass Ross
CHDmbination pH probe. The following equations (Harvey 1957) were used to

calculate free 002 and HCO3' concentrations, respectively.

H2 (1)
KIN-I + 2K2)

 

«B; where p =

X
ll

aarxci

2H (2)

H+2K2

 

135

where:

 

 

H = [H‘]
a = [HC03'] + 2 [CO,’] = Measured alkalinity
x = [HZC'O3] , includingfree CO2
y = [HC03‘]
[HCO3'] [H‘]
R; =
[H2C03]
CO'Z 0
K2 = [ 3 H.111
[HC'03]

X-ray Dispersive Microanalysis
To investigate the effects of HCO3' on the internal ion balance (Na,
Si, Cl, Ca) of D. magna, several 48-h tests in which NaHCO3 was used as a
toxicant were initiated with 6 or 7-d old, sub-adult D. magna. Three
experiments were conducted with NaHCO3. Additional experiments were
conducted with pore water collected from one of the original AOC sediment
samples (Location UG-9) due to its high alkalinity and calculated HCO3'
Concentration, and with NaSCN, a metabolic inhibitor of Cl' uptake
(Epstein et a1. 1973). Live D. magna were removed from the tests for
Subsequent X-ray analysis after 16-20 h exposure to NaHCO3 or Location UG-
9 pore water. Organisms were removed for analysis after 2 h of exposure
to NaSCN. Forty—eight hour exposures to the highest concentrations of
either NaHCO3 or pore water were lethal to D. magna. Organisms to be
analyzed by X-ray microanalysis were exposed for a shorter period of time
because the variable of interest was change in elemental peak to
lba-Cflmround (P/B) ratios in living organisms as a result of exposure to

Ncho3, pore water or NaSCN.

l A

136
Table 14. Toxicity to D. magna and C. dubia of Na+ as NaCl or NaHCO3 and

of HCO3' as NaHCO3.

 

 

Mean 1 SD
48 h LCSO LC50
Species Age Compound (mmoles/L) (mmoles/L)
D. magna <24 h NaCl 85.7
81.3 t 6.2
D. magna1 <24 h NaCl 76.9
D. magna2 4th instar/ NaCl 57.9 --
adult
C. dubia <24 h NaCl 14.3
13.5 t 1.2
C. dubia <24 h NaCl 12.6
D. magna <24 h NaHCO3 16.6
15.1 1 2.2
D. magna <24 h NaHCO3 13.5
D. magna 6 d NaHCO3 21.2 --
D. magna 7 d NaHCO3 26.3
20.6 i 8.1
D. magna 7 d NaHCO3 14.9
C. dubia <24 h NaHCO3 13.8
12.8 i 1.5
C. dubia <24 h NaHCO3 11.7

 

Mount and Anderson-Carnahan 1988.

2 Dowden 1961.

137

Whole D. magna (generally 5 per treatment) were removed from the
tests and mounted on carbon planchets. A 1:1 (v/v) mixture of Tissue Tek
and graphite was used to mount the daphnids on the carbon planchets and to
attach the carbon planchets to electron microscopy stubs. The Tissue Tek-
graphite mixture did not interfere with subsequent X-ray analysis of
D. magna (R. Hoke, unpublished data). For preliminary experiments with
post-NaHCO3 exposure organisms, traditional cryogenic preparation and
fracturing techniques (EM Scope 2000) were used with etching at -20°C for
15 min if ice crystals were present in the sample (Boekestein et al. 1980,
Marshall 1988). Because these fracturing techniques were cumbersome and
frequently displaced organisms from the planchets, in subsequent
experiments organisms were mounted on planchets and then plunged into
liquid nitrogen for 15 sec. The carapace of each D. magna was then gently
sliced away with a razor blade to reveal the internal body structure and
facilitate penetration of the X-ray beam.

X-ray microanalyses were performed with a JEOL JSM-3SC scanning
electron microscope equipped with a cryostage and a Tracor Northern Series
II X-ray microanalysis system with software. Each measurement was
acquired at an accelerating voltage of 20.0 Kev for a preset scanning time
of 60 s. Results are reported as P/B ratios (Boekestein et al. 1980,
Marshall 1988, Roomans 1988b). Elemental measurements were conducted at
five randomly chosen locations within the head and upper body area for
each of five D. nmgna from control and either, low and high, or low,
medium and high experimental treatments. .Absolute peak height or emission
intensity data were not used due to the lack of analytical standards in an

appropriate matrix. The same analytical constraints were operative among

138
all measurements conducted as part of each experiment; therefore, P/B
ratios and changes in these ratios as a result of experimental treatments,

relative to controls, were the data of interest in all experiments.

Data Analysis

X-ray dispersive microanalysis results were analyzed with a nested,
fixed effects ANOVA design (Petersen 1985) using the Statistical Analysis
System (SAS 1985) software. For the ANOVA, the x-ray microanalysis
measurements were nested within experimental units (i.e. individual
Daphnia magna). In the variance component analysis, three separate
aspects of experimental variance were examined: 1) treatment variance, 2)
organism (among daphnid within treatment) variance and measurement (among

measurement within daphnid) variance.

Results

Toxicity of Na+ and Hco3'

Forty-eight hour LC50 values for D. magna and C. dubia tests with
NaHCO3 and NaCl are presented in Table 14. Based on mean LC50 values,
Qeriodaphnia dubia neonates (< 24-h old) were approximately six times more
sensitive to Na+ as NaCl than D. magna neonates of similar age. However,
24-h old C. dubia neonates were only slightly more sensitive (~ 1.2 X)
than D. magna neonates to Na+ as NaHCO3. Little difference was observed
in the toxicity of the two compounds to <24-h old C. dubia neonates.
Daphnia magna neonates were approximately 5.4 x more sensitive to Na+ as

NaHCO3 than as NaCl. Based on literature values for adults (Dowden 1961),

139
D. magna neonates were approximately 1.4 X less sensitive than adults to
NaCl while adults were approximately 1.4 X less sensitive to NaHCO3

(Table 14).

Calculation of Free C02 and HCO3'

The results of theoretical equilibrium concentration calculations
for C02 and HCO3' in the AOC sediment pore water samples are presented in
Table 13. Calculated concentrations of C02 ranged from 0.0-1.5 mmoles/L

while calculated HCO3' concentrations ranged from 1.2-29.9 mmoles/L.

X-ray Dispersive Microanalysis

Mean P/B ratios (1 SD) for Na+, Si+4, Cl", and Ca+2 in daphnids
analyzed from each experiment are presented in Table 15. Within an
experiment and within the individual treatments comprising an experiment,

+2

relative levels of Ca were the most variable parameter measured based on

standard deviation of the means while Si+4 was the least variable. The
variability of relative levels of Na’r and Cl' among experiments was
similar but Cl‘ levels were generally more variable within an experiment.
The results of variance component analyses from the ANOVA’s of the x-ray
microanalysis data are presented in Table 16. The proportion of the total
variance accounted for by among treatment effects is presented as well as
the variance accounted for by among organism (among daphnid) effects
within an experimental treatment and the among measurement (measurements

within an individual daphnid) variance. Variation in the Cl" P/B ratio

among treatments was statistically significant in each experiment. P/B

140

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experiment 1 due to inadequate replication.

* denotes a significant ANOVA F value at a =

Variance component analysis was not possible on the

0.05.

Table 16. Variance component analyses of results from analysis of
variance of X-ray microanalysis data.
% Total Variance2
Experiment No.1,

Toxicant Source df Na+ Si+4 Cl’ a+2
No. 2, Total 29 100.0 100.0 100.0 100.0
NaHCO3 Treatment 2 9.8 0.1 86.9* 7.3

Organism 3 0.0 0.0 0.6 10.1
Measurement 24 90.2 99.9 12.5 82.6
No. 3 Total 74 100.0 100.0 100.0 100.0
NaHCO3 Treatment 2 18.7* 0.0 40.5* 21.5*
Organism 12 19.3* 19.5* 24.6* 11.6
Measurement 60 62.0 80.5 34.9 66.9
No. 4 Total 74 100.0 100.0 100.0 100.0
UG-9 Pore Treatment 2 1.1 19.3* 24.8* 0.0
Water Organism 12 5.4 0.0 18.3* 40.2*
Measurement 60 93.5 80.7 56.9 59.8
No. 5 Total 99 100.0 100.0 100.0 100.0
NaSCN Treatment 3 13.8 0.0 53.2* 5.2
Organism 16 17.5* 44.9* 8.3* 0.0
Measurement 80 68.7 55.1 38.5 94.8
1

results of

143
ratios for Na+ and Ca++ also were significantly different among treatments
in the No. 3- NaHCO3 experiment, as was the Si+4 P/B ratio in the Location
UG-9 pore water experiment. Variation among organisms treated alike also
was significantly different for Na+, Si+4 and C1' in the No.3- NaHCO3 and
NaSCN experiments and for C1" and Ca++ in the Location UG-9 pore water
experiment" Treatment effects accounted for 1.1-18.7, 0.0-19.3, 24.8-86.9
and 0.0-21.5 % of the total experimental variation in P/B ratios for Na+,

++, respectively. With the exception of experiment No. 4

Si+4, Cl' and Ca
with the UG-9 pore water, treatments effects accounted for the majority of
the observed variation in each experiment. Among daphnid variation in P/B
ratios within a given experimental treatment always was the smallest
component of the observed experimental variation. Variance due to
differences in measured P/B ratios within an individual daphnid generally

accounted for 12-57% of the total observed experimental variation in P/B

ratios.

Discussion

Sediment pore water has been hypothesized to be the primary route of
exposure to sediment contaminants for benthic macroinvertebrates (Nebeker
et al. 1984, Knezovich and Harrison 1988). Pore water also has been
demonstrated to be a reasonable surrogate test fraction for the assessment
<If bulk sediment toxicity (Ankley et al. 1991). Little research has been
conducted, however, on the physical or chemical properties of pore water,

Other than contaminants, which potentially could affect the results of

pare water toxicity tests. Numerous characteristics of pore water (major

144
ion content, ratios, etc.) potentially can confound the interpretation of
results from pore water toxicity assays.

From these results, it appears that HCOB’ (or bicarbonate
alkalinity) is one factor which can affect toxicity test results with some
sediment pore waters, irrespective of the effects of other substances
present in the sample. Several indirect lines of evidence lead to this
conclusion. The first is that HCO3' itself may be partially responsible
for the toxicity observed in NaHCO3 reference toxicant assays, as
evidenced by the observation that Na+, as NaHCO3, is approximately six
times more toxic to <24-h old D. magna than is Na+ as NaCl. This
observation assumes that Cl' does not protect against Na+ toxicity. This
assumption is supported by recent research which indicates that the
toxicity of sodium compounds is due to the anion and not Na+. Therefore,
not only is Cl‘ not protective, but it appears that it is responsible for
the toxicity of NaCl (D. Mount, ENSR, Inc.,Ft. Collins, CO, pers. comm.,
Gulley et al. 1991). Recent research by Cowgill and Milazzo (1991) also
has reported on alkalinity toxicity to cladocerans. These authors
conducted a series of experiments with NaHCO3 as a reference toxicant to
determine the chronic toxicity of alkalinity, as mg/L CaCO3, to D. magna
and C. dubia. They also reported alkalinity 48 h LCSO values for D. magpa
and C. dubia of approximately 800 and 500 mg/L as CaCO3, respectively.
These values compare well with the mean 48-h D. magna and C. dubia LCSO
values of 921 mg/L (15.1 mmoles/L) and 781 mg/L (12.8 mmoles/L) as CaCO3,
respectively (Table 2) reported in this study.

A second line of evidence is based on theoretical equilibrium
calculations of HCO3’ concentrations in the AOC pore waters. Based on

these calculations and the cladoceran LCSO values for H003” determined in

145

this study, concentrations of H003" sufficient to produce observable
toxicity were present in several of the AOC sediment pore water samples
(Locations UG-8,9,11,12). Based on the lack of effects on C, dppia at 002
concentrations of 0.014 M/L (600 mg/L) in acute tests (Mount and Anderson-
Carnahan 1988), the calculated C02 concentrations in the AOC sediment pore
water samples were not great enough to cause observable effects. The
final evidence for the toxic effects of HCO3', and its proposed mode of
action, comes from the results of the X-ray microanalysis.

The choice of X-ray dispersive microanalysis of elemental
concentrations in D. magna tissues was based on the hypothesized mode of
action for HCOB'. It is well known that freshwater fish actively regulate
internal concentrations of several ions by exchanging them for other ions
contained in the external medium (Maetz and Garcia-Romeu 1964, Garcia-
Romeu and Maetz 1964, Garcia-Romeu et al. 1969, de Renzis and Maetz 1973,
de Renzis and Bornancin 1977, Payan 1978). The existence of a similar
active regulatory mechanism in D. magna for metabolically-produced H+ and
external Na+ was proposed by Potts and Fryer (1979). They hypothesized
that freshwater crustaceans obtain a large portion of their Na+ by active
uptake from water. Higher rates of Na+ turnover (0.163 h'l) were found
for D. magna by Potts and Fryer (1979) compared to rates for freshwater
fish or crayfish weighing 10-100 9 (0.001-0.005 h '1) but rates were
believed to be consistent with the greater cladoceran surface/volume
ratios. Similar Na+ turnover rates for D. magna were observed by Stobbart
et al. (1977) in experiments in which they demonstrated that 1) Na+ influx
in D. magna followed a Michaelis-Menten relationship, 2) the Na+ transport
mechanism can be saturated at Na+ concentrations 20.5 mM/L, and 3) the

metabolic inhibitors KCN and dinitrophenol (DNP) inhibited Na+ uptake at

 

146

less than acutely lethal concentrations. They also suggested that the
presence of a Michaelis-Menten type uptake relationship for Na+ and the
effects of KCN and DNP on Na+ influx indicated that Na+ uptake is actively
regulated in D. magna. Although the presence of actively-regulated H003'-
Cl' exchange similar to that found in freshwater fish has never been
demonstrated in D. magna, it has been suggested that the NH4+-Na+ and
HCO3'-Cl' exchange processes may occur frequently in freshwater animals
(Maetz and Garcia-Romeu 1964).

The results of the experiments with NaHCO3, AOC Location UG-9 pore
water and the metabolic inhibitor NaSCN can be interpreted as supportive
of the hypothesis that the exchange process for HCO3'-Cl' is actively
regulated in D. magna and that the mode of action for HCO3’ toxicity is
the inhibition of this exchange process which results in decreased
internal concentrations of Cl". This study and others (Gulley et al.
1991, Cowgill and Milazzo 1991) have demonstrated that H003" can be toxic
to D. magna and C. dubia. This study also has demonstrated that Q. magpa
from HCO3' treatments which caused acute toxicity exhibited decreased
internal Cl‘ concentrations, relative to controls, based on mean P/B
ratios from X-ray microanalysis. Calculations of the theoretical
equilibrium concentration of HCO3' in AOC Location UG-9 pore water
demonstrated that the sample contained sufficient HCO3' to cause acute
toxicity based on the 48-h LCSO values determined in the Ncho3 reference
toxicant assays. x-ray microanalysis demonstrated that organisms from the
AOC Location UG-9 pore water toxicity test exhibited decreased internal
Cl' concentrations, relative to control, in the 32% and 100% pore water
exposures. The metabolic inhibitor NaSCN also caused statistically

significant (a = 0.05) reductions in internal concentrations of Cl' across

 

147
all NaSCN treatments; indicating that the uptake of Cl' from the external
environment is actively regulated in D. magna.

The nested ANOVA experimental design facilitated the evaluation of
the variance in measured elemental P/B ratios due to 1) experimental
treatment, 2) among daphnid differences within a treatment, and 3) among
measurement differences within an individual daphnid (Table 4). With the
exception of experiment No. 4 with AOC Location UG-9 pore water, treatment
effects accounted for the majority of the variation in measured P/B ratios
for C1" in D. magna. In experiment No 4. with the Location UG-9 pore
water, treatment effects accounted for 25% of the observed variation in
C1“ P/B rations. Among measurement differences within the same daphnid
accounted for the next greatest proportion of the experimental variance in
P/B ratios for C1“ while among daphnid differences always accounted for
the smallest proportion of the experimental variance in P/B ratios for C1-
. This analysis highlights the importance of the experimental treatments
on P/B ratios for C1. in D. magna. It further suggests that in future
investigations of this type, greater experimental precision could be
achieved by decreasing the number of daphnids analyzed per experimental
treatment but increasing the number of x-ray microanalysis measurements
conducted on an individual daphnid.

Additional circumstantial support for the hypothesis that HCO3'-Cl'
exchange is actively regulated in other aquatic organisms may be found in
the field studies of Montezuma Well, Arizona conducted by Blinn and
Sanderson (1989) and Dehdashti and Blinn (1991). In this ecosystem, high
concentrations of C02 (>550 mg/L) and alkalinity (>600 mg/L as CaCO3) were

hypothesized to be responsible for the lack of fish and benthic

macroinvertebrates of the orders TrichOptera, Lepidoptera, Megaloptera and

148

Anisoptera. Limited collections of benthic macroinvertebrates in the
orders Chironomidae and Ephemeroptera were attributed to the same
phenomenon (Blinn and Sanderson 1989). Planktonic rotifers and
cladocerans also were absent from this ecosystem (Dehdashti and Blinn
1991). Additional field studies in the Intermountain Region of the U.S.
Forest Service concluded that the number of macroinvertebrate taxa in 164
samples from 20 different streams was inversely correlated (r = -0.67, p
= 0.001) with ambient stream alkalinity as CaCO3 (Winget and Mangum 1979).
Similarly, McCarraher and Thomas (1968) observed that in alkaline lakes of
Nebraska, fathead minnows (Pimephales promelas) flourished when carbonate
alkalinity was below 800 ppm and total alkalinity was below 1,800 ppm.
However, carbonate or total alkalinity in excess of these values was
reported to greatly impair reproduction and abundance of this species.

The proposed existence in D. magna of an actively-regulated exchange
process for HCO3' and Cl' and the hypothesis that disruption of this
process is the mode of action for bicarbonate toxicity have several
practical implications. Bicarbonate concentrations should be routinely
measured during toxicity tests of aqueous samples, such as pore waters or
effluents. If’ bicarbonate concentrations are sufficient. to jproduce
toxicity, this observation should be considered when interpreting the test
results. The importance of bicarbonate toxicity also needs to be
addressed in toxicity identification evaluation (TIE) procedures (Mount
and Anderson-Carnahan 1988, 1989, Mount 1989), both for effluents and
sediment pore waters, because current TIE methodology would not identify
high bicarbonate concentrations as the specific cause of observed
toxicity. Instead, the observed toxicity could be wrongly ascribed to the

effects of total dissolved solids (TDS). This would implicate cations or

149
cation ratios in the sample as the cause of observed effects when, in
fact, the effects were due to the anion HCO3'. Finally, since active
regulatory processes for HCO3'-Cl‘ also exist in freshwater fish and
amphibians, and may exist for other aquatic invertebrates, TIE procedures
performed with these species also may be affected by high bicarbonate

concentrations in the test solution.

Acknowledgements

Jane Thompson, Jane Norlander and Debra Williams typed various
drafts of the manuscript. Mary Schubauer-Berigan, Teresa Norberg-King,
Gary’ Ankley, Howard. McCormick, Don. Mount and an anonymous reviewer

provided helpful comments on earlier versions of the manuscript.

150

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CHAPTER 4

Mutagenicity and 2,3,7,8-Tetrachlorodibenzo-p-dioxin Equivalents in

Organic Solvent Extracts of Sediments from the Grand Calumet River,

Indiana

158

3‘

Introduction

The importance of sediments as both a sink and source for chemicals
of environmental concern has become obvious during the preceding 20 years.
As a result, an increasing amount of research effort has been devoted to
investigations of the fate and effects of chemicals in sediments. Most
research on the effects of chemicals in sediments has focused on the
potential acute toxicity of in-place pollutants to benthic species. A
lesser amount of research effort has been devoted to the potential for
chronic toxicity or bioaccumulation of sediment-associated chemicals and
the potential mutagenicity of bulk sediments or various extracts of
sediments. The driving force behind research on the mutagenic effects of
pure chemicals or complex environmental mixtures in the past has been the
potential for human health effects. Although humans have little direct
exposure to sediments or extracts of sediments, we are potentially exposed
to these complex Inixtures indirectly via, aquatic food chains (i.e.
ingestion of aquatic species exposed to contaminated sediments).
Therefore, any chemical in sediments which is mutagenic and can be
bioaccumulated could potentially cause both ecological and human health
effects. Research also has been completed or is on-going which suggests
that, for certain chemicals, concentrations of mutagenic chemicals in the
environment which may be protective of human health may produce mutagenic

effects in aquatic biota and organisms which consume aquatic biota.

159

 

 

160

Chemical analysis can only provide a limited idea of the potential
effects (i.e. toxicity, mutagenicity, etc.) of exposure to a complex
mixture such as a bulk sediment or extract of a sediment, due to the
potential for interactive effects (antagonism, synergism); the potential
for effects at concentrations of chemicals which are below the detection
limits for the analytical method; and the incompleteness of the chemical
analysis (i.e. it is not possible to analyze for all chemical compounds
potentially present in a sample). Therefore, the best. measure of
potential biological effects is a direct measure of some form of
biological activity after exposure to the complex mixture (Durant et al.
1992, Tillett et al. 1991).

Evaluations of the mutagenicity of extracts of soils or bulk
sediments have been conducted with the Ames assay (Donnelly et al. 1991,
Maccubbin et al. 1991, Ersing 1987, Maccubbin 1986, West et al. 1986 a,b,
1988, Durant et al. 1992, Fabacher et al. 1988, Sabo et al. 1983 and
Maccubbin and Ersing 1991), a human lymphoblast assay (Durant et al. 1992)
and the MutatoxO assay (Kwan et al. 1990, Dutka et al. 1991). The study
by Durant et al. (1992) compared mutagenicity of organic extracts of
sediments in both the Ames and human lymphoblast assays. To date, no
comparison of mutagenicity in complex environmental samples such as
organic extracts of sediments has been made between the Ames and Mutatoxo
assays although Johnson (1992) investigated the relative mutagenicity of
pure chemicals in the two assays.

The Ealmonella typhimurium mutagenicity assay, more commonly known
as the Ames assay (Ames et al. 1975, Maron and Ames 1983) is a rapid
method of screening for potential mutagenicity, and thus carcinogenicity,

of chemicals and complex mixtures. The Ames assay is based upon the use

161

of selected mutant strains of the bacterium S. typhimugium and the
particular vulnerability of these strains to mutations at the histidine
locus. These bacteria carry a mutation for histidine dependency and are
reverted back to histidine prototropy by the mutagenic chemical. The
inclusion of liver microsomes containing biochemically induced enzymes is
required for activation of certain chemicals to their reactive forms.
Bacterial strains TA-98 (sensitive to frame-shift mutations) and TA-100
(sensitive to base-pair substitution) have separate: mutations which
increase the assay sensitivity to a greater range of chemical agents when
a sample is tested with both bacterial strains (Maron and Ames 1983).

The Mutatox® assay employs a dark mutant of the luminescent marine
bacterium, Photobacterium phosphoreum which undergoes a forward mutation
after exposure to a variety of mutagenic compounds. The MutatoxO assay is
based on a similar assay developed by Ulitzur (1986) and is capable of
detecting base substitutions (point mutations), base additions or
deletions (frameshift mutations), intercalations of DNA, or the inhibition
of DNA synthesis. Over 100 compounds representing a dozen chemical
classes, including volatile chemicals and complex mixtures, have been
tested using the dark mutant strain of E. phosphoreum (Ulitzur 1986). A
number of these compounds also have been tested and categorized by the
National Toxicology Program (Tennant et al. 1987). For these chemicals,
correlative data from long-term animal studies and four short-term assays
(Ames assay, Chinese hamster ovary chromosome aberrations and sister
chromatid exchange, mouse lymphocyte assay) exist for comparative
purposes. As in the Ames assay, it also is possible to conduct the
Mutatox‘ID procedure with and without metabolic activation to compare

mutagenic activity of parent compounds and potential metabolic products.

EL.

.._ 4' ;

A- - s 21.—_—
a
u

162

Although analytical techniques exist for the detection of minute
quantities of polychlorinated hydrocarbons (PCH), these procedures can be
extremely costly and time-consuming, particularly when samples may
theoretically contain up to 209 different polychlorinated biphenyl (PCB)
congeners and 75 different polychlorinated dibenzodioxin (PCDD) or
dibenzofuran (PCDF) isomers and congeners (Safe 1987). Thermost toxic PCB
and PCDD congeners are those which are planar, or nearly planar (Greenlee
and Neal 1985). The toxic properties of different planar PCH compounds
appear to be expressed via a common mode of action, and therefore, it is
possible to calculate the biological potencies of complex mixtures of PCHs
by expressing their toxicity relative to the most toxic PCH known,
2,3,7,8-TCDD (Bradlaw and Casterline 1979, Eadon et al. 1986, Safe 1987).
TCDD-equivalents can be assigned to complex mixtures of PCBs by measuring
the ability of the PCH mixture to induce cytochrome p-450-dependent
ethoxyresorufin-o-deethylase (EROD) activity in H4IIE rat hepatoma cell
cultures and expressing the magnitude of the response relative to
induction observed with TCDD (Bradlaw and Casterline 1979, Casterline et
al. 1983, Safe 1987).

The H4IIE rat hepatoma cell assay is also useful because it serves
as both a qualitative check and a supplement to the results of the Ames
assay. Several investigators have reported a suppression of mutagenic
potential in the Ames assay between PAHs and other components of crude oil
(Hermann et al. 1981, Hermann 1980, Petrilli et al. 1981, Carver et al.
1985, Haugen and Peak 1983). The cause of the suppression of mutagenic
potential was demonstrated to be inhibition of the hepatic microsomal

monooxygenase (MO) system (Carver et al. 1985). Therefore induction of

163
EROD activity would preclude suppression of mutagenic potential due to M0
inhibition in samples also tested in the Ames assay.

The objectives of this study were: 1) to compare the mutagenicity
of organic solvent extracts of bulk sediments with the Ames and Mutatoxo
assays, 2) to evaluate the mutagenicity of aqueous extracts of sediments
in the MutatoxO assay, 3) to evaluate EROD induction in the H4IIE assay
after exposure to organic solvent extracts of bulk sediments and 4) to
compare assay results to a limited set of organic chemical analyses of

bulk sediment.

Materials and Methods

Sample Collection

Sediment samples were collected from the Grand Calumet River, IN on
11 October and 22 November 1988; 10 March, 24 May, 30 October and 13
November 1989; and 12 May 1990. A Ponar grab sampler was used to collect
sediment samples from 10 locations along the Grand Calumet River and three
locations in the Indiana Harbor ship canal (Figure 10). At the time of
sample collection, study sites were located by triangulation of local
landmarks. The sample from each location was a composite of approximately
80-100 L of wet sediment from multiple Ponar grabs. Multiple grab samples
were collected and composited to ensure sufficient sample volume for all
necessary sub-sample collection. Compositing and homogenization of

composite samples were done in a large stainless steel pan with stainless

 

164

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165
steel tools. Large debris was removed from the composite samples and two,
1-L aliquants (sub-samples) for quantification of metals and organic
compounds in bulk sediments were placed in clean l-L glass jars capped
with solvent rinsed aluminum foil under the lid. Samples for toxicity
testing or pore water extraction were placed in coolers or plastic buckets
lined with food-grade plastic bags. After collection, sediment samples
were placed on ice in coolers and transported to the laboratory, where
they were maintained in a walk-in cooler at 4° C until processing and

analysis.

Pore Water Extraction

Pore water was extracted from sediments by a combination of
centrifugation and filtration as described by Hoke et al (1992a). The
pore water extracts were placed in clean, solvent-rinsed glass bottles,
the bottles capped with aluminum foil-lined lids, and the pore water

maintained in the dark at 4° C until used for assays (< seven days).

Ames Assay

The Ames test was conducted according to the methods and procedures
described by Maron and Ames (1983). The complete sample extraction and
Ames test methods used in this study have previously been described by
Maccubbin and Ersing (1991), however, a brief summary is presented below.
A subsample of each sediment sample was air-dried at 80° C to determine
moisture content. A wet sediment sub-sample was homogenized and a 30-50
9 sample mixed with anhydrous sodium sulphate to absorb the water from the
sample. The sample was then placed in a cellulose extraction thimble and

extracted for two successive 24 h periods with 300 mL isopropyl alcohol

 

166

and dichloromethane, respectively; The extracts were combined, the volume
reduced to 50 mL and residue content determined by drying three, one mL
aliquants of the combined extracts in tared aluminum weigh boats. An
appropriate amount of extractant was then solvent-exchanged into dimethyl
sulfoxide (DMSO) to give a final organic residue content of 10 mg/ml
(Maccubbin and Ersing 1991).

Sediment extracts were diluted to provide a dose series of 1000 pg,
600 pg, 200 pg, 100 pg and 60 pg residue per plate for testing. One
hundred pl of organic extract from each sediment were mixed with 100 pl of
an overnight culture of bacteria (tester strain TA98 or TA100) and 2 ml of
melted agar containing 5 mM histidine and biotin (Ames et al. 1975, Maron
and Ames 1983). Molten top agar was then poured onto a minimal glucose
agar base plate and the plates incubated at 37° C for 2 days. The
existence of compounds requiring metabolic activation was evaluated by
adding 0.5 ml of buffer solution containing rat liver homogenate (89 from
Aroclor treated rats, Litton Bionetics, Charleston, SC.) and co-factors to
the top agar prior to plating. Spontaneous mutation rates, solvent,and S9
effects were evaluated with plates containing bacteria only, DMSO only,
and DMSO + $9 only as negative controls. Tester strain sensitivity was
monitored with benzo(a)pyrene, sodium azide and daunomycin as positive
controls. Each extract and control treatment was tested in triplicate.
The number of His + revertant colonies/pg residue were determined after
incubation and converted to His+ revertants/mg dry weight sediment based
on the residue content of each sediment. Toxic effects were evaluated by
decreases in mutagenicity with increases in dosage and loss of the
background "lawn" of cell growth that occurred due to the small amount of

histidine present in the culture media. Dose response data were evaluated

 

 

167
over the linear portion of the dose response curve according to the two-
fold rule (Chu et al. 1987). Sediment was classified for mutagenic

potential based on dose dependent increases in the reversion rate.

Mutatox® Assay

The Mutatox® assay was conducted on aliquots of the organic sediment
extracts tested in the Ames and H4IIE tests. The dry MutatoxO medium was
reconstituted with distilled water and the pH adjusted to 6.8 with 5 N
KOH. Direct test medium was prepared by adding one mL of the
reconstituted dark mutant P. phosphoreum to 100 mL of reconstituted
medium. TWenty-six cuvettes (10 sample plus 3 control X 2 replicates
each) were set-up, labelled from 1-13 (A and B) and one mL of test medium
added to each number one tube. One-half (0.5) mL of test medium was added
to each tube numbered 2-13. A 0.2 mL aliquot of the organic sediment
extract was added to the number 1 tubes, the solution mixed and serially
diluted by transferring 0.5 mL of mixed solution from cuvette 1 through
cuvette 10 for each set of replicates. The samples were then placed on a
rotary shaker set at 70 i 10 rpm and incubated at 23 t 2° C for 16 h.
Light output of the dark mutant (i.e. mutagenesis) was measured at 16, 18
and 20 h for both the direct and S9-activated mutagenesis tests. The
measurements of light emission used the same methods as those described
for the Microtox® test (Bulich 1981). A positive mutagenic response was
classified as a light peak of 100 or more at three times the reagent
control light output. A non-reactive chemical also was expected to

exhibit a negative response in both direct and S9-activated tests.

 

168
H4IIE Assay

The H4IIE rat hepatoma cells were obtained from the American Type
Culture Collection (ATCC No. CRL 1548). The cell culture medium used was
a supplemented Dulbecco's Modified Eagle's Medium (D-MEM) (Tillett et al.
1991). Phenol red was not used in the medium because it may adversely
affect liver cell cultures and may potentiate induction in the H4IIE
cells. H4IIE cells were kept in continuous culture in the laboratory and
the cultures regularly restarted from frozen cells to insure the cultures
were not altered. Standard sterile tissue culture techniques were used to
maintain the integrity of the cell line and prevent contamination of the
cells.

Cell culturing and harvesting were performed according to the
procedures described in Tillitt et al. (1991). The cells were dosed with
aliquots of the same organic extracts of the sediment samples used in the
Ames and Mutatox® assays. DMSO was used as the carrier solvent with the
same quantity (mass) of extract delivered to separate plates in different
volumes (<100 pl) of DMSO. There was no effect on the basal EROD activity
of the H4IIE cells of DMSO volumes between 10 and 100 pL/plate.

Triplicate plates were tested at each of 4-5 doses of the sediment
extracts along with a concurrent TCDD standard curve (4 doses in
triplicate). Bioassay to bioassay variations in experimental technique
were taken into account with a standard TCDD curve analyzed with the
sediment extracts. The calculation of TCDD-equivalents does not reflect
these among-experiment variations (Tillitt et al. 1991). Plates were
incubated for 72 hours and the cells rinsed with PBS and scraped from the
plates while in a Tris-sucrose (0.05 - 0.2 M) buffer. The collected cells

were centrifuged for 10 minutes, resuspended in the Tris-sucrose buffer,

 

 

 

 

169

and duplicate analyses performed of protein content (Lowry et al. 1951)
and enzyme induction. The indirect EROD assay was chosen to monitor
induction of cytochrome P-4501A1 monooxygenase activity (Pohl and Fouts
1980). The EROD assay has good sensitivity, is specific for the induction
of P-4501A, and correlates well to aryl hydrocarbon hydroxylase (AHH)
activity in this cell line (Bandiena et al. 1984). EROD assays were
conducted in polycarbonate tubes in. a shaking' water* bath incubator
maintained at 37° C. The assay protocol followed that described by
Tillitt et al. (1991).

Fluorescence of samples was determined on a SLM 4800
spectrofluorometer at emission and excitation wavelengths of 585 and 550
nm, respectively. The machine was calibrated with a standard rhodamine B
solution and fluorescence in samples read relative to this internal
standard. Each reading was an average of 20 automatic scans. A resorufin
standard was also run on each experiment for calibration to a resorufin
standard curve and calculation of specific enzyme activities. EROD
activity was calculated and reported as pdcomoles resorufin/mg

protein/minute for each treatment combination.

Chemical Analysis

A suite of 106 organic chemical compounds and metals were analyzed
for in sediments from the Grand Calumet River and Indiana Harbor, IN. The
complete analytical methods and results for all analytes have previously

been reported by Hoke et al. (1992).

 

 

170
Statistical Analysis
All data analyses were performed with SAS statistical software (SAS
1985). Relative potencies of the samples were estimated with the slope-
ratio assay (Finney 1978). The relative potency was derived from the
slope on the linear portion of the dose-response curve for the sample in
comparison to the slope of the TCDD standard curve. Sample potency was

calculated as:

Sample Slope/TCDD Slope (l)

with final units of pg TCDD/uL of extract. TCDD-equivalents (E9) on a
pg/gm dry wt sediment basis were estimated based on the known mass of
residue extract on a dry wt sediment basis. TCDD-EQ attributable to
measured total PCBs (as Aroclor 1248) and TCDD in each sediment sample
were calculated and compared to EROD TCDD-EQ. For calculation purposes,
one pg of total PCBs (as Aroclor 1248) was assumed to be equivalent to 10
pg TCDD-EQ (Tillitt et a1. 1991). Pearson product moment correlation
analysis (SAS 1985) was used to compare the results of both the Ames and
Mutatox® mutagenicity assays with the results of chemical analyses of bulk

sediments.

Results

Mutatox® Assay
Mutagenicity in the Mutatox test without 59 activation only occurred
in the organic extract of sediment from site UG-l and was equivocal in the

extract from site UG-3 (Table 17). Organic extracts of sediments from all

 

a.

 

171
sites were umtagenic when tested with S9 activation (Table 17). The
extracts from sites UG-3 and UG-7 were the most and least mutagenic with
S9 activation, respectively. In general, organic extracts of sediments
from the western branch of the Grand Calumet River were less mutagenic
when tested with S9 activation than were extracts from the eastern branch.
No mutagenicity was observed in any sample when aqueous extracts (pore

O assay with

waters) from study site sediments were tested in the Mutatox
and without S9 activation. The relationship of the results of the
MutatoxQ assays of organic solvent extracts of sediments also was
evaluated in the relationship to the results of the organic chemistry of
the sediments which have previously been presented elsewhere (Hoke et al.

1992). No significant correlations were observed between the results of

the Mutatox® assay and the organic chemical analyses of the sediments.

Ames Assay

Organic extracts of sediments from sites UG-3 and 8 were mutagenic
when analyzed. without 59 activation using tester strain. TA98 while
extracts from sites UG-3 and 5 were mutagenic without S9 activation when
analyzed with tester strain TA100 (Table 17). Organic extracts of
sediments from all sites were mutagenic when analyzed with S9 activation
using both tester strains TA98 and TA100 (Table 17). With metabolic
activation, the number of strain TA98 revertants per mg dry wt sediment
ranged from 1 at site UG—4 to 102 at site UG-3 while the number of TA100
revertants ranged from 10 per mg dry wt sediment at site UG-4 to 1710 at
site UG-3 (Table 17).

A number of statistically significant correlations (p S 0.05) were

observed between the Ames assay results and the organic chemical analyses

 

 

 

 

 

 

 

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174
of the sediment (Table 18). No significant correlations were observed
between the results of the Ames assays and concentrations of metals in

bulk sediment (data not shown).

H4IIE Assay

Based on EROD induction, the greatest number of TCDD-EQ (pg/gm dry
wt sediment) were observed for extracts of sediments from sites UG-l, 3
and 8 (Table 19). TCDD-EQ for other sites were generally an order of
magnitude less. Concentrations of TCDD-EQ in bulk sediments from the
study sites ranged from approximately 4,700 to greater than 500,000 pg/gm
dry wt sediment.

The proportion of the total TCDD-EQ from the H4IIE assay attributable
to measured concentrations of total PCBs (as Aroclor 1248) and TCDD in
bulk sediments from the study sites was small (i.e. <0.5%). Pearson
product moment correlation analysis of TCDD-EQ and results of chemical
analyses of the bulk sediments indicated significant correlations existed
between TCDD-EQ and bulk sediment concentrations of 2,4-and 2,6-
dichlorophenols (r == 0.70, 0.66); 2,4,6-trichlorophenol (r as 0.66);
acrylonitrile (r == 0.66); p-dichlorobenzene (r == 0.66); 1,3-

dichloropropene; ethylbenzene (r = 0.76) and fluoranthene (r = 0.69).

Discussion

Although investigations of the mutagenicity of sediments from the
Great Lakes have been limited, several authors have reported that organic
solvent extracts of sediments from various locations were mutagenic when

tested in the Ames assay (Maccubbin 1986, Maccubbin and Ersing 1991,

‘0

'0

Table 18.

175
Pearson product moment correlation coefficients from analyses
of Ames assays with S9 activation and organic chemical
analyses of sediment from the Grand Calumet River, IN.

Correlations reported were statistically significant at p S

 

 

0.05.

Parameter Tester Strain

TA98 + TA100+

Hydroquinone -- 0.83
Mercaptobenzothiazole 0.64 0.79
3,4-Dichloroaniline 0.86 0.91
Heptachlor 0.62 0.82
p-Dichlorobenzene 0.68 --
1,2—Dichloropropane -- 0.74

2,4-Dinitrotoluene 0.63 0.68

 

 

 

 

176

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177
Maccubbin et al. 1991, West et al. 1986a,b, 1988, Fabacher et al. 1988).
Several authors also have reported that solvent extracts of sediments were
mutagenic in the Mutaton assay (Kwan et al. 1990, Dutka et al. 1991) but,
to date, little or no comparative data exist for results of Ames and
MutatoxO assays performed on solvent extracts of the same sediments.

The data from this study indicate that both the Ames and MutatoxO
assays yield similar results when organic solvent extracts of sediments
are tested with S9 microsomal metabolic activation. Correlation of the
number of TA98 or TA100 revertants with the Mutatox® results in samples
tested with S9 activation was weak (r = -O.35 to -O.40). Little
concordance was observed in the determination of mutagenicity for solvent
extracts of sediments tested without 89 activation in both assays although
the extract from site UG-3 was mutagenic in all direct assays. Ames
assays using tester strains TA98 and TA100 without activation indicated
extracts from sites UG—3 and 8, and UG-3 and 5 were mutagenic,
respectively. MutatoxQ assays without activation indicated the extract
from site UG-l was mutagenic while the results for the extract from site
UG-3 was equivocal.

Differences in observed results among assays and tester strains
during assays without metabolic activation are most likely due to the
individual strain or assay sensitivity to different chemicals and the type
of genetic mutation caused by the chemical. Ames assay tester strain TA98
detects mutagens which cause frameshift mutations while tester strain
TA100 detects those causing point mutations (Maron and Ames 1983). The
Mutatox® assay detects compounds which cause frameshift or point.mutations
as well as compounds which inhibit DNA synthesis and DNA-intercalating

agents (Johnson 1992).

178

Previous investigations have reported the detection of direct acting
mutagens in complex mixtures by the MutatoxQ assay (Kwan et al. 1990). In
the only direct comparison of the Ames and Mutatoxo assays conducted to
date, Johnson (1992) tested the sensitivity of the two assays with
metabolic activation to eight progenotoxins. Both assays exhibited
similar sensitivities to the eight compounds with lowest detected
concentrations in the low microgram range for both assays. However,
pyrene was not detected by the Ames assay but caused a strong positive
response in the Mutatox® assay (Johnson 1992).

Interestingly, no statistical correlations were observed between the
results of the Mutatox® assay' with activation and. organic chemical
analyses of sediments from the study sites. However, a number of
statistically significant correlations were observed between the organic
chemical analyses and the results of Ames assays with tester strains TA98
and/or TA100. Chemicals for which concentrations in bulk sediments were
significantly correlated with Ames assay results with tester strain TA98
included mercaptobenzothiazole; 3,4-dichloroaniline, heptachlor, p-
dichlorobenzene and 2,4-dinitrotoluene. Ames assay results with tester
strain TA100 were significantly correlated to chemical concentrations in
bulk sediments of hydroquinone; mercaptobenzothiazole; 3,4-
dichloroaniline; heptachlor; 1,2-dichloropropane and 2,4-dinitrotoluene.
In pure compound assays, mercaptobenzothiazole, hydroquinone; 2,4-
dichlorophenol and heptachlor have been reported as non-mutagenic in Ames
assay while 1,2-dichloropropane and 2,4-dinitrotoluene have been reported
to be mutagenic in the assay (Ashby and Tennant 1991, Tennant and Ashby

1991).

179

Extracts of sediments from the Detroit River (Maccubbin.et al. 1991),
the Buffalo River (Ersing 1987) and the Black River (Fabacher et al. 1988,
West et al. 1986a,b, 1988) within the Great Lakes basin have previously
been reported to elicit mutagenicity in the Ames assay. PAHs have been
the most frequently implicated compounds in the observed mutagenicity.
West et al. (1986a,b) identified polycyclic aromatic hydrocarbons (PAH),
nitrogen heterocycles of PAHs (PANH) and alkylated forms of both PAH and
PANHs as the primary causes of the observed mutagenicity. Maccubbin et
al. (1991) also indirectly implicated PAHs as the potential cause of
mutagenicity observed in organic solvent extracts of sediment from the
Detroit River tested with the Ames assay. When these investigators
fractionated the solvent extracts of the samples and then tested the
individual fractions of the original extract, they observed a higher total
number of revertants for the fractions than for the original extract.
This phenomenon suggests that mutagenicity was inhibited or suppressed,
possibly due to the presence of other PAHs (Hermann et al. 1981, Hermann
1980, Haugen and Peak 1983, Petrilli et al. 1981, Carver et al. 1985).
The mechanism for this suppression or inhibition was demonstrated to be
related to inhibition of the hepatic microsomal MO system in the s9
(Carver et al. 1985). Additional S9 was recommended for Ames assays of
complex mixtures containing petroleum hydrocarbons to ensure adequate
metabolic activation of PAHs (Carver et a1. 1985).

The Ames assay protocol used to test solvent extracts of sediments
from the Grand Calumet River and Indiana Harbor, IN included the addition
of a large volume of S9 mixture to each extract (Maron and Ames 1983).
Therefore, suppression of mutagenicity due to PAI-Is should have been

minimal. Hoke et a1. (1992) reported that individual PAH concentrations

 

a
. i
'-

 

180
in the bulk sediments were as great as 100 mg/kg. This may account for
the large number of revertants observed for the samples.

Solvent extracts of other sediments from the Great Lakes have been
reported to cause 80-12,000 revertants/gm dry wt of sediment (Ersing 1987,
Maccubbin et al 1991) while extracts of Grand Calumet River sediments
tested with metabolic activation caused from 1000-1,710,000 revertants/gm
dry wt sediment. Direct acting mutagens in Grand Calumet River sediments
caused 2000-45,000 revertants/gm dry wt sediments.

The importance of PAHs in the Grand Calumet River and Indiana Harbor,
IN. can also be observed in the results of the H4IIE assays. Significant
amounts of TCDD-EQ'were detected in solvent extracts of sediments from the
study area. Measured concentrations of other compounds which can cause
EROD induction in environmental samples (e.g., PCDDs, PCDFs, PCBs) did not
account for the observed TCDD—EQ from the H4IIE assay (i.e. <0.5%). It
must be noted, however, that total PCBs as Aroclor 1248 and TCDD were the
only representatives of the PCDDs, PCDFs and PCBs measured in bulk
sediments from the Grand Calumet River. .A stronger, albeit
circumstantial, case can be made that PAHs in bulk sediments from the
study area were responsible for the observed TCDD-EQ. PAHs are known to
cause EROD induction in the H4IIE assay (Whitlock et al. 1974, Xu and
Bresnick 1990, Corcos and Weiss 1988) and large concentrations of several
PAHs have been measured in sediments from the study area (Hoke et al.
1992). It is also likely that large concentrations of both unmeasured
parent PAHs and their degradation products exist in sediments from the
study area. PAHs also are likely responsible for a portion of the

mutagenicity observed in the Ames and Mutatoxo assay even though no

181
statistically significant correlations were observed between assay results
and PAH concentrations in bulk sediments.

As evidenced in the results of the Ames and MutatoxO assays, numerous
mutagenic compounds exist in sediments from the Grand Calumet River and
Indiana Harbor, IN. Black et al (1985) demonstrated that neoplasia in
fish could be induced by repeated applications of solvent extracts of
sediments from the Buffalo River, NY which contained numerous mutagens
including PAHs. However a long latency period was necessary to
demonstrate evidence of carcinoma. Thus a need arose for short-term tests
of genotoxicity such as the Ames and Mutaton tests. Although these
assays can provide an indication of the potential mutagenicity of complex
environmental mixtures, more research needs to .be conducted. on ‘the
representativeness of the exposure route used in these assays (solvent
extracts). The absence of mutagenicity in Mutatox® assays of pore waters
from the study site sediments indicates that short—term direct human
exposure to mutagens in pore waters and sediments are likely to be non-
problematic. Greater concern is required for the potential for both
ecological and human health effects due to bioaccumulation
(bioconcentration and biomagnification) of mutagenic compounds present in

sediments from the Grand Calumet River and Indiana Harbor, IN.

 

Acknowledgements

Mutatox® assays were conducted by Dr. Jim Tung of Microbics, Inc.,
Carlsbad, California and Ames assays were conducted by Dr. Lex Maccubbin
of the Grace Cancer Research Institute, Buffalo, New York. G. Ankley and

several anonymous reviewers provided comments which improved earlier

versions of the manuscript.

182

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Ames, B.N., McCann, J. and Yamasaki, E. 1975. Methods for detecting
carcinogens and nmtagens with the Salmonella/mammalian
microsome mutagenicity test. Muta. Res. 31 347-364.

Ashby, J. and R.W. Tennant. 1991. Definitive relationships among
chemical structure, carcinogenicity and mutagenicity for 301
chemicals tested by the U.S. NTP. Muta. Res. 257:229-306.

Bandiera, S., T. Sawyer, M. Romkes, B. Zmudka, L. Safe, G. Mason, B. Keys
and S. Safe. 1984. Imnychlorinated dibenzofurans (PCDFs):
effects of structure on binding to the 2,3,7,8-TCDD cytosolic
receptor protein, AHH induction and toxicity. Toxicology 32:
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SUMMARY

At the present time, the GCR—IHC system has severely degraded
sediments which contain a multitude of chemicals. The most important
acute toxicity problems for benthic macroinvertebrates within the system
appear to be petroleum hydrocarbons, metals, ammonia and PAHs. Remedial
action plans for the area which propose to foster the development of
healthy instream communities of invertebrates and fish must of necessity
provide a strategy for dealing with both source control and existing

concentrations of these contaminants in sediments.

191