United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
EPA/600/R-93/055
March 1993
Interim Report on
Data and Methods for
Assessment of 2,3,7,8-
Tetrachlorodibenzo-p-
dioxin Risks to Aquatic
Life and Associated
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EPA/600/R-93/055
March 1993
Interim Report on Data and Methods for
Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks
to Aquatic Life and Associated Wildlife
Environmental Research Laboratory
Office of Research and Development
United States Environmental Protection Agency
Duluth, MN
Printed on Recycled Paper
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DISCLAIMER
Procedures set forth here are intended as guidance to Agency and other government
employees. They do not constitute rulemaking by the Agency, and may not be relied
on to create a substantive or procedural right enforceable by any other person. The
Government may take action that is at variance with the procedures stated in this
document.
Mention of trade names or commercial products does not constitute endorsement or
recommendations for use.
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CONTENTS
CONTENTS i
TABLES ill
FIGURES iv
ACKNOWLEDGMENTS v
LIST OF SELECTED ABBREVIATIONS vii
EXECUTIVE SUMMARY ix
1. INTRODUCTION . ... 1-1
1.1 BACKGROUND 1-1
1.2 GOAL AND SCOPE OF THE INTERIM REPORT 1-2
2. EXPOSURE 2-1
2.1 PHYSICAL/CHEMICAL PROPERTIES OF TCDD AND RELATED
CHEMICALS 2-1
2.2 ANALYTICAL LEVELS OF DETECTION 2-5
2.3 DISTRIBUTION IN WATER, SEDIMENTS AND FOOD CHAINS 2-6
2.4 EXPOSURE ROUTES FOR AQUATIC ORGANISMS . . 2-9
3. BIOACCUMULATION 3-1
3.1 CONCEPTUAL FRAMEWORK AND DEFINITIONS 3-1
3.1.1 Bioconcentration 3-1
3.1.2 Bioaccumulation 3-5
3.1.3 Biota-Sediment Relationships 3-7
3.2 TCDD BIOCONCENTRATION FACTORS 3-9
3.3 TCDD BIOACCUMULATION FACTORS 3-12
3.4 TCDD BIOMAGNIFICATION FACTORS 3-17
3.5 BIOTA-SEDIMENT ACCUMULATION FACTORS . 3-19
4. EFFECTS 4-1
4.1 COMPARATIVE TOXICOLOGY . 4-1
4.2 EFFECTS OF TCDD ON AQUATIC LIFE 4-4
4.2.1 Toxicological Information 4-4
4.2.2 Epidemiological Information 4-33
4.2.3 Effects Profile 4-38
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4.3 EFFECTS OF TCDD ON AQUATIC-ASSOCIATED WILDLIFE 4-43
4.3.1 Toxicological Information 4-43
4.3.2 Epidemiological Information 4-49
4.3.3 Effects Profile 4-51
5. RISK CHARACTERIZATION METHODOLOGY 5-1
5.1 SUMMARY OF EXPOSURE AND EFFECTS INFORMATION 5-1
5.1.1 Exposure 5-1
5.1.2 Bioaccumulation 5-3
5.1.3 Effects 5-4
5.2 APPLICATION OF INFORMATION TO RISK CHARACTERIZATION . . 5-6
5.2.1 Fish Contamination in the United States: Risk to Aquatic
Life and Associated Wildlife 5-6
5.2.2 Lake Trout Reproduction in Lake Ontario 5-8
5.2.3 Environmental Concentrations Associated with TCDD
Effects 5-10
6. RESEARCH NEEDS FOR REDUCING UNCERTAINTIES 6-1
6.1 EXPOSURE 6-1
6.1.1 Octanol/Water Partition Coefficient 6-1
6.1.2 Detection Limits and Water Concentrations in Natural
Systems 6-1
6.2 BIOACCUMULATION 6-1
6.3 EFFECTS 6-2
6.3.1 Occurrence of Ah Receptor 6-2
6.3.2 Aquatic Life 6-2
6.3.3 Wildlife 6-3
6.3.4 Epidemiology 6-3
7. REFERENCES 7-1
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TABLES
Table E-1. Environmental concentrations associated with TCDD risk to
aquatic life and associated wildlife xx
Table 2-1. Minimum levels of detection (MLD) for routine high
resolution gas chromatography/high resolution mass
spectrometry analyses of TCDD in samples from aquatic
ecosystems , 2-5
s.
Table 3-1. Summary of TCDD bioconcentration factor determinations
for fish 3-11
Table 3-2. Steady-state TCDD bioaccumulation factors for lake trout,
calculated from estimated Lake Ontario water
concentrations in 1987 3-16
Table 3-3. Steady-state biota/sediment accumulation factors (BSAF)
for TCDD 3-20
Table 3-4. Bioaccumulation equivalency factors (BEF) derived for
PCDDS and PCDFs from Lake Ontario lake trout and
sediment data 3-23
Table 4-1. Summary of the toxic effects of TCDD to aquatic life and
wildlife. 4-7
Table 4-2. The base set of mammals and birds included in the
literature search for TCDD wildlife toxicity data 4-43
Table 5-1. Environmental concentrations associated with TCDD risk to
aquatic life and associated wildlife 5-11
in
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FIGURES
Figure 3-1. Fraction of organic chemical freely dissolved in water (fd) if
the total organic carbon binding factor TBFOC=1.5 and log
1^^4,5,6,7, or 8 3-14
Figure 3-2. Predicted TCDD concentration response to loading
reduction in Lake Ontario 3-18
Figure 5-1. Whole fish TCDD concentration versus cumulative
percentile from two national surveys 5-7
IV
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ACKNOWLEDGMENTS
Authors:
The following staff scientists at the Environmental Research Laboratory-Duluth
within the Office of Research and Development were responsible for the preparation of
this report:
Philip M. Cook
Russell J. Erickson
Robert L Spehar
Steven P. Bradbury
Gerald T. Ankley
Reviewers:
The authors acknowledge the input of a number of reviewers. The preliminary
draft of this document was reviewed by the following scientists in September, 1992:
Douglas Endicott, Steven Hedtke, James McKim, John Nichols, Charles Stephan and
Gilman Veith of the U.S. EPA Environmental Research Laboratory-Duluth; Diane
Black, Jack Gentile, David Hansen, Richard Pruell and Norman Rubenstein of the U.S.
EPA Environmental Research Laboratory-Narragansett; Richard Bennett and Anne
Fairbrother of the U.S. EPA Environmental Research Laboratory-Corvallis; Kenneth
Stromborg of the U.S. Department of Interior, U.S. Fish and Wildlife Service, Green
Bay, Wl; and Richard Peterson of the University of Wisconsin-Madison.
A subsequent draft was reviewed in November, 1992 by the following experts
outside of EPA:
William Adams Michael Denison
ABC Laboratories University of California, Davis
Columbia, MO Davis, CA
Lawrence Curtis Gary Heinz
Oregon State University U.S. Department of the interior
Corvallis, OR U.S. Fish and Wildlife Service
Laurel, MD
Peter deFur David Hoffman
Environmental Defense Fund U.S. Department of the Interior
Washington, DC U.S. Fish and Wildlife Service
Laurel, MD
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Robert Hugget Ross Norstrom
College of William and Mary Environment Canada
Glouchester Point, VA Canadian Wildlife Service
Hull, Quebec, Canada
Richard Kimerle Richard Sijm
Monsanto Company University of Utrecht
St. Louis, MO The Netherlands
Derek Muir Glenn Suter II
Fisheries and Oceans Oak Ridge National Laboratory
Freshwater Institute Oak Ridge, TN
Winnipeg, MB, Canada
VI
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LIST OF SELECTED ABBREVIATIONS
Ah
AHH
BAF
BAF,
BCF
BCF,
BEF
BKME
BMP
BSAF
BSSAF
DOC
EC50
ELIZA
ER50
EROD
FAF
oe
HPLC
HRGC/HRMS
IHNV
K
"\DW
LD50
LOAEL
LR50
MFO
MLD
MOA
NOAEL
PCB
PCDD
PCDF
PCH
PeCDD
Aryl hydrocarbon
Aryl hydrocarbon hydroxylase
Bioaccumulation factor
Lipid-normalized bioaccumulation factor
Bioconcentration factor
Lipid-normalized bioconcentration factor
Bioaccumulation equivalency factor
Bleached kraft pulp mill effluent
Biomagnification factor
Biota-sediment accumulation factor
Biota-suspended solids accumulation factor
Dissolved organic carbon concentration in water
Concentration causing a 50% effect
Enzyme-linked immunosorbent assay
Residue concentration causing a 50% effect
7-Ethoxyresorufin-O-deethyIase
Food accumulation factor
Fraction of chemical freely dissolved in water
Fraction of lipid in organism
Fraction of organic carbon in sediment or suspended solids
High pressure liquid chromatography
High resolution gas chromatography/high resolution mass
spectrometry
Infectious hematopoietie necrosis virus
Aquatic organism/water equilibrium partition coefficient
Chemical uptake rate constant
Chemical elimination rate constant
Organic carbon/water partition coefficient
Octanol/water partition coefficient
Lethal dose to 50% of exposed organisms
Lowest observed adverse effect level
Residue concentration causing 50% lethality
Mixed-function oxygenase
Minimum level of detection
Mode of action
No observed adverse effect level
Polychlorinated biphenyl
Polychlorinated dibenzodioxin
Polychlorinated dibenzofuran
Polychlorinated hydrocarbons
Pentachlorodibenzodioxins
vn
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PeCDF Pentachlorodibenzofurans
POC Participate organic carbon concentration in water
Rw Aquatic organism/sediment disequilibrium ratio
Raw Aquatic organism/water disequilibrium ratio
R^, Water/sediment disequilibrium ratio
Sm Molar solubility in water
ssBAF Steady-state bioaccumulation factor
ssBCF Steady-state bioconcentration factor
Sw Water solubility
TBF,,,; Total organic carbon binding factor
TCDD 2,3,7,8-Tetrachlorodibenzo-p-dioxin
TCDF 2,3,7,8-Tetrachlorodibenzofuran
TEC Toxlcity equivalency concentration
TEF Toxicity equivalency factor
VIII
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EXECUTIVE SUMMARY
INTRODUCTION
Background
In April, 1991 the Administrator of the U.S. Environmental Protection Agency
(EPA) announced that the Agency would conduct a scientific reassessment of the risk
of 2,3,7,8-tetrachIorodibenzo-p-dioxin (hereafter referred to as TCDD), and similar
chemicals, to human health and the environment. Since 1985, EPA has classified
TCDD, which it considers the most potent known animal carcinogen, as a probable
human carcinogen. Sources of TCDD in the environment were subsequently
regulated on the basis of animal cancer rates extrapolated to doses associated with
human exposures. Recently, consensus has developed that the toxic effects of TCDD
appear to be mediated by its binding to a receptor protein, the Ah cytosolic receptor.
This conclusion has led to the supposition that a receptor-based mode! is appropriate
for characterizing TCDD risk. As a consequence, EPA designed and implemented a
reassessment research plan, founded on biologically-based dose-response models for
TCDD and related chemicals, to establish a more scientific basis for credible risk
assessments.
In addition to addressing human health risks of TCDD, the reassessment plan
includes a component on the risks of this compound to aquatic life and associated
wildlife. This component is included because EPA not only recognizes that TCDD in
aquatic environments can be a major contribution to overall human exposure through
fish and shellfish consumption, but also that piscivorous fish and wildlife may be
particularly at risk due to their large exposure through aquatic food chains. The
limited available toxicological data indicate that fish, especially salmonid sac fry, and
mink (Mustela vison) are among the most sensitive animals to TCDD and related
compounds. Therefore, independent of the human health risks, there is an increased
concern for assessing ecological risks.
Research to provide the needed exposure and effects information to
characterize the risk of TCDD to aquatic life began in September, 1991, and is
continuing. The experimental design for this research was formulated on existing data
gaps and employs an approach founded on the Ah receptor-based mode of action.
The design also addresses the definition of sensitive and ecologically-relevant toxic
effect endpoints, a variety of potentially threatened species and complex exposure
relationships. Overall, the research is based on major uncertainties regarding TCDD
exposures and toxic effects, but is also designed to provide scientific data of the
quality and quantity required for the development of an aquatic life-based EPA water
quality criterion for TCDD. The goal of the EPA Office of Research and
Development's aquatic effects research plan is to develop data and models needed to
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complete a comprehensive final report on data and methods for TCDD aquatic risk
assessment for review in June, 1995.
Goal and Scope of the Interim Report
The goal for this report was to critically review and evaluate relevant published,
and in some instances unpublished, data and models currently available for analyzing
TCDD exposure to, and effects on, aquatic life and wildlife and to identify major areas
of uncertainty that are expected to limit how well related risks can be characterized. It
offers a critical analysis of currently available exposure and effects information and
outlines important principles and concepts that must be considered when these data
are later integrated to characterize risk. Specifically, this report addresses TCDD
exposure to and bioaccumulation in aquatic organisms, TCDD toxic effects on aquatic
life and wildlife, and aspects of risk characterization to exemplify approaches and
applicability of current information.
The resulting analyses have two uses. First, as discussed previously, analysis
of available data assists research planning by identifying data gaps. Second, by
collecting and organizing existing information in line with EPA's recent report
"Framework for Ecological Risk Assessment", such analyses serve as first steps in
planning a risk assessment. That is, for future risk assessments associated with
TCDD and related chemicals in aquatic food webs, the available data can be
examined specifically in terms of its suitability to form a "conceptual model", consistent
with the EPA framework for ecological risk assessment.
This interim report has a limited scope. The analyses presented specifically
address the direct toxic effects of TCDD to aquatic life and wildlife based on uptake
from aquatic prey, sediment and surface water. This report does not provide generic
or site-specific TCDD risk assessments nor TCDD risk assessments for specific
aquatic life or wildlife species. Information on polychlorinated and polybrominated
dibenzo-p-dioxins, dibenzofurans, and planar polychlorinated biphenyls, that have the
same mode of toxic action as TCDD, is not treated in detail; however, mixtures of
these compounds certainly must be considered to ultimately characterize the
ecological risk of TCDD. The final report on aquatic ecological risk assessment
elements will attempt to incorporate the contribution of TCDD-related chemicals in
assessing the risk of complex mixtures to aquatic life and wildlife.
Techniques to assess TCDD risk in aquatic ecosystems are hampered by
limited exposure and effects data as well as generic shortcomings that confront most,
if not all, ecological risk assessments for chemical stressors. Using currently available
dose-response relationships for reproductive and/or developmental impairment, the
interim risk characterization approach provides techniques to relate TCDD
concentrations in water, sediment and fish tissues to a low or high likelihood of
population failure in aquatic life and wildlife. More precise probability estimates of
aquatic life and wildlife population responses are not currently possible because of the
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limited dose-response relationships available and because, in general, validated
species-specific population dynamic models do not exist. With more refined TCDD
dose-response relationships, well-developed aquatic life and wildlife population
models, and a better understanding of the interaction of additional chemical and non-
chemical stressors, the contribution of TCDD exposure to the probability of changes in
population dynamics could be quantified with increased certainty in the future. Finally,
this report does not discuss the impact of TCDD on community or ecosystem structure
and function and the resultant interrelated effects on aquatic life or wildlife populations.
To assess the risk of TCDD to communities and ecosystems will require models, with
parameters specified appropriately for intended applications, to evaluate and forecast
responses as a function of current or future TCDD concentrations in water, sediment
and biota.
Peer Review
As outlined in the Acknowledgements Section, this report has undergone
extensive review. The report was first reviewed by U.S. EPA scientists within the
Office of Research and Development and two non-EPA scientists who are undertaking
particularly relevant research on TCDD and related chemicals. A subsequent draft of
the document was then reviewed by twelve scientists from academia, industry, non-
EPA governmental agencies and private organizations. The current document reflects
the comments and suggestions provided through this review process.
This report provides an initial base of information and analyses that EPA is
planning to use for assessing risks of TCDD to aquatic life and wildlife. This report will
provide a working document for a meeting of EPA scientists and ecological risk
assessment experts who will further evaluate the present data limitations and
uncertainties that should be incorporated into plans for completing a comprehensive
final report on TCDD risk assessment in 1995. The peer panel will also evaluate the
applicability of the data and methods for actual TCDD risk assessments following the
U.S. EPA's framework for ecological risk assessment.
EXPOSURE
In natural environments, TCDD is typically associated with sediments, biota and
the organic carbon fraction of ambient waters. This distribution of TCDD in aquatic
systems is a function of its low water solubility and extreme hydrophobicity. Water
solubility measurements for TCDD indicate that a range of 12 to 20 ng/L is likely for
colder water (4 to 12°C), but the extent to which the solubility increases at warmer
temperatures is more uncertain due to the limited data and variable results reported.
The true activity of TCDD in water (freely dissolved or bioavailable concentrations) is
uncertain because current analytical chemistry techniques and associated minimum
levels of detection are not sufficiently sensitive. Concentrations of total TCDD (freely
dissolved and organic carbon-bound) in ambient waters have also not been measured
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to date; however, lower levels of detection through improved high resolution gas
chromatography/high resolution mass spectrometry techniques may permit such
measurements in the future. Estimates for log octanol/water partition coefficients (log
K^s) and log carbon/water partition coefficients (log Kocs) range from approximately
6.5 to 8.0 and from 6.5 to 7.5, respectively. Based on an analysis of a variety of
measurements and models, the log K^ and log Koc for TCDD were both estimated to
be 7.0 for this report; however, a wide range of uncertainty exists and the "true" log
and log K^. are likely to be greater than 7.0.
Because of its lipophilicity, and low rates of chemical and biological degradation
in aquatic environments, TCDD does accumulate in biota to detectable levels. When
interpreting and comparing TCDD residue accumulation in aquatic organisms, it is
important to realize that exposure occurs through combinations of water, sediment,
and dietary routes that are influenced by species-specific differences in physiology,
bioenergetic condition and habitat, as well as site-specific TCDD bioavailability. Two
EPA national surveys undertaken during the mid 1980s indicate that most areas
across the United States had TCDD fish concentrations ranging from non-detectable
(less than 0.5 to 1 .0 pg/g) to 1 .0 pg/g on a whole body, wet weight basis. Fish from
open Great Lakes sites and some rivers and estuaries downstream from kraft paper
mills had concentrations in the range of 1 to 20 pg/g, while fish taken from specific
sites in Lake Ontario and other inland freshwater sites had concentrations in the range
of 20 to 100 pg/g. There have also been reports of occasional fish collected with
concentrations in excess of 100 pg/g. On the basis of limited samples and sites,
TCDD residues in fish appear to have decreased over the past decade.
TCDD also accumulates in sediments to measurable levels; however, there
have been no national surveys undertaken that are comparable to those for fish.
Sediment monitoring studies have tended to focus on sediments from areas known to
be contaminated. Based on dated sediment core samples from Lake Ontario sites
contaminated by loading from the Niagara River, TCDD levels were nondetectable
prior to 1940 and reached a maximum level of approximately 500 pg/g dry sediment in
1962. By 1987, the average surface sediment TCDD concentrations declined to 68
pg/g. Sediment samples taken from the Newark Bay, New Jersey estuary have had
TCDD concentrations ranging from 730 to 7,600 pg/g dry sediment in association with
herbicide production from 1948 to 1969.
BIOACCUMULATION
Lipid-normalized bioaccumulation factors (BAFts) for TCDD derived from lake
trout (Satvetinus namaycush) and sediment data from Lake Ontario combined with
chemical mass balance model predictions of TCDD concentrations in water, are
approximately 1-106 for total TCDD concentration in water and 3-1 06 for freely
dissolved TCDD, assuming a log K^ of 7.0. BAFs for whole fish or fish tissue of a
particular lipid content are determined by multiplying the BAF4 by the fraction of lipid in
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the fish. The BAFjS can be applied to waters other than Lake Ontario by adjusting for
the effect of dissolved organic carbon on the fraction of TCDD that will be freely
dissolved (fd). The BAF, for total TCDD is approximately 3-106
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Because of uncertainties in establishing the bioavailabiiity of TCDD in aqueous
solutions, measured TCDD concentrations in test organisms, as opposed to aqueous
TCDD concentrations, are a more useful metric of expressing aquatic toxicity dose-
response relationships. Largest TCDD concentrations reported not to cause effects in
algae, snails, cladocerans and bullfrogs (Rana catesbeiana) are 2,295,000, 502,000,
1,570,000 and 1,000,000 pg/g, respectively, in contrast to a concentration range of 47-
65 pg/g in lake trout eggs which can cause significant mortality in the hatched fry.
Measured concentrations in the eggs and tissues of fish indicate that young fish are
more sensitive to TCDD than older fish, and that delivered doses to eggs that elicit
adverse effects in salmonid sac fry are similar regardless of the route of egg exposure
(i.e., waterborne, egg injection or maternal transfer). There is no evidence that
adverse effects occur in any fish species if egg concentrations are less than 34 pg/g.
This likely corresponds to an accumulation in parent fish, with lipid content similar to
the eggs, of less than 50 pg/g. However, depending on fish species, effects on fish
fry survival are expected to be substantial when accumulations in the eggs are 50 to
500 pg/g, which corresponds to a range of 75 to 750 pg/g in parent fish. Substantial
mortality to older fish is expected to occur when total body accumulations are in the
range of 1,000 to 15,000 pg/g.
To date, lake trout sac fry mortality due to low concentrations of TCDD in eggs
has been one of the most sensitive and ecologically-relevant endpoints identified. This
mortality in salmonids has been associated with a stress-syndrome commonly referred
to as blue-sac disease. The dose-response curve for survival of hatched fry is very
steep. For example, a lake trout egg concentration of 34 pg/g has not been shown to
cause adverse effect but 104 pg/g causes complete mortality. Preliminary analyses
based on assumptions of TCDD bioavailabiiity during waterborne egg exposures
suggest that other fish, such as the northern pike, are nearly as sensitive, if not more
so, than lake trout. Sensitivity also appears to vary between salmonid species and
within different strains of the same species.
In juvenile and adult fish, sublethal effects such as fin necrosis, lesion
development, histopathological changes, enzyme induction (eytochrome P4501A1) and
immune suppression appear to occur in fish at TCDD tissue concentrations that
ultimately cause lethal effects. These results indicate that such responses may have
applicability as screening parameters for assessing contamination due to TCDD and
related compounds.
The laboratory toxicity information shows that fish are generally more sensitive
than aquatic plants, aquatic invertebrates and other aquatic vertebrates such as
amphibians. However, the database regarding this sensitivity is limited and the
exposures are not always readily comparable to the data on fish. It is possible that
aquatic ecosystem components other than fish are sensitive to TCDD, but simply have
not been adequately tested. Additional long-term tests that utilize a greater portion of
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the species' life cycle are needed to more definitively establish dose-response curves
for these groups of seemingly insensitive organisms.
Few TCDD epidemiological studies of aquatic life populations have been
reported. Such studies are complicated because TCDD exposures to aquatic
organisms in natural systems are confounded by the presence of other chemical and
non-chemical stressors that may add to, or otherwise modify, the effects of TCDD.
Despite these difficulties, investigations of lake trout populations in Lake Ontario
provide some insights into the relationship of TCDD exposure to reproductive success
that are consistent with laboratory findings that lake trout sac fry are very sensitive to
TCDD.
Prior to the onset of severe TCDD contamination in Lake Ontario in about 195C
lake trout populations had declined precipitously due to overfishing, sea lamprey
predation and habitat degradation. Despite a lake trout stocking and management
program (i.e., controls on the lamprey and habitat improvement) which began in 1971,
populations of naturally reproducing lake trout have not been achieved. In 1978, lake
trout eggs contained approximately 30 pg TCDD/g and a 48% incidence of blue-sac
disease, and associated mortality, was reported for hatchery-raised sac fry collected
as eggs in 1979. If all of this reported incidence of blue-sac disease were due to
TCDD and toxieologically-related chemicals, TCDD is estimated to have accounted foi
half of the toxicity equivalents. Recent reports, which indicate that some lake trout sa
fry are reappearing in Lake Ontario and that there is an absence of blue-sac disease,
suggest an improvement in reproduction may be occurring. In 1988, TCDD lake trout
egg concentrations were reduced to 10 pg/g, which is consistent with declines in
TCDD concentrations in the sediments (since the 1960s) and adult lake trout, since
their reintroduction in the 1970s, The most recent measurements of TCDD
concentrations in lake trout eggs are approximately one third of the no-effect level
derived from laboratory toxicity tests. Although there is consistency between single
chemical dose-response curves for TCDD and related compounds and the recent
history of Lake Ontario lake trout populations, further study is needed to define dose-
response curves for the specific chemical mixtures accumulating in lake trout eggs.
In a potentially complementary approach to assessing risks of chemical
mixtures, eytoehrome P4501A1 induction in some populations of fish has been
explored as an indicator of chronic exposure to TCDD and related chemicals. To
date, the biochemical indicator studies have not established a clear linkage between
toxic effects and TCDD exposure.
Wildlife
There are few TCDD toxicity studies reported for mammalian and avian wildlife
species and apparently no information exists for TCDD toxicity to reptiles.
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Based on a single dose LD50 value of 4,200 pg/g, the mink is one of the most
sensitive mammals evaluated thus far and is comparable in sensitivity to the guinea
pig. TCDD administered by intraperitoneal (i.p.) injection to newborn mink at 100 and
1,000 pg/g/day for 12 consecutive days resulted in 100% mortality within 14 days at
the high dose and 62% mortality by 19 weeks at the lower dose. No reproduction and
developmental bioassays for mammalian wildlife were found in the literature; however,
rat and Rhesus monkey (Macaca mulatta) bioassays indicated threshold effect levels
for offspring production and survival of 1 and 0.13 pg TCDD/g/day, respectively.
Based on these values, and the greater sensitivity of the mink compared to the rat
based on acute and subehronic exposures, a threshold level of 0.1 pg TCDD/g/day
was estimated for primarily piscivorous mammalian wildlife species, such as the mink
and river otter (Lutra canadensis). This oral intake value is equivalent to a dietary fish
concentration of 0.5 to 1.0 pg TCDD/g, assuming dietary consumption rates as a
percentage of body weight equal to 10 to 20%.
Based on currently available data for avian species, it appears that gallinaceous
birds are the most sensitive to TCDD. Of these birds, the ring-necked pheasant
(Phasfanus colchicus) and the chicken are among the most sensitive with mortality
rates of 80 to 100%, respectively, reported for a single dose of 25,000 pg/g. The
available evidence suggests that effects on reproduction are of particular concern;
however, only one reproduction bioassay has been reported. Results from this study,
which consisted of a 10 week i.p. injection dosing regime to female ring-necked
pheasant, indicated that a female adult dose equivalent to 140 pg/g/day resulted in
100% embryo mortality, while a dose of 14 pg/g/day appeared to be a threshold level
for no effect. Several egg injection studies completed with the chicken and ring-
necked pheasant indicate that egg concentrations of 100 to 500 pg/g are threshold
levels for embryo mortality, the chicken being the more sensitive species. Considering
uncertainties associated with the ring-necked pheasant reproduction study (i.e., study
duration and interspecies sensitivity), the adult pheasant threshold value of 14 pg
TCDD/g/day for embryo mortality was extrapolated to a value of 1.4 pg TCDD/g/day
for representative piscivorous avian wildlife species, such as raptors, wading birds and
diving ducks. This oral intake value is equivalent to a dietary fish concentration of 3 to
14 pg TCDD/g, assuming dietary consumption rates as a percentage of bird body
weight equal to 10 to 50%.
Results from limited epidemiology studies are consistent with laboratory-derived
threshold levels for TCDD impairment of reproduction in avian wildlife. Population
declines in herring gulls (Larus argentatus) on Lake Ontario during the early 1970s
coincided with egg concentrations of TCDD and related chemicals expected to cause
reproductive failure based on laboratory experiments (TCDD concentrations in excess
of 1,000 pg/g). Improvements in herring gull reproduction through the mid-1980s were
correlated with declining TCDD concentrations in eggs and lake sediments. Studies at
sites in British Columbia, Canada also showed an inverse relationship between Great
Blue heron (Ardea herodias) chick growth and incidence of edema and TCDD egg
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concentrations. The herring gull and heron data indicate that successful reproduction
exists in wild bird colonies even though TCDD is present in the eggs at concentrations
between 200 to 500 pg/g. In turn, toxic effects at the individual level are associated
with concentrations above 100 pg/g. These concentrations are consistent with
findings from laboratory-based egg injections studies, summarized previously. Based
on residue monitoring studies in Lake Ontario that relate TCDD herring gull egg
concentrations to TCDD fish concentrations, the threshold value for TCDD
concentrations in fish associated with successful herring gull reproduction in Lake
Ontario was calculated to be 24 pg/g. This value is consistent with the independently
derived fish concentration threshold value of 3 to 14 pg/g extrapolated from the ring-
necked pheasant reproduction study.
The limited data available for TCDD contribute to uncertainty in establishing the
mammalian and bird effect profiles. The estimates of fish TCDD contamination that
pose a risk to wildlife are for organisms which are completely piscivorous; other
dietary sources can alter exposures. For example, mink that do not primarily forage
on fish and aquatic invertebrates could have approximately 50% lower TCDD
exposures, while raptors that may consume fish-eating birds could have approximately
50% higher TCDD exposures. The limited toxicological databases also contribute to
uncertainties, as reflected in the use of interspecies and subchronic to chronic
extrapolation factors in the mammalian and avian profiles, respectively. Because
there are essentially no toxicokinetic and toxicodynamic data, it is very difficult to
address these extrapolation uncertainties quantitatively. Finally, the dose-response
curves currently available are based on large differences in treatment levels, which
also contributes to uncertainty for establishing the probability of effects for a given
exposure.
RISK CHARACTERIZATION METHODOLOGY
In an ecological risk assessment for a chemical stressor, risk characterization is
the phase in which the results of exposure and effects analyses are integrated to
evaluate the likelihood of adverse effects in exposed organisms, populations,
communities, or ecosystems based on actual or projected exposures of organisms to
the chemical, or suite of chemicals, in the environment. The degree to which risk is
characterized can vary markedly, but ideally involves a quantitative scale of effects
and estimation of probabilities and uncertainties. Current information is insufficient to
provide such a thorough description for TCDD risk to aquatic life and associated
wildlife, with quantification of uncertainty being particularly difficult given the limited
knowledge base. Furthermore, a thorough assessment of TCDD risk should consider
its joint action with other contaminants and non-chemical stressors and the expression
of effects on individual organisms at a population and community level; such
techniques are even less well established and must await further development.
However, the adequacy of a risk characterization depends on the nature of the
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specific problem of interest, so current information can be adequate to characterize
TCDD risk to aquatic life and associated wildlife in some cases.
The principal goal of this report was to evaluate and summarize data and
methods that are available for the assessment of TCDD risk to aquatic life and
associated wildlife, and to identify the major uncertainties that currently limit how well
risks can be characterized. A definitive risk assessment for a specific problem was
not a goal of this report. However, it is necessary to discuss how exposure and
effects information should be integrated into a description of risk and it is most
effective to do this in the context of actual problems regarding the risk of TCDD to
aquatic life and associated wildlife. The following examples are presented to illustrate
methods that can be used in risk characterizations and to identify major uncertainties
that should be of concern. The degree to which these problems were evaluated was
limited to that needed to accomplish this limited purpose, so these examples should
not be treated as complete risk characterizations.
Fish Contamination in the United States: Risk to Aquatic Life and Associated
Wildlife
In two national EPA surveys for a large number of diverse sites, fish were
analyzed for TCDD and other chemical contaminants. These surveys provide the best
data set for consideration of TCDD risk on a national basis and at different classes of
sites, although effects can only be considered on the basis of accumulation in fish and
cannot reliably be related back to environmental concentrations and source loadings.
Although there are limitations in the geographical sampling designs associated with
these fish residue data, they can be used to judge how extensive and severe risk of
TCDD to aquatic life and associated-wildlife might be on a national and regional scale.
When contrasted with TCDD concentrations in fish established in this report to be of
low to high risk to sensitive fish, birds, and mammals, the survey data suggest that
TCDD contamination is below levels of concern for aquatic life at all but a small
percentage of the sites nationwide. However, there are sites where TCDD
concentrations are high enough to pose significant risk to fish. This is especially true
if joint action with other chemicals is considered. For wildlife, significant risk is
potentially more widespread than for aquatic life, which is expected since the effects
profile developed in this report suggests completely piscivorous wildlife are much more
susceptible to TCDD than fish. In particular, TCDD residues in fish associated with a
low risk for mammals are exceeded for about half of the samples in these national
surveys. This characterization must be qualified by the recognition of uncertainties
cited earlier for wildlife effects data and uncertainties associated with actual wildlife
exposures. Nevertheless, despite these uncertainties, this comparison of fish survey
results and wildlife effects data does raise significant concerns about the present risk
of TCDD to piscivorous wildlife.
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Lake Trout Populations in Lake Ontario
Lake trout are a Great Lakes ecosystem quality indicator species and are the
subject of extensive efforts for restoration of naturally reproducing populations. Prior
to the onset of severe TCDD contamination in Lake Ontario in about 1950, lake trout
populations had declined precipitously due to overfishing, sea lamprey predation and
habitat degradation. Despite a lake trout stocking and management program, which
began in 1971, populations of naturally reproducing lake trout have not been achieved.
Laboratory-based information on TCDD residues associated with toxic effects can be
compared to TCDD concentrations observed in Lake Ontario lake trout and their eggs
over the last 15 years. In 1987, eggs collected from Lake Ontario contained about 10
pg TCDD/g wet weight. This concentration is about three-fold lower than the threshold
for sac fry mortality derived from toxicrty studies using a variety of exposure routes.
Therefore, even given the uncertainties in these bioassays, it is unlikely that lake trout
reproduction is currently at risk in Lake Ontario due solely to TCDD effects on sac fry
survival. However, this conclusion is not necessarily appropriate for past conditions in
Lake Ontario. TCDD in surficial sediments and lake trout both declined two- to three-
fold from 1978 to 1988. In 1978, lake trout had average concentrations of 78 pg
TCDD/g wet weight, high enough that TCDD alone would be expected to have effects
on reproduction as a result of sac fry mortality. Based on the sediment record, this
concentration would have been several-fold higher in the early 1960s, and well above
that which would have precluded successful reproduction.
Based on the data from 1987-1988, TCDD, in concert with other polychlorinated
dibenzo-p-dioxins, polychlorinated dibenzofurans and planar polychlorinated biphenyls,
may still be contributing to the continuing Lake Ontario lake trout reproduction
problem. TCDD concentrations in eggs are at about one-fifth the lethal residue
expected to cause 50% mortality, which could be a significant component in a
complex mixture. The joint toxicrty of TCDD and related chemicals is a major
uncertainty that needs to be addressed. There also is the issue of whether the
measurement endpoint used here, sac fry survival in a laboratory environment, is an
adequate surrogate for the assessment endpoint of interest, namely lake trout
reproduction in a natural system. Other toxicological endpoints associated with
reproductive physiology may be more relevant and sublethal effects on fry might also
affect their survival in a natural environment. These uncertainties need to be
addressed to adequately assess risk of TCDD to Lake Ontario lake trout, and
represent major uncertainties confronting aquatic life risk assessments in general.
Environmental Concentrations Associated with TCDD Effects
Some EPA regulatory activities, such as the establishment of water and
sediment quality criteria, require the specification of environmental concentrations that
are considered to represent acceptably low risk. These concentrations are set
generically and are applied to specific sites to calculate allowable discharges or set
goals for remedial actions. Such procedures do not fit the risk paradigm in which
effects and exposure information are combined into a statement of risk, after which
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decisions about managing risk are made. Rather, water and sediment quality criteria
incorporate risk management decisions before specific exposure information is
Introduced into the process. Nevertheless, the setting of such criteria still contain
major elements of risk assessment, albeit an incomplete one, and the entire regulatory
process includes all the elements of risk assessment, although somewhat intertwined
with risk management.
Information presented earlier on effects and their relationship to exposure
conditions can be used to specify concentrations in water and sediment that will be
associated with different levels of risk to aquatic life and aquatic-associated wildlife
(see Table E-1). This exercise serves as a simple demonstration of how the
Information reviewed in this document can be used and is not intended to be a
complete and definitive characterization of TCDD risk and associated uncertainties.
Consistent w'rth the presentation of aquatic life and wildlife effect profiles, the results of
these analyses indicate that completely piscivorous wildlife, and especially piscivorous
mammals, are at greatest potential risk to TCDD exposure in aquatic ecosystems.
Table E-1. Environmental concentrations associated with TCDD risk to aquatic life and associated wildlife.
Organism
Fish
Concentration
(pg/i)
Sediment
Concentration
(pg/g dry wt.)
Water Concentration
(P9/L)
roe=o,2
POC=1.0
Low Risk
Fish
Mammalian Wildlife
Avian Wildlife
50
0,7
6
60
2,5
21
0.6
0,008
0.07
3.1
0.04
0.35
High Risk to Sensitive Species
Fish
Mammalian Wildlife
Avian Wildlife
80
7
60
100
25
210
1.0
0,08
0.7
5
0.4
3.5
Fish lipld of 8% and sediment organic carbon of 3% assumed where needed.
For risk to fish, BSAF of 0.3 used; for risk to wildlife, BSAF of 0,1 used.
Low risk concentrations are derived from no-affects thresholds tor reproductive effects (mortality in embryos and young)
In sensitive species.
High risk concentrations are derived from TCDD doses expected to cause 50 to 100% mortality in embryos and young
of sensitive species.
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RESEARCH NEEDS FOR REDUCING UNCERTAINTIES
In the preparation of this interim report important and immediate uncertainties
associated with characterizing the risks of TCDD to aquatic life and wildlife have been
identified. The major knowledge gaps that are highlighted below are intended to be
specific to TCDD only and do not address uncertainties associated with mixtures of
polyhalogenated dibenzodioxins, dibenzofurans and biphenyls. The following
discussion also does not address uncertainties associated with the lack of appropriate
population, community and ecosystem level models and their linkage to toxicological
inputs. Clearly, the development of such models are needed for improving ecological
assessments of chemical stressors in general.
Exposure
Research is needed to better establish the K^ for TCDD because current
uncertainty in this parameter leads to increased uncertainty in extrapolating TCDD
BAFs across aquatic habitats and in estimating the bioavailable fraction of TCDD in
water. Current understanding of the role of organic carbon on the partitioning and
bioavailability of TCDD from water and sediment is primitive. Analytical procedures
are currently inadequate to reliably measure TCDD in water at the concentrations
expected to elicit adverse effects in aquatic life and wildlife; this situation is especially
critical for monitoring dissolved TCDD. As a consequence of this analytical deficiency,
there are very few reports of TCDD concentrations in natural waters and the values
that are reported are uncertain. With the development of new techniques and
instrumentation that lower detection limits, measurements of total and dissolved TCDD
in a variety of water systems need to be made to better establish current and future
exposures.
Bioaccumulation
There are currently very few reliable TCDD BAFs available for natural systems,
which makes predictions of TCDD concentrations in fish tissues based on TCDD
concentrations in water or surficial sediments difficult to assess. Measurements of
TCDD in biota, surficial sediments and water are needed for a variety of natural
systems to develop a database of sufficient scope to calibrate and validate predictive
bioaccumulation models. In gathering such data particular attention must be given to
organism attributes (e.g., lipid content) and water column and sediment properties
(e.g., particulate and dissolved organic matter) that might influence bioaccumulation.
Effects
Current data suggests fish are more sensitive than aquatic invertebrates and
amphibians to TCDD, perhaps due to the absence of the Ah receptor, or a comparably
sensitive receptor, A more rigorous assessment is needed to determine if the Ah
receptor is present in aquatic invertebrates, amphibians and reptiles. Based on results
from such a study specific toxicity studies could be undertaken to insure that current
conclusions on interspecies differences in TCDD sensitivity are valid. Associated with
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this testing, it is critical that reproductive studies be incorporated and that TCDD
accumulation be monitored to establish the basis to characterize risk with increased
certainty.
Current data suggest fry survival as the most critical endpoint for fish; however,
the current threshold values are based on exposures of limited durations in two
salmonid species. As a consequence there is uncertainty regarding the species
sensitivity distribution, as well as the impact of chronic exposures on reproduction.
Effects on reproductive physiology, early life stage development, and immune
response should be studied in a diversity of species to reduce these uncertainties.
Associated with such studies it is critical that TCDD accumulation be monitored and
that supportive toxicokinetic studies be undertaken to establish biologically-based
dose-response models.
With both the mammalian and avian wildlife risk characterizations, there is a
lack of quality reproduction bioassays and toxicokinetic information to establish well-
defined dose response relationships. For the mammalian assessment there were no
reproduction bioassays available for a representative piscivorous wildlife species (e.g.,
the mink) and for the avian assessment, a ring-necked pheasant reproduction
bioassay of limited duration and incorporating an i.p. exposure regime was the only
study reported in the literature. Finally, there are apparently no data available to
assess the toxicity of TCDD to reptiles. To reduce current uncertainties in the wildlife
risk characterization, long-term (i.e., one generation) feeding studies in the mink and
an avian species are needed to more adequately assess reproductive and
developmental effects. These bioassays should be supported by toxicokinetic studies
and Ah receptor investigations to develop biologically-based dose response models to
better establish critical parameters in species extrapolations and in linking TCDD
accumulation to toxic effects. The avian study should also be designed in such a way
to better quantify the uncertainties of using egg injection studies as a source of
toxicological data in avian hazard assessments.
Finally, there is a need to assess TCDD concentrations in aquatic life and
wildlife in association with population assessments in appropriately selected natural
systems. Such studies would further refine and validate the reliability of relationships
between TCDD accumulation and toxic effects that have been developed based on
currently available laboratory and ecoepidemiological investigations.
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1. INTRODUCTION
1.1 BACKGROUND
In April, 1991 the Administrator of the U.S. Environmental Protection Agency
(EPA) announced that the Agency would conduct a scientific reassessment of the risk
of 2,3,7,8-tetrachIorodibenzo-p-dioxin (hereafter referred to as TCDD), and similar
chemicals, to human health and the environment. Since 1985, EPA has classified
TCDD, which it considers the most potent known animal carcinogen, as a probable
human carcinogen. Sources of TCDD in the environment were subsequently
regulated on the basis of animal cancer rates extrapolated to doses associated with
human exposures. Two major activities, however, prompted the decision to reassess
this approach in evaluating TCDD toxicity and its associated risks. First, an
epidemiological study of cancer mortality in U.S. chemical workers (Fingerhut et al.,
1991) by the National Institute of Occupational Safety and Health provided evidence of
TCDD-mediated human carcinogenicity. Second, a conference at the Banbury Center
of the Cold Spring Harbor Laboratory in New York in October, 1990 resulted in
general agreement that TCDD's mode of action involves the activation of a TCDD-
receptor complex and its subsequent translocation into the cell nucleus as a
necessary, but not sufficient, prerequisite for any TCDD related effects (Scheuplein et
al., 1991). It was also generally agreed that: 1) animals which possess an aryl
hydrocarbon (Ah) receptor respond similarly to TCDD; 2) multiple effects, including
enzyme induction, immunotoxicity, reproductive toxicity, developmental toxicity and
carcinogenicity, occur in all susceptible species; 3) chemicals which are isostereomers
of TCDD (i.e., polychlorinated and polybrominated dibenzo-p-dioxins, dibenzofurans)
and planar biphenyls may act through the same mode of toxic action; and 4) since all
toxic effects of TCDD appear to be mediated by the chemical's binding to a receptor
protein within the cytoplasm of a cell, a receptor-based risk assessment model for this
chemical is appropriate. Based on these findings, EPA identified a need for
biologically-based dose-response models for TCDD, and related chemicals, to
establish a scientific base for more credible risk assessments. As a consequence, the
EPA designed and implemented a TCDD reassessment research plan.
In addition to addressing human health risks of TCDD and related chemicals,
the reassessment plan includes a component on the risks of these compounds to
aquatic life and associated wildlife. This component is included because EPA not only
recognizes that these chemicals in aquatic environments can be a major contribution
to overall human exposure, through fish and shellfish consumption, but piscivorous
fish and wildlife may be at risk due to their exposures through aquatic food chains.
The limited available toxicological data also indicate that fish, especially salmonid sac
fry (Walker et al., 1991; Cook et al., 1991), and mink (Mustela visori) (Aulerich et al.,
1988; Hochstein et al., 1988; Auterich et al., 1985; Bleavins et al., 1980; Aulerich and
Ringer, 1977) are among the most sensitive animals to TCDD and related compounds.
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Research to generate data and models for characterizing the risk of TCDD and
related compounds to aquatic life (Cook et al,, 1992) began in September, 1991 and is
continuing. The experimental design is based on a toxicological perspective founded
on the Ah receptor-based mode of action and addresses existing ecotoxieological data
gaps. The design is formulated on the definition of sensitive and ecologically-relevant
toxic effect endpoints, a variety of species and complex exposure relationships.
Overall, the plan not only treats major uncertainties regarding exposures and toxic
effects, but is also designed to provide scientific data of the quality and quantity
required for the development of an aquatic life-based EPA water quality criterion for
TCDD. The goal of the EPA Office of Research and Development's aquatic effects
research plan is to develop data and models needed to complete a comprehensive
final report on data and methods for TCDD aquatic risk assessment for review in
June, 1995. This report will contain the data and scientific interpretation of TCDD-
related risk elements for aquatic life and associated wildlife that are needed by a
number of EPA program and regional offices for their regulatory activities.
1.2 GOAL AND SCOPE OF THE INTERIM REPORT
The goal for this report was to critically review and evaluate relevant published,
and in some instances unpublished, data and models currently available for analyzing
TCDD exposure to, and effects on, aquatic life and wildlife and to identify major areas
of uncertainty that limit how well this risk can be characterized. It offers a critical
analysis of currently available exposure and effects information and outlines important
principles and concepts that must be considered when integrating these data to
characterize risk. This interim report addresses TCDD exposure to and
bioaccumulation in aquatic organisms, TCDD toxic effects on aquatic life and wildlife,
and aspects of risk characterization to exemplify approaches and applicability of
current information. The report is intended for those individuals familiar with the
scientific literature concerning TCDD; however, key reviews, in addition to primary
references, are cited throughout the document for those readers who wish to consult
additional background material.
The resulting analyses have two uses. First, as discussed previously, analysis
of available data assists research planning by identifying data gaps. Second, by
collecting and organizing existing information in line with EPA's recent report on a
framework for ecological risk assessment (U.S.EPA, 1992c), such analyses serve as
first steps in planning a risk assessment. That is, for future risk assessments
associated with TCDD and related chemicals in aquatic food webs, the available data
can be examined specifically in terms of its suitability to form a "conceptual model",
consistent with the EPA framework for ecological risk assessment.
This interim report has a limited scope. The analyses presented specifically
address the direct toxic effects of TCDD to aquatic life and wildlife based on uptake
from aquatic prey, sediment and surface water. This report is not intended to provide
1-2
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detailed generic or site-specific TCDD risk assessments nor TCDD risk assessments
for specific aquatic life or wildlife species. Information on polychlorinated and
polybrominated dibenzo-p-dioxins, dibenzofurans, and planar polychlorinated
biphenyls, that have the same mode of toxic action as TCDD, is not directly treated in
this interim report; however, mixtures of these compounds certainly must be
considered to ultimately characterize ecological risk. The final report on aquatic
ecological risk assessment elements will attempt to incorporate the contribution of
TCDD-related chemicals in assessing the risk of complex mixtures to aquatic life and
wildlife.
Techniques to assess TCDD risk in aquatic ecosystems are hampered by
limited exposure and effects data as well as generic shortcomings that confront most,
if not all, ecological risk assessments for chemical stressors. Using currently available
dose-response relationships for reproductive and/or developmental impairment, the
interim risk characterization approach provides techniques to relate TCDD
concentrations in water, sediment and fish tissues to a low or high likelihood of
population failure in aquatic life and wildlife. More precise probability estimates of
aquatic life and wildlife population responses are not currently possible because of the
limited dose-response relationships available and because, in general, validated
species-specific population dynamic models do not exist. With more refined TCDD
dose-response relationships, well-developed aquatic life and wildlife population
models, and a better understanding of the interaction of additional chemical and non-
chemical stressors, the contribution of TCDD exposure to the probability of changes in
population dynamics could be quantified with increased certainty in the future. Finally,
this report does not discuss the impact of TCDD on community or ecosystem structure
and function and the resultant inter-related effects on aquatic life or wildlife
populations. To assess the risk of TCDD to communities and ecosystems will require
physical and mathematical models, with parameters specified appropriately for
intended applications, to evaluate and forecast responses as a function of current or
future TCDD concentrations in water, sediment and biota.
Peer Review
As outlined in the Acknowledgements Section, this report has undergone
extensive review. The report was first reviewed by U.S. EPA scientists within the
Office of Research and Development and two non-EPA scientists who are undertaking
particularly relevant research on TCDD and related chemicals. A subsequent draft of
the document was then reviewed by twelve scientists from academia, industry, non-
EPA governmental agencies and private organizations. The current document reflects
the comments and suggestions provided through this review process.
This report provides an initial base of information and analyses that EPA is
planning to use for assessing risks of TCDD to aquatic life and wildlife. This report will
provide a working document for a meeting of EPA scientists and ecological risk
1-3
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assessment experts who will further evaluate the present data limitations and
uncertainties that should be incorporated into plans for completing a comprehensive
final report on TCDD risk assessment in 1995. The peer panel will also evaluate the
applicability of the data and methods for actual TCDD risk assessments following the
U.S. EPA's framework for ecological risk assessment.
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2. EXPOSURE
The distribution of TCDD and other planar halogenated aromatic chemicals in
aquatic ecosystems and their accumulation in specific animal tissues susceptible to
TCDD intoxication must be understood to assess risks to aquatic life and wildlife. A
residue-based approach is an essential component of ecological risk assessment and
water quality criteria development for TCDD and related chemicals because long term
chemical accumulation in tissues of fish and other aquatic organisms is the best
correlate of dose for evaluating toxicity. Thus, exposure information that impacts
bioaccumulation, tissue-residue, and wildlife dietary exposure relationships used for
this interim report on TCDD aquatic risk assessment are emphasized in this section.
Information on physical-chemical properties; chemical sources; environmental
concentrations in water, sediment and organisms; and aspects of chemical fate and
transport will be included only to the extent necessary to complete a generic TCDD
aquatic risk characterization. More detailed information concerning all aspects of
exposure assessment for chemicals with a toxic mode of action like TCDD can be
obtained from the EPA draft report "Estimating Exposure to Dioxin-Iike Compounds"
(U.S. EPA, 1992a) which presents procedures for conducting site-specific exposure
assessments. The occurrences and magnitudes of TCDD residues in aquatic
organisms are reviewed in this section, but TCDD residues in avian and mammalian
wildlife are not included because TCDD risks to wildlife in this report are estimated
from exposure through ingestion of aquatic organisms, rather than from accumulation
of TCDD in the tissues of wildlife.
2.1 PHYSICAL/CHEMICAL PROPERTIES OF TCDD AND RELATED CHEMICALS
Since fate and transport modeling of TCDD and related chemicals is not within
the scope of this report, only those properties that influence bioavailability during
exposure of aquatic biota are described here. TCDD has a molecular weight of
321.977 daltons and is a crystalline solid at ambient temperatures with a melting point
of 305°C (Boer et al., 1972). The water solubility (SJ of TCDD is approximately 12-
20 ng/L for cold water (ca. 4-12°C). Solubility measurements based on water
equilibration with a thin film of TCDD on glass (Marple et al., 1986a) indicate a range
in Sw of 12.5 to 19.3 ng/L at 22°C. A generator column experiment indicated 12.9
ng/L at 4.3°C but 480 ng/L at 17.3°C (Lodge, 1989). Increase in Sw with water
temperature is expected but the 17.3°C result of Lodge appears to be inconsistent
with the measurements of Marple et al. (1986a) at 22°C. Water with excess TCDD
allowed to equilibrate for 7 months with soil particles and flask surfaces contained only
8 ng/L at 20 to 22°C (Adams and Elaine, 1986). However, a Sw of 420 ng/L at 22.7
°C has been reported for 2,3,7,8-tetrachlorodibenzofuran (TCDF) based on a
generator column experiment (Friesen et al., 1990). Friesen et al. (1990) reported
equivalent correlations of log Sw with chlorine number for both polychlorinated dibenzo-
p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) using generator
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column data obtained at warmer water temperatures and which included the
measurement for TCDD at 17.3 °C from Lodge (1989).
The most important physical-chemical parameter used in TCDD exposure
assessments is the oetanol/water partition coefficient (K^). The log Kw is also
referred to as log P and is used as a measure of the hydrophobicity and lipophilicity of
organic chemicals. Measurement of K,,w introduces a complication into its definition,
which is the need to recognize the distribution of water into octanol and octanol into
water when equilibrium between the phases is achieved (Miller et al., 1985). Thus;
ow - -^ -
^wo Vow
where C^, and Cwo are the concentrations of the solute in octanol saturated with water
and water saturated with octanol, respectively, and y is the activity coefficient of the
solute in each solvent. If octanol in the water does not significantly decrease the
solute activity, as suggested by Miller et al. (1985), the activity coefficient for pure
water (I.e., Tw=Ywo) anc3 the measured Kow (assuming equilibrium is reached; the solute
is not present as micelles, etc.) can be used as a first approximation of the activity of
the solute in water in comparison to organic carbon phases in water, sediments and
biota. For DDT, yw(/yw was measured to be 2.8 (Chiou et al., 1982). Since TCDD is
very Insoluble in water and more hydrophobic than DDT, the possibility for
overestimation of yw should be considered. The value of octanol as a surrogate for
naturally occurring organic phase interactions with chemicals such as TCDD may also
have limitations (Miller et al., 1985). The importance of interactions between organic
solutes and organic solvents is indicated by the observation of increase in free energy
of solution for chlorinated dibenzo-p-dioxins with increasing degree of chlorination due
to an increase of the free energy of vaporization with increasing solute size (Gobas
and Zhang, 1992).
Only two directly measured K^ values for TCDD have been reported. A slow
water phase stirring technique with mutually presaturated octanol and water phases
reached equilibrium within one week and a log K^ range of 6.54 to 6.95 was
measured (Marple et al., 1986b). Slow stirring with an extract of municipal incinerator
fly ash for 9 days gave a TCDD log Kow estimate of 6.4 (Sijm et al., 1989). After 21
days of slow stirring, Sijm et al. (1 989) observed decreases in all KQW measurements
possibly, due to slow micelle formation in the water phase. The degree to which the
KOT, measured at 9 days was so influenced is unknown. A generator column method
was used to measure a log K^ of 6.2 for another tetrachlorinated congener, 1 ,2,3,4-
tetrachlorodibenzo-p-dioxin (Shiu et al., 1988).
Reverse-phase high pressure liquid chromatography (HPLC) with liquid
chromatography/mass spectrometry detection was employed to estimate a log Kow of
7.02 for TCDD from an HPLC retention time - log K^, linear regression equation
2-2
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derived from seven chemicals with measured K^, values (Burkhard and Kuehl, 1986).
Burkhard and Kuehl (1986) also estimated a log K^ of 7.2 for both 1,2,3,4- and
1,3,6,8- tetrachlorodibenzo-p-dioxin. However, HPLC retention was used with three
different calibration chemicals with measured Kow values to estimate log Kow values of
8.6 and 8.7 for 1,2,3,4- and 1,3,6,8-tetrachlorodibenzo-p-dioxin, respectively (Webster
etal., 1985).
CLOGP version 3.54 (Leo and Weininger, 1989) predicts a log Kow of 7.31 from
the molecular structure of TCDD. Since water solubility of TCDD is probably more
reliably measured than K.OVJ, the relationship between molar solubility in water (Sm),
melting point (MP) and K^, (Yalkowsky et a!., 1983) has been used to estimate
IOgSm . .01MP - 0.323 (2.2)
Using TCDD's melting point of 305°C and Sm = 6-10"11 (20 ng TCDD/L), log K^ is
calculated to be 7.94.
Related to K^, is the organic carbon/water partition coefficient, Koc:
s* /->
_ ^ ^
oc
IS _ 00 _ S /ft n\
Koc - j - V2'3)
where Coc is the concentration of chemical in organic carbon associated with
suspended solids or sediment, C^ is the concentration of freely dissolved chemical in
water (theoretically, the chemical activity), Cs is the concentration of chemical in
suspended solids or sediment (dry weight) and foc is the fraction of organic carbon in
suspended solids or sediment (dry weight).
Desorption of TCDD from contaminated soils, followed by 0.45 u,m membrane
filtration to remove particles and associated TCDD from the soil leachates, resulted in
measured log sediment organic carbon - water partition coefficients (log K^) of 7.4 to
7.6 (Jackson et al., 1985). Although data were not presented to demonstrate that an
equilibrium was reached between soil and water, the consistency of results for a single
extraction compared to three successive extractions of fresh soil samples with the
same water suggests that equilibrium was probably achieved. Sorption experiments
using [14C]TCDD with two uncontaminated soils by batch shake testing and isolation of
the water phase by centrifugation (3500 rpm for 30 min) resulted in an estimated log
Koc of 6.66 (Walters et al., 1989). One to five successive prewashings to remove
nonsettleable matter resulted in an observation of ^ dependence on soil water
contact period and/or prewashing. No measurements of the amount of TCDD
associated with organic carbon remaining in the centrifugates were provided. In a
related study, sorption of TCDD to soil from water/methanol mixtures gave an
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estimated log K^ of 6.6 based on co-solvation theory (Walters and Guiseppi-Elie,
1988).
Only one attempt to measure Koc on the basis of freely dissolved TCDD has
been reported (Lodge and Cook, 1989). In this experiment, the equilibrium distribution
of TCDD between water containing decreasing amounts of non-settled organic matter
and settled sediment was measured with the intent to approach the lowest solids
concentration in water that would still allow detection of TCDD. These data
established a minimum value of K^ and, through extrapolation to a pure water
condition, allowed an estimate of K^ based on TCDD activity in water. At 10°C
extrapolation to zero solids concentration in water gave an estimated log Koc of 7.6.
Subsequent measurements have confirmed the attainment of equilibrium between
solids and water and the probability that log K^ for TCDD is equal to or exceeds 7.3
(Lodge, 1992).
"Apparent" log K,,,. values for seventeen PCDD and PCDF congeners were
calculated on the basis of filtered particulate and filtrate ("apparently dissolved")
samples obtained from 1.5 to 2.2 m3 samples of Baltic Sea water (Bronnan et al.,
1991). The log K^ for TCDD was 6.8 with values of 6.8 to 7.9 for congeners with
greater chlorination. The effect of residual colloidal organic carbon in the filtrates is to
increase the fraction of chemical measured as apparently dissolved and this effect is
greater for the more hydrophobia congeners. Thus both individual congener log Koc
values and the range of log K,,,. values for PCDDs and PCDFs would be greater if the
freely dissolved concentrations of each chemical determined on the basis of residual
organic carbon in the filtrate is used to calculate the log K^. Many of the congeners,
including TCDD, were not detectable in the filtrates and the log K^s were calculated
on the basis of half the analytical detection limit. The detection limits for TCDD in
filtrates were much greater than for TCDD in filtered particulate samples so the half
detection limit approximation may overestimate actual TCDD concentrations and
contribute to underestimation of the
Correlations between Kow and Koc such as that of Karickhoff (1981) are
frequently used to estimate K^. Increasingly, chemical partitioning models set Koc
equal to K^ (DiToro, 1985). Since chemical fate and bioaccumulation models used
for this report depend on estimates of the activity or freely dissolved concentration of
the chemical in water, this risk assessment will consider Kow equal to Koc with a
possible range of 107 to 108 for TCDD. This range exceeds the average values
frequently selected for modeling TCDD fate and transport because inaccuracies
associated with measurements of K^ and K^ for very hydrophobic chemicals such as
TCDD tend to result in underestimates. If headspace analysis of hydrophobic organic
chemicals in air equilibrated with water solutions containing co-solvents or sorbents
(Resendes et al., 1992) can be applied to TCDD, uncertainties for TCDD activity or
freely dissolved concentrations may be reduced in the future. Regardless of the
elusive true activities of TCDD and related chemicals in water and organic carbon
2-4
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phases, exposure modeling errors can be minimized by consistent use of a single best
estimate value of K^, for each chemical in all components of the models and their
application.
2.2 ANALYTICAL LEVELS OF DETECTION
Large uncertainties associated with estimating aquatic exposures to TCDD and
related chemicals could be greatly reduced if minimum levels of detection (MLD) could
be lowered to allow measurements of concentrations in water. High resolution gas
chromatography/high resolution mass spectrometry (HRGC/HRMS) is used to provide
the most sensitive and reliable measurements of TCDD concentrations in sample
extracts. MLD generalized in this discussion are based on the minimum amount of
chemical present in a sample that allows attainment of a 3:1 signal/noise ratio. Table
2-1 summarizes the present routinely achievable MLD for TCDD in the various sample
matrices analyzed for exposure assessments. These estimates are based on typical
sample sizes, established sample preparation methods and routine HRGC/HRMS
detection of 20 pg TCDD (Marquis et al., 1992). Large reductions in these MLD can
now be achieved with use of the best research HRGC/HRMS instrumentation
available. Through large volume sampling, optimum sample clean-up and maximum
instrument sensitivity, MLD for water samples may be lowered by as much as a factor
of 104. Lesser but significant reductions in MLD could occur for sediment and tissue
samples.
fable 2-1. Minimum levels of detection (MLD) for routine high resolution gas
chromatography/high resolution mass spectrometry analyses of TCDD
in samples from aquatic ecosystems.
Sample Matrix
Tissue (pg/g)
Sediment (pg/g)
Water-solids (ng/L)
Water-dissolved (ng/L)
Environmental
Range
0-1000
0-10,000
0-0.002
0-0.0002
MLD
0.5
1.0
0.005
0.0005
2-5
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2.3 DISTRIBUTION IN WATER, SEDIMENTS AND FOOD CHAINS
TCDD and other persistent, hydrophobia non-polar organic chemicals partition
appreciably into organic matter in water and sediments and into biota. Despite the
very low solubility and great hydrophobicity of these chemicals, water acts as the
medium for their transport and partitioning within aquatic systems. Equilibrium
distributions are approached to the extent that rates of transfer to, from and within an
ecosystem allow. Thus when chemical loading is constant for an extended period of
time, steady-state distributions can occur. Concentrations of TCDD in the water
column respond much faster to changes in chemical loading than do concentrations in
sediments. Ecosystems like the Great Lakes may have elevated concentrations in
water for decades after loading reductions due to the redistribution of chemicals from
contaminated sediments. Chemical mass balance models are required in order to
determine the fraction of chemical in water and biota from active external loadings in
comparison to sediment sources.
Only one report of detectable ambient TCDD concentrations in water can be
found. Very large volume (1.5 - 2.2 m3) water samples from the Baltic Sea were
separated into filtered solids and filtrate fractions for determination of concentrations of
PCDDs and PCDFs (Broman et al., 1991).' Since some colloidal organic carbon could
pass through the 0.45 \un filter used, the filtrate TCDD concentration was regarded as
"apparently dissolved" and not 100% freely dissolved. Only two of nine water filtrate
samples were reported to have detectable TCDD. The concentrations were 0.0002
and 0.0003 pg/L (0.2 and 0.3 parts per quintillion). Four of nine filtered solids samples
were reported to have detectable TCDD at 0.0002 pg/L. These levels of detection, if
applied to Lake Ontario water, would result in extraordinary new insight into TCDD
partitioning model accuracy since concentrations 10OOx greater than these Baltic Sea
results are expected (Endicott et al., 1990).
Sediments can contain measurable concentrations of PCDDs and PCDFs
because of the transport and deposition of hydrophobic organic chemicals on
waterborne biotic and abiotic solids. The distribution of TCDD throughout the surface
sediments of Lake Ontario after several decades of loading through the Niagara River
is remarkably uniform, especially if the differences in organic carbon distribution
between depositional basins and other regions are considered (Short et al., 1990).
The Lake Ontario mean surface sediment concentration of TCDD was determined to
be 68 pg/g dry sediment for samples collected In 1987. One centimeter increments of
sediment cores collected with the Lake Ontario surface sediments in 1987 were dated
by radionuclide methods and analyzed by HRGC/HRMS for PCDDs, PCDFs,
polychlorinated biphenyls (PCBs) and other organochlorine chemicals (Cook et al.,
1993a). TCDD levels rose from non-detectable before 1940 to maximum sediment
concentrations of approximately 500 pg/g dry sediment in about 1962, and then
steadily declined to the 1987 surface sediment conditions measured in conjunction
with the Lake Ontario TCDD Bioaccumulation Study (Short et al., 1990). The
2-6
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sediment core profiles for other PCDDs and PCDFs were similar. The PCBs and
other organochlorine chemicals appeared to follow the same general historical pattern
but maximum sediment concentrations occurred in different years with different rates
of change.
The sediments of Lake Ontario contain relatively large concentrations of TCDD
in comparison to other PCDDs and PCDFs because of past chlorophenol production
as a source of TCDD to the lake. A similar pattern was found in the Newark Bay,
New Jersey estuary where surface sediments were found to contain up to 730 pg
TCDD/g overlying mid-1960s sediments with 7,600 pg TCDD/g (Tong et al., 1990).
The highly contaminated Newark Bay sediments are associated with 2,4,5-
trichlorophenoxyacetic acid production between 1948 and 1969. Contaminated
sediments from many other locations have less TCDD and relatively more TCDF and
other PCDFs, frequently in association with large concentrations of PCBs. New
Bedford Harbor, Massachusetts sediments were found to contain TCDF concentrations
up to 1,100 pg/g dry sediment in association with large concentrations of PCBs but
with TCDD concentrations <4 pg/g dry sediment (Pruell et al., 1990). Another
example is the lower Fox River in Wisconsin in which the sediments (dry weight)
contained 1 to 7 pg TCDD/g in comparison to 3 to 61 pg TCDF/g and 0.31 to 6.57 u.g
total PCBs/g (Ankley et al., 1992a). Fox River sediment concentrations of each
chemical varied more in relation to sediment organic carbon concentration distribution
than to a downstream distribution gradient associated with transport from
contaminated sediments. The upper Great Lakes sediments have PCDD and PCDF
congener distributions that are indicative of atmospheric deposition from regional
combustion sources since the 1940s (Czuezwa and Hites, 1984; Czuczwa et al.,
1985).
Fish are good detectors for monitoring the concentrations of TCDD and similar
bioaccumulative chemicals in aquatic systems. The National Dioxin Survey initiated by
EPA in 1983 included a nationwide survey of TCDD residues in fish from 395 sites
(U.S. EPA, 1987; Kuehl et al., 1989). TCDD concentrations in fish (reported
throughout this report as pg TCDD/g wet weight of whole organism unless specified
otherwise) ranged from below detection at 1 pg/g (72% of all samples) to a maximum
concentration of 85 pg/g. TCDD was detectable (>1 pg/g) in whole fish samples from
17 of 90 randomly selected sites but 23 of 29 Great Lakes sites (U.S. EPA, 1987).
Highest concentrations among all fish sampled were associated with the open Great
Lakes and river sites downstream from kraft paper mills. In another study (U.S. EPA,
1992b) TCDD was detected in fish from 70% of sites sampled across the United
States in 1986-1988 for a wide range of organic chemical residues. The maximum
TCDD concentration was 204 pg/g with an average of 6.8 pg/g. Although the
selection of sites in this study was biased toward sites where TCDD and other PCDDs
and PCDFs were likely to be found, the survey probably provides the most extensive
fish residue data set for PCDDs and PCDFs based on a single analytical effort.
2-7
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Although sediment samples were not collected, it is likely that sediments associated
with sites having contaminants in fish are similarly contaminated.
TCDD (O'Keefe et al., 1983) and other PCDDs and PCDFs (Zacharewski et al.,
1989) are ubiquitous contaminants of Great Lakes fish. Whereas concentrations of
congeners such as TCDF vary little in fish among the lakes (16 pg/g for Lake Superior
versus 35 pg/g for Lake Michigan), Lake Ontario's fish have much greater
concentrations of TCDD; 40 pg/g for Lake Ontario versus 1 pg/g for Lake Superior
(Zacharewski et al., 1989). TCDD residues in Lake Ontario lake trout (Salvelinus
namaycush) have decreased to approximately 25% of the levels present in 1977
shortly after lake trout populations were reestablished through stocking. These
reductions in TCDD and other PCDD and PCDF residues parallel the changes
recorded in sediment profiles from depositional basins (Cook et al., 1993a). In
contrast to the findings of Cook et al. (1993a), Whittle et ai. (1992) has reported a
significant increase in mean TCDD concentrations over the period of 1981 to 1990,
despite evidence for decreased contaminant loading to Lake Ontario.
Striped bass (Morone saxatilus) from the lower Hudson River and its estuary
were found to contain TCDD concentrations up to 120 pg/g in comparison to 1-4 pg/g
in striped bass from Chesapeake Bay (O'Keefe et al., 1984). Two striped bass
collected from Newark Bay contained 84 and 734 pg TCDD/g fillet (Rappe et al.,
1991). The large difference in TCDD concentrations in these samples was not
paralleled for other PCDD and PCDF congeners which differed by a factor of only two,
Rappe et al. (1991) commented that the striped bass with the greater TCDD
concentration probably resided in the contaminated area for a longer period. TCDD
concentrations in crab and lobster meat averaged 111 and 5.5 pg/g, respectively,
while hepatopanereas concentrations were 4,900 and 435 pg/g, respectively.
Adult bullfrogs, collected in 1984-1985 near a former 2,4,5-
trichlorophenoxyacetie acid production site along Rockey Branch Creek in Arkansas,
contained 640 - 48,000 pg TCDD/g of liver (Korfmaeher et al., 1986a). Bullfrog
muscle tissue samples contained less than 10% of the liver TCDD concentrations
(Korfmaeher et al., 1986b). Fish collected in 1979-1981 from Bayou Meto, which
receives Rockey Branch Creek drainage, contained 37 to 480 pg TCDD/g (Mitchum et
al., 1980). The largest TCDD concentrations found in bullfrogs indicate the potential
for TCDD residues in fish, if present in a highly contaminated stream like Rockey
Branch Creek, to exceed 2,000 pg/g.
Mussels (Elliptic complanats) were placed in cages to detect TCDD in water
associated with pulp and paper mills and other sources in the Rainy River, Ontario
watershed. The mussels accumulated up to 10 pg/g after 21 days (Hayton et al.,
1990). Invertebrate species, which do not appear to biotransform PCDDs, PCDFs and
PCBs, can be used to monitor distribution patterns for complex mixtures of these
chemicals that are not detectable in water through direct chemical analyses.
2-8
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2.4 EXPOSURE ROUTES FOR AQUATIC ORGANISMS
TCDD and related chemicals are accumulated by aquatic organisms through
exposure routes that are determined by the habitat and physiology of each animal.
Site-specific differences in exposure can occur as a result of temporal and spacial
variation of the distribution of the chemical between the water, sediment and food that
organisms contact. With log Kow>5, exposure through ingestion of contaminated food
becomes an important route for uptake in comparison to respiration of water
(Thomann, 1989). Thus all biosignificant PCDDs, PCDFs and PCBs are significantly
accumulated through food ingestion, although the net result for TCDD and other
PCDDs and PCDFs, because of metabolism, is to approach, rather than exceed,
tissue levels that are at equilibrium with their activities in water (Opperhuizen and Sijm,
1990).
The relative contributions of water, sediment and food to uptake of TCDD by
lake trout in Lake Ontario was examined in a laboratory study by exposing yearling
lake trout to a diet prepared from Lake Ontario rainbow smelt (Osmerus mordax) and
to sediment from Lake Ontario along with water at a TCDD concentration simulated to
be at equilibrium with the sediments (Batterman et al., 1989). Each exposure route
was studied individually and in combination with the other routes for two 10-fold
different TCDD exposure levels. Food ingestion was found to contribute approximately
75% of total TCDD uptake (Cook et al., 1990). The amount of TCDD uptake from
water in the absence of contaminated sediment was negligible in relation to uptake
from food. TCDD uptake associated with sediment exposure was thought to be
exaggerated due to suspended sediment concentrations much greater than trout
normally experience, and the presentation of food to the trout at the sediment-water
interface. In the water exposures, TCDD was introduced continuously to the water
flowing into the aquaria at twice the estimated sediment equilibrium concentration in
an attempt to accommodate sorption to organic material not at steady-state. There
may still have been significant loss of bioavailable TCDD to organic carbon that
accumulated in the water exposure system. However, when TCDD was added to
water in exposures with contaminated sediment, no increase in TCDD uptake (over
TCDD uptake associated with sediment alone) occurred. Three interpretations of this
result are possible: (1) desorption of TCDD from sediment was sufficiently rapid to
allow maintenance of equilibrium between TCDD in sediment and the water flowing
through the aquaria and thus uptake in the sediment exposures could be from both
sediment and water; (2) the TCDD added to water in the presence of sediment was
rapidly sorbed to the suspended sediment because the assumed TCDD log K^ of 7.0
was an underestimation, and thus uptake was due to direct exposure to sediment and
the reduced TCDD concentration in water at equilibrium with the sediment; or (3) the
TCDD added continuously to the water increased the availability of TCDD from water
but the water was a minor source for uptake in comparison to TCDD on sediment.
2-9
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Included with the Lake Ontario sediment study (Cook et al., 1990) were
exposures of lake trout to Lake Ontario sediments with a TCDD concentration ten-fold
greater (10X sediments) than the 1987 lake wide average. Measurements of large
volume samples of filtered water from the 10X exposures indicated that freely
dissolved TCDD concentrations in the water were less than 0.3 pg/L. Cook et al.
(1990) concluded that direct ingestion of sediment, and possibly gill contact with
suspended sediment were probably more important routes of uptake than ventilation of
water containing freely dissolved TCDD resulting from desorption from suspended
sediment. An independent analysis (Barber et al., 1991) of data selected from a
preliminary report (Batterman et al., 1989) of the 10X sediment study, and an
assumed freely dissolved TCDD concentration of 0.75 pg/L in equilibrium with
sediment, resulted in the conclusion that gill uptake of freely dissolved TCDD in water
associated with the Lake Ontario sediments was primarily responsible for the TCDD
accumulation observed in lake trout by Cook et al. (1990). The analysis of Barber et
al. (1991) did not consider the alternative interpretations (2) and (3) above or the
water analysis data presented in Cook et al. (1990).
The availability of TCDD and related chemicals to aquatic organisms through
uptake by ingestion or contact with sediments is an important consideration for this
risk characterization. Sediment dwelling invertebrates and bottom feeding fish may
have exceptional exposures that can increase human or wildlife exposures, increase
concentrations in aquatic food chains and cause sensitive fish species to be at
increased risk. Many aquatic organisms, including fish, ingest sediment and/or
suspended particles either as a source of food or inadvertently while feeding. The
diets of minnows and suckers include large quantities of detritus as a major food
source. Gizzard shad (Dorosoma cepedianum) can digest 50 to 66% of the organic
matter in ingested sediments which are consumed at a rate of up to 20% of their wet
weight in dry sediments each day (Mundahl, 1991). No studies of fish gut assimilation
efficiency from ingested sediment have been reported for TCDD or similar nonpolar
organic chemicals. Carp (Cyprinus carpio) (250 g) were exposed by oral gavage to
500 mg of sediment contaminated with PCDFs but this was an insufficient amount of
sediment to allow detection of chemical assimilation (van der Weiden et al., 1989).
Since trout gut assimilation efficiency for TCDD in natural food is about 40%
(Cook et al., 1990) to 50% (calculated from the data of Kleeman et al., 1986a) and
some fish appear to digest organic matter associated with ingested sediment, TCDD
uptake from ingested sediments may be an important route of exposure for some
species of fish. A recent nationwide survey of organic chemical residues in fish
indicated that TCDD and other bioaccumulative chemicals were more frequently
detected in bottom feeding fish, such as carp, and that the average concentrations in
the bottom feeders were greater than for other species (U.S. EPA, 1992b). However,
comparison of lipid-normalized residues is needed in order to confirm the tendency for
greater bioacccumulation by bottom feeding fish. Carp exposed to TCDD only through
water do not appear, on a lipid-normalized basis, to accumulate more than rainbow
2-10
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trout (Oncorhynchus mykiss) and fathead minnows (Pimephales promelas) (Cook et
al., 1991) but do appear to have five-fold greater TCDD bioaccumulation in
environmental exposures where sediment ingestion can occur (Kuehl et al., 1987).
Accumulation of PCBs by sandworms (Nereis wrens) from ingestion of and
contact with contaminated sediments was estimated to exceed accumulation possible
from water alone, even with maximum water concentrations of PCBs from sediment
elutriates (Rubinstein et al., 1983). TCDD, TCDF and PCBs in sediment from the
Passalc River, New Jersey were accumulated by sandworms, clams (Macoma nasuta)
and shrimp (Palaemonetes pugio) in laboratory exposures involving flowing filtered
water over the sediment (Rubinstein et al., 1990). The Passaic River sediment
exposure data indicate that each invertebrate species at steady-state would
accumulate each chemical to a level expected for equilibrium partitioning between
sediment organic carbon and organism lipid. Only a few measurements of gut uptake
efficiency for any hydrophobic organic chemicals on ingested sediments have been
made for aquatic invertebrates. Assimilation efficiencies for 2,4,5,2',4',5'-
hexachlorobiphenyl on Lake Michigan sediments ingested by ollgochaetes ranged from
15 to 36%, depending on gut clearance times (Klump et al., 1987). Efficiency of gut
uptake of hexachlorobenzene from ingested sediment by the deposit-feeding clam
(Macoma nasuta) was measured to be 38 to 56% (Lee et al., 1990).
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3. BIOACCUMULATION
Relating information on effects of TCDD to environmental concentrations is
made difficult by several factors. A variety of routes and durations of exposure are
employed in laboratory tests, which are usually quite dissimilar from what occurs in
natural systems. In natural systems, multiple and variable routes of exposure also
must be considered. Association of TCDD wrth organic matter can affect its
availability to organisms both in laboratory and natural systems. Effects of TCDD are
often expressed in terms of concentrations in the diet or In the test animal itself, rather
than environmental concentrations. Because of these considerations, the amount of
TCDD accumulated by an organism generally will be an important component in
integrating exposure and effects information. This chapter will review concepts and
information on the bioaceumulation of TCDD and consider how it should be used in a
risk assessment.
3.1 CONCEPTUAL FRAMEWORK AND DEFINITIONS
Discussions of TCDD bioaceumulation frequently are complicated by use of
poorly defined terms and parameters. As a prelude to a review of specific information
on TCDD bioaceumulation, key concepts and terms will be defined.
3.1.1 Bioconcentration
For aquatic organisms, bioconcentration refers to the accumulation of a
chemical from exposure via water only. Bioconcentration is a dynamic and often
complicated process, involving uptake via gills and skin; elimination of the chemical via
gills, skin, urine, and feees; and metabolic transformation of the compound. For
chemicals that only slowly come to steady state, growth of the organism can also
dilute tie accumulated chemical and affect the apparent bioconcentration.
Additionally, organisms consist of multiple tissues and organs with different
toxicokinetic characteristics, which can complicate the time course of accumulation.
A useful approximation often employed for bioconcentration is to consider the
organism as a single compartment at internal equilibrium and to describe the change
in chemical concentration with time by a first-order kinetic expression such as the
following:
Cw -
where: Cw and Ca are chemical concentrations in the water and aquatic organism; k^
and kus are uptake rate constants via gill and skin; k^, kes and k^ are elimination rate
3-1
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constants via gill, skin, and feces; km is a metabolic transformation rate constant; and
g Is the proportional growth rate of the organism. The rate constants k, and k2 are the
uptake and loss constants that would be derived from net uptake versus time in
classical bioconcentration experiments. As indicated here, k, and Iq are the
summation of a variety of constants. In some bioconcentration measurements, k> is
adjusted for growth and therefore does not include the parameter g.
A bioconcentration factor (BCF) is defined as the ratio of the chemical
concentration in the aquatic organism (CJ to that in the water (Cw):
BCF = $L (3-2)
Cw
A BCF can apply at any time during the process of bioconcentration and is premised
on a reasonably constant chemical concentration in water. After sufficiently long time,
the concentration of chemical in the organism will tend to approach a constant value,
termed the steady-state bioconcentration factor (ssBCF). For highly hydrophobic
chemicals, steady-state takes a long time to reach and often cannot be closely
approached In practical experiments, especially for larger organisms which tend to
have slower uptake rates and longer half-lives for elimination. ssBCFs must then be
estimated from the kinetic rate constants for uptake and loss, which for the single-
compartment, first-order case is:
- (3-3)
k^g
The driving force for bioconcentration is the difference in activities of the
chemical between the water and organism. At equilibrium, these activities are equal
and C,/CW would equal Kaw, an equilibrium constant for partitioning of the chemical
between the aquatic organism and water. Because exchange of chemical across gills
and skin is a two-way, diffusive process driven by the activity differences, keg=kug/Kaw
and Kos-kuj/Kaw under the model used here of a single compartment organism at
internal equilibrium. The relationship between ssBCF and Kaw therefore is:
ssBCF = K- " - - /.("l - Ť
-------
concentration due to growth can also cause steady-state estimates to fall below the
expected equilibria if the growth rate is appreciable compared to rates of elimination.
Bioconcentration can depend on the composition of the water and organism
since this composition can affect the relationship between chemical activity and the
total concentration of chemical. For highly hydrophobic organic chemicals, a
significant fraction of the chemical concentration in the water can be associated with
suspended particles and dissolved organic matter and be less available for uptake by
an organism. The total chemical concentration in water (C^) can be expressed as:
Ci = Ci * POC-Cpoc + DOC-C^ (3-5)
where C^ refers to the concentration of chemical which is freely dissolved and is
considered here as equivalent to the chemical activity; POC and DOC are the
concentrations of particulate and dissolved organic carbon in the water; and Cpoc and
CDOC are the concentrations of chemical associated with the particulate and dissolved
organic carbon (on an organic carbon weight basis). If these different constituents in
the water are at equilibrium with each other, this expression can be rewritten as:
Cd
c* - c* -0 + POC-KPOC + DOC-KOOC) = -=. (3-6)
'a
where Kpoc and KDOC are the equilibrium constants for partitioning of the chemical
between water and POC and DOC, respectively, and fd is the fraction of chemical
which is freely dissolved (i.e., [1+POCťKPOC+DOCťKDOO]"1). To relate the freely
dissolved concentration just to POC, this expression will be further modified for use
here as:
f =
d ^ ^ TBFoc.POC.Kow -
where TBF00 ("total binding factor for organic carbon") is an empirical adjustment to
express binding by all organic carbon In water in terms of POC equivalents (e.g., a
TBFOC=2 indicates that the chemical is as much associated with DOC is POC) and Kow
is used as a surrogate for KpOC.
Bioconcentration factors therefore should specify what water concentration
basis is used, total (BCF1, C^) or freely dissolved (BCFd,CŁ):
3-3
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BCFd - -t (3-8)
Similarly, the uptake rate constants (kug, kus) and equilibrium partition coefficient (Kaw)
must clearly be based on either dissolved or total water concentration.
BCF1 is expected to be largely independent of site water characteristics and
thus to be similar among different sites. BCF* between two sites (x and y) therefore
should follow the relationship:
**** tf (3-9)
For sites in which fdŤ1 and in which TBFOC is similar, this can be further simplified to:
POC*
Analogous to the expression above (equation 3-5) for water, the total
concentration of chemical in an aquatic organism (Cla) can be expressed as a sum
across different constituents of an organism's body. For example, Barber et al. (1991)
considered three constituents - water, lipid, and nonlipid organic matter:
ci - vc/ + ft-ct + f^q,
where C^ is the concentration of chemical freely dissolved in water in the organism
and, as for C^, is assumed to be equivalent to chemical activity; C4 and Cn, are,
respectively, the concentration of chemical in the lipid, and nonlipid phases; and fw, f,,
and frt are the fraction of these phases on the basis of an organism's wet weight. The
proportion of chemical bound to a receptor such as the case for TCDD and the Ah
receptor is incorporated into fnt. If the organism is at internal equilibrium, this can be
expressed as:
where Kt and K^ are the lipid/water and nonlipid/water partition coefficients for the
chemical. For hydrophobic chemicals, lipids are thought to dominate partitioning in
many cases and this expression is often simplified to the following:
3-4
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where the octanol/water partition coefficient (K^) is used as an estimate for the
lipid/water partition coefficient. The equilibrium partitioning between an organism and
chemical freely dissolved in the water is therefore expected to be:
Kd * f-K (3-14)
AŤW w 'I AOMr
Because of the importance of the lipid phase, it is sometimes useful to express
BCFs on the basis of organism lipid content in order to reduce variability among
organisms and to better contrast activities. In this case, the chemical concentration in
the organism is denoted C4 and is calculated as the mass of chemical in an organism
divided by the lipid content rather than the wet weight of the organism:
BCFt = A = °*lf* = Ž2E (3-15)
Cw cw ft
This expression can be based on either total or freely dissolved water concentrations.
The equilibrium partitioning between the lipid-normalized concentration in an organism
and chemical freely dissolved in the water is expected to be approximately equal to
K^ (i.e., a ssBCFj will equal Kow in the absence of growth dilution, metabolism, and
elimination other than across the gills and skin).
3.1.2 Bioaccumulation
For aquatic organisms, bioaccumulation refers to the net accumulation of a
chemical from exposure via food and sediments as well as water. The kinetic
expression for bioaccumulation is simply that for bioconcentration with an additional
term for chemical absorbed in the gastrointestinal tract:
where C, is concentration in the food and ku, is a rate constant for absorption of
chemical from food; this rate constant depends on the amount and nature of the food
consumed. Under the assumption of a single-compartment organism at internal
equilibrium, the other rate constants should be the same for bioconcentration and
bioaccumulation, although in practice different internal distribution of chemical
3-5
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absorbed via the gastrointestinal tract might produce somewhat different elimination
rate constants.
A bioaccumulation factor (BAF) is the ratio of the chemical concentration in the
organism to that in the water. As such, the basic definitions for BCFs above apply
also to BAFs. The previous discussion regarding the importance of specifying the
form of the chemical in the water and the value of lipid normalization also applies.
Thus, there are four important BAF variations, and care must be taken to clearly
specify which variation is used:
BAF1 = -3" BAFd = -% BAF; = -5. BAFd = -% (3-17)
The difference between BAFs and BCFs is primarily in the routes of exposure involved
and the levels of accumulation therefore attained. BCFs must be measured for water
exposure alone in the laboratory or from environmental data with the assumption that
chemical uptake by the organism from water is the predominant route of accumulation.
BAFs are usually determined from measurements of chemical concentration in water
and organism tissue samples.
The steady state BAF has the following form under the model used here:
If +lc +lc -FAF
ssBAF = f us " (3-18)
where FAF is the ratio of the chemical concentration in food to that in water. BAFs
will therefore exceed BCFs to the extent that uptake via the gastrointestinal tract is
significant compared to uptake from water via the gills and skin. In fact, whereas
ssBCFs must be at or below equilibrium with the water, ssBAFs can exceed the
equilibrium expected based on water and food concentrations because the removal of
lipids and other organic material from food during digestion may increase the activity
of chemical in the gut contents and promote uptake (Connolly and Pedersen, 1988).
Thomann (1989) defined the condition of BAF4/BCF{ > 1.0 as "food chain
accumulation". The increase in the concentration of chemical in an organism relative
to its food has also been termed biomagnification, with the ratio of these
concentrations being the biomagnification factor (BMF). Ideally, BMFs should be
calculated on the basis of lipid-normalized food and organism concentrations of the
chemical so that they actually represent differences in chemical activity, rather than
just a response to different lipid contents and thus different affinities for the chemical.
For chemicals that are not highly hydrophobic, the distinction between
bioconcentration and bioaccumulation is of little consequence because accumulation
rates via water are much greater than those via food (kug+kusťkufťFAF). This is due
3-6
-------
to the rate constants being roughly proportional to the volume of material processed.
For food intake, this is generally a few percent of body weight a day, whereas water
passing over gills this can vary from about two hundred to several thousand times the
organism weight per day, depending on the size and type of organism, its activity, and
the temperature. For food uptake to be as important as uptake via water, the FAF
therefore must be a few thousand to as much as one hundred thousand
(corresponding to a K of 104 to more than 105}, depending on the organism and its
environment. For chemicals with lower hydrophobicities, measured accumulations
would therefore be similar in water only and water+food exposures.
Although the complex kinetics of exchange make it unlikely that a highly
hydrophobic chemical in an organism will be at equilibrium with the chemical in the
surrounding water, this equilibrium is still a useful reference point in interpreting and
applying BAFs. A measure of the disequilibrium between an aquatic organism and
water is:
_
*iaa ~
_ O _ BAF^ _ BAF
**iaa ~ _ j ~~ ť *
aw
ic
*^
3.1.3 Biota-Sediment Relationships
For many extremely hydrophobic chemicals such as TCDD, reliable
measurements of ambient water concentrations, especially dissolved concentrations,
are not available. Therefore, accumulation of chemical by an organism cannot be
referenced to a water concentration as required for a BCF or BAF. However,
concentrations are generally measurable in sediments as well as In organisms
because these chemicals distribute predominantly in association with organic carbon.
Analogous to equations 3-5 and 3-9, presented previously for water and aquatic
organisms, the relationship of total chemical concentration jn the sediment (Cls} to
freely dissolved chemical in pore water (C^) can be described as follows:
C/ - fm-C; + /<Ł = C/.C/U + f^/U . C/.f^/^ (3-20)
where fws is the fraction of water in the sediment, foc is the fraction organic carbon in
the sediment, C^js the concentration of chemical bound to sediment organic carbon
(this is different from the organic carbon normalized sediment concentration, Coc,
which equals C^/f00 and thus includes both bound and free chemical), and Koc is the
equilibrium constant for partitioning of chemical between organic carbon and water.
The relationship between chemical concentrations in organisms and sediment is
defined by Ankley et al. (1992b) as the biota-sediment accumulation factor (BSAF):
3-7
-------
/
(3-21)
Concentration of chemical in the organism is normalized to a lipid basis and in the
sediment it is normalized to an organic carbon basis to make the BSAF more
independent of the effect of these factors on chemical partitioning and more indicative
of activity differences. The BSAF has previously been named the bioavailability index
(Bl) (Kuehl et al., 1987), the accumulation factor (AF) (Lake et al., 1990) and the
biota-sediment factor (BSF) (Thomann et al., 1992). It can be used to quantify
bioaccumulation relationships in the field and to help evaluate risk. Selection of
surface sediment samples that quantitatively represent the average
sediment/water/food chain exposure environment of an organism is difficult.
Underestimation of the concentration of chemical in thin sediment surface layers
acting as sources of the chemical to water and food chains may cause large
overestimates of BSAFs.
BSAF measurements reflect disequilibrium between the organism, water and
sediment. The form of the aquatic organism/water disequilibrium factor (Raw) in
equation 3-17 can be applied to disequilibrium between water and sediment (Rws) and
organism and sediment (R^). Thus:
(3'22)
If fish and sediment are at equilibrium (Ras = RawťRws = 1), the BSAF is expected to
equal the ratio of the lipid-normalized partition coefficient for aquatic organisms to the
organic-carbon normalized partition coefficient for sediment. Because K^ is of similar
magnitude to and varies proportionally with K^, the expected equilibrium BSAF is
unity or slightly greater and is insensitive to uncertainty in Kow. The degree to which
the BSAF deviates from the expected value is indicative of the degree of the
disequilibrium between fish and sediment. Such disequilibrium can occur due to
factors such as the following: (1) kinetic limitations for chemical transfer from
sediments to water (Rws<1 for systems which have a net flux of chemical from
sediment); (2) surface sediment concentrations of the chemical that have not reached
steady-state with water (Rws>1); (3) organic carbon diagenesis in sediments which
commonly results in C!j > CŁ (R^xl) ; and (4) biological processes which cause
accumulation to be lower (Raw<1) such as for biotransformation by the organism or
greater (Raw>1) such as for biomagnification.
3-8
-------
Analogous to the BSAF, and perhaps more directly applicable to
bioaccumulation by fish if the chemical can be detected on suspended solids, is the
biota-suspended solids accumulation factor (BSSAF):
(3-23)
l'a
ssac
where C^ is the concentration of chemical in suspended solids and f^ is the fraction of
organic carbon in suspended solids. Like the BSAF, this term should be of the order
of one at equilibrium and disequilibria among system components will result in
deviations from this expected value. For the BSSAF, the important disequilibrium is
between the organism and water, because the disequilibrium between sediment and
overlying water does not affect this parameter and because the extent of
disequilibrium in water between the free chemical and that bound to POC and DOC
should be much less than between the organism and water. DOC need not be
explicitly considered in this relationship since it should have chemical activity similar to
POC; also, DOC tends to be correlated with suspended solids concentrations and
partitioning of the chemical to DOC is probably less than to POC. The BSSAF should
not vary with differences in average suspended solids concentration or foc between
ecosystems, but field measured BSSAFs should fluctuate some due to the slow
response of C, to changes in suspended solids and foc.
Since the BSSAF and the BSAF do not vary significantly with Kow, the great
uncertainty existing for the Kow of TCDD is not incorporated into these bioaccumulation
factors. Unfortunately, the K^ uncertainty still is important when the application of
either BSSAF or BSAF involves the prediction of Cpoc or C^ on the basis of chemical
loading to the water. If the K^s for both sediment organic carbon and suspended
organic carbon in water are similar, the difference between the BSAF and BSSAF for
TCDD should be attributable to the disequilibrium present between surface sediment
and the overlying water body. The difference between a measured BSSAF and the
theoretical equilibrium partitioning value (perhaps 1 to 2) could be attributable to the
combined effects of metabolism, biomagnifieation and choice of a nonequilibrium Cpoc.
Unfortunately, there are almost no C^ data available for TCDD at this time. When
available, these data could provide a better water-based bioaccumulation relationship.
3.2 TCDD BIOCONCENTRATION FACTORS
The expected equilibrium value for ssBCF, is K^, which is approximately 107
for TCDD (Burkhard and Kuehl, 1986). In practice, the extreme hydrophobicity of this
chemical and the inability to determine C* requires bioconcentration factors to be
calculated only on the basis of total water concentrations and to be significantly less
than this expected value. The absence of food uptake can also reduce accumulation
and keep it below equilibrium values. However, if bioconcentration factors are
3-9
-------
measured and used with these limitations in mind, they can still provide useful
estimates of bioaeeumulation potential.
A summary of TCDD ssBCFj determinations for fish (Table 3-1) indicates a
range of 81,300 to 4,300,000. The variability of these results seems to be mostly due
to differences in the uptake rate constants (kt), since the elimination rate constants
(kg) differ only by a factor of 7. The large variability in k, is likely due to reduced
bioavailability of TCDD associated with organic carbon buildup and the limited mass of
TCDD added in static exposure systems. In these types of experiments, available
TCDD is removed through bioconcentration by the fish, leaving only TCDD that is
primarily absorbed to organic matter in the water. The results suggest that C* may be
further reduced from equilibrium with fish and absorbents due to a desorption rate
slower than the rate of gill uptake. Renewal exposures are probably characterized by
rapidly declining TCDD uptake between periodic additions of TCDD. The exposure
reported for rainbow trout (Branson et al., 1985) is 27 times greater than the solubility
of TCDD in water at 10°C (Lodge, 1989). Although the water concentration used for
this BCF determination was reduced to 107 ng/L to adjust for bioavailability, it is likely
that the actual fraction of TCDD available for bioconcentration during the 6-hour static
exposure was only a small fraction of this value.
Fish size, age, and lipid content influence bioconcentration kinetics, but the
impact of exposure water conditions on TCDD bioavailability is most difficult to
evaluate because only total TCDD in the water can be measured. Steady-state BCFjs
for the flow-through exposures of rainbow trout (Mehrle et al. 1988), carp (Cook et al.
1991) and fathead minnows (ibid) ranged from 510,000 to 837,000 (mean of 728,000).
These exposures involved TCDD added to the water with a solvent carrier and are
expected to have a significant reduction in bioavailability of TCDD due to both organic
carbon binding and a co-solvation effect due to the solvent carrier. The recently
reported ssBCFj of 4,300,000 for medaka (Oryzias latipes) (Schmieder Št al., 1992) is
the only determination using a generator column for adding dissolved TCDD to the
water. If log K^ for TCDD is 7 and log BCF* is equal to log Kow in the absence of a
metabolism or enhanced elimination rate, this result indicates that most of the TCDD
in the water was dissolved (or at least bioavailable) and metabolism of TCDD by the
fish was too slow to greatly reduce bioconcentration. However if the log K^ for
TCDD is 8, the combined impact of low bioavailability and metabolism results in a
BCFj that is approximately 20 times less than the theoretical equilibrium BCF* for
TCDD.
The differences in kgS for the ssBCF determinations (Table 3-1) are primarily
attributable to differences in size, life-stage and lipid content of the test species.
Unfortunately, all of the BCF determinations involved measurements of kg for fish with
toxic symptoms. Also, the fathead minnow kj, reported by Adams et al. (1986) was
calculated without growth correction of TCDD concentrations in the fish and thus may
overestimate the true l^. The rainbow trout kg values appear large for a cold water
3-10
-------
CO
Tai
Die 3-1. Summary of TCDD steady-state bioconcentration factor determinations for fish.
Study
Branson et al.
(1985)
Mehrie et al,
(1988)
Adams et al.
(1986)
Cook et al,
(1991)
Cook et al.
(1991)
Schmieder et
a!. (1992)
Species
rainbow
trout
rainbow
trout
fathead
minnow
fathead
minnow
carp
medaka
Initial Size
(g)
35
0.38
fry
0.5-1.0
juvenile
1.0
young adult
15
0.175
%Upld
11
n.r
(est. 5)
n.r.
(est, 7)
19
9
8
Temp
(C)
10
11
25
35
25
25
TCDD cone.
(pgrtiter)
320,000
38
1,000
49-67
62
101
Exposure
period
(days)
0.25
static
28
flow-through
28 static
renewal
71
flow-through
71
flow-through
12
flow-through
no solvent
Depuration
period
(days)
139
none
20
61
61
175
Uptake
k,
(nWg/day)
108
1852
381
1280-1890
700
2306
Elimination
K*
(1/day)
0,012
0.047
0.048
0.012-0.013
0.010
0.0067
SSBCF
9,270
39,000
7,900
97,000-
159,000
66,000
344,000
ssBCF,*
81,300
780,000
113,000
510,000-
837,000
733,000
4,300,000
* Where not reported (n.r.), % lipld is estimated on baste of fish speolts, size, and age.
-------
species (Table 3-1). The Ic, of 0.047 reported by Mehrle et al. (1988) was calculated
on the basis of gill uptake kinetic data only. Lake trout (approximately 40 g), exposed
to environmental, non-toxic levels of TCDD for 120 days via combinations of food,
sediment, and water (TCDD from a generator column), eliminated TCDD at a slower
rate with kj, values in the range of 0.002-0.011 without growth correction (Cook et al.,
1991). The nominal water exposure concentration for lake trout was only 0.7 pg/L and
limited water analyses indicated that almost all of the TCDD was bound to suspended
solids. On the basis of the nominal TCDD water concentration and a k., of 10, the
ssBCFj for these lake trout is only 104,000 despite the slow elimination rate observed.
TCDD elimination was measured in adult carp with long-term exposure through
water, food and sediment (Kuehl et al., 1987). Carp (1.5 kg) with 16% lipid were
netted from the Wisconsin River and maintained in Lake Superior water for up to 336
days. The t1/2 for TCDD elimination was approximately 320 days, with a k,, of 0.0022
days"1, in comparison to a t1/2 of 63 days and a kg of 0.011 days"1 for smaller carp
(Cook et al., 1991), The fivefold slower elimination rate for Wisconsin River adult carp
may be attributable to size and lipid differences between the two groups of carp,
without invoking consideration of long-term environmental exposure versus short-term
laboratory water exposure.
The kj, of 0.0067 for medaka (Schmieder et al., 1992) is the smallest reported
for a small, warm water species. The medaka TCDD concentrations were corrected
for growth and no solvent was present that might increase TCDD elimination across
the gills. Insufficient kinetic data exist for induction of TCDD metabolism over time of
exposure and depuration to allow a determination of the possible influence of the short
12 day exposure period on lc>.
3.3 TCDD B1OACCUMULATION FACTORS
Measurements of BAFs are usually based on presumed steady-state or
pseudo-steady-state field exposure conditions. The ideal BAF,s, for risk assessments
in general and water quality criteria in particular, would be calculated on the basis of,
and applied to, concentrations of freely dissolved or bioavailable organic chemical in
water averaged over a time period greater than that required to reach 90% of steady-
state (approximately one year for TCDD in fish). This would, reduce variability due to
fluctuations in water concentrations experienced by an organism and its food chain as
a result of seasonal, meteorological, chemical loading or biota migrational patterns.
Weather-induced fluctuations in water flow probably have considerable site-specific
variation in the magnitude of influence on chemical concentration. This variability may
be reduced through use of an annual mean concentration estimated on the basis of all
factors influencing the concentration of the chemical which is bioavailable to the
organism.
3-12
-------
Great uncertainty exists for the estimation of such water concentrations,
especially for very hydrophobic chemicals. In some cases, such as for TCDD and
lexicologically related chemicals, ambient water concentrations can not be measured
even on the basis of the total chemical present (see section 2.2). This makes direct
validation of predicted water concentrations for these chemicals impossible at this
time. Surface sediment contamination levels in systems approaching steady-state can
be measured and thus provide an indirect and partial check on fate and transport
model predictions of chemical concentrations in water.
Although ssBAF^ should more readily extrapolate across sites than ssBAFj,
both factors will be estimated here because present EPA regulatory and risk
assessment actions routinely estimate bioaccumulation on the basis of total chemical
concentration in water. If BAFj is estimated on the basis of water having small
concentrations of DOC and suspended solids, the BAF1, will overestimate
bioaccumulation of TCDD in systems with greater DOC and suspended solids.
However, site-specific BAFjs for fish can be readily calculated on the basis of the
fraction of chemical freely dissolved (fd) in the water:
BAF! = BAF? - fd (3-24)
Equation 3-7 and its derivation describe fd. Figure 3-1 demonstrates how fds and thus
BAFjs vary with POC and log
Variability in TCDD BAF,s, that may be attributable to interspecies differences
in biomagnification, bioenergetics and chemical biotransformation kinetics, should be
considered in addition to site-specific bioavailability conditions. These factors can only
be Included in bioaceumulation calculations through the use of site-specific
bioaccumulation models requiring extensive calibration or through development of
generic bioaccumulation models that allow a choice of BAFs based on site-specific
food web and ecosystem characteristics (Cook et al., 1991). All BAF measurements
or estimates also have variability and inaccuracy associated with movement of
organisms over time through regions with different concentrations of chemical in water,
sediment and food chain.
The best data for calculating a bioaccumulation factor for TCDD are provided
by EPA studies of TCDD bioaccumulation for fish in Lake Ontario (Carey et al., 1990;
Cook et al., 1990) coupled with estimates of sediment and water concentrations of
TCDD (Endicott et al., 1990). Five species of fish were sampled at ten different
locations throughout the lake in 1 987. Rainbow smelt and slimy sculpin (Cottus
cognatus} were also sampled in 1986 to determine lake trout food exposure (Cook et
al., 1990). Since water concentrations of TCDD were expected to be nondetectable,
the bioavailability index (Bl = BSAF) approach (Kuehl et al., 1987; Ankley et al.,
1992b) was attempted for a direct measure of bioaccumulation. Surface sediment (0-
3 cm) samples were collected from 60 locations throughout the lake. In general, the
3-13
-------
Figure 3-1. Fraction of organic chemical freely dissolved in water (fd) if the total organic carbon binding factor
TBFoc=1.5 and log KoW=4, 5, 6, 7, or 8.
1 + TBF^POC-K^
TBFť - 1 +
POC (mg/L)
log Kow = 4 log Kow = 5 log Kow = 6 log Kow = 7 log Kow = 8
S^^^___ _A ./'V * * -- ..-JIB ,__.
""" Bť ..... lŁ^ . . - Ť Ť ŤŤ HH^I^MK^M m. - -mm- -1_-
-------
TCDD distribution in surface sediments was found to follow organic carbon distribution.
TCDD concentrations in fish did not reveal any distinct association with location (Carey
et al., 1990). Lipid-normalized TCDD concentrations in different fish species ranged
only ą 50% from the mean, except for older fish which had greater concentrations.
There was a tendency for deeper water fish (lake trout, sculpin, smelt) to have greater
lipid-normalized TCDD concentrations than near-shore, shallower water species,
including brown trout (Salmo trutta), yellow perch (Perca flavescens) and smailmouth
bass (Micropterus salmoides). White perch (Morons americana), an introduced
species with a less known habitat and feeding preference in Lake Ontario, were found
to have the greatest concentration of TCDD, at least in part due to their age. The lake
wide average BSAFs found for each species were; lake trout - 0.07, brown trout -
0.03, yellow perch- 0.03, white perch - 0.20, smailmouth bass - 0.05, smelt - 0.06 and
slimy sculpin - 0.12.
Lake Ontario lake trout TCDD residues declined approximately 65% between
1978 and 1988 (Cook et al., 1993a). The problem of estimating the Lake Ontario
water concentration of TCDD associated with the measured TCDD residues in fish and
wildlife can only be approached from the sediment contamination record. Surface
sediment contamination in depositional basins has steadily decreased to the present
10% of the peak level that occurred around 1960 (Cook et al., 1993a). A mass
balance model for TCDD loading of Lake Ontario (Endicott et al., 1990) estimated that
1 kg TCDD/year would result in a steady-state average surface sediment
concentration of 55 pg TCDD/g dry weight sediment for a log K^ of 6.5 and a foc of
0.03. The 1987 lake wide average sediment concentration of TCDD measured for
Lake Ontario was 68 pg/g (Short et al., 1990) with 2.34% organic carbon (2,906 pg
TCDD/g sediment organic carbon). Since the sediment record indicates that Lake
Ontario is not at steady-state due to large reduction in TCDD loading since 1962,
TCDD concentrations in the water and the pelagic food web are likely to be controlled
by contaminated sediment interaction with the lake water, principally through
resuspension of sediment.
The dynamic mass balance model for Lake Ontario (Endicott et al., 1990)
predicted that a longterm 100% reduction in TCDD loading causes the water
concentration of TCDD to decline for decades in proportion to the decrease in surface
sediment concentrations. The ratio between water and sediment TCDD
concentrations (CyCJ associated with reduced TCDD loading was predicted to be
40% of the ratio associated with a steady-state condition. This model, with slight
changes in organic carbon partitioning parameters (Endicott et al., 1993), was used for
this report to estimate freely dissolved TCDD (C^) and total TCDD (C^,) concentrations
under either steady-state or major load reduction conditions.
Three TCDD loading scenarios were simulated for the purpose of estimating the
1987 Lake Ontario average concentration of TCDD in water and BAFjS for lake trout:
(1) steady-state; (2) a 90% reduction in TCDD loading for 20 years; and (3) a 100%
3-15
-------
reduction in TCDD loading for 20 years. The lake wide average TCDD concentration in
the model's 1.8 cm active surface sediment layer was set at 110 pg/g dry sediment or
3,667 pg/g organic carbon (1^=0.03). This sediment concentration was calculated
from 29 surface sediment samples (0 to 3 cm) collected in 1987 (Short et al., 1990)
from depositional basins in Lake Ontario. The model conservatively assumes an
average suspended solids retention time of 43 days in a completely mixed water
column and instantaneous equilibrium of TCDD in water with suspended solids. The
suspended solids concentration is set at 1.2 mg/L with a foc of 0.15 and the amount of
dissolved TCDD associated with colloidal or dissolved organic carbon is assumed to
equal half the amount of dissolved TCDD associated with suspended particle organic
carbon (TBF^. = 1.5). Although the ^,c can be approximated to equal the measured
K^ of 107 (Burkhard and Kuehl, 1986), K^ and Koc > 107 may be more accurate
(Lodge and Cook, 1989). Thus a K^ of 10s is a reasonable upper bound for TCDD
partitioning and is included in this estimation of C^ and C^ associated with the 1987
Lake Ontario lake trout average TCDD concentration (Ct =194 pg/g lipicl) and
sediment active layer average TCDD concentration (C00 = 3667 pg/g organic carbon).
Table 3-2 contains estimated concentrations of dissolved and total TCDD in
Lake Ontario water for 1987 TCDD contamination levels under the three TCDD
loading scenarios. Figure 3-2 shows the expected time course of the ratio of water to
sediment concentrations for the two load reductions. The water concentrations only
vary as a function of the loading scenario and K^. The BAF*[ and BAFj are calculated
for lake trout. BAFts for other fish species can be calculated by dividing their average
lipid-normalized TCDD residues by Cw. The same result is obtained by multiplying the
lake trout BAF,s by the ratio of the species' BSAF to the lake trout BSAF. BAFj
changes slightly when K^ is increased from 107 to 108 but BAF, is strongly influenced
rable 3-2. Steady-state TCDD bioaccumulation factors for lake trout, calculated
from estimated Lake Ontario water concentrations in 1987.
Loading Scenario
Steady-state
Steady-state
20 y-90% load red.
20 y-90% load red.
20 y-1 00% load red.
20 y-1 00% load red.
KOB
107
10s
107
108
107
108
rd
Cw
pg/i-
.102
.0102
.064
.0068
.050
.0057
rvt
Cw
pg/L
.376
.286
.235
.190
.185
.160
fd
.27
.036
.27
.036
.27
.036
BAP?
1.90X108
1.90 X107
3.03X106
2.86 X107
3.86 X 10s
3.40 X 107
BAP*,
5.16 X 10s
6.78 X 10s
8.26 X 10s
1.02X109
1.05X108
1.21 x10e
3-16
-------
by K^, This coincides with a ten-fold difference in the fraction of TCDD dissolved in
water (fd). The choice of loading scenario influences the estimated water/sediment
TCDD concentration ratio (CJJ/Cs) but only by a factor of two. Uncertainty associated
with the choice of 20 years as the period of Lake Ontario response to TCDD loading
reductions which probably began in the 1960s does not influence C^/C^ for the zero
loading scenario (Figure 3-2). The 90% TCDD loading reduction scenario, however,
results in initial reduction in C|/C*, followed by many years of slow increase toward a
new steady-state condition for the reduced TCDD loading.
On the basis of the sediment core record (Cook et al., 1993a), the 90% TCDD
load reduction scenario for Lake Ontario is most probable and thus is used here to
select the best estimate of lake trout BAF(. Koc and ^ are both estimated to be 107;
however, the uncertainties for the true values of K,,w and K00 remain. On this basis,
the BAFf for lake trout is estimated to be 3.0-106 and the BAF, is estimated to be
8.3-10s. The latter value is close to an earlier BAF, estimate of 1 -104/1% lipid used
for human health and wildlife exposure assessments in the Great Lakes Water Quality
Initiative (U.S. EPA, 1991 a). The BAF^ estimate for TCDD may be less than Kow as
the net result of biomagnification potential in the aquatic food web countered by a slow
rate of TCDD metabolism in fish which can be shown to significantly decrease organic
chemical bioaccumulation (de Wolf et al., 1992). If the Lake Ontario model
overestimates the magnitude of sediment resuspension on lake water TCDD
concentrations, the BAFs reported here are too small.
The applicability of the Lake Ontario TCDD BAF,s to other aquatic systems is
primarily a question of partitioning or bioavallability differences. Because the range of
BAF,s across fish species seems small, especially if age and sediment feeding habits
are considered, the Lake Ontario BAFj estimate may be the best predictor of TCDD
residues in other systems jf G* can be estimated accurately. The Lake Ontario BAFJ
estimate is based on a C^ estimate that is strongly influenced by suspended solids
and DOC concentrations and thus could not be used directly to estimate
bioaccumulation in aquatic systems having greater suspended solids and DOC
concentrations. A site-specific fish BAFj for TCDD, assuming a log Kow of 7.0, is
3.0-106*fd, where fd is the fraction freely dissolved. Alternatively, since fd is low even
in Lake Ontario water, this can be further simplified to approximately 0.2-106/POC
using equation 3-10. The extent to which fish accumulate super-hydrophobic organic
chemicals like TCDD when bound to suspended solids in turbid water systems which
have less than 1% of waterborne chemical in a freely dissolved state is uncertain.
3.4 TCDD BIOMAGNIFICATION FACTORS
Lake Ontario provides a good TCDD aquatic exposure site for studying BMFs
because of the relatively uniform spatial distribution of TCDD and similar chemicals
and the availability of monitoring data. Biomagnification of TCDD appears to be
significant between fish and fish-eating birds but not between fish and their food
3-17
-------
:igure 3-2. Predicted TCDD concentration response to loading reduction in Lake
Ontario.
V)
O
5
O
1.00e-6
9.00e-7 -
S.OOe-7 -
7.000-7 -
6.00e-7 -
5,OOe-7 -
4.00e-7
zero load
90% bad reduction
log Kow = 7.0
5 0 5 10 15 20 25 30
years following load reduction
CO
Q
O
1.00e-7
9.00e-8
B.OOe-8 -
7.00e-8 -
6.00e-8 -
S.OOe-8
1
w zero load
90% load reduction
log Kow = 8.0
-5 0 5 10 15 20
years following load reduction
25 30
3-18
-------
(Carey el al., 1990). When calculated for older predaceous fish such as lake trout
eating young smelt, the BMP can equal 3. The lack of apparent biomagnification
among fish is probably due to the influence of biotransformation of TCDD by the fish.
Limited data for the base of the Lake Ontario lake trout food chain indicates little or no
biomagnification between zooplankton and forage fish (Whittle et al., 1992), BMFs
based on fish consuming invertebrate species probably are close to 1.0 because of
the TCDD biotransformation by forage fish. BMFs greater than 1.0 may exist between
some zooplankton species and their food due to the lack of TCDD biotransformation in
invertebrates. The TCDD BMFs for herring gulls (laws argentatus) and herring gull
eggs in comparison to alewife (Alosa pseudoharengus) are 32 and 21, respectively
(Braune and Norstrom, 1989). This compares to BMFs of 93 and 32 for total PCBs.
A pelagic [phytoplankton-zooplankton-herring (Clupea harengus)-cod (Gadus
morrhua)] food chain and a littoral [phytoplankton/detritus-blue mussel (Mytilus edulis)-
juvenile eider duck (Somateria mollissima)] food chain from the northern Baltic Sea
were examined for biomagnification of PCDDs and PCDFs (Broman et al., 1992). The
authors claimed that the combined concentrations of TCDD, 2,3,4,7,8-
pentachlorodibenzofuran and 1,2,3,7,8-pentachlorodibenzo-p-dioxin increased with
trophic level; however, the concentrations when lipid normalized do not appear to
demonstrate biomagnification. TCDD was not detectable in zooplankton and gave a
BMF a 1.3 for the blue mussel-eider duck relationship and a BMP of 0.84 for the
herring-cod relationship.
3.5 BIOTA-SEDIMENT ACCUMULATION FACTORS
If the range in TCDD BSAFs for different fish species from different ecosystems
with different food chain structures, contaminant distribution patterns and bioavailability
conditions is relatively small, the TCDD BAFf estimates obtained from Lake Ontario
data should have general application to fish in other systems. BSAFs carefully
determined for a wide range of bioaccumulative organic chemicals found in lake trout
from Lake Ontario correlate well with BSAFs measured for brown bullheads (Ictalurus
nebulosus] in the Fox River near Green Bay, Wl (P. Cook, ERL-Duluth, unpublished
data). BSAFs for PCBs in both oligochaetes and bullheads in the Fox River (Ankley et
al., 1992b) demonstrate that these organic chemicals in the sediments are
bioaccumulated to levels expected for equilibrium partitioning. In both Lake Ontario
and the Fox River, BSAFs for TCDD in fish are approximately 20-fold than for
PCBs. Thus, despite the large differences in ecosystem characteristics and fish
species, the BSAFs consistently predict each chemicals bioaccumulatipn potential.
Agreement in BSAFs measured for TCDD and other bioaccumulative organic
chemicals at different sites, however, does not eliminate the large uncertainty in BAF^s
associated with accuracy of Kow measurements or estimates that translate directly into
uncertainty for the estimated
3-19
-------
Beside the TCDD BSAFs reported for fish in Lake Ontario, there are only a few
studies in which sediment sampling and chemical analyses were performed so that
BSAFs may be calculated and used as reliable quantitative estimates of
bioaccumulation potential. BSAFs obtained from data sets having sediment TCDD
concentrations likely to represent average surface sediment conditions associated with
the region inhabited by the organisms are summarized in Table 3-3.
Table 3-3. Steady-state biota/sediment accumulation factors (BSAF) for TCDD.
Species
Brown trout
Lake trout
Small, bass
White perch
Yellow perch
Smelt
Sculpin
Herring gull
Carp
Bullhead
Sandworm
Clam
Shrimp
Location
Lake Ontario
a
ft
it
H
11
n
i*
Wisconsin River
Fox River
Passalc River
II
11
Sed. Char,
Lakewide mean
' ť
ti
n
il
n
n
II
Reservoir mean
Segment mean
Laboratory mean
n
n
BSAF
0.03
0.07
0.05
0.20
0.03
0.04
0.12
0.43
0.27
0.05
0.48
0.93
0.73
Fleference
1
1
1
1
1
2
2
1,3
4
5
6
6
6
References: 1. Carey et al. (1990).
2. Batterman et al. (1989).
3. Braune and Norstrom (1989).
4. Kuehletaf. (1986).
5. Cook (unpublished).
6. Rubinstein et al. (1983).
The larger fish BSAFs in Table 3-3 appear to result from factors such as age
(white perch), association with sediments having TCDD concentrations that exceed the
average used to calculate the BSAF (sculpins), or sediment ingestion exposure (carp).
The Lake Ontario herring gull BSAF reflects biomagnification of TCDD. The saltwater
benthic invertebrates have larger BSAFs due to direct exposure to sediment and the
3-20
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lack of metabolism of TCDD. Additional BSAFs for TCDD were reported for suckers
(Catostomus spp.), mountain whitefish (Prosopium williamsoni) and Lake whitefish
(Coregonus clupeaformis) in Canadian rivers associated with bleached kraft paper
mills (Muir et al., 1992a); however, the sediment samples were not collected for the
specific purpose of obtaining the mean BSAF values, which ranged from 0.14 to 1.88.
It is possible that the feeding habits of these fish and/or slight disequilibrium between
sediment and water (R^) caused the approximately ten-fold greater BSAFs than
reported above in Table 3-3 for fish but it is also possible that the sediment sample
analyses underestimated the amount of TCDD available for bioaccumulation in the
river water and food to which the fish were exposed.
In summary, few BSAF values that quantitatively measure the degree to which
fish approach equilibrium with surface sediment have been reported. For the systems
and fish species studied, the BSAF appears to vary within the 0.03 to 0.3 range.
Larger BSAFs, which have been reported on the basis of less certain average surface
sediment characterizations, may be biased by measurement of less contaminated
layers than the average sediment/water interface. BSAFs much less than one indicate
that fish generally are well below equilibrium with respect to the sediment (at least ten-
fold for BSAF=0.1, or even more if KOW/KOC>1). Low BSAFs might reflect disequilibrium
between water and sediment (Rws<1) due to the effects of decreasing anthropogenic
inputs into the water coupled with slow sediment resuspension; sediment diagenesis
which increases chemical activity in the sediment; and loss processes from the water
which keep TCDD concentrations depressed relative to sediment. The low BSAFs
might also reflect effects of growth dilution and metabolism (in the entire food chain)
which keep tissue concentrations below equilibrium values with the water (Raw<1).
Comparison of BSAFs for TCDD to BSAFs for PCBs which are approximately 20-fold
greater suggests that metabolism of TCDD by fish is a major contributor to the overall
TCDD disequilibrium (RasŤ1). The importance of metabolism is slightly uncertain due
to uncertainty for the relative differences in biomagnification potential and Rws between
TCDD and PCBs.
BSAFs may be used to calculate BAFs for other PCDD, PCDF and PCB
congeners that could contribute to Ah receptor mediated toxic effects. This is
particularly important when toxicity equivalency factors (TEFs) are used to estimate
the combined toxic potential of PCDDs, PCDFs and PCBs, as a toxicity equivalence
concentration (TEC), in the exposure of fish, wildlife or humans through aquatic food
chains:
TEC = Ł, (Cij-BAFf-TEF,) (3-25)
The selection of BAFj for each congener is an important step in the TEC calculation.
Assuming that the relative amount of each organic chemical bound to sediment
approximates the relative amount bound to suspended solids, bioaccumulation
3-21
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-------
4. EFFECTS
4.1 COMPARATIVE TOXICOLOGY
Hundreds of publications have focused upon the toxicology of TCDD in
laboratory mammals and birds. In addition to being one of the most toxic synthetic
molecules known, TCDD represents the prototypical compound for a variety of
structurally-similar contaminants of environmental concern that appear to act via the
same mode of action (MOA), which include other 2,3,7,8-substituted PCDDs and
PCDFs, and several non-and mono-ortho-substituted (planar) PCBs (for reviews, see
Goldstein, 1980; Poland and Knutson, 1982; Greenlee and Neal, 1985; Whitlock,
1987, 1990; Safe, 1990). The initial step by which TCDD is thought to exert its toxicity
is through binding to the cytosolic Ah receptor (Poland and Glover, 1980).
Endogenous ligands for the Ah receptor have not been identified, and some
researchers have hypothesized that the functions of the Ah receptor may be regulated
by exogenous materials (for reviews, see Nebert et al., 1981; Okey, 1983; Nebert and
Gonzalez, 1987). After initial binding, the ligand-receptor complex is translocated to
the nucleus of the cell where it becomes associated with DMA thereby causing
initiation of transcription of one or more target genes (Okey et al., 1979; Gonzalez et
al., 1984; Denison et al., 1989; Nebert, 1990; Whitlock, 1990). The subsequent suite
of physiological effects observed are somewhat species-specific, but surprisingly
consistent across vertebrate phylogenetic lines. In addition to lethality, common
effects include weight loss ("wasting syndrome"), decreased immunocompetence,
subcutaneous edema, reproductive effects (fetotoxicity, teratogenesis), alterations in
lipid metabolism and gluconeogenesis, thymic atrophy, and induction of certain
enzyme systems, most notably cytochrome P4501A1 (Goldstein, 1980; Poland and
Knutson, 1982; Greenlee and Neal, 1985; Safe, 1990). Characteristic of TCDD-
induced toxicity in mammals and birds is a delayed onset of mortality, even at
relatively large doses (see following sections on the effects of TCDD on aquatic life
and wildlife). For this reason, it is inappropriate to refer to the acute versus chronic
effects of TCDD; these terms are often used to refer to both exposure duration and
time-to-effects. In the case of TGDD, more accurate terminology would be short-
versus long-term exposure and lethal versus sublethal effects.
Species-specific factors such as uptake, disposition and metabolism of TCDD,
as well as interspecies differences in concentration, tissue distribution and ligand
affinity of the Ah receptor, all likely play a role in determining the relative sensitivity of
organisms to TCDD. However, the presence of the Ah receptor clearly appears to be
a necessary prerequisite for TCDD (and related compounds) to exhibit toxicity.
Moreover, the relative affinity of planar PCBs, and 2,3,7,8-substituted PCDFs and
PCDDs for the Ah receptor appears to dictate the relative toxicity of these compounds
to different test species (Poland and Glover, 1977; Bandiera et al., 1984; Mason et al.,
1986). When sensitive techniques have been used, all mammals (including man) and
birds investigated thus far have exhibited detectable concentrations of Ah receptor in a
4-1
-------
number of different tissues, though often at quite different levels. This suggests that it
is valid to extrapolate TCDD toxicity/receptor models developed for standard
laboratory species, such as the rat or chicken (Gallus domesticus), to avian and
mammalian wildlife. Results of initial studies using sucrose density gradient
centrifugation to detect the Ah receptor in fishes and amphibians were somewhat
ambiguous (Denison et al., 1985; 1986). However, subsequent experiments using
more sensitive techniques to detect and quantify the Ah receptor conclusively
demonstrated its presence in teleost and elasmobranch fishes and teleost cell lines
(Lorenzen and Okey, 1990; Hahn et al., 1992). The Ah receptor has not been
detected in some primitive fishes (hagfish, lamprey; class Agnatha), and has not been
found in nine species of invertebrates representing eight classes of four phyla (Hahn
et al., 1992). Some studies suggest the presence of a protein similar to the Ah
receptor in certain terrestrial invertebrates (Bigelow et al., 1985; Muehleisen et al.,
1989); however, it appears that the structure (and function) of that protein differs from
the Ah receptor present in vertebrates (Hahn and Stegeman, 1992). The possible
presence of the Ah receptor in amphibians or reptiles remains uncertain, as relatively
sensitive detection techniques have not been applied to these animal classes.
The presence of the Ah receptor in fishes, and lack of the receptor in aquatic
invertebrates, is consistent with the relative sensitivity of the two groups of species to
TCDD and structurally-similar compounds. For example, TCDD has been shown to be
lethal to a number of fish species when administered either through the diet or the
water (Miller et al., 1973; 1979; Norris and Miller, 1974; Hawkes and Morris, 1977;
Helder, 1980; 1981; Adams et al., 1986; Kleeman et al., 1988; Mehrle et al., 1988;
Spitsbergen et al., 1988a; 1988b; Walker, 1991; Cook et al., 1991; Walker et al.,
1993). Moreover, exposure of fishes to TCDD results in effects similar to those seen
in mammals, such as delayed mortality (see numerous references in the following
section on the effects of TCDD on aquatic life), a "wasting" syndrome (Hawkes and
Norris, 1977; Miller et al., 1979; Kleeman et al., 1988), reproductive toxicity (Walker,
1991; Walker et al., 1993), histopathologic alterations (Helder, 1980; Johnson et al.,
1986; Spitsbergen et al., 1988a; Cook et al., 1991), possible immunosuppression
(Spitsbergen et al., 1988c) and induction of cytochrome P450-dependent
monooxygenases (Pohl et al., 1975; Vodicnik et al., 1981; Janz and Metcalfe, 1991;
van der Weiden et al., 1992). Conversely, long-term exposures of a number of
invertebrate species (snails, worms, daphnids and mosquito larvae) and aquatic plants
to TCDD failed to cause discernable toxicity (Miller et al., 1973; Isensee and Jones,
1975; Isensee, 1978; Yockim et al., 1978; Adams et al., 1986). In other instances,
exposure of invertebrates to 3,3,'4,4'-tetrachlorobiphenyl, a PCB congener with a MOA
similar to TCDD, also did not cause toxicity, even in relatively long-term exposures
(Borgmann et al., 1990; Dillon et al., 1990). More extensive studies, however, are
required to fully evaluate the potential for TCDD to elicit adverse toxicological
responses in aquatic invertebrates, either via protein(s) with properties similar to the
Ah receptor or via another MOA.
4-2
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The Ah receptor has been detected in most vertebrates and preliminary studies
indicate a high degree of structural similarity among species (Denison et al., 1991;
Bank et al., 1992). However, because of other factors (e.g., metabolism, distribution,
differential gene expression, etc.) which may influence the toxicity of TCDD and
related compounds, interspecies extrapolations of potential toxicity based only on
presence of the Ah receptor should be made with caution. For example, although
certain mono-ortho substituted PCBs (e.g., 2,3,3',4,4-pentachlorobiphenyl, 2,3',4,4',5-
pentachlorobiphenyl) strongly bind to the Ah receptor and induce cytochrome P4501A1
in mammals, these PCB congeners were found not to be potent inducers of
cytochrome P450 in fishes (Gooch et al., 1989). Similarly, some evidence suggests
that planar PCBs may be less potent in fishes than in mammals (Walker and
Peterson, 1991).
The planar PCBs and 2,3,7,8-substituted PCDFs and PCDDs appear to act via
the same MOA as TCDD, suggesting that their toxicities would be additive. In fact, a
number of studies suggest that for certain endpoints, the effects of mixtures of PCBs,
PCDFs and PCDDs are additive (Sawyer and Safe, 1985; Weber et al., 1985; Vecchi
et ai., 1985; Eadon et al., 1986; Bimbaum et al., 1987; Pluess et al., 1988). This is of
concern because in most instances complex mixtures of the chlorinated hydrocarbons
are simultaneously present in environmental samples. Thus, although an individual
PCB, PCDF or PCDD congener may not be present at a toxic concentration, the
combination of these compounds could result in toxicity. For this reason, the potential
toxicity of mixtures of PCBs, PCDFs and PCDDs in environmental samples has been
evaluated using different methods to determine sample "toxic equivalents" relative to
TCDD. Two basic methods have been used for this. In the first, concentrations of
individual PCB, PCDF and PCDD congeners are determined and multiplied by TEFs
which then are summed to express potential toxicity in TCDD-equivalents (TCDD-EQ)
(Sawyer and Safe, 1985; Eadon et al., 1986; Kannan et al., 1988; 1989; Niimi and
Oliver, 1989; van Zorge et al., 1989; Bellward et al., 1990; Olafson et al., 1990; Safe,
1990; Ankley et al., 1992a). The TEFs may be derived from in vitro or in vivo studies
evaluating the potency of individual congeners relative to TCDD. It should be noted
that many of the TEFs currently employed for risk assessments were derived from
induction of cytochrome P4501A1 in laboratory mammals or mammalian cell lines, and
there remains not only some question as to exact relationships between P450
induction potency and in vivo toxicity of PCBs, PCDFs and PCDDs, but also
uncertainly as to the validity of extrapolations from mammalian systems to
nonmammalian species.
The second method for determining sample TCDD-EQ employs techniques to
extract the total PCB/PCDF/PCDD mixture from environmental samples. The extract
then is tested for potency, relative to TCDD, using a standard biological response as
an endpoint. A commonly used biological system for the latter approach to
generating TCDD-EQ is induction of cytochrome P4501A1 (and associated
monoxygenase activities) in the H4IIE rat hepatoma cell line (Bradlaw and Casterline,
4-3
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1979; Bradlaw et al., 1980; Casterline et al., 1983; Safe et al., 1987; 1989;
Zacharewski et al., 1989; Ankley et al., 1991; Hanberg et al., 1991; Tillitt et al., 1991 a;
1991b; 1992). Other biological systems also are potentially capable of measuring the
overall potency (or toxicrty) of complex mixtures of poiychlorinated hydrocarbons. For
example, the salmonid egg injection system described by Walker et al. (1992) may
prove to be a useful method for this type of analysis. Overall, a biological approach to
generating TCDD-EQ may in some instances be superior to the analytical technique
described above, because the assumption of an additive model of toxiclty for the
complex mixtures of planar and nonplanar PCBs, PCDFs and PCDDs found in
environmental samples may not always be appropriate (Birnbaum et al., 1985;
Bannister et al., 1987; Bannister and Safe, 1987; Biegel et al., 1989). However, if a
biological approach to measuring TCDD-EQ is to be used for quantitative risk
assessment, it is important to calibrate the biological system used with specific
toxicological endpoints in the species of concern.
4.2 EFFECTS OF TCDD ON AQUATIC LIFE
4.2.1 Toxicological Information
A review of the literature concerning the toxic effects of TCDD to aquatic life
was obtained from both computerized and manual searches. Aquatic toxicity and
accumulation data for freshwater and saltwater organisms from both laboratory tests
and aquatic model ecosystem studies were included in the review. Most of the
studies contained in previous reviews (Kenaga and Morris, 1983; U.S. EPA, 1984; U.S.
Department of Interior, 1986; Cooper, 1989; and U.S. EPA, 1990) and in the present
literature review, involved short-term exposure periods, where organisms were
exposed to TCDD through several exposure routes followed by a depuration period
where the organisms were held in clean water to observe delayed effects. The
following sections provide both a narrative description of each of these studies and a
summary of the toxicity results for some of the species tested (Table 4-1).
4.2.1.1 Freshwater Plants
The few data that are available for the effects of TCDD on freshwater aquatic
plants are from aquatic model ecosystem studies and indicate that plants are less
sensitive to this chemical than most species of aquatic animals after similar periods of
exposure (Table 4-1). Measured water concentrations (as determined from initially
exposing the bottom soils with TCDD) up to 1,330 ng/L had no observable effect on
algae (Oedogonium cardiacum} and duckweed (Lemna minoi) during 33 days of
exposure (Isensee and Jones, 1975). TCDD residues in algae were 2,295,000 pg/g
after the 33 day period. TCDD residues measured in algae exposed to up to 4.2 ng/L
in other studies by these authors (Isensee, 1978; Yockim et al., 1978) ranged from
1,300 to 5,000 pg/g during a 32-day period. Residues of TCDD in duckweed exposed
to 0.05 to 7.13 ng/L ranged from 200 to 30,700 pg/g (Isensee and Jones, 1975).
4-4
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4.2.1.2 Freshwater Invertebrates
Only a limited number of experiments have been conducted to determine the
effects of TCDD on freshwater aquatic invertebrates (Table 4-1). Isensee and Jones
(1975), Isensee (1978) and Yockim et al. (1978) exposed invertebrates to measured
water concentrations of TCDD in aquatic model ecosystems. After approximately 33
days of exposure, these authors found that concentrations ranging from 0.05 to 1,330
ng/L were not toxic to snails (Physa sp.) or daphnids (Daphnia magna), as determined
by reproductive activity, feeding and growth. TCDD residues measured in snails and
daphnids exposed to 1,330 ng/L were 502,000 and 1,570,000 pg/g, respectively, after
the test period (Isensee and Jones, 1975; Isensee, 1978). In other studies by Isensee
(1978) and Yockim et al. (1978), TCDD residues in snails and daphnids exposed to
lower water concentrations of 2.4 to 4.2 ng/L ranged from 2,500 to 9,700 and 6,800 to
17,100 pg/g for these species, respectively, during the 32-day test period. In another
study, Adams et al. (1986) found no effect on different age groups of Daphnia magna
during a 7-day period following exposure for 48 hours to TCDD concentrations of 0.2
to 1,030 ng/L. Residue concentrations in daphnids were not measured in this study.
In other experiments, Miller et al. (1973) indicated that an initial nominal
concentration of 200 ng/L of TCDD did not cause mortality to snails (Physa sp.),
worms (Paranais sp.) or mosquito larvae (Aedes aegypfy during 36, 55 and 17 days of
exposure in static tests, respectively. Although this concentration appeared to
decrease reproductive success in snails after 48 days of observation and the total
number of worms produced after 55 days, these decreases were not statistically
significant. In addition, no effect on the growth of worms or the pupation rate of
mosquitos were seen at this concentration after 55 and 40 day observation periods
with these organisms. Residue concentrations were not measured in these studies.
4.2.1.3 Freshwater Fish
Because of the large number and diversity of fish studies, the following text is
organized into subsections dealing with the route through which TCDD was
administered to the test species. Fish were exposed to TCDD via water, egg injection,
intraperitoneal (i.p.) injection and the diet. The studies included all life stages (i.e.,
eggs, larvae, juveniles and adult). A summary of the toxic effects of TCDD to some of
these species is provided in Table 4-1.
Waterborne Exposure
Isensee and Jones (1975), Isensee (1978) and Yockim et al. (1978) found that
mosquito fish (Gambusia affinis) and channel catfish (Ictalurus punctatus) exposed to
continuous measured water concentrations of 2.4-4.2 ng TCDD/L died after 15-20
days of exposure in aquatic model ecosystem studies. Death was accompanied by
nasal hemorrhaging, fin necrosis and listless swimming. Measured residues of
4-5
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[14C]TCDD In fish that died ranged from 4,400 to 7,200 pg/g wet weight. Other studies
by Miller et al. (1973), indicated that initial mortality of coho salmon (Oncorhynchus
kisutch) did not occur until 5 to 10 days after the beginning of the exposure period and
that mortality often extended over a 2 month period. Fish exposed to 56 to 100 ng/L
did not feed which resulted in significant decreases in growth after 80 days. Most fish
exposed to 100 ng/L for 24 hours died after 60 days. After a 24-hour exposure
period, the lowest test concentration of 0.056 ng/L caused 12% mortality to young
coho salmon in 60 days compared to a 2% control mortality rate. However, the lowest
concentration did not appear to cause significant effects in later studies (Miller et al.
1979). As with other fish species, signs of toxicity in salmon included skin
discoloration, fin necrosis, and lethargy prior to death.
Later experiments (Miller et al., 1979} showed that an initial, measured
concentration of 0.56 ng/L had no effect on food consumption, weight gain, or survival
of young coho salmon but 5.6 ng/L reduced survival and growth 114 days after an
exposure of 12 hours and caused 50% mortality to salmon in 56 days after exposure
for 96 hours. Percent survival at 114 days decreased with increased exposure
duration beyond 12 hours, indicating to the authors that a critical exposure period for
long-term survival in these studies was between 12 and 24 hours. However, duration
of exposure appeared to be less important than exposure concentrations in
determining mean survival time. Based on these results, the reported waterborne, no-
effect and effect TCDD concentration for young coho salmon (based on survival and
growth) was between 0.56 and 5.6 ng/L. After exposures for 1.5, 3, 6, 12 and 96
hours, TCDD residues measured in whole body extracts of coho salmon selected at
random from those surviving for 114 days indicated that residue levels in these fish
increased with both water concentration and duration of exposure. Fish exposed to
TCDD concentrations of 0.001 to 1.053 ng/L for 96 hours contained 0 to 125 pg/g and
those exposed to 10.53 ng/L for 1.5 to 96 hours contained 68 to 2,170 pg/g after 114-
days. Morris and Miller (1974) exposed different sized guppies (Poecilia reticulata)
from 9 to 40 mm in static tests to nominal TCDD concentrations of 100, 1,000 and
10,000 ng/L for 120 hours. After exposure, fish were transferred to TCDD- free water
and observed for up to 37 days. Eight percent of the fish exposed to 100 ng/L died
during exposure, but all fish exposed to this and higher concentrations died during 37
days in clean water. Test results showed that smaller fish were more sensitive than
larger fish. Similar results were observed for coho salmon in previous studies by
Miller et al. (1973). Later tests by Miller et al. (1979) also showed that a 24-hour
exposure to a much lower nominal water concentration of 0.1 ng/L caused a
significant increase in fin necrosis during 42 days in guppies 8 to 12 mm in length.
The severity of this effect was lessened after 69 days indicating that sublethal effects
of TCDD may be reversible in fish. No fin disease was observed in fish exposed to
0.01 ng/L.
Eggs of northern pike (Esox luclus) and eggs, yolk sac fry and juvenile rainbow
trout were exposed to four nominal concentrations of TCDD (0.1, 1, 10 and 100 ng/L)
4-6
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Table 4-1. Summary of the toxic effects of TCDD to aquatic life and wildlife.
Test
Species
Algae,
Oedogonium cardiacum
Vascular plant
Duckweed,
Lemna minor
Duckweed
Lemna minor
Annelid
Worm,
Paranais sp.
Mollusc
Snail (adult),
Physa sp.
Snail (adult),
Physa sp.
Arthropod
Mosquito (larvae),
Aedes aegyptl
Cladoceran (adult),
Daphnia magna
Test
Method
Model
ecosystem
Model
ecosystem
Model
ecosystem
Water
(static)
Model
ecosystem
Water
(static)
Water
(static)
Model
ecosystem
Water
Cone.
(ng/L)a
1,330
1,330
7.13
200°
1,330
200°
200°
1,330
Organism Duration
Cone.
(P9/9)b Exposure Observation
AQUATIC LIFE
Freshwater Species
2,295,000 33-d
33-d
30,700 33-d
55-d
502,000 33-d
36-d 12-d
17-d 23-d
1,570,000 33-d
Effect
No toxic effect
No toxic effect
No toxic effect
No decrease in
reproductive success
No toxic effect
No decrease in
reproductive success
No effect on pupation
No toxic effect
Reference
Isensee and Jones,
1975; Isensee, 1978
Isensee and Jones,
1975; Isensee, 1978
Isensee and Jones,
1975
Miller etal., 1973
Isensee and Jones,
1975; Isensee, 1978
Miller etal., 1973
Miller eta!., 1973
Isensee and Jones,
1975
4-7
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Test
Species
Cladoceran (1-21 d),
Daphnla magna
Fish
Coho salmon,
Oncorhynchus klsutch
Juvenile (3.5 g)
Juvenile (3.5 g)
Rainbow trout,
Oncorhynchus myklss
Eggs
Eggs
Eggs
Eggs
Eggs
Eggs
Eggs
Test Water
Method Cone.
(ng/L)'
Water 1,030
(renewal)
Water 0.56
(static)
Water 5.60
(static)
Water 0.10°
(renewal)
Water 1°
(renewal)
Water 10°
(renewal)
Egg
injection
Egg
injection
Egg
injection
Egg
injection
Organism
Cone.
(pg/g)b
230
(in eggs)
240
(in eggs)
374
(in eggs)
488
(in eggs)
Exposure
48-h
96-h
96-h
96-h
96-h
96-h
Single
injection
Single
injection
Single
injection
Single
injection
Duration
Observation
7-d
114-d
56-d
160-d
160-d
40-d
Fertilized egg to
swim-up
Fertilized egg to
swim-up
Fertilized egg to
swim-up
Fertilized egg to
swim-up
Effect
No toxic effect
No toxic effect
50% mortality
Delayed development,
reduced growth of fry
Reduced growth,
mortality in sac fry
100% mortality in sac
fry
LR50d (sac fry of
McConaughy strain)
LR50d (sac fry of Erwin
strain)
LR50d (sac fry of Artee
strain)
LR50d (sac fry of Eagle
Lake strain)
Reference
Adams etal., 1986
Miller etal., 1979
Miller etal., 1973,
1979
Holder, 1981; 1982a,b
Helder, 1981;1982a,b
Helder, 1982a,b
Walker and Peterson,
1991
Walker and Peterson,
1991
Walker and Peterson,
1991
Walker and Peterson,
1991
4-8
-------
Test
Species
Rainbow trout (cont.)
Eggs
Eggs
Eggs
Sac fry
Sac fry
Sac fry
Sac fry
Swim-up fry (0.38 g)
Swim-up fry (0.38 g)
Swim-up fry (0.38 g)
Swim-up fry (0.38 g)
Juvenile (0.85 g)
Test
Method
Egg
injection
Water
(renewal)
Water
(renewal)
Water
(renewal)
Water
(renewal)
Water
(renewal)
Water
(renewal)
Water
(flow-thru)
Water
(flow-thru)
Water
(flow-thru)
Water
(flow-thru)
Water
(renewal)
Water
Cone.
(ng/L)a
1°
10°
12.2°
1.83°
0.176
0.0011
0.038
0.046
10
Organism
Cone.
(pg/g)b
421
(in eggs)
279
(in eggs)
439
(in eggs)
3,220
21 9
7659
Exposure
Single
injection
48-h
48-h
96-h
96-h
96-h
96-h
28-d
28-d
28-d
28-d
96-h
Duration
Observation
>48-h to post
swim-up
>48-h to post
swim-up
>48-h to post
swim-up
160-d
10-d
16-d
21 -d
28-d
28-d
28-d
28-d
68-d
Effect
LR50d (sac fry of Fish
Lake strain)
Significant mortality in
sac fry
LR50d (sac fry)
Reduced growth,
mortality
100% mortality
100% mortality
LC50
95% mortality
NOAEL"
LOAEL' (45% mortality)
LC50
Reduced growth,
mortality
Reference
Walker etal., 1992
Walker etal., 1992
Walker etal., 1992
Helder 1981; 1982a,b
Helder 1982a,b
Helder and Seinen,
1985
Bol et al., 1989
Mehrleetal., 1988
Mehrleetal., 1988
Mehrleetal., 1988
Mehrleetal., 1988
Helder 1 981 ;1982a,b
4-9
-------
Test
Species
Rainbow trout (cont.)
Juvenile (0.85 g)
Fingeriing (25-45 g)
Fingerting (25-45 g)
Fingeriing (25-55 g)
Fingeriing (25-45 g)
Fingeriing (35 g)
Fingeriing (7,8 cm)
Fingerting (7,8 cm)
Fingerting (3-7 g)
Fingerting (8 g)
Yearling (100-1 50 g)
Test Water
Method Cone.
(ng/L)a
Water 100
(renewal)
i.p.
injection
i.p.
injection
i.p.
injection
i.p.
injection
Water 107
(static)
Diet
(3.29 ng/g)
Diet
(1,700
ng/g)
Diet
(0.494
ng/g)
i.p.
injection
Organism
Cone.
(pg/g)b
1,000°
5,000°
5,000C
10,000C
650-2,580
314
276,000
250
10,000°
Duration
Exposure Observation
96-h 23-d
Single 25-d
injection
Single 20-d
injection
Single 11-12-wk
injection
Single 80-d
injection
6-h 42-1 39-d
71 -d
71 -d
13-wk 13-wk
Single 2-4-wk post
injection exposure
Effect
100% mortality
Significant
hematological changes
20% mortality
20% mortality,
increased liver weight
LD50
Mortality, fin rot,
increased liver weight
No effect on survival
and growth
100% mortality
No toxic effect
Fin necrosis, no effect
on immune
Reference
Holder 1981; 1982a,b
Spitsbergen et al.,
1988a
Spitsbergen et al.,
1988a
van der Weiden et al.,
1990
Spitsbergen et al.,
1988a; Kleeman et al.,
1988
Branson et al., 1985
Hawkes and Norris,
1977
Hawkes and Norris,
1977
Kleeman et al., 1986a
Spitsbergen et al.,
1986; 1988C
suppression
4-10
-------
Test
Species
Rainbow trout (cont.)
Juvenile (46 g)
Juvenile (46 g)
Immature Adult (300-
400 g)
Lake trout,
Salvelinus namaycush
Eggs
Eggs
Eggs
Eggs
Eggs
Adult
Adult
Test Water
Method Cone.
(ng/L)a
i.p,
injection
i.p.
injection
i.p.
injection
Water
(renewal)
Water
(renewal)
Water
(renewal)
Water
(renewal)
Egg
injection
Diet"
Diet"
Organism
Cone.
(pg/g)b
300-3,060
790
640°
34
(in eggs)
40
(in eggs)
55
(in eggs)
65
(in eggs)
47
(in eggs)
59
(in eggs)
104
(In eggs)
Duration
Exposure
Single
injection
Single
injection
Single
injection
48-h
48-h
48-h
48-h
Single
injection
90-d
90-d
Observation
6-12-wk
3-wk
72-h
>48-h to post
swim-up
>48-h to post
swim-up
>48-h to post
swim-up
>48-h to post
swim-up
Fertilized egg to
swim-up fry
Eggs to swim-up
fry
Eggs to swim-up
fry
Effect
Fin hemorrhage, spleen
histopathology, EROD
induction, P4501A1
induction
ED50 for EROD
Induction
ED50 for AHH
induction
NOAEL8
23% mortality in sac fry
LOAEL1 (sac fry
mortality)
LR50d (sac fry)
LR50d (sac fry)
LR50d (sac fry)
100% mortality to sac
fry
Reference
van der Weiden et al.,
1992
van der Weiden et al.,
1992
Janz and MetcalfŠ,
1991
Walker etal., 1991
Spitsbergen et al.,
1991
Walker etal., 1991
Walker etal., 1991
Walker etal., 1992
Walker, 1991;
Walker etal., 1993
Walker, 1991;
Walker etal., 1993
4-11
-------
Test
Species
Northern pike,
Esox ludus
Eggs
Eggs
Eggs
Caip,
Cyprinus caiplo
Juvenile (20 g)
Adult
Zebrafish,
Brachydanio revio
Adult
Adult
Test Water
Method Cone.
(ng/L)'
Water 0.1°
(renewal)
Water 1.0
(renewal)
Water 10.0
(renewal)
i.p.
injection
Water 0.060
(flow-thru)
Diet
(1.7ng/g)
Diet
Ł8.3 ng/g)
Organism Duration
Cone.
(pg/g)b Exposure Observation
96-h 72-d
96-h
96-h
3,000° Single 80-d
injection
2,200 71 -d 61 -d
Single dose 22-d
Single dose 1-2 Spawnings
Effect
Delayed hatch, reduced
growth of fry
53% mortality to fry
99% mortality to fry
LD50
Mortality and pathology
No effect
Reduced eggs per
spawn,
Reference
Holder, 1980; 1982a,b
Helder, 1980; 1982a,b
Helder, 1980; 1982a,b
Kleeman et al., 1988
Cooketal., 1991
Wannemacher et al.,
1992
Wannemacher et al.,
1992
Fathead minnow,
Pimephales promelas
Juvenile (1.0 g)
Juvenile (0.5-1.0 g)
Water 0.049-
(flow-thru) 0.067
Water 1.7
(renewal)
71-d
28-d
61-d
100% larval mortality
Mortality and pathology Cook et al., 1991
LC50
Adams et al., 1986
4-12
-------
Test
Species
Fathead minnow (cont.)
Juvenile (1. 0-2,0 g)
Bullhead,
Ictalurus me/as
Juvenile (6 g)
Channel catfish,
Ictalurus punctatus
Fingeriing
Japanese medaka,
Otyzlas latlpes
Eggs
Japanese medaka (cont.)
Eggs
Eggs
Eggs
Test Water Organism
Method Cone. Cone.
(ng/L)a (pg/g)b
Water 7.1
(static)
i.p. 5,000°
injection
Model 2.4-4.2
ecosystem
S,M 3.5-6.0
S,M 9.0-13.0
S,M 140-15,0
S,M 14.0-17.0
Duration
Exposure Observation
24-h 60-d
Single 80-d
Injection
15-20-d
Fertilized egg
to 3-d post
hatch
Fertilized egg
to 3-d post
hatch
Fertilized egg
to 3-d post
hatch
Fertilized egg
to 3-d post
hatch
Effect Reference
40% mortality Adams et al., 1986
LD50 Kleeman et a)., 1988
100% mortality Yockim et al., 1978
EC50 (embryos with Wisk and Cooper,
lesions) 1990a,b
LC50 Wisk and Cooper,
1990a,b
EC50 (embryos with Wisk and Cooper,
severe lesions) 1990a,b
EC50 (prevent hatch) Wisk and Cooper,
1990a,b
4-13
-------
Test
Species
Japanese medaka (cont.)
Eggs
Mosquito fish,
Gambusla affinls
Quppy,
Poecllia retlculata
Juvenile (8-12)
Juvenile (9-40 mm)
Bluegill,
Lepomls mactochlrus
Juvenile (30 g)
Largemouth bass,
Mtcropterus salmoides
Juvenile (7 g)
Yellow perch,
Perca flavescens
Juvenile (3-6 g)
Test Water
Method Cone.
(ng/L)Ť
Water
(static)
Model 2.4-4.2
ecosystem
Water 0.1°
(static)
Water 100°
(static)
i.p.
injection
i.p.
injection
Diet
(0.494
ng/g)
Organism
Cone.
(pg/g)b
240
fm
embryos)
16,000°
11,000°
143
Duration
Exposure Observation
Fertilized egg
to 3-d post
hatch
15-d
24-h 42-d
120-h 37-d
Single 80-d
injection
Single 80-d
injection
13-wk 13-wk
Effect
ER501 (embryos with
lesions)
100% mortality
Significant increase In
fin necrosis
100% mortality
LDSO
LDSO
No toxic effect
Reference
Wisk and Cooper,
1990b
Yocklm et al., 1978
Miller etal., 1979
Morris and Miller, 1974
Kleeman et al., 1988
Kleeman et al., 1988
Weeman et al., 1986b
Juvenile
injection
3,000° Single 80-d
injection
LDSO
Spitsbergen et al.,
1988b;
Weeman et al., 1988
4-14
-------
Test
Species
Amphibian
Bullfrog,
Rana catesbeiana
Tadpole
Adult
Rays
Little skate,
Raja erinacea
500-1 ,1 00 g
500-1, 000 g
Fish
Mummiehog,
Fundulus heteroclitus
Eggs
Test Water
Method Cone.
(ng/L)a
injection
injection
injection
i.p.
injection
Water 200°
(static)
Organism Duration
Cone.
(pg/g)b Exposure Observation
1,000,000° Single 50-d
injection
500,000° Single 35-d
injection
Saltwater Species
1,000° Single 10-d
injection
4,500° Two 7-1 2-d
injections
Fertilized egg
to hatch
Effect
No effect on
metamorphosis
No toxic effect
Increased enzyme
activity
10-fold increase in
enzyme activity
20% mortality and 50%
lesions in embryos
Reference
Beatty et al., 1976
Beattyetal., 1976
Bendetal., 1974
Pohl et al., 1974
Cooper, 1989; Prince
and Cooper, 1989
Winter founder,
Pleuronectes americanus
250 g
Oral dose
4,500°
Two doses 8-d
Increased enzyme
activity
Pohl et al., 1974
4-15
-------
Test
Spades
Mink,
Mustela vlson
Newborn
Newborn
Adult
Test Water
Method Cone.
(ng/L)*
i.p.
injection
i.p.
injection
Oral doss
Organism
Cone.
(pg/g)b
1,CX30S
1,000°
4,200°
Duration
Exposure
WILDLIFE
Mammals
Daily for 12 d
Daily for 12 d
Single dose
Observation
133-d
133-d
28-d
Effect
100% mortality after
14-d
62% mortality after
133-d
LD50
Reference
Aulerich et al,, 1988
Aulerich 0t a!., 1988
Hochstein et al., 1988
Bobwhite quail,
Collnus vlrglnlanus
Oral dose
Birds
15,000° Single dose 37-d
LD50
Hudson et al., 1984
Mallard,
Anus platytynchos
Oral dose
>108,000° Single dose 37-d
LD50
Hudson et al., 1984
Ringed turtle dove,
Stretopelia risoria
Oral dose
>810,000° Single dose 37-d
LD50
Hudson etal., 1984
4-16
-------
Test
Species
Ring-necked
pheasant,
Phasianus colchicus
Eggs
Eggs
Eggs
Eggs
Eggs
Eggs
Adult hen
Adult hen
Adult hen
Test Water
Method Cone.
(ng/L)a
Egg
injection,
yolk
Egg
injection,
yolk
Egg
injection,
yolk
Egg
injection,
albumin
Egg
injection,
albumin
Egg
injection,
albumin
i.p.
injection
i.p.
injection
i.p.
injection
Organism
Cone.
(pg/g)b
2,100°
10,000°
1,000°
1,400°
1,000°
100°
100,000°
25,000°
6,250°
Duration
Effect
Exposure Observation
Single dose 28-d post hatch LD50
Single dose 28-d post hatch LOAEL' mortality
Single dose 28-d post hatch NOAEL
Single dose 28-d post hatch LD50
Single dose 28-d post hatch LOAEL' mortality
Single dose 28-d post hatch NOAEL0
Single dose 77-d 100% mortality after
42-d
Single dose 77-d 80% mortality
Single dose 77-d 0% mortality
Reference
Noseketal., 1992c
Noseketal., 1992c
Noseketal., 1992c
Noseketal., 1992c
Noseketal., 1992c
Noseketal., 1992c
Noseketal., 1992a
Noseketal., 1992a
Noseketal., 1992a
4-17
-------
Test
Species
Test Water Organism
Method Cone. Cone,
(ng/L)* (pg/gf Exposure
Duration
Effect
Observation
Reference
Ring-necked pheasant (cont.)
Adult hen
injection
1,000° Weekly 7-wk post dose
Injections for
10 wk
Adult hen
100"
Injection
Weekly 7-wk post dose
injections for
10 wk
LOAEL1
hen mortality
hen weight
egg production
embryo mortality
NOAEL8
Noseketal., 1992a
Noseketal,, 1i92a
Measured TCDD concentration in water.
Measured TCDD concentration in organism (wet weight).
Unmeasured TCDD concentration in water or organism (wet weight).
LR50 (corrected for control mortality) term defined in this report as the measured residue concentration in eggs that caused 50% mortality to sac fry.
NOAEL = No observed adverse effect level.
LOAEL = Lowest observed adverse effect level.
NOAEL and LOAEL values (based on mean measured wet weight organism concentrations) were calculated for this report.
Diet consisted of 22 ng/g peiletized feed followed by fathead minnows injected with 500 pg/flsh.
ER50 = Term defined in this report as the measured residue concentration in eggs that caused 50% effect.
4-18
-------
that were renewed every 24 hours for a period of 96 hours (Helder, 1980; 1981;
1982a,b). After the exposures, eggs, fry and juveniles were held in TCDD-free water
for extended periods of time. The lowest concentration of 0.1 ng TCDD/L caused
embryo underdevelopment and delayed hatching in pike and caused significant growth
retardation of fry of both species in 72 days. Larvae that hatched were also stunted
and hemorrhaging and the onset of edema was observed. Higher concentrations
increased the frequency of these effects. A concentration of 1 ng/L caused 53%
mortality and significantly decreased the growth of rainbow trout fry exposed initially in
the yolk sac stage. The next higher concentration of 10 ng/L killed all yolk sac fry and
retarded growth of juvenile trout 68 days after the 96-hour exposure period. For
juvenile trout, a 100 ng/L exposure for 96 hours was required to cause 100% mortality
after 27 days indicating that older fish were less sensitive than yolk sac fry. TCDD-
intoxication of pike and trout was generally characterized by growth retardation,
delayed mortality and histopathological changes in the stomach, pancreas and liver.
No TCDD residue information was obtained in these studies.
The toxicity and bioconcentration of [3H]TCDD in fathead minnows were
examined by Adams et al. (1986) using two exposure concentrations and duration
regimes; concentrations of TCDD in both water and tissues were based on measured
[3H]TCDD equivalents. Toxicity tests included exposing juvenile fathead minnows to 0,
0.12, 0.72, 7.14 and 81.8 ng/L, each for 1, 2, 3 and 4 days, and to 0, 1.7, 6.7, 63 and
82 ng/L continuously for 28 days. At the end of each of these experiments, fish were
transferred to clean water and observed for 150 and 20 days, respectively. Test
results provided evidence that toxicity from acute exposure to TCDD was dose
dependent and occurred as a function of body burden. TCDD was also found to be
slow acting and to cause delayed mortality (up to 40%) for several weeks following a
1-day exposure to a concentration as low as 7.1 ng/L. Delayed mortality was typically
complete 44 days after exposure which was consistent with studies by Helder (1980;
1981). The effect of longer exposure was a shorter time to 100% mortality;
continuous exposure to 6.5 ng/L produced complete mortality in 23 days, whereas
short exposures to similar concentrations produced 40 to 60% mortality 60 days after
exposure. Continuous exposure to 1.7 ng/L produced 53% mortality in 28 days,
whereas exposures to 0.1 and 0.7 ng/L for 4 days resulted in no significant mortality in
fish during the 150-day obseivation period. Concentrations of 63 to 82 ng/L eventually
produced complete mortality with exposures as short as 1 day. A 28-day LC50 of 1.7
ng/L for fathead minnows was calculated from the results.
Analysis of fathead minnows that died in these toxicity tests indicated that
whole body residues were a function of both the length of exposure prior to death and
exposure concentration. The minimum measured body burden observed in dead fish
obtained from the toxicity studies was 16,700 pg/g and the maximum was 2,042,000
pg/g. Residue information from these experiments was not used by the authors to
determine residue-based effect and no-effect concentrations for TCDD and this
species.
4-19
-------
Adult carp and fathead minnows were exposed continuously to TCDD for 71
days and then were placed in clean water for 61 days to determine the toxicity, uptake
and elimination of TCDD over a period of 132 days (Cook et al., 1991). Mortality and
other toxic symptoms occurred during the exposure and depuration period to both
species. A water concentration of 0.06 ng/L (measured throughout the exposure
phase) was toxic to carp causing mortality, fin erosion, hemorrhaging, cranial
deformation, edema, exopthalmia, tumors and abnormal swimming behavior.
Histological analysis of the liver, spleen, gill and gastrointestinal tract revealed
extensive pathology in these fish (Johnson et al., 1986). Concentrations of 0.049 and
0.067 ng/L caused similar, but less pronounced, toxic effects in fathead minnows
during this time period. Whole body analysis of carp resulted in a peak residue-effect
concentration of 2,200 pg/g after 71 clays and then declined to approximately 900 pg/g
after the depuration period.
Rainbow trout sac fry were exposed to TCDD concentrations ranging from 0 to
50 ng/L and a fly ash extract for 96 hours using renewed water solutions (Helder and
Selnen, 1985). After exposure, fish were held in clean water for 16 days to observe
hemorrhages, edema or death. A comparison of TCDD toxicity with components of
the incinerator fly ash showed that increased concentrations of TCDD caused
increased mortality in less time to trout sac fry than lower concentrations. Cumulative
mortalities ranged from 2% in the lowest concentration of 1.6 ng/L to 100% at 12.2
ng/L and higher concentrations in the test with TCDD alone. Residues in fish were not
measured in these experiments.
Bol et al. (1989) also studied the interactive effects of PCDDs, PCDFs and
PCBs using early life stage tests with rainbow trout. Newly hatched sac fry were
exposed to TCDD and other compounds in water that was renewed every 24 hours for
96 hours. Fish were then held in a continuous clean water flow for three weeks for
observation. Nominal concentrations of TCDD ranged from 0.61 to 20.09 ng/L.
Results, similar to those found by Helder and Seinen (1985), indicated that higher
concentrations caused increased mortality in less time to sac fry than lower
concentrations. Observations on the interaction of TCDD with several compounds
showed that additivity occurred, but that there was a lesser effect in fish exposed at
higher temperatures. The authors also hypothesized that the synergistic effects were
not sufficiently explained on the basis of using a single receptor model, but that a
multiple receptor model consisting of an Ah receptor and a second receptor was more
suitable for explaining the observed "dioxin-like" toxicity. The LC50 calculated for
TCDD alone was 1.83 ng/L.
The toxicity and bioconcentration kinetics of TCDD in rainbow trout were
studied by Branson et al. (1985). Fingerling trout were exposed for 6 hours to an
initial, total measured [14C]TCDD water concentration of 107 ng/L (dissolved TCDD
based on filtration) and then placed in clean water for a period up to 139 days. TCDD
water concentrations did not decline significantly over the 6-hour exposure period,
4-20
-------
Trout appeared healthy for several weeks but delayed mortality was observed on days
78, 136, 137 and 139 after this period. No deaths were observed in the controls.
Increased liver weights of fish were noted from day 42 through the end of the
observation period. After day 64 of the observation period, clinical evidence of fin rot
and the accompanying inflammatory reaction was also observed. Measured residue
concentrations of [14CjTCDD in whole fish decreased from 2,580 pg/g (wet weight)
after the 6-hour exposure period to 780 pg/g after 64 days. Residues further
decreased to 650 pg/g after the 139 day observation period.
Mehrle et al. (1988) exposed juvenile (post swim-up) rainbow trout continuously
to measured [3H]TCDD concentrations of 0.0011 (control), 0.038, 0,079; 0.176, 0.382
and 0.789 ng/L for 28 days in a flow-through system to determine toxicity and
bioconcentration. After the exposure period, fish were transferred to clean water for
another 28-day period for observation and residue analysis. Significant mortality in
trout was observed during 14 days exposure to the highest concentration of 0.789
ng/L and there was a trend toward increased mortality in fish exposed to 0.176 and
0.382 ng/L Significant mortality was not observed in the two lowest concentrations of
0.038 and 0.079 ng/L during the 28-day exposure although reduced weight and
abnormal behavior (lethargic swimming, feeding inhibition and lack of response to
external stimuli) were observed. However, a dose-dependent increase in mortality
was observed throughout the depuration period. Significant mortality (45%) and fin
erosion was recorded at 0.038 ng/L after 28 days of depuration. The next highest
concentration of 0.079 ng/L caused 83% mortality and 95-100% of the fish were dead
at concentrations of 0.176 ng/L and above after this time. Feeding inhibition and other
behavioral changes were not reversed during the depuration period at these
concentrations. A 56-day LC50 of 0,046 ng/L was calculated from the mortality data
for the combined exposure and depuration periods. The no observed adverse effect
level (NOAEL) and lowest observed adverse effect level (LOAEL) for TCDD and
rainbow trout, based on mortality, growth and behavior, were 0.0011 (control) and
0.038 ng/L, respectively.
Tissue analysis of the rainbow trout showed that whole body residues in fish
exposed to the lowest effect concentration of 0.038 ng/L were fairly constant
throughout the exposure and depuration periods (380, 710, 960 and 930 pg/g wet
weight in pooled fish after 7, 14, 21 and 28 days of exposure and 780 pg/g in fish
after 28 days of depuration, respectively). Residue concentrations in control fish
exposed to 0.0011 ng/L (likely due to volatilization and translocation of TCDD from
treated aquaria) ranged from 12 to 27 pg/g (detection limit 6 pg/g) during the exposure
period. If the mean values from these results were used to calculate no effect and
effect values (based on tissue residues) for this study, the NOAEL and LOAEL for
TCDD and rainbow trout using mortality, growth and behavior as the endpoints would
be 21 and 765 pg/g, respectively.
4-21
-------
Experiments were conducted by Spitsbergen et al. (1991) to study the
pathological alterations in early life stages of the lake trout exposed to TCDD as
fertilized eggs. Newly fertilized eggs of lake trout were exposed to graded nominal
concentrations of pH]TCDD in water for 48 hours and were then transferred to clean
water for up to 6 weeks. Water solutions were renewed during the exposure period to
ensure stable solutions. Eggs exposed to 0.1, 1.0, 10, and 100 ng/L contained
measured concentrations of 0, 0, 40 and 400 pg/g wet weight (detection limit for
residue analysis was 0.7 pg/g), which remained constant over the 1- to 6-weeR post
exposure period. Embryos developed normally in all groups until one week prior to
hatch. At this time, a sharp increase in mortality occurred in embryos containing 400
pg/g where approximately 56% of the embryos died during the hatching process.
Eggs containing 40 pg/g also showed an increase in embryo death during hatching.
Cumulative mortality at swim-up of 23 and 100% was significant for embryos
containing 40 and 400 pg/g, respectively. Retrobulbar, meningeal, subcutaneous and
pericardial hemorrhages were evident in many live, morbid and dead embryos and sac
fry containing 400 pg/g of TCDD. By two weeks after the onset of hatching,
approximately 2% of the sac fry containing 40 pg/g developed lesions similar to those
associated with the blue-sac disease syndrome (lesions with petechial hemorrhages,
disruption of the vitelline circulation, cessation of circulation to the tail, head and gills,
subcutaneous edema of the yolk sac, hypopigmentation and arrested development of
skeletal and soft tissue). Mild yolk sac edema was observed in approximately 30% of
this group 2 to 4 weeks after hatching but fry showed no gross lesions at the time ,of
swim-up.
Newly fertilized lake trout eggs were exposed to nominal water concentrations
of 10, 20, 40, 62 and 100 ng TCDD/L for 48 hours in another study (Walker et al.,
1991). After exposure, the eggs were transferred to clean water to observe the toxic
effects of TCDD on hatching and fry survival over a period of 180 days. Measured
concentrations of [3H]TCDD in eggs of 34, 55, 121, 226 and 302 pg/g (wet weight)
remained constant through the egg and sac fry stages of development and were used
as the basis for determining residue-effect levels. Cumulative mortality of embryos
prior to the onset of hatching was unaffected by TCDD. However, hatchability was
significantly less in eggs containing 226 and 302 pg/g. Toxicity was manifested in the
sac fry stage where fry developed subcutaneous yolk sac edema, reflected by an
increase in sac fry weight that preceded and paralleled fry mortality. Neither
cumulative mortality nor wet weight gain in fry was affected by TCDD after swim-up.
The second lowest concentration in eggs of 55 pg/g caused significant mortality to sac
fry 120 days after the exposure period. Increased fry mortality occurred as
concentrations increased. No adverse effects were observed in fry exposed to the
lowest concentration of 34 pg/g in eggs. From these results, the NOAEL and LOAEL,
based on residues, were 34 and 55 pg/g egg, respectively. The LR50 (term defined in
this report as the measured residue concentration in eggs that caused 50 percent
lethality to sac fry) for TCDD in lake trout was calculated to be 65 pg/g egg (corrected
for control mortality) based on mortality of sac fry. Kinetic studies showed that residue
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content in eggs and sac fry were constant during this period of life stage development,
but in post swim-up fry, residue content of TCDD decreased in a first order manner
(tV2 was 35 and 37 days in eggs containing 34 and 55 pg/g, respectively). The
authors noted that the decrease in TCDD concentration per fish, during the fry stage
of development, was due to a combination of whole body TCDD elimination and
dilution from growth.
Wisk and Cooper (1990a) compared the toxicities of several PCDDs (including
TCDD) and TCDF and determined the stage specific toxicity of TCDD to different age
embryos of the medaka. A time interval from fertilization to 3 days post hatch was
used as the measure of survival In these studies. Exposure to graded, measured
water concentrations of TCDD showed a concentration dependent increase in the
percentage of embryos with lesions, embryos with severe, life-threatening lesions and
embryos that were dead by 3 days post hatch. Serious hemorrhaging and pericardial
edema were classified as severe lesions because animals with these lesions died prior
to hatching or at hatch. Generally, the sequence of lesions began with hemorrhages
in various areas including, the caudal region, periorbital, around the pectoral fins, in
the viteiline veins and in the posterior brain area. Concomitantly, pericardial edema
occurred which resulted in the collapse of the yolk sphere and prevented the heart
from undergoing normal chamber formation.
The calculated EC50s and LCSOs for the studied effects, in order of decreasing
sensitivity, were: EC50 of 3.5-6.0 ng/L for embryos with lesions, LC50 of 9-13 ng/L for
survival to 3 days post hatch, EC50 of 14-15 ng/L for embryos with severe, life-
threatening lesions and an EC50 of 14-17 ng/L which prevented hatching. An ER50
(defined in this report as the measured residue concentration in embryos causing 50
percent effect) was calculated to be 240 pg/g based on concentrations analyzed in
embryos that developed lesions (Wisk and Cooper, 1990b). Studies using embryos at
different stages of development indicated that the most sensitive period for toxicity
was during the time of liver formation on day 4 or 5 of development.
Egg Injection
A series of experiments were recently conducted to measure the effects of
chemicals on early life stages of rainbow trout using an egg injection technique.
Walker and Peterson (1991) determined the potencies of PCDDs, PCDFs and PCBs
relative to TCDD for producing early life stage mortality in the trout to help assess the
risk of these chemicals to early life stage mortality of lake trout in the Great Lakes.
Fertilized rainbow trout eggs were injected with graded doses of TCDD or other
PCDD, PCDF and PCB congeners and were then held in clean water until swim-up.
Calculated LR50s (corrected for control mortality) for TCDD (the most potent in
producing early life stage mortality) ranged from 230 to 488 pg/g and were determined
based on the egg dose that caused mortality from hatching onset to swini-up. LR50s
were dependent on the strain of rainbow trout studied (Table 4-1). Signs of toxicity
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from TCDD exposure consisted of a low incidence of half-hatching, while the greatest
TCDD-related mortality occurred in sac fry preceded by hemorrhages and severe fluid
accumulation beneath the yolk sac epithelial membrane.
In a more recent publication, Walker et al. (1992) published the results of the
egg injection technique used to expose fertilized rainbow trout eggs to TCDD (Walker
and Peterson, 1991) and compared this method with waterborne exposure
experiments. Egg injection data on lake trout were also compared to waterborne
exposures with this species from previously published work (Spitsbergen et al., 1991;
Walker et al., 1991). Eggs were either injected with graded doses of [14C]TCDD or
exposed In static tests for 48 hours to various concentrations of [3H]TCDD and then
held in clean water for post swim-up observation. Analysis showed that TCDD
concentrations in rainbow trout eggs were constant through 22 days of development.
Egg mortality of this species was not affected by TCDD following either route of
exposure. Significant increases in mortality of rainbow trout occurred from hatching
onset to swim-up and occurred at egg TCDD doses of 291 pg/g following injection and
279 pg/g following waterborne exposure. Egg mortality in controls in these studies
was 46 and 35 percent for injected fish and those exposed in waterborne experiments,
respectively. The LRSOs (corrected for control mortality) for rainbow trout calculated
from egg injection was 421 pg TCDD/g and from waterborne exposure was 439 pg
TCDD/g (Walker et ai., 1992).
Combined hatching and sac fry mortality was significantly increased at lake
trout egg TCDD doses of ^ 58 and 55 pg/g (Walker et al., 1991) following injection
and waterborne exposure, respectively. Sac fry of lake trout and rainbow trout were
the most sensitive stage tested, whereas no mortality occurred to post swim-up fry.
As in previous studies, toxicity was characterized by half hatching and by sac fry
mortality associated with pericardia! edema, subcutaneous edema of the yolk sac,
exopthalmia and subcutaneous hemorrhages, resembling blue-sac disease. LRSOs
(corrected for control mortality) for lake trout calculated from injection was 47 pg/g and
from waterborne exposure was 65 pg/g (Walker et al., 1991); however, high mortality
(73%) following injection of control lake trout eggs as compared to 15% for control fish
in waterborne exposures somewhat obscured the interpretation of the egg injection
results for this species.
I.P. Injection
Studies to determine the effects of TCDD on several species of fish using i.p.
injection have also been conducted. Spitsbergen et al. (1988a) measured the effects
of TCDD on survival, growth and morphological lesions in juvenile rainbow trout. Two
strains of trout were anesthetized and injected i.p. with single, graded doses of TCDD;
lethality was assessed daily for up to 80 days post treatment. A dose of 1,000 pg/g
did not cause mortality or reduce growth, whereas, 5,000 pg/g caused 20 percent
mortality and depressed body weight gain by week five after treatment. Doses of
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25,000 and 125,000 pg/g caused 90 and 95 percent mortality after 80 days. An 80-
day LD50 of TCDD for rainbow trout was calculated to be 10,000 pg/g (95% C.L.
7,000-15,000 pg/g). Gross and microscopic lesions were evident in trout treated with
10,000 pg/g but not in fish treated with 1,000 or 100 pg/g. However, significant (p <
0.05) hematological changes (leukopenia and thrombocytopenia) occurred in trout
exposed to 1,000 pg/g after 25 days post treatment.
Other studies (Spitsbergen et al., 1986) showed that an injection of 10,000 pg
TCDD/g did not significantly alter humoral immune responses of yearling rainbow trout
(100 to 150 g) two weeks after treatment. Subsequent studies (Spitsbergen, 1988c)
also showed that this dose did not significantly affect mortality or mean time to death
of juvenile rainbow trout following challenge with Infectious Hematopoietic Necrosis
Virus (IHNV). However, within 2 to 4 weeks post-treatment, both studies showed that
fish with this dose became less active, exhibited necrosis of the fin margins and
consumed less food than fish administered lower doses of TCDD. At early times
following virus challenge, histopathologic lesions due to virus disease were more
severe and occurred more frequently in virus-challenged fish which received TCDD
than in virus-challenged control fish. Although TCDD exacerbated IHNV-induced
disease in these fish, the authors suggested that it was not enough to overcome the
battery of defense mechanisms to alter mortality. Spitsbergen et al. (1986) discussed
that the general failure of TCDD to suppress immune responses in rainbow trout at
doses below those causing toxicity parallels findings regarding the effects of a number
of other pharnnacologic immunomodulators in this species. Cyclophosphamide was
indicated to suppress humoral immune responses in rainbow trout only at doses that
approach those that caused lethality. Only corticosteroids were reported to have
suppressed immune responses of rainbow trout at sublethal doses. Even in vitro,
rainbow trout lymphocytes seem to be relatively more resistant to certain
immunosuppressive agents than are lymphocytes of other animals such as mammals.
Although it was suggested that the defense mechanisms against disease in rainbow
trout are fundamentally similar to those of mammals, the processes involved in
immunoregulation in this species seem to be somewhat different. It was suggested
that additional work is needed to clarify immunoregulatory and effector processes in
fish to elucidate the ways TCDD-like chemicals interact with these mechanisms.
van der Weiden et ai. (1990) conducted TCDD dose-response studies with
rainbow trout in an effort to correlate enzyme induction (cytochrome P450) with
toxicological effects such as mortality, growth inhibition and histopathological changes
in the liver and spleen. Trout received single TCDD i.p. injections of 10, 50, 100, 500,
1,000 or 5,000 pg/g. Growth inhibition and 20% mortality were observed 11 and 12
weeks after the administration of 5,000 pg/g. Fish exposed to this concentration were
characterized by fin necrosis and abnormal behavior (head up swimming with
hyperactivity followed by periods of immobility). Relative liver weight was also
significantly increased after six weeks at 5,000 pg/g. Changes in spleen weight did
not appear to be related to dose or exposure time, however, histopathological lesions
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were observed. Increases in 7-ethoxyresorufin-O-deethylase (EROD) activity and total
cytochrome P450 content in the liver were TCDD dose-related and persisted above
control levels for 6 weeks at 500 pg/g and 12 weeks for the 1,000 and 5,000 pg/g
dose levels. This EROD activity response pattern paralleled effects found on growth
and survival. However, both EROD induction in the liver and histopathological
changes in the spleen were observed at the lower dose of 500 pg/g, a dose that did
not cause toxic effects to these animals. No TCDD residues in liver or whole fish
were measured.
van der Weiden et al. (1992) conducted further studies to determine if a
correlation or concurrence between cytochrome P4501A1 induction and toxic
parameters could be established in fish. Juvenile rainbow trout (46 g) were given
single Lp. injections of 6, 30, 60, 300, 600 or 3,060 pg/g body weight, respectively,
and were observed for 1, 3, 6 and 12 weeks post treatment. DosŠ levels of 300 pg/g
in trout caused hemorrhages in fins and the skin after 6 weeks. After 9 weeks, trout
dosed w'rth 3,060 pg/g showed discoloration, became lethargic and did not react to
external stimuli. After 12 weeks, 20% mortality occurred at the highest dose but
significant growth inhibition was not observed. Relative liver weights of trout did not
show pronounced changes at any of the doses; however, histopathological evaluations
revealed inflammations, necrosis and sinusoidal dilatation in the liver at a dose of 600
pg TCDD/g after 3 weeks. Examination of the spleen showed histopathological
damage at the next lowest dose of 300 pg TCDD/g after this same time period.
The 300 pg/g dose of van der Weiden et al. (1992) was also the lowest dose
which caused a significant induction of EROD activity in rainbow trout, although EROD
activity due to this dose decreased to near control levels after 12 weeks post
treatment. An ED50 of 790 pg TCDD/g body weight was calculated for rainbow trout
based on EROD activity in the third week. The dose of 600 pg/g also significantly
increased the total cytochrome P450 content in trout but decreases, similar to that
found in trout EROD activity, were observed after 12 weeks. EROD activity for the
3,060 pg/g dose did not decline after 12 weeks suggesting that at this dose sufficient
TCDD was available in the liver to maintain EROD activity. An overall correlation
between EROD activity and cytochrome P450 content suggested that the increase in
cytochrome P450 content might be caused by the increase in the cytochrome
P4501A1 isoenzyme. Good correlations between enzyme induction and toxicological
effects in rainbow trout showed that cytochrome P4501A1 induction may have
applicability as a screening parameter for assessing aquatic pollution.
Janz and Metcalfe (1991) injected immature rainbow trout (300 to 400 g) with
TCDD and 3,3',4,4'-tetrachlorobiphenyi (PCB 77) separately and in combination to
investigate aryl hydrocarbon hydroxylase (AHH) induction potency in this species.
Trout received nominal doses of 118, 704 and 1,440 pg TCDD/g by i.p. injection and
were sacrificed 72 hours post injection for AHH determination. The ED50 calculated
for AHH induction was 640 pg/g. Expected AHH activities were compared to observed
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AHH activities from fish dosed with equivalent toxic units of the TCDD and PCB 77
mixtures. Observed activities were significantly greater than expected at the two
lowest doses tested indicating that mixtures of TCDD and PCB 77 were greater than
additive based on the AHH response.
Survival, growth and morphologic lesions in juvenile yellow perch were also
studied after treatment with graded single i.p. injected doses of TCDD between 1,000
and 125,000 pg/g body weight (Spitsbergen et al., 1988b; Kleeman et al., 1988).
Perch were then kept in clean water for the same 80-day post-treatment period.
TCDD treatment caused a dose-dependent increase in cumulative mortality in perch.
Doses of 25,000 or 125,000 pg/g caused over 95 percent mortality by the 28th day
post treatment, whereas, approximately 80 percent mortality occurred 80 days after a
single injection of 5,000 pg/g. Cessation of growth followed by loss of body weight
began 2 to 3 weeks after perch received this dose. This dose affected only 20
percent of rainbow trout five weeks post treatment (Spitsbergen et al., 1988a)
indicating that perch survival was more sensitive to TCDD than rainbow trout survival.
As with the trout, no mortality was observed in perch at a dose of 1,000 pg/g or in the
controls. The 80-day LD50 of TCDD in perch was calculated to be 3,000 pg/g (C.L
2,000-4,000 pg/g). Although lesions in the liver tissues occurred in fish exposed to
1,000 pg/g, these and other histologic changes were not sufficient to explain mortality
in perch. Residues of TCDD in these fish were not measured.
The toxicity and biotransformation of TCDD in carp, bullheads (Ictalurus me/as),
largemouth bass (Micropterus salmoides) and bluegills (Lepomis macrochirus) as well
as in rainbow trout (Spitsbergen et al., 1988a) and yellow perch (Spitsbergen et al.,
1988b) were evaluated by Kleeman et al. (1988). Fish were treated with a single
nominal i.p. injection of 1,000, 5,000, 25,000 or 125,000 pg TCDD/g body weight and
were again observed for the same 80-day post-treatment period. Cumulative mortality
was more than 50 percent for all species at the highest doses of 25,000 and 125,000
pg/g. However, at 5,000 pg/g, mortality was greater than or equal to 50% for only the
more sensitive species and no significant mortality was observed in any species at
1,000 pg/g. After an i.p. treatment of 125,000 pg/g, the time to produce 50% mortality
in rainbow trout, largemouth bass, bluegill, and bullheads was 16 to 22 days. In
contrast, yellow perch reached 50% mortality after this same dose in only 7 days,
whereas carp required 44 days. In trout and bluegill, a decrease in body weight was
observed at 1,000 and 5,000 pg/g, respectively, whereas in perch it was observed
only at 5,000 pg/g. Perch treated with 25,000 or 125,000 pg/g exhibited the shortest
delay period prior to death (about 7 to 21 days). The 80-day LD50s were: 3,000,
3,000, 5,000, 10,000, 11,000 and 16,000 pg/g for yellow perch, carp, bullheads,
rainbow trout, largemouth bass and bluegills, respectively. Analysis of gall bladder bile
following a 60,000 pg/g i.p. injection of [14C]TCDD showed the presence in all species
of TCDD metabolites in addition to the parent compound. However, bile samples for
yellow perch indicated that the major metabolite was different from that of the other
fish species. Although the authors did not define what the metabolites were in these
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studies, further analysis suggested that two or more of the biliary TCDD metabolites
were giucuronide conjugates. Because the retention times for the major metabolites
were similar to those observed in mammals, it was suggested that fish and mammals
may have similar pathways for TCDD biotransformation.
Dietary Exposure
Three species of fish in different experiments were fed diets containing TCDD
to determine long-term adverse effects. Hawkes and Norris (1977) fed fingerling
rainbow trout food containing nominal TCDD concentrations of 2.3, 2,300 and
2,300,000 pg/g for 6 days each week for 105 days. Measured values from a single
sample of fish food for each dietary level were 10 (detection limit), 3,290 and
1,700,000 pg/g, respectively. Only the highest concentration in food caused adverse
effects during the test. Feeding activity declined after 10 days and significant
differences in growth were noted in fish after 30 days. After 33 days, the first death
occurred at this concentration and by day 61, 50% of the fish had died. All but two
fish had died by day 71. Fin erosion, the development of fungal growth and
degenerative histological changes were also apparent in these fish during exposure.
No effects on survival, feeding activity, growth or histopathology were noted in fish fed
food containing lower TCDD concentrations or the control diet. Residue analysis of
fish exposed to 3,290 and 1,700,000 pg/g indicated that TCDD was not concentrating
significantly in body tissues after 65 and 105 days. The measured residue
concentrations in single fish exposed to these no effect and effect concentrations,
were 314 and 276,000 pg/g (wet weight), respectively, based on applying a dry to wet
weight ratio of 0.2 to the author's dry weight values.
The effects of dietary TCDD on the reproduction and oogenesis of zebrafish
(Brachydanio rerio) were studied by Wannemacher et al. (1992). After establishing a
stable spawning cycle (10 weeks), female zebrafish weighing 0.6 g (mean weight)
were exposed to nominal TCDD doses of 1,000, 5,000, 10,000 or 20,000 pg added to
a chow diet of TetraMinl Single TCDD doses of 1,670 pg/g did not result in
significant changes in the number of eggs/spawning, adult body weight, larval survival
or histological development. In contrast, fish treated with TCDD at >8,300 pg/g
showed a rapid decrease in the number of eggs per spawning, and after 1 to 2
spawning cycles, spawning was arrested completely. This dose also induced a
significant loss of adult body weight and reduced active movements 20 to 24 days
after exposure. This same dose caused all hatched larvae to develop cranial and
thoracal edema, notochord malformations and killed all larvae 2 to 3 days after these
initial symptoms occurred. Histological examinations of the ovaries revealed no
differences between fish in the controls and those exposed to 1,670 pg TCDD/g. After
TCDD treatment with 8,300 pg/g, the percentage of immature, previtellogenic oocytes
(stages Mil) significantly increased while the number of mature, vitellogenic follicles
(stages IV and V) decreased, and numerous atretic follicles were observed.
Preliminary data from other feeding experiments with zebrafish using [3H]TCDD
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suggested that a single TCDD dose of 2,500 pg/g (which did not affect reproductive
functions) led to a body concentration of approximately 670 and 330 pg/g (assuming a
similar fish weight as above) after 15 and 45 days, respectively.
Kieeman et ai. (1986a,b) studied the accumulation, tissue distribution, and
depuration of TCDD in fingerling rainbow trout (3-7 g) and yellow perch (3-6 g) fed a
diet containing [3H]TCDD at 494 pg/g for 13 weeks followed by the same diet without
TCDD for 13 weeks. This exposure, which resulted in an accumulation of about 250
and 143 pg TCDD/g expressed on a whole fish basis for trout and perch, respectively,
did not cause effects such as fin necrosis, cutaneous hemorrhage or increase in
mortality during the 26-week period. When dietary exposure was stopped, TCDD
residues were siowiy eliminated. Visceral fat, carcass, skin, liver, skeletal muscle,
pyloric caeca and all fatty tissues, accounted for greater than 90% of the TCDD in fish
after the 13 week exposure period.
Increased hepatic EROD activity in rainbow trout was reported to occur within
two days following single doses of 60 - 2,000 pg TCDD/g delivered orally via gelatin in
capsules (Parrott et al., 1992). EROD activity was estimated to increase significantly
above controls at an oral dose of 100 pg TCDD/g, which corresponded to 1,000 pg
TCDD/g of liver. However, no whole fish tissue concentrations of TCDD were
reported. If 50% of the TCDD was assimilated by the trout, the oral dose threshold
level for EROD induction corresponded to 50 pg TCDD/g of fish. Single i.p. injections
of TCDD in rainbow trout required doses approximately five times greater to
significantly raise EROD activity (van der Weiden et al., 1990) in comparison to the
single oral dose threshold for EROD activity found by Parrott et al. (1992). Insufficient
data exist to determine if greater concentrations of TCDD in the rainbow trout livers
were associated with the apparent greater sensitivity for EROD activity as a result of
single dietary exposures in comparison to single i.p. injections. The magnitude and
duration of EROD responses for continuous dietary exposures are most relevant for
assessing exposures and effects in the environment. Hepatic EROD activity in
rainbow trout appeared to decline faster than whole fish TCDF residues (and probably
liver residues) following cessation of a 30 day dietary exposure (Muir et al., 1992b).
The same phenomenon appears to occur for i.p. injection exposure of rainbow trout
with TCDD at dose below 3,060 pg/g, but at doses of 3,060 and 5,000 pg/g no change
in EROD activity was observed for 12 weeks post-exposure (van der Weiden et al.,
1990; van der Weiden et al., 1992).
Walker (1991) and Walker et al. (1993) fed adult lake trout TCDD contaminated
food in an effort to determine adverse effects of TCDD on early life stages via
maternal transfer. In the first phase of the study, adult female lake trout were fed
pelletized food (controls), or graded amounts of pelletized food containing 22,000 pg
TCDD/g food (same concentration of TCDD per dose) during the third month prior to
spawning (early oocyte maturation). The diet was changed to live fathead minnows
(controls), or graded numbers of fathead minnows injected with the same dose of
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TCDD (500,000 pg/minnow) for the remaining period prior to spawning (tate oocyte
maturation). All oocytes (unfertilized eggs) produced were fertilized with sperm from
unexposed mates. Although high, medium and low TCDD exposures were attempted,
females that were aggressive in feeding obtained more TCDD than less aggressive
feeding females in each group. This resulted in sexually mature lake trout with a wide
range of TCDD whole body burdens and associated TCDD egg burdens. In the
second phase of this study, lake trout eggs were injected with graded amounts of
TCDD to compare this route of exposure with effects resulting from maternal transfer
and from waterborne exposure (Spitsbergen et al., 1991; Walker et al., 1991).
Measured TCDD concentrations in the eggs from both phases of the study were used
as the basis for determining relationships between residue and effect concentrations in
early life stages.
TCDD concentrations in oocytes resulting from maternal transfer were: <1 pg/g
(3 control egg samples), 1 to 23 pg/g (4 TCDD egg samples), 50 to 152 pg/g (3 TCDD
egg samples) and 233 to 387 pg/g (3 TCDD egg samples). Because the egg TCDD
dose following maternal transfer varied within and among the 3 exposure groups, no
relationship could be determined between the targeted dietary exposure levels of
TCDD fed to female lake trout and the egg TCDD dose spawned from these fish;
however, TCDD concentrations in eggs were approximately 50% of the whole fish
TCDD concentrations. TCDD doses > 233 pg/g in the eggs resulted in inviable
oocytes and failure of fertilization. Oocytes and ovarian fluid obtained from females
exposed to the greatest TCDD doses were cloudy when released from the female.
Following attempted fertilization, the eggs turned opaque and died when placed in
water. However, eggs with 152 pg TCDD/g following maternal deposition were viable
and successfully fertilized, but all sac fry that successfully hatched from these eggs
died with a blue-sac disease syndrome (Spitsbergen et al., 1991) prior to swim-up.
Therefore, egg TCDD doses lower than those that reduced oocyte viability arid
prevented fertilization caused 100% sac fry mortality. In contrast, egg TCDD doses >
226 pg/g egg following waterborne exposure for newly fertilized lake trout eggs did not
result in an increase in egg mortality (Spitsbergen et al., 1991; Walker e\ al, 1991).
The difference in egg/oocyte viability between these routes of exposure suggested that
reduced oocyte viability and failure of subsequent fertilization might be secondary to
an effect of TCDD on maternal oocyte formation.
Mortality observed from maternally-derived TCDD, injected TCDD or waterborne
TCDD exposure was similar and always associated with subcutaneous yolk sac
edema and hemorrhages; the greatest TCDD-related mortality occurred during the sac
fry stage. The TCDD LR50s for lake trout fry (based on egg residue concentrations)
following maternal transfer, waterborne exposure and injection were 59, 65 and 47
pg/g, respectively, indicating that the route of exposure of TCDD to this life stage may
not be as important as the delivered dose (Walker, 1991; Walker et al., 1991, 1992,
1993).
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4.2.1.4 Amphibians
Only one study has been reported that examined the effects of TCDD on
amphibians (Table 4-1). Beatty et ai. (1976) injected (i.p.) tadpoles and adult bullfrogs
(Rana catesbeiana) with TCDD at doses of 25,000-1,000,000 and 50,000-500,000
pg/g body weight, respectively. Through 50 days post-dose, no tadpole mortality was
attributed to TCDD at any of the doses, indicating that these animals are more tolerant
to TCDD than fish. AH surviving tadpoles were observed to successfully complete
metamorphosis with no observable morphological abnormalities. In addition,
histopathological examination of the liver, heart, kidney, lung and reproductive organs
of these animals, shortly after the completion of metamorphosis, revealed no lesions.
No mortality occurred in any of the treatment groups of adult bullfrogs through 35 days
post-injection. Although food intake was initially lessened in animals exposed to the
highest dose of 500,000 pg/g, food consumption by these animals was not different
from that of the controls. Histopathological examination of several organs revealed no
significant lesions at any of the TCDD doses administered.
4.2.1.5 Saltwater Fish
Information concerning TCDD toxicity was found for only three saltwater
species (Table 4-1). As with freshwater species, some delayed adverse effects
occurred in these studies after short-term exposure (oral dose or i.p. injection). No
information was found on the long-term exposure of TCDD to saltwater species.
Bend et al. (1974) exposed little skates (Raja erinacea) to TCDD (at an oral
dose of 1,000 pg/g) and reported increases in renal and hepatic microsomal
benzo(a)pyrene hydroxylase activity 10 days after treatment. Pohl et al. (1975)
repeated these experiments with the little skate at higher doses (two doses at 4,500
pg/g wet weight administered by i.p. injection) and noted a ten-fold increase in hepatic
benzo(a)pyrene hydroxylase activity 7 to 12 days after treatment. The same study also
reported that winter flounder (Pleuronectes americanus), given two administrations of
TCDD at 4,500 pg/g by stomach intubation, did not show an increase in the activity of
benzo(a)pyrene hydroxylase; however, a significant increase in EROD activity was
observed 8 days after treatment. Residue concentrations of TCDD in these organisms
were not measured.
Studies have also been conducted to determine the effects of TCDD on killifish
(Fundulus heteroclitus) embryos (Cooper, 1989; Prince and Cooper, 1989). Eggs and
sperm were stripped from adult feral fish collected from a site known to be
contaminated with TCDD (Newark Bay, NJ) and a site not impacted by TCDD (Long
Island, NY). Eggs and sperm from fish collected at these two locations were exposed
from fertilization through hatch to solutions containing various concentrations of TCDD.
Major lesions (tubed heart/collapsed yolk sphere) and mortality of the embryos were
observed. At an exposure concentration of 200 ng/L, 55 and 5.5% of the Long Island
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and Newark Bay embryos, respectively, had major lesions. Twenty percent of the
embryos from the Long Island fish and 12% of those from the Newark Bay fish died.
The results suggested that previously exposed fish were less sensitive than fish that
had not been previously exposed to TCDD. Residues were not measured in embryos
from these studies.
4.2.1.6 Toxicokinetics
It is generally understood that persistent, lipophilic organic chemicals
accumulate in fish in proportion to the lipid content and age of each animal
(Gutenmann et al., 1992). Aquatic ecological risk assessments for TCDD and related
chemicals that bind to the Ah receptor will be strongly facilitated when more specific
information concerning the relationships between chemical concentrations in tissues
associated with toxic effects and whole organism exposure levels is available. TCDD
distribution between organs in carp was found to be proportional to lipid content (Kuehl
et al., 1987). With time to reach steady-state, TCDD appeared to distribute in lake
trout tissues in proportion to the total lipid content of the tissues (Cook et al., 1993b).
Exceptions were the testes and ovaries with double and half, respectively, the lipid
normalized TCDD concentration of other organs.
Comparison of [14C]TCDD distribution in rainbow trout to cod with whole-body
autoradiography and liquid scintillation counting was reported to reveal a substantial
Snterspecies difference in TCDD distribution (Hektoen et al., 1992). The TCDD in
rainbow trout followed the expected organ lipid distribution pattern but in cod the
TCDD was predominantly associated with the central nervous system and liver. The
cod has its lipid reservoir in the liver but the cod brain has no more lipid than the brain
of the rainbow trout. Insufficient data were presented to allow comparison of lipid-
normalized TCDD concentrations. It is possible that the lipid distribution pattern in cod
may explain the observed differences in TCDD distribution.
Although TCDD is eliminated slowly from fish, it is metabolized (biotransformed)
to an extent that increases the rate of elimination. TCDD metabolites were measured
in the bile of rainbow trout (Kleeman et al., 1986a) and yellow perch (Kleeman et al.,
1986b) but were not detectable in tissues. Other PCDDs and PCDFs were found to
have faster elimination rates than TCDD due to faster rates of metabolism
(Opperhuizen and Sijm, 1990). Also it has been demonstrated that the influence of
biotransformation on bioaccumulation increases as a function of the K^ of the
chemical (de Wolf et al., 1992). Measurements of the rate of TCDD metabolism in
fish have not been reported; however, comparison of measured elimination rates to
theoretical elimination rates with no metabolism could provide such estimates. An
alternative approach utilized piperonylbutoxide exposure of rainbow trout to examine
the effect of partial inhibition of biotransformation on the elimination and tissue
distribution 1,2,3,7-TCDD, 1,2,3,4,7-PeCDD and 2,3,4,7,8-PeCDF (Sijrn et al., 1990).
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Some variation in the effect of metabolism on TCDD elimination rate may occur
as a function of gill versus digestive tract uptake. For example, dietary preexposure of
toadfish (Opsanus tau) to benzo[a]pyrene was shown to facilitate metabolism of the
compound and was associated with induction of intestinal AHH activity (van Veld et
al., 1988). Unless lymphatic transport of lipids from fish gastrointestinal tracts exists
as an alternative route, TCDD absorbed by fish from food in the intestine must pass
through the portal vein to the liver before entering general circulation and transport to
other organs. First-pass metabolism of TCDD entering the fish through the digestive
tract could result in relatively faster elimination of TCDD than observed in
bioconcentration studies in which the chemical is entering general circulation through
gill uptake. If this phenomenon is only present during the exposure phase of
laboratory bioaccumulation studies, its primary impact would be to reduce the
measured assimilation efficiency and uptake rate constant (k,). As reported earlier,
net assimilation efficiencies reported for TCDD were less than for PCB congeners of
similar K, but with lower rates of metabolism.
Additional metabolism-related uncertainties in bioaccumulation modeling and
predictions of tissue distribution of PCDDs and PCDFs in fish involve the use of
nonsteady-state measurements to predict steady-state conditions. The dependence of
metabolism rate on TCDD dose and length of exposure is not well understood but
time-course studies of P450 induction in rainbow trout by p-napthoflavone demonstrate
that different responses can occur over time depending on the frequency and duration
of exposure (Zhang et al., 1990). Conditions of exposure were shown in mice to alter
the toxic effects of TCDD. TCDD-induced suppression of antibody response was
enhanced approximately 10-fold following subchronic versus acute exposures to the
same cumulative doses of TCDD by oral gavage (Morris et al., 1992). These
exposure complexities indicate that residue-based aquatic toxicity hazard testing
should involve long-term, continuous exposures to the maximum extent possible until
more is known about long-term versus short-term exposure induced responses.
4.2.2 Epidemiological Information
There have been many efforts undertaken to assess responses in fish,
mammal, and bird populations in relation to residue concentrations of TCDD and
similar acting compounds. Clearly, a fundamental challenge in interpreting the results
of such studies lies in the elucidation of plausible cause and effect relationships, given
the variety of chemical and nonchemical stressors that can impact populations. Fox
(1991) discussed criteria by which "ecoepidemiological" studies can be judged and
these include: (a) probability, (b) chronological relationship, (c) strength of association,
(d) specificity, (e) consistency of association upon replication, (f) predictive
performance, and (g) coherence (i.e., biological plausibility, presence of a dose-
response relationship). It should be stressed that the validity of epidemiologically-
based associations increase when they can be supported by independent and
controlled experiments. Epidemiological data are reviewed in this report to insure that
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TCDD residue-toxic effect threshold levels established from toxicological studies are
not contradicted by reports of the existence of fish or wildlife population declines
associated with TCDD exposures. The epidemiological data are not used to establish
prima facie cause and effect relationships for TCDD exposures.
Residues of TCDD and related polychlorinated aromatic chemicals in fish have
not been reported to approach levels associated with lethality to adults, but population
declines through loss of adult organisms may have occurred through undocumented
exposures in small ponds, streams or embayments directly contaminated by nearby
sources of TCDD and/or mixtures of related chemicals. The extreme sensitivity of
trout sac fry (Walker et al., 1991) indicates that early life stage lethality is more likely
to cause fish population declines as a result of TCDD exposure of spawning adult fish.
Insufficient information exists concerning sublethal effects that could contribute to
decreased survival. Immunosuppression and associated biomarkers are particularly
deserving of future epidemiological investigation.
4.2.2.1 Biochemical Changes Observed in Native Fish Populations
Field investigations have been conducted concerning eytochrome P4501A1
induction in fish exposed to TCDD and related compounds. Complicating factors in
these studies are the possible presence of other contaminants or biological factors
that could have caused the effects (Monosson and Stegeman, 1991), uncertainty
regarding the level of exposure to TCDD and related chemicals and lack of
established correlations between these levels of enzyme induction and chemical dose
or toxic effects in fish. Hepatic EROD activity in fish has been used as a general
screening technique for exposure to chemicals which induce eytochrome P4501A1
(Vindirnian et al., 1991). EROD activity has also been used in the rat hepatoma cell
bioassay as a correlate of the overall TCDD toxic potency of complex mixtures of
PCBs, PCDDs and PCDFs in extracts from environmental samples (Tillitt et al.,
1991b). TEFs based on these data did not appear to predict mortality in one study of
Lake Michigan chinook salmon (Oncorhynchus tschawytscha) eggs and fry following
fertilization (Williams and Giesy, 1992) but an inverse correlation between total PCB
concentration in Lake Michigan salmon eggs and hatching success was found in a
second study (Ankley et al., 1991). The relatively larger range of reproductive success
observed by Ankley et al. (1991) may explain why they found a correlation between
chemical exposure and hatching success for salmon eggs that Williams and Giesy
(1992) did not find.
AHH activity in sac fry hatched from eggs obtained from Lake Michigan lake
trout was 3.5 to 8.6-fold greater than hatchery controls (Binder and Lech, 1984).
Dose-response experiments and Lake Michigan trout PCB residue data indicated that
the induction of hepatic mixed function oxidase (MFO) activity in trout embryos and fry
was attributable to PCBs and possibly other xenobiotic organic chemicals.
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EROD activity has been found to be a sensitive biomarker for complex mixtures
of organochlorine chemicals in chlorine bleached kraft pulp mill effluents (BKME)
(Andersson et al., 1988). Rainbow trout in cages located in waters receiving BKME
effluents for three weeks had up to seven-fold greater EROD activity than controls
(Lindstrom-Seppa and Oikari, 1990).
The sediments and biota of Jackfish Bay, Lake Superior are contaminated with
PCDDs and PCDFs attributable to BKME (Sherman et al., 1990). Concentrations of
TCDF in sediments are reported to reach 3,000 pg/g dry weight and white suckers
(Catostomus commersoni) contained 44 pg TCDF/g wet weight and at 7 pg TCDD/g
wet weight. Suckers collected in August, 1988 were found to have elevated hepatic
MFO activity and altered steroid profiles (Munkittrick et al., 1991). Further study
(McMaster et al., 1991) of Jackfish Bay suckers, in comparison to two non-BKME
exposed control populations, indicated increased liversomatic indices, elevated MFO
activity, lower gonadosomatic indices, increased age to maturity and severe reductions
in plasma steroid levels in both males and females. The reduced plasma steroids
included testosterone and 17, 20-dihydroxyprogesterone in both sexes as well as 11-
ketotestosterone in males and 17-estradiol in females. The females contained fewer
eggs at maturity and males were reported to have reduced development of secondary
sexual characteristics. Study of the reproductive performance of the suckers,
however, did not reveal effects on fertilization potential, hatchability of eggs, larval size
at hatch or MFO activity in larval fish (McMaster et al., 1992), although BKME
exposed male suckers had reduced spermatozoan motility.-
Hepatic MFO activity was reduced while gonadal steroid reductions persisted in
Jackfish Bay fish just two weeks after a mill shutdown in September, 1990 (Munkittrick
et al., 1992). Munkittrick et al. (1992) concluded that MFO induction was not related
to accumulation of persistent compounds such as TCDF, however cessation of
exposure of juvenile rainbow trout to TCDF appears to result in rapid reduction of
EROD activity while TCDF residues in the fish decline slowly (Muir et al., 1992b).
Thus the contribution of PCDDs and PCDFs in BKME effluents to observed
biochemical and physiological changes in exposed fish populations is unclear at this
time.
4.2.2.2 Fish Population Declines in Lake Ontario
The largest freshwater body documented to have widespread TCDD
contamination is Lake Ontario (U.S. EPA, 1990). Here the extinction of major fish
populations such as the lake trout and burbot (Lota lota) occurred before 1950
(Hartman, 1988) after which anthropogenic organochlorine chemical inputs to Lake
Ontario increased dramatically. Commercial over-fishing and sea lamprey
(Petromyzon marinus) predation are thought to be the primary causes of lake trout
and burbot disappearance (Christie, 1972), although lamprey existed in Lake Ontario
since the 1830s. Habitat degradation including eutrophication may also.have been a
4-35
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factor in the lake trout decline (Christie, 1974). Low dissolved oxygen and potentially
toxic levels of ammonia in interstitial water of spawning beds may have contributed to
the reproductive failure of lake trout in Lake Ontario (Sly, 1988).
The loss of a third deep water species, the fourhorn, or deep water sculpin
(Myoxocephalus quadricornis) from Lake Ontario can not be explained (Christie,
1974). These small, bottom-dwelling fish which range to extreme depths were
important forage for lake trout and burbot but their decline occurred during the 1950s
after their predators had been extirpated. Lack of predation on sculpins has been
suggested as an explanation for the disappearance of the fourhorn sculpin (Brandt,
1986). The slimy sculpin, under this theory, out-competed the fourhorn sculpin,
however the slimy sculpin has not extended its range to the deep water habitat of the
fourhorn sculpin. Since the fourhorn sculpin population disappeared at a time when
sediments in Lake Ontario depositional basins were rapidly approaching the severe
anthropogenic organic chemical contamination of the 1960s (Cook et al., 1993a), a
toxic chemical etiology is warranted as an alternative theory. Limited data indicate
that slimy sculpins present in Lake Ontario tend to bioaccumuiate TCDD and other
organic chemicals in proportion to the contamination levels in the surface sediments
where they reside rather than on a lakewide average exposure basis as found for
pelagic species. The direct toxicity of sediment contaminants such as TCDD and
related chemicals; reproductive failure due to chemical exposure of parents, eggs or
fry; or loss of food supply due to sediment contamination effects on benthic
invertebrates are possible explanations for fourhorn sculpin extinction.
More than two decades after the extirpation of lake trout in Lake Ontario and
after a lamprey control program was initiated in 1971, the restoration of a naturally
reproducing lake trout population became a Lake Ontario management objective under
the 1978 Great Lakes Water Quality Agreement (Edwards et al., 1990). Stocking of
lake trout began in 1973 but natural reproduction was not achieved. Eggs collected
from Lake Ontario lake trout in 1981-1984 showed a very high incidence of blue-sac
disease and associated mortality (Symula et al., 1990). Blue-sac disease in salmonid
sac fry is known to be caused by different physical and chemical stressors. The
partial reduction in blue-sac mortality with reduced water temperature during
incubation was interpreted as evidence for involvement of a pathogen such as
myxobacteria rather than simple toxicity from chemical contamination of eggs prior to
spawning. Fry that did not succumb to blue-sac did not exhibit the DDT-associated
"feeding fry syndrome" (Skea et al., 1985) after swim-up. Swim-up fry mortality in
Lake Michigan lake trout appears to be different than blue-sac syndrome and has
been associated with exposure to organic chemicals that have a similar mode of
action to that of DDT (Mac and Edsall, 1991). Lake Ontario lake trout eggs collected
after 1984 are reported to have been successfully reared in U.S. Fish and Wildlife
Service hatcheries but the data on the percentage of blue-sac incidence are not
reported (Marsden et al., 1988). Seventy-five lake trout fry were captured in Lake
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Ontario in 1986 (ibid). This provided the first evidence of natural reproduction since
stocking began in 1973.
The strong association now established between TCDD exposure of lake trout
eggs, either through water exposure (Spitsbergen et al., 1991; Walker et al., 1991),
injection (Walker et al., 1992) or maternal transfer (Walker, 1991; Walker et a!., 1993),
and blue-sac syndrome-associated mortality of sac fry compels consideration of
TCDD-related blue-sac disease as a primary cause of the failure to reestablish natural
reproduction of lake trout in Lake Ontario. The egg TCDD residue range for this effect
is 30 to 100 pg/g wet tissue (0-100% blue-sac associated mortality of sac fry prior to
swim-up), TCDD analysis of 1987 Lake Ontario lake trout eggs indicated 7 to 16 pg/g
wet weight (Cook et al., 1993a). The addition of toxic contributions from other PCDD,
PCDF and PCB congeners is estimated to give a TCDD toxicity equivalency
concentration (TEC) in eggs of 14 to 32 pg/g wet weight. This is consistent with
observation of some natural reproduction in 1986 (Marsden et al., 1988). Only a
doubling of the TEC would be predicted to result in up to 50% mortality from blue-sac
disease without consideration of environmental factors, chemical synergism or
additional toxic chemical contributions that may increase the response. Older lake
trout in Lake Ontario have up to three-fold greater whole body TCDD residues and
probably have egg TECs of 42 to 96 pg/g wet weight, and therefore may be ineffective
spawners.
Lake Ontario lake trout collected in 1978 had average TCDD concentrations of
78 pg/g wet weight whole fish and therefore could have produced eggs with sufficient
TCDD and related chemical residues to cause the average 48% blue-sac disease
incidence observed for Lake Ontario lake trout eggs collected in 1979 (Symula et al.,
1990). During the period of 1978-1988 Lake Ontario lake trout TCDD residues
decreased proportionally to decreases in TCDD concentrations in Lake Ontario surface
sediment indicated by analysis of radionuclide-dated 1 cm increments of sediment
cores from depositional basins of the lake (Cook et al., 1993a). The core analysis
indicates that the greatest sediment TCDD contamination occurred around 1962 at
approximately eight times the 1987 level. Since lake trout TCDD residues appear to
be declining at least as fast as surface sediment TCDD levels and bioaccumulation of
TCDD by lake trout relative to sediment contamination would be greater when peak
TCDD loadings occurred, lake trout egg residues in 1960 (if lake trout had been
stocked then) would have probably exceeded 70 to 160 pg/g wet tissue. This level of
exposure to TCDD alone would be sufficient to cause complete sac fry mortality due
to the blue-sac syndrome. Since other Ah-active PCDDs, PCDFs and co-planar PCBs
appear to have decreased less in fish and sediments since 1960, the TEC was
probably more than twice the TCDD concentration in lake trout eggs during the 1960s
and 1970s. Thus reproductive failure of lake trout during the 1970s could well have
occurred due to TCDD and related chemical bioaccumulation by lake trout.
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Because of (1) the strong association of TCDD with blue-sac disease observed
in laboratory exposures and the effect in sac-fry from eggs collected from Lake
Ontario lake trout; (2) the historical record of Lake Ontario TCDD and related chemical
exposure to lake trout; and (3) the consistency of the predicted toxicity with the field
sampling record of no natural reproduction until 1986, the probability is high that
TCDD toxicity alone could have prevented lake trout reproduction during the post-1950
period. Attainment of a self-sustaining population of lake trout in Lake Ontario must
still depend on further reduction in TCDD and related chemicals, control of the sea
lamprey, maintenance of adequate spawning beds and water quality, and introduction
of appropriate lake trout strains for Lake Ontario conditions and food web. More
information is needed on the effect of TCDD on other fish populations such as the
deep water sculpin before similar conclusions for rehabilitation can be made.
4.2.3 Effects Profile
4.2.3.1 Use of Bioaccumulation to Characterize Risk
A major issue in using laboratory information to characterize risk to organisms
in natural systems is extrapolating effects information among different exposure
conditions. This is especially true for highly bioaccumulative chemicals such as
TCDD. The information for TCDD effects on aquatic organisms presented above
exemplifies a number of important considerations. Various routes of exposure were
used, Including waterborne, i.p. injection and diet. For waterborne exposures, the
duration of exposure varied from six hours to several weeks. Because TCDD
accumulates slowly, the exposure concentrations needed to elicit effects change
greatly over this range of durations. Among those studies using exposure via water,
bioavailability probably varied due to the effects of different amounts and types of
solvent carriers and natural organic matter in the test systems. Buildup of organic
matter would be of particular concern for static exposures, which also would have
exhibited declining TCDD exposure concentrations with time. Finally, because of
delays in responses to a toxic TCDD dose, it is sometimes unclear to what magnitude
and duration of exposure an organism is actually responding.
As explained in earlier sections, such difficulties in relating effects among
different exposure conditions can be partially addressed by using toxicant
accumulation as a reference point for effects. Because exposure duration and
bioavailability are manifested in terms of how much chemical is accumulated,
expressing toxicity on the basis of accumulation should largely adjust for the effects of
these factors. In particular, this should better account for the fact that stopping an
exposure has little immediate impact on response, since the accumulated TCDD
concentration at the site of action persists for a long time due to this chemical's slow
elimination. Expressing effects on the basis of accumulation also allows for
waterborne, injection and dietary exposures to be directly compared and for better
interpretation of field data on accumulation by organisms.
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Expressing effects on the basis of accumulation does not solve all difficulties in
extrapolating among exposures. Under typical laboratory exposures, accumulation will
change with time. To the extent that effects are delayed or are cumulative, there can
be some uncertainty regarding the accumulation associated with a particular level of
effect. Also, the relationship between toxicity and accumulation is not well established
and is not necessarily simple. This is particularly true of accumulation on a whole
body basis, which would not reflect the different internal distribution pathways
associated with different routes of exposure and rates of accumulation, and how this
might alter response. Nevertheless, expressing toxicity on the basis of accumulation
will reduce the uncertainty in comparing and applying toxicity information and will be
the basis for the discussion of risk here.
Expressing effects on the basis of accumulation does not eliminate the need to
address issues of bioavailability and kinetics of accumulation. Rather, it simply
removes the confounding influence of these factors from the interpretation of effects
data. Bioaccumulation relationships then must be explicitly considered in relating
accumulation-based effects assessments to environmental concentrations. This
integration of exposure, bioaccumulation, and effects will be considered in a later
section on risk characterization (see section 5). The remainder of this section will
provide a summary of effects to aquatic life, emphasizing accumulation as the basis
for describing effects.
4.2.3.2 Relationship of Risk To TCDD Accumulation
Effects information presented earlier (section 4.2.1) indicates that fish are
substantially more sensitive to TCDD than are plants, aquatic invertebrates and
amphibians. Fish populations will therefore be the focus here of the effects profile for
aquatic life. Available data also indicate that early life stages of fish are substantially
more sensitive than older fish. Therefore, this section will summarize TCDD
accumulations associated with (1) no evidence for risk to fish populations, (2)
impairment of early life stage fish development and (3) effects on growth and survival
in juvenile and older fish.
Although no study has evaluated the effects of TCDD on the entire reproductive
process in fish, several studies discussed in section 4.2.1 examined the effects of
TCDD exposure of fish eggs and have shown fry survival and development to be an
important, sensitive endpoint. Wannemacher et al. (T992) also examined effects on
oogenesis by exposing actively spawning female zebrafish to a single oral dose.
Effects on oogenesis were only evident at doses which also affected embryo survival.
Thus, there is as yet no evidence for endpoints significantly more sensitive than early
life stage development, but definitive studies with chronic exposures are still needed.
For lake trout, comparisons among three routes of TCDD exposure to eggs
(waterborne, injection, and maternal transfer) have shown 50% mortality of fry to be
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associated with similar egg residues, ranging from 47 to 65 pg TCDD/g wet weight
(Walker, 1991; Walker et al., 1991, 1992, 1993). For the waterborne exposure, a
residue of 34 pg/g did not exhibit significant effects relative to controls, suggesting that
it is a level associated with no or limited risk. However, concentrations ;> 40 pg/g
caused significant mortality indicating that the dose/response curve is very steep, with
the response ranging from nondetectable to complete mortality over a threefold range
of accumulation. For rainbow trout, the median lethal accumulation was somewhat
higher, ranging from 230 to 490 pg TCDD/g egg depending on the strain of trout
(Walker and Peterson, 1991; Walker et al., 1992). Results again were similar for
waterborne exposure .(for 48 hour post fertilization) and injection of eggs.
Waterborne exposure of medaka eggs and fry to TCDD from fertilization
through 3 days post hatch (approximately two weeks) resulted in LG50s for fry of 13
ng/L (Wisk and Cooper, 1990a) and 9 ng/L (Wisk and Cooper, 1990b). Based on
measurements of TCDD in 11 day old embryos after similar exposures, accumulation
on a whole egg basis was approximately 100 times the water concentration (Wisk and
Cooper, 19iOb). The median lethal accumulation therefore would be expected to be
about 1,000 pg/g. However, this likely is an overestimate of the accumulation needed
to elicit this effect, since accumulation was probably increasing throughout the test
period. This is in contrast to the salmonid studies cited above, in which water
exposure continued for only 48 hours after fertilization and TCDD concentrations in
eggs and fry were constant after that until fry began feeding. For the medaka, the
observed effects would have depended in large part on lower accumulations earlier in
the developmental period, especially since Wisk and Cooper (1990b) did demonstrate
that the sensitivity of newly hatched fry depended on exposure in the first few days
after fertilization. Thus, fry mortality in this species would occur at lower egg
accumulations, probably within the range of the salmonid studies discussed above.
Helder (1980; 1982a,b) exposed rainbow trout and northern pike eggs for 96
hours to TCDD in water, but did not report any data on accumulation. The rainbow
trout showed a slight increase in fry mortality at 1 ng/L, up to about 25% mortality at
10 ng/L, and complete mortality at 100 ng/L. These results are in the range of, or
slightly more sensitive than, those of Walker et al. (1992) for rainbow trout (e.g., the
48-hour LC50 of about 80 ng/L was estimated from the results for the present report).
The northern pike showed reduced fry survival at 0.1 ng/L, 50% mortality at 1 ng/L,
and almost no survival at 10 ng/L. This suggests even greater sensitivity than for lake
trout (for which a 48-hour LC50 of about 25 ng/L was estimated from the results of
Walker et al. (1991)). However, it is quite possible that pike eggs absorb TCDD faster
than the salmonid eggs, as do medaka, so whether these results indicate greater
sensitivity is unclear. Even if the northern pike eggs bioconcentrated as much TCDD
in their 4 day exposure as medaka eggs do in 11 days, the median lethal
accumulation would be only 100 pg/g. Therefore, this study still provides evidence
that the early life stage of another fish is nearly as sensitive, if not more so, than that
of lake trout. Miller et al. (1979) also demonstrated adverse effects (fin necrosis) on
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young gupples from a 24-hour exposure at water concentrations as low as 0.1 ng/L.
Helder (1982a) reported delays in development when rainbow trout eggs were
exposed to 0.1 ng/L and 20% mortality when newly hatched fry were exposed to 1
ng/L These studies provide direct evidence that even acute waterbome exposures of
fish eggs to TCDD would need to be limited to much less than 1 ng/L to prevent
adverse effects to early life stages of fish.
Wannemacher et al. (1992) reported no effects on oogenesis and fry
development when female zebrafish received a single, nominal oral dose of 1,670
pg/g, whereas a dose of 8,300 pg/g severely reduced oogenesis and caused severe
malformations and complete mortality in those fry that did hatch. Based on separate
experiments that documented accumulation in the adult fish, the concentration of
TCDD 15 days post dose would be about 670 pg/g in fish exposed to the no effect
dose. If egg concentrations were similar or moderately lower, this species may be
more resistant than the salmonids discussed above, but should still show substantial
early life stage effects for egg concentrations <1,000 pg/g.
The accumulation of TCDD in eggs should largely reflect maternal transfer,
except in the unlikely situation that eggs are laid where they have high exposures
while the parents have negligible exposure. Effects based on the concentration of
TCDD in eggs must therefore be related to concentrations in the parents, to in turn
relate these concentrations to environmental concentrations. In laboratory tests with
maternal transfer, eggs were determined to have about 40% of the TCDD
concentration (on a wet weight basis) as parent fish. For fish collected from Lake
Ontario, this percentage was about 30%. However, the Lake Ontario data were for
adult fish with 18% Hpid whereas the eggs had about 8% lipid (In the laboratory, adult
fish and eggs both had about 8% lipid). On a Hpid normalized basis, the field data
indicated the TCDD concentration of eggs to be about 65% of that in the adults. The
fact that this percentage is greater than that found in the laboratory probably reflects
continuous exposure over a longer time. For the purpose of the present analysis, this
latter factor will be used to relate TCDD concentrations in eggs to that in adults with
similar lipid contents. Using this value, the NOAEL of 34 pg TCDD/g for eggs
corresponds to about 50 pg TCDD/g in the parent fish. The median lethal
accumulation for lake trout eggs reported above would correspond to about 75 pg ,
TCDD/g in parent fish. If it is assumed that similar parental/egg relationships exist for
rainbow trout as for lake trout, the egg concentrations resulting in 50% rainbow trout
fry mortality would correspond to about 350 to 750 pg TCDD/g in parent fish.
The growth and survival of older fish appear to be less sensitive to TCDD than
that of early life stages. As summarized in Table 4-1, significant effects do not appear
to occur at accumulations less than 500 pg TCDD/g wet weight for any of the studies.
Median lethal accumulations range from about 1,000 to 15,000 pg TCDD/g, depending
on species, lifestage, and route of exposure. Effects on growth are generally evident
from 1,000 to 3,000 pg/g.
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Based on the above information, the effects and associated uncertainties of
TCDD on aquatic life are summarized in Box 1.
Boxl. Effects Prof He for Aquatic Life
Ť/'"-'" "'''=* -" = ' ' ; r, ť ,'v ', *-
*' ; * ' ' .. '
The most sensitive effects ihtlsW estairfiihed for aquatic organisms are in fish fry
exposure of eggs before or shortly after fertilization. There is no definite evidence of
effects in any of th6 ftsh species, iesled II SecurmJlafiorj in eggs Is less than 34 pg TGDD/g,. the
highest no observed effect level for lake trotrtfry, This likely torrasponds to an accumulation
fn parent fish, with llpld content to the Ť0gst of test than SO pg fC0f3/g, These values
do not Incorporate any uncertainty factors'that address Issues such as greater sensitivity of
untested fish species and andpaintSi the relalkmshlj) of "TCDD in egg$4o parents or the effects
of llptd content* Fuithermorit.there Is no reason at thfe point to sŤppp'se thatjhls'number Is
low simply because lake trout are exc-eptlonallj sensitive, 'Tests on other species', both Ťarfy
life stage and older fish, 'suggest thafsome might he" as, or more, sensitive than lake trout.
Because of the importance of this endpojrtt ancj the severity of response at exposures not
much higher than this exposure, this estimate for tow risk srtoufci'hof be considered
conservative or having. Bmjteof applicability. There is also the possibility of the ekistence of' -'
more sensrfiVa endpofnts that have not yettiian'tested/ ' ,- >
'"^^ ^ 5 and transfer fo the eggs via ,
maternal transfer). In other Words', proteclfert"of early life stages would require restricting
accumulation by parent fish In all circumstances, not just at the time 'and ptaee for ^ - / '
reproduction* ' vv" "'" '!'"''',' ' ' T ,
Available laboratory toxicity Information Indlcat&Jhaf aquatic lnvertebr^esf plants and
amphibians are substantially less sensitive to TCDD than ftshs% However, th'e database
regarding this sensitivity is limited and the'exposures are not always readily comparable to the
data on fish. It is possible that" aquatic ecosystem components other than fish are sensitive to
TODD but simply hav& not been adequately tested ' -
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4.3 EFFECTS OF TCDD ON AQUATIC-ASSOCIATED WILDLIFE
4.3,1 lexicological Information
Wildlife toxicology literature on TCDD was obtained, in part, through searches
on a base set of mammals and birds listed in Table 4-2. Because very few bird
studies were available, the literature search was expanded to include domestic fowl,
such as the chicken, and upland game birds, such as the ring-necked pheasant
(Phasianus colchicus) and bobwhite quail (Colinus virginianus). Attempts to obtain
TCDD toxicology literature for marine mammals and reptiles were unsuccessful.
Table 4-2. The base set of mammals and birds included in the literature search
for TCDD wildlife toxicity data.
Common Name
Beaver
Fisher
Marten
Mink
Muskrat
Racoon
River Otter
Black Duck
Mallard
Northern Pintail
Red-Breasted Merganser
Common Merganser
Canada Goose
Common Loon
Double-Crested Cormorant
Great Blue Heron
Herring Gull
Ring-Billed Gull
Caspian Tem
Common Tem
Forster's Tem
Brown Pelican
American White Pelican
Red-Winged Blackbird
Bald Eagle
Peregrine Falcon
Osprey
Mammals
Birds
Scientific Name
Castor canadensis
Mattes pennanti
Martes am&ricanus
Mustela wson
Ondatra zlbethlcus
Procyon lotor
Lutra canadensis
Anas rvbrlpes
Anas platyrhynchos
Anas acuta
Mergus senator
Mergus merganser
Branta canadensis
Gavia immer
Phalacivcorax auritus
Ardea herodlas
Larus argentatus
Lams delawarensis
Sterna caspla
Sterna hlrunda
Sterna forsteri
Pelecanus occidentalls
Pelecanus erythrorhynchos
Agelalus phoenlceus
Haliaeetus leucocephalus
Falco peregrinus
Pandion hallaetus
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Mammalian and avian wildlife toxicity data were collected without regard to
study duration, number of doses, or route of exposure. In general, however, only
those studies that incorporated exposure durations of sufficient length to establish
dose-response curves for reproductive and developmental effects were summarized
for this report.
4.3.1.1 Mammals
No reproduction and developmental studies with mammalian wildlife were found
in the literature; however, studies by Aulerich et al. (1988) and Hochstein et al. (1988)
provided information on the toxicity of TCDD to mink following exposures of one to
twelve days.
Hochstein et al. (1988) administered TCDD as a single oral dose (0, 2,500,
5,000, and 7,500 pg/g body weight) to adult male mink and observed the animals for
28 days. No animals died at the lowest two doses; however, between 14 and 17 days
post-exposure, 100 and 75% mortality was observed at doses of 7,500 and 5,000
pg/g, respectively. A 28-day LD50 of 4,200 pg/g was calculated. Food consumption,
body weight, and adipose tissue were significantly reduced in mink receiving doses of
5,000 and 7,500 pg/g. Food consumption in mink receiving TCDD at the low dose of
2,500 pg/g was depressed by approximately 25 to 40% during the first two weeks
post-exposure (not statistically significant), but increased to control levels after three
weeks. Body weights were significantly (p < 0,05) depressed 11.4% in mink exposed
to TCDD at the lowest dose. In mink exposed to the two higher TCDD doses, gross
necropsy revealed mottling and discoloration of the liver, spleen, and kidneys and
enlarged brain, kidneys, heart, and thyroid and adrenal glands. Animals that survived
TCDD exposure at the lowest dose showed no alterations in hematological and thyroid
hormone measurements.
Aulerich et al. (1988) administered TCDD at doses of 100 and 1,000 pg/g by
i,p. injection to newborn mink for 12 consecutive days. Newborn mink exposed to
TCDD at a dose of 1,000 pg/g died within 14 days, whereas mortality reached 62% by
19 weeks in mink exposed to 100 pg/g. Body weight gains were initially reduced in
mink exposed to the low dose; however, by 19 weeks of age there was no significant
difference in body weights between survivors and control animals. No discernible
effects on the time of eyelid opening, the time of tooth eruption, and hair growth were
attributed to TCDD exposure in the survivors.
Studies by Hochstein et al. (1988) and Aulerich et al. (1988) indicated that the
mink Is among the mammalian species most sensitive to TCDD intoxication. Based
on a 28-day LD50 of 4,200 pg/g (Hochstein et al., 1988), the mink seems to be less
sensitive than the guinea pig, for which LD50s of 600 to 2,000 pg/g have been
reported, but more sensitive than the rat (LDSOs of 22,000 to 45,000 pg/g), rabbit
(LD50 of 115,000 pg/g), mouse (LD50s of 114,000 to 284,000 pg/g), and hamster
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(LDSOs of 1,157,000 to 5,000,000 pg/g) (see reviews of Kociba 1982a,b; Schwetz et
al., 1973).
4.3.1.2 Birds
Several studies have been undertaken to determine the iethal potency of TCDD
to avian wildlife. Hudson et al. (1984) reported 37-day LDSOs of 15,000, > 108,000
and >810,000 pg/g for the bobwhite quail, mallard, and ringed turtle dove (Stretopelia
risoria), respectively, following a single oral administration of TCDD. In comparison,
Grieg et al. (1973) reported that chickens given single oral doses of TCDD at 25,000
to 50,000 pg/g died within 12 to 21 days post treatment. In a longer study Schwetz et
al. (1973) orally administered TCDD to 3-day-old white leghorn chickens for 21 days at
doses of 0, 10, 100, 1,000 and 10,000 pg/g/day and reported a NOAEL for mortality of
100 pg/g/day. Nosek et al. (1992a) treated ring-necked pheasant i.p. with single
doses of TCDD at 0, 6,250, 25,000 and 100,000 pg/g and observed the animals for 11
weeks post treatment AH the birds treated at the high dose of 100,000 pg/g died
within six weeks of exposure. Onset of mortality in birds exposed to 25,000 pg/g was
observed six weeks post treatment and at 11 weeks post exposure, 80% mortality was
observed. No birds died in the 6,250 pg/g exposure group during the course of the
study. Common to all of the above mentioned studies, a dose-dependent decrease in
food consumption and body weight were reported to precede death.
Nosek et al. (1992a) also investigated the effects of TCDD on reproduction in
ring-necked pheasant. Hens were administered TCDD by i.p. injection once a week at
doses of 0, 10, 100 or 1,000 pg/g for 10 weeks. During the final two weeks of
exposure, hens were paired with roosters and kept in egg production for an additional
9 to 13 weeks. No deaths were observed in the control group or with hens exposed
to TCDD at the lowest two doses; however, a 57% mortality rate occurred in the group
administered TCDD at 1,000 pg/g/week. A significant decrease (p < 0.05) in adult
body weight and egg production, compared to controls, was also associated with the
highest dose, but not with the lowest two doses. For these two endpoints there was
no indication of a dose-response relationship between 0, 10 and 100 pg/g/week.
There was a trend towards increasing embryo mortality rates with increasing dose to
the hens. At the highest dose, 100 ą 2% of the embryos died, whereas mortality rates
of 15 + 5 and 25 + 15% were reported for hens dosed at 10 and 100 pg/g/week,
respectively (percent mortality corrected for lethality in control eggs; values cited
based on interpretation of Figure 6 in Nosek et al., 1992a). A dose of 450 pg/g/week
was calculated to elicit a 50% increase in embryo mortality above the control rate.
The embryo mortality rates at 10 and 100 pg/g/week were, however, not significantly
different nor were they different from control mortality (p < 0.05).
Several studies have also been undertaken to assess the toxicity of TCDD to
birds following In ovo treatment; however, there were only a limited number of studies
available with wildlife species. Nosek et al. (1992c) injected fertile ring-necked
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pheasant eggs with graded TCDD doses of 0.01 to 100,000 pg/g. Injections were
administered into the albumin (0.01 through 100,000 pg/g) in one experiment and into
the yolk (10 to 10,000 pg/g) in another. Mortality was assessed in chicks through 28
days post hatch. Embryo mortality rates of 50 and 37.5% were reported for control
eggs injected with 1,4-dioxane, which served as a carrier, in the yolk and albumin,
respectively. Nosek et al. (1992c) reported that mortality in the range of 30 to 50% is
within the historical range for embryo mortality with uninjected fertile eggs obtained
from the hatchery used in the study. A dose dependent increase in embryo mortality
was observed following both albumin and yolk injections and LD50s of 1,400 (99 -
3,700 pg/g; 95% C.I.) and 2,100 (522 - 4,393 pg/g; 95% C.I.) pg/g, respectively, were
reported. Based on injections into the yolk, the LOAEL for mortality was 10,000 pg/g,
and the NOAEL was 1,000 pg/g (p < 0.05). At 10,000 pg/g, 97.5% mortality was
observed, whereas at 1,000 pg/g a mortality rate of 60% was noted. At 10 and 100
pg/g, 52.5 and 47.5% mortality were recorded, respectively. Results from albumin
injection studies indicated a LOAEL of 1,000 pg/g (57.7% mortality) and a NOAEL of
100 pg/g (38.7% mortality). Mortality rates of 35.0, 30.0, 32.5, and 31.2% were
recorded for doses of 0.01, 0.1, 1, and 10 pg/g, respectively. At doses of both 10,000
and 100,000 pg/g, 97.5% mortality was observed. Martin et al. (1989), as cited by
Nosek et al. (1992c), reported 100% embryo mortality in Eastern bluebirds (Sialia
stalls) following injection into the albumin of TCDD at a dose of 10,000 pg/g and no
mortality at 1,000 pg/g.
The chicken may be more sensitive than the pheasant or bluebird. A LD50 of
240 pg/g has been reported by Alfred and Strange (1977) following injection into the
airspace of fertile eggs. Cheung et al. (1981) reported up to approximately 30%
mortality in chicken eggs injected at 0.05 to 450 pg/g (assuming a 55 g egg) in the
albumin; however, a significant linear log dose-response relationship was not observed
(p > 0.05; approximately 20% mortality in control eggs).
In the same egg injection study outlined previously, Nosek et al. (1992c) also
assessed growth, histological effects and immune response in pheasant chicks
through 28 days post hatch. Injections of TCDD into the albumin at 1, 10, 100, or
1,000 pg/g had no effect on growth, nor on carcass morphometrics, cardiac
morphometrics, absolute organ weights, or relative organ weights in 1-day old
hatchlings and 28-day old chicks compared to chicks exposed in ovo to the vehicle.
There was also no significant TCDD effect on histology of the liver, spleen, heart,
Bursa of Fabricius, or thymus. No TCDD related increase in the incidence of ascites
or subcutaneous, pleural, or pericardia! edema was observed. Immune response, as
monitored by serum liters of IgG, IgM, and total antibody in 28-day old chicks injected
with washed sheep erythrocytes, was not significantly affected by TCDD exposure.
At non-lethal, in ovo, doses both Nosek et al. (1992c) and Martin et al. (1989),
as cited by Nosek et al. (1992c), did not report any evidence of TCDD related effects
on chick growth, histopathological abnormalities, or the incidence of edema, ascites, or
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hydropericardium in pheasants or Eastern bluebirds. The role of TCDD and related
PCDDs, PCDFs, and PCBs in producing a syndrome of edema and histological
alterations in chickens has been reviewed extensively (e.g., see review of Giibertson
et al., 1991). For example, Sawyer et al. (1986), Flick et al. (1972) and Cheung et al.
(1981) reported edema, involution of the Bursa of Fabricius, and cardiovascular
malformations in chickens following i.p., dietary or in ovo exposure, respectively, to
TCDD. Cheung et al. (1981) reported concentrations of approximately 5.8 pg/g egg
(assuming a 55 g egg) associated with the combined occurrence of four
cardiovascular malformations (combined occurrences of ventricular septal defect,
aortic arch anomaly, aortic arch anomaly plus ventricular septal defect, and
conotruncal malformation) in 50% of chicken embryos. There was approximately a
30% incidence rate for combined malformations in control groups, for which a similar
rate in uninjected, sham-injected, and vehicle injected eggs was observed. There
were, however, no significant relationships between TCDD exposure and the
percentage of embryos having cardiac malformations when each of the four defects
were analyzed separately, rather than combined. These differences between the
chicken, pheasant and bluebird might reflect variations in routes of exposure and/or
species sensitivity.
Only limited studies are available on the toxicokinetics of TCDD in avian
species. Nosek et al. (1992b) reported that the bioavailability of TCDD was 30, 33,
41, and 58%, respectively, when adult pheasants were orally administered
suspensions of treated earthworm, soil, paper mill sludge,, and cricket homogenates.
Martin et al. (1989), as cited by Nosek et al. (1992b), found that the oral bioavailability
of TCDD to European starlings (Sturnus vulgaris) was 14, 17, 37, and 44% from
suspensions of earthworms, paper mill sludge, soft-bodied invertebrates, and hard-
bodied invertebrates, respectively. These results were consistent with the longer
gastrointestinal retention time for hard-bodied insects (Nosek et al., 1992b).
Elimination studies of [3H]TCDD by adult non-egg producing ring-necked
pheasant (Nosek et al., 1992b) and European starlings (Martin et al., 1989, as cited by
Nosek, et al., 1992b) provided whole-body half-lives (total TGDD on a whole-body wet-
weight basis) of 380 and 7.2 days, respectively. Nosek et al. (1992b) also derived a
half-life of 13 days for pheasant chicks exposed in ovo. The differences in half-lives
between adult pheasants and pheasant chicks and starlings were attributable, in part,
to differences in mass-specific metabolic rates. Differences in species- or age-specific
TCDD metabolism was also thought to contribute to the variability in elimination. In
adult pheasant carcass samples, TCDD was detected but no metabolites of TCDD
were recovered, indicating little or no biotransformation in adult hens. However, fecal
material and chick carcass samples were not analyzed for TCDD or metabolites, thus
it is not possible to assess the role of age-specific differences in metabolism to
differences in TCDD elimination. Nosek et al. (1992b) also suggested that declining
lipid content in developing chicks, as the yolk lipids are depleted, coupled with the
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developing gastrointestinal tract and high lipid content of chick starter feed could result
in a redistribution of TCDD leading to an increased elimination rate.
Nosek et al. (1992b) also assessed the elimination of TCDD in egg-producing
pheasant hens. Birds were exposed by i.p. injection to [3H]TCDD at a rate of 100 pg/g
once a week for 10 weeks and it was reported that a mean of 1.1% of the cumulative
TCDD dose was eliminated in each of the first 15 eggs laid by a hen. The tritium in
pheasant eggs was exclusively associated with the parent compound in the yolk.
Martin et al. (1989), as cited by Nosek et al. (1992b), estimated that approximately 5%
of a TCDD body burden in an Eastern bluebird would be eliminated in each egg.
Based on an average of 30.5 eggs laid per pheasant hen over a 7-week post
treatment period, it was estimated that approximately 35% of the total administered
dose was eliminated in the eggs (Nosek et al., 1992b). At the time of sacrifice, an
average of 30% of the administered TCDD remained in the hens. Although not
actually measured, it was speculated that the remaining 35% of the cumulative TCDD
dose was excreted. The greatly increased rate of TCDD elimination in laying hens, as
opposed to non-laying hens, suggested that egg laying is an important route of
elimination and that changes in metabolism and/or lipid distribution in the adult bird
during egg development and reproduction might be responsible for increased fecal
elimination (Nosek et al., 1992b).
Consideration of the pheasant toxicokinetic information with the results of the
toxicity studies provides some limited insights into the relationship between female
exposure, egg residue levels and embryotoxic effects of TCDD. Based on in ovo
exposures, Nosek et al. (1992c) reported that NOAELs for embryo mortality ranged
from 100 to 1,000 pg/g egg. In an extended reproduction study with TCDD exposure
to the hens, a NOAEL for embryotoxic effects was associated with a total accumulated
body burden of 1,000,000 pg in a 1.0 kg hen, whereas a LOAEL was associated with
a 10,000,000 pg body burden (Nosek et al., 1992a). Results from the toxicokinetic
studies (Nosek et al. 1992b) suggest that approximately 1.0% of a hen's body burden
is translocated to each egg laid. Therefore, assuming a pheasant egg weight of 30 g
(Nosek et al., 1992b), a cumulative NOAEL body burden of 1,000,000 pg TCDD and a
LOAEL of 10,000,000 pg TCDD in a hen would correspond to NOAEL and LOAEL
TCDD egg concentrations of approximately 300 and 3,000 pg/g, respectively. These
derived pheasant egg effect levels are within the range of those levels obtained
empirically from in ovo pheasant studies. For several reasons, comparisons between
maternal and in ovo exposures must certainly be viewed with caution before any
definitive conclusions can be made on the defensibility of solely using in ovo
exposures to assess the impact of TCDD on avian reproduction. First, the available
dose-response curves are based on regimes that incorporate 10-fold exposure
Increments, which makes quantitative comparisons uncertain. Secondly, in ovo
exposures fail to incorporate any possible adverse effects that may result from TCDD-
mediated alterations in the adult female or male reproductive physiology and
endocrinology.
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4.3.2 Epidemiological Information
As was discussed in the context of fish populations (see section 4.2.2), the
criteria discussed by Fox (1991), which include: (a) probability, (b) chronological
relationship, (c) strength of association, (d) specificity, (e) consistency of association
upon replication, (f) predictive performance, and (g) coherence (i.e., biological
plausability, presence of a dose-response relationship), should be considered when
assessing "ecoepidemiological" studies for wildlife populations. The elucidation of
plausible cause and effect relationships, given the variety of chemical and
nonchemical stressors that can impact populations, is obviously a fundamental
challenge in interpreting the results of wildlife studies. Certainly the validity of
epidemiologically-based associations increase when they can be supported by
independent and controlled experiments.
Using the criteria of Fox (1991), by far the most convincing case linking a
specific suite of adverse biological effects (embryonic and chick mortality, edema,
growth retardation, deformities) to residue concentrations of TCDD-like chemicals can
be made for fish-eating birds (herring gulls, Forster's terns, double-crested cormorants,
Caspian terns) from the Great Lakes (Gilbertson et a!., 1991 and references cited
therein; Tillit et al., 1992). One confounding factor in the interpretation of studies with
populations of Great Lakes birds is the co-occurrence of a number of other potentially
toxic contaminants (e.g., DDT, dieldrin, mercury) with TCDD and toxicologically related
PCDDs, PCDFs, and RGBs. However, because of reasonably consistent exposure
profiles, studies with herring gulls in Lake Ontario (Mineau et al., 1984; Environment
Canada, 1991a,b) and Great Blue herons in British Columbia (Bellward et al., 1990;
Hart et al., 1991) provide some data relating reproductive success to egg
concentrations of TCDD itself.
The productivity of herring gull colonies at several islands in Lake Ontario was
monitored from 1972 through 1984, as summarized by Mineau et al. (1984) and
Environment Canada (1991a). From 1971 through 1975, productivity, defined as the
number of young reaching 21 days of age per nesting adult, ranged from 0.06 to 0.21,
which was well below the range of 0.8 to 1.0 that is required for population stability.
However, in 1977 productivity values exceeded 0.8 and through 1984 (the last year of
reported data), ranged from 0.86 to 2.13. The population of gulls increased from 520
pairs in ten colonies in 1976 to 1,540 pairs in 15 colonies in 1987. During the period
of poor reproductive performance, a suite of adverse effects that are commonly
associated with TCDD and toxicologically-related compounds (Gilbertson et al., 1991),
was reported. However, egg shell thinning of only 4 to 8% was noted in this DDE-
resistant species, which strongly suggested that total DDT exposure was not
associated with the low productivity (Environment Canada, 1991 a).
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During the period of time that productivity was assessed at gull colonies
associated with Snake, Muggs and Scotch Bonnet Islands in Lake Ontario, TCDD
levels were also monitored in gull eggs sampled from these sites (Environment
Canada, 1991b). In 1971 and 1972, TCDD levels of approximately 2,000 to 2,400
pg/g in gull eggs were reported for Scotch Bonnet Islands (1972 productivity of 0.12);
In 1974, levels were down to approximately 900 pg/g (1973 and 1975 productivities of
0.06 and 0.15) and in 1977 and 1978, were at 500 pg/g (productivities of 1.10 and
1.01). In 1982, levels were down to 204 pg/g (1981 productivity of 2.13).
Concentrations of TCDD in gull eggs from Snake Island dropped from approximately
175 pg/g in 1981 (productivity of 1.73) to 90 pg/g in 1989. At Muggs Island, TCDD
concentrations in gull eggs have remained relatively constant at approximately 25 to
50 pg/g from 1984 through 1989 (1981 and 1984 productivities of 1.40 and 1.17). The
reduction in TCDD levels in gull eggs is consistent with discontinued TCDD inputs
from waste sites and chemical manufacturing plants and declining concentrations in
water, sediments and fish (Hallet and Brooksbank, 1986; Cook et al., 1993b;
Environment Canada, 1991b). It should also be noted that during the period of 1974
to 1977/78, when productivity sharply improved, total PCB, mirex, hexachlorobenzene
and dieldrin concentrations in gull eggs also dropped from approximately 100,000,000
to 35,000,000; 7,000,000 to 1,000,000; 500,000 to 100,000; and 500,000 to 100,000
pg/g, respectively (Environment Canada, 1991b).
In the studies by Bellward et al. (1990) and Hart et al. (1991), survival and
growth of Great Blue heron chicks were monitored in colonies from three sites
(Nicomekl, Vancouver and Crofton) in British Columbia which varied in PCDF and
PCDD concentrations. The predominant congeners present in Great Blue heron eggs
from these sites were TCDD, 1,2,3,7,8-pentaehlorodibenzo-p-dioxin and 1,2,3,6,7,8-
hexachlorodibenzo-p-dioxin. However, based on mammalian TEF models (Safe,
1990), the most toxic of the congeners was TCDD. Mean TCDD concentrations (ą
SEM) in eggs from the Nicomekl, Vancouver, and Crofton sites were 10 ą 0.9, 135 ą
49.6 and 211 ą 33.7 pg/g, respectively. Hatching success did not differ significantly
across sites but there was an apparent increase in the incidence of edema in chicks
with increasing TCDD concentration (0 of 11, 2 of 13, and 4 of 12, from the Nicomekl,
Vancouver and Crofton colonies, respectively). In addition, there was an inverse
correlation between TCDD egg concentrations and various growth measurements
made on the heron chicks. In a more recent study, preliminary results suggest there
may also be an inverse association between TCDD egg concentrations and
morphometric changes in the brains of these hatchlings (Henshel et al., 1992).
The herring gull and Great Blue heron data indicate that successful
reproduction exists in wild colonies (i.e., maintenance of a stable population), even
though TCDD is present at concentrations in eggs between 200 to 500 pg/g. In turn,
toxic effects at the individual level are associated with concentrations above 100 pg/g.
These data are consistent with the results of laboratory toxicity studies for birds
described previously (see section 4.3.1.2), which indicated that a threshold for embryo
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mortality would be expected in the range of 100 to 500 pg/g. The causal connection
between TCDD and the observed field effects is, however, by no means proven,
especially since a number of other contaminants were also present in the egg
samples. Further, the generalizations concerning adverse effect levels of TCDD in
eggs, drawn from the field studies, should be considered conservative because PCBs,
PCDDs, PCDFs, and, in the case of the herring gulls in Lake Ontario,
hexachlorobenzene, dieldrin and mirex, may also have contributed to the observed
responses. As a consequence, it is impossible to quantitatively extrapolate these
results to a more definitive TCDD-specific threshold value. These results clearly
illustrate the need for validated methods to assess TCDD equivalents in avian models
to establish defensible approaches for evaluating mixtures of PCBs, PCDFs and
PCDDs, as well as the need to develop techniques to assess chemical mixtures in
general.
4.3.3 Effects Profile
4.3.3.1 Approach for Summarizing Effects on Wildlife
In cases where studies have related TCDD effects to TCDD accumulation in
wildlife, the approach for summarizing effects will be the same here as that used
earlier for fish. However, much of the toxicity information for wildlife presented
previously, especially for sensitive endpoints, cannot be expressed on this basis. This
is due to the lack of wildlife accumulation measurements and because there is
uncertainty about how the route and duration of exposure would affect accumulation.
Rather wildlife-effect information Is more conveniently expressed on the basis of TCDD
dose via oral consumption. This dose metric has the added advantage of being able
to relate effects on both wildlife and fish to TCDD accumulation in fish.
The approach used here to express wildlife effects profiles is similar to that in
the proposed Great Lakes Water Quality Initiative procedure for deriving criteria for the
protection of wildlife (U.S. EPA, 1991b) and the analysis of risk of selenium to wildlife
(Peterson and Nebeker, 1992). These efforts were intended to identify the highest
aqueous concentrations of toxicants that would not cause unacceptable reduction in
the growth, reproduction, or viability of representative mammalian and avian species
which ingest surface water or aquatic life taken from surface waters. For this analysis,
concentration in the food, rather than water, will initially be the focus to allow
comparisons of risk between wildlife and fish. Water as a route of TCDD exposure to
wildlife can be ignored because the BAF for TCDD in food organisms is generally 104
or greater; therefore food consumption, which in wildlife is of similar magnitude to
water consumption, would provide nearly all TCDD exposure.
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"For the analyses here, the following equation will therefore be used:
FC EL ^ EL
EFS*EFC FA EFS-EFC-RA
where:
FC= Food concentration associated with an effect level (pg/g)
EL= Effect level from a toxicity study (pg/g/d)
WtA= Weight of the organism of interest (g)
EFS= Extrapolation factor for relative species sensitivity
EFC= Extrapolation factor for subchronic to chronic exposure
FA= Food consumption for the organism of interest (g/d)
RA= Food consumption as fraction of weight (1/d)
The following analyses will use the above equation and consider risk to
mammalian and avian wildlife species that have a diet consisting solely of fish, or
other aquatic macrofauna that would have TCDD levels comparable to fish.
Representative mammalian wildlife species would include the river otter and mink.
Representative avian wildlife would include bald eagle, osprey, kingfishers, terns,
herons, diving ducks, mergansers, and loons. For the mammals, the food
consumption rate (RA) would be expected to be in the range of 10-20% of body weight
per day (Aulerich et al., 1973; Lauhachinda, 1978; Bleavins and Aulerich, 1981;
Linscombe et a!., 1982; Toweill and Tabor, 1982; Newell et al., 1987). For birds, the
value can range from 10 to 50% (Alexander, 1977; Fry, 1980; Stalmaster and
Gessaman, 1982, 1984; Bortolotti, 1984; Newell et al., 1987; Nagy, 1987; Craig et al.,
1988; Palmer, 1988; Poole, 1989).
4.3.3.2 Selection of Effect Levels - Mammals
As discussed in section 4.3.1.1, a 28-day LD50 for adult mink after a single oral
dose was estimated to be 4,200 pg/g, with some effects on growth at 2,500 pg/g
(Hochstein et al., 1988). The extent of absorption of this dose is not known, but
assuming absorption is substantial but not necessarily complete (e.g., 25 to 75%), this
suggests the median lethal accumulation would be 1,000 to 3,000 pg/g. For newborn
mink, exposed via i.p. injection for 12 days, a dose of 100 pg/g/day (total dose of
1,200 pg/g) resulted in a mortality of 62% over an observation period of 19 weeks
(Aulerich et al., 1988). The median lethal accumulation would be near 1,000 pg/g,
assuming that the half life of this material is greater than the 12-day exposure period.
These data indicate that mink are among the most sensitive mammals tested and that
their sensitivity is within the range exhibited by fish. However, because of inadequate
knowledge of the half life of TCDD in this species, this lethal accumulation data cannot
be related to food or environmental concentrations over a more prolonged exposure.
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As also discussed in section 4.3.1.1, there are no subchronic or chronic studies
available for mammalian wildlife species. Numerous chronic studies are, however,
available for other mammalian species. These investigations have been summarized
in reports previously published by EPA in the mid-1980s (e.g., see U.S. EPA, 1984).
Of these investigations, the study of Murray et al. (1979) with Sprague-Dawley rats
was considered the most complete for the mammalian wildlife analysis. Subsequent
to the investigation of Murray et al. (1979), two noteworthy studies with Rhesus
monkeys (Macaca mulatta) have also been published (Bowman et al., 1989a,b).
Murray et al. (1979) maintained three generations of Sprague-Dawley rats on
TCDD treated diets equivalent to doses of 0, 1, 10, and 100 pg/g/day. Decreases in
F0 generation fertility and F., generation litter size were reported for rats in the 100
pg/g/day treatment group. Furthermore, rats in the 10 pg/g/day group exhibited
decreases in fertility in the F, and F2 generations. Other effects observed included
decreased litter size at birth, decreased gestational survival and decreased neonatal
growth and survival. Murray et al. (1979) reported that the reproductive capacity of
rats in the low dose group was not significantly affected in any generation and a
NOAEL of 1 pg/g/day can be inferred. It should be noted that there has been debate
in the literature regarding whether or not a statistically significant reduction in offspring
survival occurred at the low dose (Nisbet and Paxton, 1982; Kimmel, 1988); however,
this dose probably reasonably reflects the threshold for risk in this species.
Bowman et al. (1989a,b) assessed the reproductive success of Rhesus
monkeys maintained on TCDD treated diets equivalent to doses of 0, 0.13, and 0.67
pg/g/day. Reproductive success was monitored in two cohorts produced during a
four-year exposure period and in a third cohort produced ten months post-exposure.
No reproductive effects were observed in the third cohort. At a dose of 0.13 pg/g/day,
no significant effects of TCDD on pregnancy rate, abortion rate, still birth rate, and
survival through one year were observed in the first two cohorts. At the higher dose,
significant effects on reproduction were observed. Results from these studies suggest
a NOAEL of 0.13 pg/g/day for reproductive success in the Rhesus monkey, which is
approximately eight-fold lower than the value reported for rats.
For the present risk characterization, the study of Murray et al. (1979) was
considered the most relevant for assessing adverse effects in wildlife populations
because it incorporated a multigenerational exposure regime. However, as discussed
previously, the mink appears to be one of the mammals most sensitive to TCDD
intoxication and based on a comparison of LD50 values is about an order of
magnitude more sensitive than the rat. An inter-species extrapolation factor (EFS) of
10 will therefore be used. Applying this factor to the rat NOAEL results in a value
roughly equivalent to using the NOAEL from the Rhesus monkey study with no inter-
species extrapolation factor.
4-53
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This analysis therefore suggests that TCDD poses no demonstrable risk to
mammalian wildlife if daily intake does not exceed 0.1 pg/g/day, based on a NOAEL of
1 pg/g/day and a interspecies extrapolation factor (EFS) of 10. For a consumption rate
(R^ of 10-20%, this daily intake corresponds to concentrations in fish of 0.5-1.0 pg/g.
For sensitive organisms, substantial effects on reproduction would be expected at
concentrations approximately ten-fold higher.
4.3.3.3 Selection of Effect Levels - Birds
As discussed in section 4.3.1.2, LD50s vary markedly among birds. For a
single oral dose followed by 37 days of observation, an LD50 for quail was 15,000
pg/gť but was greater than 100,000 pg/g for mallards and greater than 800,000 pg/g
for turtledoves. After a single i.p. dose, pheasants had an LD50 between 10,000 and
20,000 pg/g. Chickens that were orally dosed for 21 days died at a total dose of
21,000 pg/g, but not at 2,000 pg/g. These data suggest that gallinaceous birds may
be the most sensitive to TCDD, and for acute lethality, this group of birds seems to be
somewhat more resistant than sensitive mammals and fishes. As was the case for
mammals, there is an absence of good toxicokinetic information (especially half life),
so rt is difficult to link these numbers to more chronic dietary exposure or to
environmental concentrations. If it is assumed that these acute median lethal doses
(LD50ft) can be approximately equated to a lethal body burden under chronic exposure
(LR50), the median lethal chronic dose (LD50C) can be estimated as LDSO3-!^. If the
half life of ca. 1 year for non-egg laying adult pheasant is used
(k2=0.693-*-t1/2=0.002/day), an LR50 of around 15,000 pg/g corresponds to an chronic
absorbed oral dose of about 30 pg/g/day. If the half life of the starling is used (7.2
days), the chronic dose would be about 1,400 pg/g/day.
Based on the summary of TCDD wildlife toxicity studies in section 4.3.1.2, the
report of Nosek et al. (1992a) on reproductive effects in ring-necked pheasants was
selected to establish effect profiles for birds. Ring-necked pheasants were dosed
weekly (i.p.) for 10 weeks with TCDD at rates equivalent to 1.4, 14 and 140 pg/g/day.
A significant decrease in egg production and 100% mortality in embryos was observed
in hens treated at 140 pg/g/day. The numbers of eggs produced by the control hens
and hens in the 14 and the 1.4 pg/g/day groups were not significantly different. There
was a tendency for embryo mortality to increase with the cumulative TCDD dose to
the hens; however, the dose/response slope was very shallow and mortality in the two
lower dose groups was not significantly different from that observed in the control
group. A value of 14 pg/g/day will be used here as being associated with a low level
of risk for reproductive effects in this species. The steepness of the response curve
above this level is uncertain, but complete failure of reproduction would occur at
concentrations tenfold higher.
In applying these data to characterizing risk for birds, two considerations must
be made regarding whether extrapolation factors are needed. First, although studies
4-54
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using a route of exposure comparable to that employed by Nosek et al. (1992a) are
not available, the results of egg injection studies described previously indicate that the
chicken embryo is more sensitive than the pheasant, based on overt lethality. In
addition, histopathological alterations have been observed in the chicken, but not in
the pheasant. These data suggest that gallinaceous birds other than the pheasant
could be more sensitive to TCDD intoxication. However, the pheasant is still one of
the most sensitive birds tested thus far and there is no clear need for extrapolating to
the chicken (which seems to be the most sensitive member of a sensitive group) in
order to assess risk to piscivorous avian wildlife. Therefore, an EFS of 1 will be used
in this analysis. Second, based on the half life of nearly a year reported earlier for the
elimination of TCDD from non-egg laying adult pheasant, the subchronic exposure (10
weeks) in the study of Nosek et al. (1992a) would have resulted in achieving only 13%
of steady-state accumulation. A truly chronic exposure could presumably have had
nearly an order of magnitude tower concentration in the food and still elicited the same
tissue levels and effects. Thus, a subchronic to chronic extrapolation factor (EFC) of
10 will be employed.
An exposure associated with low risk to avian wildlife of 1.4 pg/g/day is
therefore calculated here using an EL of 14 pg/g/day based on the NOAEL derived
from the pheasant study and an EFSťEFC of 10 based on considerations of inter-
species sensitivity and exposure duration. For a food consumption rate of 10-50% of
body weight per day, this daily intake would correspond to a TCDD concentration in
fish of 3 to 14 pg/g. Substantial effects on the reproduction of sensitive birds would
be expected at chronic exposures approximately ten times higher.
As discussed in section 4.3.2, a threshold TGDD concentration of 500 pg/g in
gull eggs was associated with effects on herring gull production on Lake Ontario.
Since Braune and Norstrom (1989) reported a biomagnification factor of 21 for TCDD
in gull eggs relative to alewives in Lake Ontario, this threshold is associated with a
concentration in alewives of approximately 24 pg TCDD/g. It should be noted that
these wild populations are subject to a variety of other stressors, including other
hydrophobic chemicals (PCDDs, PCDFs and PCBs) that could act jointly with TCDD.
Therefore the value of 24 pg TCDD/g likely underestimates a no effect exposure if
TCDD was the sole chemical stressor responsible for the adverse effects.
Based on the above information, the effects and associated uncertainties of
TCDD on aquatic-associated wildlife are summarized in Box 2.
4-55
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BOX
correspondst&iish
anajysfs baseti'on ' .:
significant risk a^soci^e|f|'|||::|[?|ftf|^|||i||f||||
cortsumptiort rate:vS^'^ :":: :-x::':::':-r^.--:
W?:!Ť.:;;:::^
^^
2,000 pg/g. .For.
sensitivities.1 ^^^f^<^^jjjl^^^^i^^^^^^^^^^^^^^^^^^^^j^^^^^^^^^^^^.
the lack.of good ^ tpxfcohMi§;;|n|||||^|^
may
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effects woŤW be
0fci^on:CT
S'isg;&ftgiS!2if &mmmmm&^^i-^m^m^mMmmm&&s^^mK^'m
4-56
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5. RISK CHARACTERIZATION METHODOLOGY
In an ecological risk assessment for a chemical stressor, risk characterization is
the phase in which the results of exposure and effects analyses are integrated to
evaluate the likelihood of adverse effects in exposed organisms, populations,
communities, or ecosystems based on actual or projected exposures of organisms to
the chemical, or suite of chemicals, in the environment (U.S. EPA, 1992c). The
degree to which risk is characterized can vary markedly, but ideally involves a
quantitative scale of effects and estimation of probabilities and uncertainties. Current
information is insufficient to provide such a thorough description for TCDD risk to
aquatic life and associated wildlife, with quantification of uncertainty .being particularly
difficult given the limited knowledge base. Furthermore, a thorough assessment of
TCDD risk should consider its joint action with other contaminants and non-chemical
stressors and the expression of effects on individual organisms at a population and
community level; such techniques are even less well established-and must await
further development. However, the adequacy of a risk characterization depends on
the nature of the specific problem of interest, so current information can be adequate
to characterize TCDD risk to aquatic life and associated wildlife in some cases.
The principal goal of this report was to evaluate and summarize data and
methods that are available for the assessment of TCDD risk to aquatic life and
associated wildlife, and to identify the major uncertainties that currently limit how well
risks can be characterized. A definitive risk assessment for a specific problem was
not a goal of this report, and there was consequently no explicit risk assessment
problem identified earlier in the report. However, it is necessary to discuss how
exposure and effects information should be integrated into a description of risk and it
is most effective to do this in the context of actual problems regarding the risk of
TCDD to aquatic life and associated wildlife. Therefore, this section will summarize
the exposure and effects information already presented and then offer examples of
how this information can be applied to the characterization of risk. These examples
are intended only to illustrate methods that can be used in risk characterization and to
identify major uncertainties that should be of concern. The level of detail presented
will be limited to that needed to accomplish this purpose and will be less than that
which would be provided in complete risk characterizations.
5.1 SUMMARY OF EXPOSURE AND EFFECTS INFORMATION
5.1.1 Exposure
As discussed earlier in this report (section 2 and references cited therein), the
high bioaccumulation potential and toxicity of TCDD result in water concentrations of
concern that are below ordinary analytical detection limits. Thus, there are few
measurements that reliably quantify the concentration of TCDD in natural waters. The
5-1
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one example cited in Section 2.3 indicated that concentrations in the Baltic Sea are
below a detection limit of 0.0002 pg/L in most samples, but some samples showed
concentrations in filtrates of 0.0002 to 0.0003 pg/L and similar amounts associated
with particulate matter. These data document the general magnitude of TCDD
concentrations in this specific system and demonstrate the importance of particulate
phases in TCDD distribution, even in waters with very low suspended solids.
However, there are as yet no water measurements that will support risk
characterization at the types of sites that are of concern to EPA regulatory activities.
In contrast, TCDD in sediments is more readily measured at exposure
concentrations of concern and there are measurements of sediment contamination
from several sites. Concentrations range from less than 1 pg TCDD/g dry sediment in
relatively uncontaminated sites to several hundreds and thousands of pg/g in the
highly contaminated Newark Bay, NJ. Concentrations above 10 pg/g commonly
appear at sites that have some industrial impact, but the sites evaluated thus far are
so limited and selective that firm conclusions about the general extent of sediment
contamination are not possible. Concentrations within a particular system also tend to
be variable, in part due to the variation in the organic content of sediments. Surficial
sediments in Lake Ontario were found to contain an average of 70 pg TCDD/g dry
weight in 1987, and appear to have been as high as 500 pg/g in the early 1960s. On
an organic carbon normalized basis, the concentrations average approximately 2,500
pg TCDD/g organic carbon and tend to be similar throughout the depositional zones of
the lake,
TCDD contamination in fish has been much more widely measured than in
water or sediments, and results of systematic surveys of large number of sites in the
U.S. have been published (U.S. EPA, 1987; 1992b). While such data do not define
environmental exposure concentrations, they do provide information on the potential
exposure of fish-eating wildlife and also can help assess risk to fish when compared to
accumulation-based effects information. These data will be discussed at length in
section 5.2.1 below.
TCDD risk characterizations are subject to a fundamental, and nonquantifiable,
uncertainty in the aforementioned lack of essentially any data on water concentrations
and only a handful of data on sediment concentrations. This precludes any
characterization based on measured environmental concentrations except at a very
few sites, and forces reliance on accumulation in organisms as a measure of exposure
and/or on fate models to estimate exposure. Improvements in analytical methods and
expanded efforts to measure TCDD in water and sediment are needed to address
these uncertainties. The database of concentrations in fish can be used to assess a
variety of sites or classes of sites, but there is considerable uncertainty when these
concentrations must be related back to water and sediment concentrations or to
loadings of TCDD to an ecosystem.
5-2
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5.1.2 Bioaccumulation
As discussed in previous chapters, bioaccumulation relationships for TCDD in
fish are an important tool for integrating exposure and effects information for highly
hydrophobic, bioaccumulative chemicals. For TCDD, accumulation in fish is both the
primary referent for exposure of fish-eating wildlife and a better basis for assessing
effects in aquatic life than using water concentrations directly. Steady-state BCFs
determined in the laboratory vary by over an order of magnitude (Section 3.2, Table 3-
1). This variability is likely due largely to incomplete characterization of exposure
concentrations or experimental shortcomings, including partitioning onto organic matter
in test systems, oversatu ration of TCDD, and time-varying concentrations in static
systems. On a lipid normalized and total water concentration basis, steady-state
bioconcentration factors (ssBCF|) would appear to be at least 10s even for systems
with uncertain exposures and would appear to more likely be of the order of 106.
Because these factors do not reflect exposure via food, they should underestimate
accumulation under more natural exposures. However, the relative bioavailability in
laboratory versus natural systems and the estimation of steady-state through
extrapolation results in uncertainties which might lead to either underestimation or
overestimation of BAFs in the field.
calculated for lake trout from Lake Ontario data are based on water
concentrations from chemical mass balance model calculations, but also seem to
suggest values of about 106 on a total water TCDD basis .(Table 3-2). The BAFj for
lake trout is estimated to be 2-1 06 to 3-1 07; uncertainty in the K^ of TCDD results in
this variability of predicted partitioning of TCDD between fish tissue and the freely
dissolved fraction. In contrast, BAFjs do not vary much with KQW because the
partitioning both within the water (to organic matter) and the organism are roughly
proportional to KQW. As discussed in section 3, the best current estimate for
extrapolating this Lake Ontario BAF to other situations is to equate BAFj to 3-106-fd,
which should be approximately 0.2-1 06/POC.
As discussed in section 3.4 of this report, BSAFs for TCDD appear to vary
within the 0.03 to 0.3 range for a variety of systems and fish species. Although
BSAFs for specific sites can always help to interpret problems and address remedial
actions at those sites, a concern here is how well these TCDD BSAF measurements
apply to other sites. A key issue in this extrapolation is whether TCDD BSAFs
invariably are below the expected equilibrium value, which is 1 or slightly higher. Low
BSAFs might reflect disequilibrium between water and sediment (Rws<1), possibly due
to the effects of decreasing anthropogenic inputs into the water, sediment diagenesis
which increases the freely dissolved chemical concentration in the sediment, and/or
loss processes from the water which keep the chemical concentration in water
depressed relative to the sediment. The low BSAFs might also reflect effects of
growth dilution and metabolism (in the entire food chain) which keep tissue
concentrations below equilibrium values with the water (Raw<1). Whatever the reason
5-3
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for the lower than expected BSAFs for TCDD, based on appropriate measurements for
some rather diverse sites and species, there is no evidence that these factors are
greater than 0.3 or iess than 0.03, with the best point estimate being on the order of
0.1. If these values are in some part due to water/sediment disequilibrium, then such
an extrapolation would underestimate risk at sites with Rws closer to 1. However, this
error is unlikely to be more than two- to four-fold, given the observed range of BSAFs.
The paucity of exposure data from natural systems also affects understanding
of bioaccumulation because there are no directly measured BAFs and only a limited
number of measured BSAFs. For BAFs, it is difficult to quantify the uncertainty, but it
could conceivably be several-fold. This uncertainty is associated in large part with the
variation in estimates for Kow and by a lack of a good empirical database on the
effects of organic matter on the distribution of TCDD. BSAFs are somewhat better
defined, but they do vary an order of magnitude among species and sites in a manner
that is inadequately understood; extrapolation of these BSAFs to new situations could
arguably be uncertain by two- to four-fold.
5.1.3 Effects
As discussed in section 4.2 and the references cited therein, fish appear to be
much more sensitive than other aquatic organisms and early life stages of fish appear
to be significantly more sensitive than older fish. Studies with lake trout currently
provide the most useful information on an important, sensitive endpoint, with fry
survival being little impacted for egg accumulations up to 34 pg TCDD/g, but being
severely affected at concentrations only 1.5- to 2-fold higher. Current information
suggests that accumulation is higher in maternal fish than in eggs, but only by about
50% on a Hpid normalized basis. Thus, using these lake trout as a surrogate for
sensitive fish species would suggest that low risk would be associated with
accumulations of up to 50 pg/g in fish with the same lipid content (about 8%) as eggs
(about 600 pg/g lipid) and high risk would be associated with accumulations above 80
pg/g (about 1000 pg/g lipid). Based on the variability among tested species discussed
In section 4.2, more tolerant organisms would be at high risk at exposures
approximately ten-fold higher than for sensitive organisms.
The most sensitive ecologically important endpoints established for both
mammals and birds are associated with reproduction. For mammals, observed
reproductive effects in tests on the rat and Rhesus monkey, combined with the higher
sensitivity of mink than rats based on lethality tests, suggest the risk is low for
mammalian wildlife which consume less than 0.1 pg TCDD/g/day. For a food
consumption rate of 15% of body weight per day (a median value for the species
reviewed) this low risk level corresponds to fish with a TCDD concentration of up to
0.7 pg/g. A similar analysis suggests that there is no evidence for significant risk to
avian wildlife which consume less than 1.4 pg TCDD/g/day. For a food consumption
rate of 25% of body weight per day, this corresponds to fish with up to 6 pg TCDD/g.
5-4
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Based on epidemiological evidence for herring gulls in Lake Ontario, there is no
evidence for significant risk to gulls associated with fish concentrations of up to 24 pg
TCDD/g. Substantial effects on reproduction for both sensitive, birds and mammals
are likely at exposures 10-fold higher than specified above for low risk, but the dose-
response curves are uncertain because available studies are limited and employed
large differences between treatment levels. The range of tolerance among all species
is also uncertain, but for birds probably exceeds an order of magnitude.
The limited database concerning wildlife effects makes these values uncertain.
The use of effects data from relatively resistant species (e.g., the rat) and/or
subchronic exposures (e.g., the pheasant study) necessitated the use of extrapolation
factors' to estimate risk to more sensitive species and longer exposures. In addition,
the number of species studied is limited, TCDD accumulation was usually not
monitored, and studies used widely-spaced exposure concentrations. The net result is
very little information on variability among species and exposure concentrations and a
poorly defined dose-response curve. Uncertainties again might be several-fold and
the underlying deficiency is a limited database on reproductive effects of TCDD to
organisms of interest.
Effects data for aquatic life provide somewhat more information on
uncertainties. If the focus of a risk characterization is solely on the early life stage of
lake trout as the receptor of interest, uncertainties in egg accumulation associated with
different levels of risk are probably about a factor of 2 to 3. This uncertainty is small
because of the steep response curve and the good agreement in LR50s among
different experiments and routes of exposure. This uncertainty range also includes
consideration of possible sensitivity differences in stocks of fish, which can have a
range of about two based on experiments with different strains of rainbow trout. On
the basis of accumulation in adult fish, the uncertainty would be slightly higher (e.g., a
factor of 3 to 4) because of uncertainty in the relationship of TCDD in eggs to that in
maternal fish.
However, the definition of lake trout early life stage survival as the
measurement endpoint for aquatic life risk assessments is based on limited testing.
Effects of TCDD on fish reproduction have not been evaluated after exposure of males
and females through a complete reproductive cycle. On the basis of accumulations in
eggs (which is necessary for application of these data), sensitivities of early life stages
have been established for only two species in Salmonidae, with tests on other species
providing indications that their sensitivities are not dissimilar to the salmonids, but may
span a somewhat broader range. The range of aquatic species other than fish that
have been tested is limited. Also, many tests do not document accumulation, so there
is no ability to extrapolate to different exposure conditions. These factors add
additional uncertainty to this analysis which is difficult to quantify.
5-5
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5.2 APPLICATION OF INFORMATION TO RISK CHARACTERIZATION
5.2,1 Fish Contamination in the United States: Risk to Aquatic Life and
Associated Wildlife
Two national surveys by the EPA (U.S.EPA, 1987; 1992b) analyzed fish from a
large number of diverse sites for TCDD and other chemicals. These surveys provide
the best set of data for .consideration of risk on a national basis and at different
classes of sites, although effects can only be considered on the basis of accumulation
in fish and cannot be reliably related back to environmental concentrations and source
loadings. This section will consider how this information can be applied to evaluations
of the risk of TCDD to aquatic-life and associated wildlife.
The National Dioxin Survey (U.S.EPA, 1987) reported results from fish sampled
at 395 sites. Ninety of these sites were randomly located throughout the United
States at U.S. Geological Survey (USGS) National Stream Quality Accounting Network
(NASQAN) and Benchmark Network sites. In these statistically located sites, TCDD
was not detected in almost 85% of the samples (detection limits ranged from 0.2 to 4
pg TCDD/g and were typically about 1 pg/g) and exceeded 5 pg/g in only about 6% of
the samples, with a maximum concentration of 19 pg/g. In sites selected based on
regional concerns with certain discharges, 75% of the samples had no detectable
TCDD, while only 10% were >5 pg/g and 2% were >25 pg/g. Greater contamination
was generally observed for 29 sites sampled in the Great Lakes, where TCDD was
detected in 80% of samples, exceeded 5 pg/g in 60% of the samples, and exceeded
25 pg/g in about 6% of the samples. The results for whole fish from all sites in this
survey are plotted on the top portion of Figure 5.1 as cumulative percentile versus log
concentration in whole fish. This plot also includes designation of the 10th, 50th, and
90th percentiles for each of the three subsets of sites.
Fish analyses from 388 sites were reported as a result of the National Study of
Chemical Residues in Fish (U.S.EPA, 1992b). This study emphasized sites influenced
by point and nonpoint sources (314), but also included a sample of USGiS NASQAN
sites (39) and selected background sites (35). Concentrations in fish were less than 1
pg/g or not detectible in nearly 50% of the samples, exceeded 10 pg/g in about 13%
of the samples, and reached a maximum of about 200 pg/g (bottom of Figure 5-1).
This survey noted significant differences among classes of sites, and Figure 5-1
depicts the 10th, 50th, and 90th percentiles for several of these classes.
U.S. EPA (1992a) summarizes results from several other studies which are
largely consistent with these distributions. However, it should be noted that there are
a variety of locations with contamination significantly higher than noted in Figure 5-1,
which would not be expected to be included in general surveys. For example, U.S.
EPA (1992a) noted concentrations reaching several hundred pg/g at some locations.
5-6
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Figure 5-1, Whole fish TCDD concentration versus cumulative percentlle from two national
surveys (U.S.EPA, 1987; 1992b), Filled circles denote measurements on individual
samples of fish. Vertical lines in cumulative plot depict one or more samples in which
TCDD was below the detection limit. Diamonds on horizontal lines denote 10th, 50th,
and 90th percentiles for subset of samples. Thick bars denote estimated range for
low and high risk of TCDD to sensitive species in group of organisms.
o
QŁ
UJ
0.
100
80
60
40
20
=>
O
<
I
z
UJ
o
UJ
Q_
UJ
>
O
U.S.EPA (1987)
Great Lakes
Reg i onal Si tes
Statistical Sites
0.1 0,3 1 3 10 30 100
TCOO CONCENTRATION IN WHOLE FISH (pg/g)
FISH
'BIRDS
High
RISK TO
SENSITIVE
ORGANISMS
Low
MAMMALS
Low
100
80
60
40
20
U.S.EPA (J992b)
High
Paper Mills
Using Chlorine
Industrial/Urban
frNASQAN Si tes
-Background S i tes
I I I I
0,1 0.3 1 3 10 30 100
TCDO CONCENTRATION IN WHOLE FISH (pg/g)
5-7
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Interpretation of this data must also be tempered by several considerations.
First, the datasets in their entirety are not random samples of U.S. waters, but rather
skewed to some extent to waters affected by anthropogenic activities. The percentiles
on Figure 5-1 must be interpreted accordingly. Additionally, fish surveys inherently
reflect the life history and sensitivity of fish, which can bias their use as exposure
indicators for a particular site. At sites where TCDD contamination would preclude
self-sustaining fish populations, fish might not be available or the sampled fish might
be immigrants that do not completely reflect exposure at that site. At the more
contaminated sites, fish would also be expected to reflect those species which are
less sensitive and/or do not readily accumulate TCDD because of their life history and
behavior. The database has not been evaluated for trends that would reflect such
considerations and, in fact, the inadequacies of current understanding of species
sensitivity and TCDD accumulation would make any definite conclusions difficult.
Nevertheless, these data can support some characterization of the risk of
TCDD to aquatic life and associated-wildlife on a national and regional scale. Figure
5-1 depicts the TCDD concentrations in fish associated with low and high risk to
sensitive fish and wildlife. The risk ranges are as developed in sections 4.2.3 and
4.3.3 and summarized in section 5.1.3.
Relative to these risk ranges, these survey data suggest that TCDD
contamination is below levels of concern for aquatic life at all but a small percentage
of the sites nationwide. However, there are a variety of sites in these surveys, and
from other studies, where TCDD concentrations are high enough to pose significant
risk to fish. This is especially true if joint action with other chemicals is considered.
For wildlife, Figure 5-1 suggests significant risk is more widespread than for
aquatic life, which is expected since the effects profile developed in this report
suggested piscivorous wildlife are more susceptible than the fish they consume. In
particular, the low end of the risk range for mammals is exceeded in about half of the
samples in these surveys. This characterization must be qualified by the uncertainties
cited earlier for wildlife effects data. Furthermore, it is unknown whether sensitive
piscivorous wildlife are actually or potentially present at these sites and if their diet is
actually comprised primarily of contaminated fish. Such questions would need to be
resolved in any specific risk assessment, but are beyond the scope of this discussion.
Nevertheless, this comparison of fish survey results and wildlife effects data does
raise significant concerns about the risk of TCDD to piscivorous wildlife. Research to
address the uncertainties which limit current abilities to characterize risks to wildlife is
clearly warranted.
5.2.2 Lake Trout Reproduction in Lake Ontario
A good example of TCDD risk characterization for a particular assessment
endpoint and site is that of lake trout reproduction in Lake Ontario. This system has
5-8
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been the subject of considerable study, both in documenting problems in fish
reproduction and in measuring and modeling levels of TCDD and other organic
contaminants. Also, lake trout have been the subject of laboratory studies on the
effects of TCDD. The Lake Ontario lake trout problem has already been discussed
soemwhat in section 4,2.2 and a more definitive review of this problem is the subject
of a separate effort (Cook et al., 1993a). The following discussion is designed to
briefly illustrate how the Information and methodologies presented in this report can be
applied and is not intended to be a complete discussion of this problem.
As discussed in section 4.2.2, lake trout populations in Lake Ontario declined in
the early part of this century and were severely depleted by 1950, before significant
contamination by TCDD and other organochlorine chemicals occurred. Commercial
fishing, lamprey predation, and/or degradation of spawning habitats from conventional
pollutants probably were largely responsible for the original decline. However, even
after the reduction of these stresses, a stocking program failed to establish a natural-
reproducing population, apparently at least in part due to blue-sac syndrome in lake
trout sac fry. The association of blue-sac syndrome with chemical stress in laboratory
studies raises the possibility that a chemical, or combination of chemicals, might have
contributed to continued reproductive failure of Lake Ontario lake trout.
As summarized earlier, laboratory studies have established that lake trout sac
fry survival is severely affected by TCDD, with a steep dose-response curve which
ranges from little or no mortality at approximately 30 pg TCDD/g wet weight of egg to
complete mortality at about 100 pg/g. In separate experiments with different routes of
exposure, 50% mortality occurred at accumulations in the eggs of 47 to 65 pg/g.
Parent fish from Lake Ontario have approximately three-fold higher TCDD
concentrations than their eggs, presumably in large part due to a higher lipid content
(18% in parent fish versus 8% in eggs), so the threshold concentration in eggs would
correspond to about 90 pg/g wet weight in fish and the LR50s in eggs to about 140-
200 pg/g in the adult fish. In addition to some experimental variability and some
question on the egg/parent relationship, these numbers are also uncertain because
they are based on a limited stock of fish. Experiments with different strains of rainbow
trout showed over a two-fold variability in LRSOs, so it is quite possible that some lake
trout in natural systems are as much as twofold more or less sensitive.
Information on TCDD residues associated with toxic effects can be compared to
TCDD concentrations observed in Lake Ontario lake trout and eggs. In 1987, eggs
collected from Lake Ontario contained about 10 pg TCDD/g wet weight. This is about
three-fold lower than the threshold cited-above. Therefore, even given the
uncertainties in the effect relationships, it is unlikely that lake trout reproduction is
currently at risk in Lake Ontario due solely to TCDD effects on sac fry survival.
However, in the past, the impact of TCDD to Lake Ontario lake trout may have
been significant. TCDD in surficia! sediments and lake trout both declined two- to
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three-fold from 1978 to 1988. In 1978, lake trout had average concentrations of 78 pg
TCDD/g wet weight, near the threshold for effects from TCDD alone. Based on the
sedimentary record, this concentration would have been several-fold higher in 1962,
well above that which would have precluded successful reproduction.
Even based on the data from 1987-1988, TCDD might still be contributing to
continuing Lake Ontario lake trout reproduction problems in concert with other PCDDs,
PCDFs, and planar PCBs. TCDD concentrations in eggs are at about one-fifth of the
LR50, which could be a significant contribution to total effects in a complex mixture.
As mentioned throughout this report, risk of TCDD cannot be adequately evaluated in
isolation from chemicals with which it often occurs. The joint toxicity of TCDD and
related chemicals is a major uncertainty that needs to be addressed. There also is an
issue of whether the measurement endpoint used here, sac fry survival in a laboratory
environment, is an adequate surrogate for the assessment endpoint of Interest,
namely lake trout reproduction in a natural system. Other toxicological endpoints
associated with reproductive physiology might be more relevant and sublethal effects
on fry might also affect their survival in a natural environment. These uncertainties
need to be addressed to adequately assess risk of TCDD to Lake Ontario lake trout,
and to aquatic life in general.
5.2.3 Environmental Concentrations Associated with TCDD Effects
Some EPA regulatory activities, such as the establishment of water and
sediment quality criteria, require the specification of environmental concentrations that
are considered to represent acceptably low risk. These concentrations are set
generically, and are applied to specific sites to calculate allowable discharges or set
goals for remedial actions. Such procedures do not fit the risk paradigm in which
effects and exposure information are combined into a statement of risk and
uncertainty, after which decisions about managing risk are made (U.S.EPA, 1992c).
Rather, water and proposed sediment quality criteria incorporate risk management
decisions before specific exposure information is introduced into the process.
Nevertheless, the setting of such criteria still involve major elements of a risk
assessment, albeit an incomplete one, and the entire regulatory process includes all
the elements of risk assessment, although somewhat intertwined with risk
management.
In this section, the information presented earlier on effects and their relationship
to exposure conditions will be used to associate concentrations in water and sediment
with different levels of risk to aquatic life and aquatic-associated wildlife. This
integration provides a simple, limited demonstration of how the information reviewed in
this document can be used and is not intended to be a complete and definitive
characterization.
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Concentrations associated with two levels of risk will be calculated here. "Low
risk" will be equated to the highest concentration that is unlikely to cause significant
effects to sensitive organisms. "High risk to sensitive organisms" will be associated
with the lowest exposure concentration that will likely cause severe effects (an EC50
or LC50, or a worse effect if a median effect concentration is not available). Risk to
more tolerant organisms will not be explicitly included in this discussion. Where
relevant, calculations will be based on fish with 8% lipid content and sediment with 3%
organic carbon. Concentrations for other conditions can be estimated based on
procedures outlined in Section 3.
Table 5-1 lists TCDD concentrations in fish associated with low and high risk to
sensitive fish, mammals, and birds. These concentrations were developed in the
effects profiles in section 4 and are summarized in section 5,1.3. For fish, the survival
of lake trout sac fry exposed as eggs was used as the measurement endpoint For
fable 5-1. Environmental concentrations associated with TCDD risk to aquatic
life and associated wildlife.
Organism
Fish
Concentration
(pg/g)
Sediment
Concentration
(pg/g dry wt.)
Water
Concentration
(pg/L)
POC=0.2
POC=1.0
Low Risk
Fish
Mammalian
Wildlife
Avian
Wildlife
50
0.7
6
60
2.5
21
0.6
0.008
0.07
3.1
0.04
0.35
High Risk to Sensitive Species
Fish
Mammalian
Wildlife
Avian
Wildlife
80
7
60
100
25
210
1.0
0.08
0.7
5
0.4
3.5
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mammals, the effect concentrations are based on rat reproduction studies, adjusted
for the apparent greater sensitivity of the mink. For birds, the effects levels are based
on a pheasant reproduction study, adjusted for the absence of a chronic exposure.
To translate fish concentrations to sediment concentrations, a BSAF of 0.3 was
used for assessing risk to fish. This upper end of the BSAF range was used (rather
than an explicit value for lake trout) because as far as is currently known, some
sensitive fish may be at the high end of the accumulation range. For wildlife, a
midrange BSAF value of 0.1 was used to reflect the fact that wildlife would consume a
variety of fish with a range of accumulations. It should be recalled that BSAFs are
based on a very limited database of field values and pertain to the surficial sediment
layers that Interact with the organisms and overlying water.
To translate risk based on fish accumulations to water accumulations, BAFs
based on Lake Ontario data were used, adjusting for the possible effects of POC
using the relationships developed in section 3.4 and summarized in section 5.1.2.
Different BAFs were not used here for sensitive fish and for fish that are wildlife food
sources because current information does not support specifying any differences. For
a POC of 0.2 mg/L, equal to that in Lake Ontario, the numbers in Table 5.1 reflect a
BAF1, of 10s. For a POC of 1 mg/L, effect concentrations are five-fold higher to reflect
the presumed greater binding by organic matter. As discussed previously, this
relationship is tentative and needs further investigation.
Relative to the limited information on environmental exposures available as
summarized in section 5.1.1, the environmental concentrations listed in Table 5-1 do
suggest that significant risk of TCDD to aquatic life and associated wildlife can be
expected in some situations. This table also exemplifies the higher risk expected to
occur for piscivorous wildlife that are exposed to TCDD contaminated fish. To better
characterize these risks, improvements in data and methodologies are needed to
reduce the uncertainties discussed thus far in this report.
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6. RESEARCH NEEDS FOR REDUCING UNCERTAINTIES
Throughout this report it has been emphasized that there are important and
Immediate uncertainties associated with characterizing the risks of TCDD to aquatic
life and associated wildlife. As outlined below, these uncertainties lead to a number of
major research needs regarding both TCDD exposures to and effects in aquatic
ecosystems. Consistent with the objective of this interim assessment to focus on
TCDD only, the following discussion does not address needed studies to evaluate the
appropriateness of using existing TEF values, based on human health and other
toxicoiogicai endpoints, for quantifying aquatic life and wildlife effects resulting from
exposures to mixtures of PCDDs, PCDFs, and PCBs. The following discussion also
does not address the development of population and community level models and
their linkage to toxicoiogicai inputs. Clearly, the development of such models are
needed for improving ecological risk assessments of chemical stressors in general.
6.1 EXPOSURE
6.1.1 Octanol/Water Partition Coefficient
The interpretation and application of data on the partitioning of TCDD in test
systems and natural ecosystems depend in part on estimates of K^, for TCDD. As
discussed in section 2, uncertainty in this parameter leads to uncertainty in
extrapolating bioaecumulation information and estimating the bioavailable fraction of
TCDD in water. Current understanding of the role of organic carbon on the
partitioning and bioavailability of TCDD from water and sediment is primitive.
Improved estimates of KM, and its relationship to TCDD partitioning onto natural
organic matter would reduce these uncertainties.
6,1.2 Detection Limits and Water Concentrations in Natural Systems
Typical analytical procedures are inadequate to reliably measure TCDD in water
at the low concentrations expected to elicit ecological effects. This analytical limitation
is especially true for dissolved TCDD. As a consequence of this analytical deficiency,
there are few reports which quantify TCDD in natural waters and the values reported
are uncertain. A high priority should be given to applying new techniques and
instrumentation to lower the detection limit for TCDD in water samples. With the
improvement of analytical techniques, total and dissolved TCDD should be measured
in a variety of aquatic systems to better establish current and future exposures.
6.2 BIOACCUMULATION
Reliable BAFs for TCDD based on measurements of water and biota in natural
systems are essentially nonexistent, which makes TCDD concentrations in fish tissues
difficult to predict in terms of TCDD concentrations in water or surficial sediments.
6-1
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Measurements of TCDD in surficial sediment, biota and water are needed to improve
the database on BAFs, BSAFs and BSSAFs. In gathering such data, particular
attention should be given to organism attributes (e.g., iipid content) and water column
properties (e.g., particulate and dissolve organic matter) which might alter
accumulation.
6.3 EFFECTS
6.3.1 Occurrence of Ah Receptor
Available toxicity data suggest that aquatic invertebrates and amphibians are
much less sensitive to TCDD than fish, perhaps due to the absence of the Ah
receptor, or a comparably sensitive receptor. A more rigorous assessment is needed
to determine if the Ah receptor is present in these taxa, as well as taxa not tested,
such as reptiles. The tissue distribution and appearance of the Ah receptor during
embryo and fry development must be determined in order to understand the TCDD
mode of action in fish. Based on the results of such studies, specific toxicity testing
on selected species could be undertaken to insure that current conclusions on
interspecies differences in TCDD sensitivity are valid.
6.3.2 Aquatic Life
Current data suggest fry survival as the most critical endpoint for fish; however,
the current threshold values are based on exposures of limited durations in a few
species. As a consequence, there is uncertainty regarding the species sensitivity
distribution within fish, as well as the impact of chronic exposures on reproduction. At
a minimum, additional early life stage tests with several fish species are needed to
establish a better sensitivity distribution for fish. There is also a need to determine
whether there are more sensitive endpoints than larval fish survival. Partial life cycle
tests should be conducted on at least two fish species in which chronic exposures
substantially precede the onset of reproduction. Effects on reproduction, early life
stage development and immune response should be examined in long-term exposures
to identify more sensitive chronic endpoints. Associated with such testing, TCDD
accumulation should be monitored to establish the basis to characterize risk with
increased certainty. These efforts would provide the basis to establish biologically-
based/dose-response models and thereby improve extrapolations across fish species
and exposure scenarios.
There is little available TCDD toxicity data for aquatic invertebrates and virtually
no information on the relationship of toxic effects on these organisms relative to TCDD
accumulation. Before reaching a final conclusion regarding invertebrate sensitivity,
additional long-term toxicity tests on a diverse set of aquatic invertebrate species
should be conducted to establish their sensitivity relative to that of fish and to identify
sensitive chronic endpoints. Additional studies will need to be conducted to define the
6-2
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relationship of these endpoints to TCDD accumulation. As mentioned previously, such
testing should be integrated with Ah receptor analyses.
6.3.3 Wildlife
With both the mammalian and bird risk characterizations, there was a lack of
quality reproduction bloassays and toxicokinetic information to establish well-defined
dose response relationships. For the mammalian assessment there were no
reproduction bioassays available for a representative piscivorous wildlife species (e.g.,
the mink) and therefore this characterization was based on an extrapolation of
bioassay results from rat and Rhesus monkey tests. For the bird assessment, a ring-
necked pheasant reproduction bioassay of limited duration and incorporating an i.p.
exposure regime was the only study available, although some limited toxicokinetic data
and in ovo toxicity studies were incorporated in the analysis. Finally, there are
apparently no data available to assess the toxicity of TCDD to reptiles.
A long-term (i.e., one generation) feeding study with mink would provide data to
more adequately assess reproductive and developmental effects. This bioassay
should be supported by toxicokinetic studies and Ah receptor investigations to develop
a biologically-based/dose-response model to better establish critical parameters in
species extrapolations and in linking TCDD accumulation to toxic effects. A long-term
(i.e., one generation) feeding study with birds would also support a more certain
assessment of reproductive and developmental effects. This bioassay must be
supported by toxicokinetic studies and Ah receptor investigations to better establish
the delivered dose to both the adults and developing embryos. Again, these studies
would contribute to a biologically-based appreciation of species extrapolation and
TCDD accumulation. These studies should also be designed in such a way to better
quantify the uncertainties of using egg injection studies as a source of toxicological
data in avian hazard assessments.
6.3.4 Epidemiology
Finally, there is a need to assess TCDD concentrations in aquatic life and
wildlife in appropriately selected natural systems. Such studies would further refine
and validate the reliability of relationships between TCDD accumulation and toxic
effects that have been established from laboratory investigations and epidemiological
studies in Lake Ontario and British Columbia. Future investigations should focus on
TCDD as well as chemicals with a mode of action similar to TCDD.
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7-25
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