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ANALYSIS OF RISKS FROM CONSUMPTION OF
QUINCY BAY FISH AND SHELLFISH
TASK IV REPORT
Prepared by:
METCALF & EDDY, INC.
Dr. Robert J. Reimold, Project Manager
Ms. Sara E. Bysshe, Project Consultant
Mr. Charles B. Cooper, Project Scientist
Ms. Mary Doyle, Project Scientist
Ms. Mildred Garcia, Project Scientific Support
Prepared under:
Prepared for:
U.S. EPA Contract No. 68-02-4357
Delivery Order No. 5
Mr. Stephen J. Silva, P.E., Project Officer
Ms. Katrina Kipp, Project Monitor
U.S. Environmental Protection Agency
Water Quality Management Section
Water Management Division
JFK Federal Building
Boston, Massachusetts 02203
In cooperation with:
U.S. EPA New England Regional Laboratory
U.S. EPA Environmental Research Laboratory,
Narragansett, RI
U.S. EPA Region I Public Health Advisor
May 1988
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TABLE OF CONTENTS
Page
I. Introduction 1
II. Approach 3
III. Hazard Identification 8
IV. Exposure Assessment 17
A. Species Selection and Characteristics 17
B. Contaminant Characterization 25
C. Estimates of Seafood Consumption..... 26
1. Commercial Catch 26
2. Recreational Catch 28
3. Consumption Estimates 30
V. Public Health Evaluation 36
A. Dose Calculation 36
B. Risk Characterization 37
C. Maximally Exposed Individual 39
D. Typical Quincy Area Resident 46
VI. Conclusions and Uncertainty 50
VII. References 66
Appendix A - Toxicity Profiles
Appendix B - Risk Calculations
Appendix C - Development of Carcinogenic Potency Factor for PCBs
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LIST OF TABLES
TABLE PAGE
1 Summarized Contaminant Levels and Hazard
Identification 9
2 Toxicity Values for Indicator Chemicals 12
3 Evidence of Carcinogenic!ty in Animals 15
4 Summary of Assumed Lifetime Consumption Levels 33
5 Risk Characterization for a Maximally Exposed
Individual from Ingestion of Quincy Bay Flounder,
Clams, Lobster and Hepatopancreas 41
6 Percent Contribution to Upper Bound Cancer Risk
by Each Indicator Chemical 44
7 Lifetime Risk Characterization for a Maximally
Exposed Individual from Ingestion of Quincy Bay
Flounder Only 45
8 Risk Characterization for a Typical Quincy Area
Individual from Ingestion of Quincy Bay Flounder
and Lobster 47
9 Risk Characterization for a Typical Quincy Area
Individual from Ingestion of Quincy Bay Flounder,
Lobster and Eepatopancreas 48
10 Upper Bound Estimated Lifetime Cancer Risks from
Quincy Bay Fisheries 51
11 Comparison of Estimated Lifetime Cancer Risks 52
12 Sources of Intake of PCBs 54
13 Comparison of PCB Levels Measured in Quincy Bay
and Boston Harbor Organism Samples 56
11
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LIST OF FIGURES
FIGURE PAGE
1 Quincy Bay Sampling Area Location of Sediment
Sampling Sites 18
2 Location of Other Trawl Fishing Transects for
Winter Flounder 21
3 Locations of Lobster Collections 22
4 Field Sampling Locations for Soft-shelled Clams.... 23
5 Effect of CPFs for PCBs (Maximally Exposed
Individual-Mixed Diet) 59
6 Effect of CPFs for PCBs (Typical Quincy Area
Resident-Mixed Diet with Tomalley) 60
7 Flounder Consumption Effects
(Sensitivity Analysis) 62
8 Tomalley Consumption Effects
(Sensitivity Analysis) 63
9 Lobster Consumption Effects
(Sensitivity Analysis) 64
111
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I. Introduction
This report is one of a series of studies being conducted by
the U.S. Environmental Protection Agency, Region I, to
investigate the types and concentrations of pollutants in
sediment deposits in Quincy Bay, Massachusetts and the potential
public health implications of consumption of seafood exposed to
these deposits.
This series of studies was mandated by Report 99-731 of the
99th Congress, 2nd Session, U.S. House of Representatives,
relative to appropriations, on page 30. Other reports in the
series which have been completed include the following:
• Types and Concentrations of Pollutants and Extent of
Sludge Deposits in Quincy Bay, Massachusetts - Draft Report
by Metcalf & Eddy to U.S. EPA Region I, October, 1987.
• A Histopathological and Chemical Assessment of Winter
Flounder, Lobster, and Soft-shelled Clams Indigenous to
Quincy Bay, Boston Harbor and an In Situ Evaluation of
Oysters including Sediment (surface and cores) Chemistry -
Report by George R. Gardner and Richard J. Pruell, U.S.
Environmental Protection Agency, Environmental Research
Laboratory, Narragansett, Rhode Island to U.S. EPA Region I,
December 1, 1987.
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These reports provided a summary of available historical
data on sediment and biological residues of contaminants in
Quincy Bay and the results of field and laboratory investigations
of concentrations of contaminants in Quincy Bay sediments and
biota conducted by the U.S. EPA in 1987. Together, these two
reports represent the results of Phases I, II, and III of the
five phases of the required studies. This report presents the
results of Phase IV, the analysis of risks of consuming seafood
which originates in Quincy Bay. As described in more detail
below, the report is based on the use of measured values of
seafood contamination obtained in the Phase II and III work
(Gardner and Pruell. 1987) in a quantitative risk assessment
conducted according to the most recently available EPA guidance
(PTI. 1987). The results of this and the previous studies are
integrated in the Phase V/Task V Summary Report.
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II. Approach
The general approach used in the conduct of this study
involved use of the data on tissue concentrations of contaminants
in Quincy Bay seafood obtained by EPA in 1987 (Gardner and
Pruell. 1987) in a quantitative risk assessment following the
latest available EPA guidance for studies of this type
(PTI. 1987). Specific aspects of the approach to components of
the risk assessment are described below.
Hazard Identification
Identification of contaminants of concern for this task was
based on inclusion of those chemical species for which residue
concentrations were documented in Quincy Bay seafood. These
included the organic and metal compounds measured by EPA in 1987
(Gardner and Pruell. 1987). The contaminants chosen for study
had the following characteristics:
• corresponding data were available for sediment and fish
tissue concentrations;
• the contaminants were those for which either an EPA
Carcinogenic Potency Factor (CPF) or a Reference Dose (RfD)
or a U.S. Food and Drug Administration (FDA) Action Level had
been published.
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As recommended by PTI, 1987, the latest available
compilations by the EPA of reference doses, carcinogenic potency
factors, and toxicity profiles were used. We relied primarily on
the EPA's Integrated Risk Information System (IRIS) data base,
but supplemented it as necessary (as described in Section III
below).
Dose-Response Assessment
As suggested in PTI, 1987, two forms of dose-response
information were used. The first was the Carcinogenic Potency
Factor (CPF), which attempts to quantify the implied finite risk
of cancer at various doses of a chemical. The second, for non-
carcinogens, was the reference dose (RfD), defined as the highest
average daily exposure over a lifetime that would not be expected
to produce adverse effects. With the exception of a congener-
specific CPF for the mix of PCBs found in the 1987 Quincy Bay
seafood samples, no new data in either category were developed.
This CPF was developed in the manner documented in Appendix C by
EPA's Office of Health and Environmental Assessment in
Washington, D.C. (USEPA. 1988a).
Exposure Assessment
The Guidance Manual (PTI.1987) suggests that two forms of
exposure assessment are appropriate, depending upon the level of
available information. Consistent with those suggestions and the
level of available information, we used the following basis for
exposure assumptions:
4
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• A dual basis for consideration of detection limit values
in fish tissue, first assuming that values below the
detection limit represent zero concentration and secondly
assuming that the values are equal to the detection limit.
• Evaluation of risks due to consumption of three species
of seafood from Quincy Bay: lobsters, flounder and
soft-shelled clams.
» Use of a standard consumption rate from among those
contained in the manual (PTI. 1987) for the hypothetical
maximally exposed individual. Other, potentially more
typical seafood consumption patterns were developed on the
basis of historical surveys of fisheries consumption in New
England (Penn State. 1985) and field interviews with persons
familiar with the Quincy Bay fishery.
• Assumption that the ingested dose is equal to the
absorbed dose of the pollutants of interest.
• Initial assumption of zero background concentration of
the pollutants in other ingested items, such as drinking
water and other foods. This is consistent with the overall
methodology for carcinogens, which assesses incremental risk
above background.
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• Use of other standard assumptions for an integrated
exposure analysis, including exposure over a 70-year lifetime
and a body weight of the exposed individual of 70 kilograms.
• Assumption that cooking has no effect on the
contaminants (as noted in Section VI, this assumption may or
may not be conservative).
Risk Characterization
Based on the guidance of PTI, 1987, two measures of risk
were examined:
1. The plausible upper limit to excess lifetime risk of
cancer;
2. The summary of non-carcinogenic risk represented by the
ratios of the estimated exposure doses to the Reference Doses
for the studied chemicals.
As suggested by the Manual in its discussion of chemical
mixtures, we evaluated the additive risks of the several
contaminants present in the seafood as follows:
• Arithmetic summation of upper limit risks for
carcinogens; and
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> Arithmetic summation of the ratios of exposure dose to
RfD for only those non-carcinogens acting on the same target
organs.
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III. Hazard Identification
To focus the public health assessment on those contaminants
likely to represent the greatest risks, the 1987 Quincy Bay
analytical data collected under Tasks II and III of this study
and toxicity information were reviewed, including analytical data
from Tasks II and III as available in the December lf 1987 draft
report (Gardner and Pruell. 1987). Maximum and mean
concentrations of contaminants detected in each sediment and in
each of the different seafood species tissues were summarized
(Table 1). The mean concentrations represent the average with
concentrations below the detection limit assumed to be equal to
the detection limit, and were used in the public health
assessment. A second mean was also calculated with contaminant
concentrations below the detection limit assumed to equal zero.
These values are included in Appendix B.
Three of the references used extensively to obtain toxicity
data were (1) the Integrated Risk Information System (USEPA.
1986a-b; 1987d-h), an EPA-maintained computer database currently
available in hard copy, (2) Health Effects Assessment Documents
(USEPA. 1984a-j) and (3) the Superfund Public Health Evaluation
Manual (USEPA. 1986c). The availability of data from the first
two sources was also summarized (Table 1). In the Superfund
Public Health Evaluation Manual, the Carcinogenic Potency Factor
(CPF) is defined as an upper 95 percent confidence limit on the
probability of carcinogenic response per unit intake of a
chemical over a lifetime. The 95 percent confidence limit
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conventionally referred to implies a greater degree of accuracy
than is currently available given the uncertainty associated with
calculating CPF values. However, the CPF is generally considered
the plausible upper bound value. The CPF is the generally
accepted approach to convert estimated intake levels directly to
estimated plausible upper bound incremental risk as described in
Section V of this report. CPFs are presented for those chemicals
considered by the EPA to be human carcinogens or probable human
carcinogens (USEPA. 1986a-b; 1986c; 1987d-h). The EPA weight of
evidence (Table 1) refers to evidence of carcinogenicity, with
Group A signifying a known human carcinogen and Group D
signifying no classification. Group B signifies probable human
carcinogenicity based on animal studies, while Group C signifies
possible human carcinogenicity. The weight of evidence
classifications are described in more detail in Appendix A. In
general, the weight of evidence is classified by EPA without
regard to route of exposure, and route specific information is
included in the CPF determination. In Table 1, for metals,
different classifications have been made for inhalation and oral
routes. A classification of Group D was input for the oral route
where no evidence of carcinogenicity by the oral route of
exposure is available.
The values for reference dose (RfD) are generated by the EPA
based on the assumption that threshold levels exist for
noncarcinogenic health effects (USEPA. 1986c). The RfD is
considered to be the level unlikely to cause significant adverse
10
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health effects associated with a threshold mechanism of action in
humans exposed for a lifetime (USEPA. 1986a-b; 1986c; 1987d-h).
The RfD is used for comparison with calculated intake levels as
discussed in Section V. The EPA toxicity rating (Table 1) is
associated with noncarcinogenic health effects where 1 is
associated with small changes due to contaminant exposures and 10
is associated with death or pronounced life shortening and
teratogenic effects. The basis for the toxicity ratings is
presented in more detail in Appendix A.
Indicator Chemicals
The majority of those contaminants recently analyzed by EPA
and found in Quincy Bay sediments and seafood (Gardner and
Pruell. 1987) have been included as indicator chemicals in the
public health evaluation. In some cases the contaminants are
grouped based on availability of toxicological information, and
on similarity of chemical properties and toxicological effects.
The subset of indicator metals and compounds considered in the
public health evaluation (Table 2) are shown with the CPFs, RfDs
and critical effects for each. Toxicity profiles for the organic
compounds and metals found in Appendix A and excerpted here focus
on chronic exposure by ingestion. While some of the metals are
considered possible or probable human carcinogens, where there is
no evidence of carcinogenicity by ingestion, no CPF or weight of
evidence values are provided in Table 2. The RfD values for an
oral exposure to metals as well as the critical effect (Table 2)
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TABLE 2. TOXICITY VALUES FOR INDICATOR CHEMICALS
Oral
Carcinogenic
Potency
Factor(CPF)
Weight Oral Ref.
of Evi-. Dose (RfD) Critical
Metal/Compound
Cadmium
Chromium
Copper
Lead
Mercury
Chlordane
Dichlorodiphenyl
trichloroethane
(DDT)
Hexachlorobenzene
(HCB)
(mg/kg/day)"1 dence*6
_
_ —
— -
- -
- -
1.3 B2
0.34 B2
1.69 B2
) (mg/kg/day)
2.90 * 10~4
o
5.00 * 10~3
M
3.70 * 10~*
1.40 * 10~3
2.00 * 10"3
5.00 * 10~5
5.00 * 10~4
8.0 * 10~4
Effect*a)
Renal
dysf unction* b)
NOEL, renal
dysfunction*^)
GI symptoms *'
-, renal.
effects*4)
-' renal
effects*5)
Liver necrosis
Liver lesions*
-, liver
changes, * 3) *9
(2
)
(1
D
)
Hexachlorocy- 6.3*°' B2
clohexane (HCH) 1.33*c) B2/C
Polycyclic aromatic 11.53 (d)
hydrocarbons (PAHs)
Polychlorinated 2.6 B2
biphenyls (PCBs)
3.0 * 10'4
NA
1.0 * 10~4
teratogenic
effects
Liver hyper-
trophy*^
Reduced
size of
offspring*8'
(a) The critical effect is the effect seen in the studies from which th
RfD is developed. The RfD is set at a level where the critica
effect is unlikely to occur. Where the study used to set the Rf
indicates a NOEL (no observable effect level), the most common!
observed effect is also noted. A "-" indicates the information i
the specific study defining the RfD is not included in this report
and the critical effects reported from other studies are included.
(b) Alpha HCH.
(c) Gamma HCH.
(d) See Table 3 for weight of evidence for PAHs
(e) For explanation of weight of evidence see Appendix A Table Al.
References:
1. USEPA. 1986a, and 1987e;g.
2. USEPA. 1984C.
3. USEPA. 1984f.
4. USEPA. 1984g.
5.USEPA.
6.USEPA.
7.USEPA.
8.USPHS.
1987a.
1987b.
1984d.
1987 for RFD.
USEPA. 1988a for CPF.
9.USEPA. 1987C.
12
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document the health effects seen at the lowest exposure
concentrations. The RfD is set by the EPA at a level where the
critical health effect is judged unlikely to occur.
Chlordane is considered here as total chlordane without
distinguishing between the alpha and gamma isomers measured by
Gardner and Pruell (1987). Many of the toxicity studies in the
referenced database were performed utilizing a technical grade
chlordane which includes both isomers. RfDs and CPFs were only
available for total chlordane (Table 2). The weight of evidence,
B2, indicates that the evidence of carcinogenicity in humans is
inadequate to consider the compound a known human carcinogen,
however, due to sufficient evidence of carcinogenicity in
animals, chlordane is considered a probable human carcinogen.
Toxicity profiles for chlordane and other organic chemicals are
provided in Appendix A.
Technical DDT (dichlorodiphenyltrichloroethane) is generally
a mixture of p,p-DDT, o, p-DDT, p,p-DDD, and traces of other
materials. Metabolites of DDT include p,p-DDE and o,p-DDD. DDT
isomers and metabolites are often found together and have similar
properties, therefore, they have been considered together as a
chemical class (Clement. 1985). The analytical data for p,p-
DDD, p,p-DDE, and p,p-DDT are presented separately in the public
health evaluation, however, the same RfD and CPF values provided
by the EPA, for DDT as a class, are used for all three compounds.
Teratogenic and carcinogenic effects have been documented
for exposure to hexachlorobenzene (HCB). Both CPF and RfD values
13
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are available for ECB. Liver effects such as hepatomegaly have
been noted in the literature. CPF values for
hexachlorocyclohexane (ECH) are available for both the alpha and
gamma isomers and are both used in this public health
evaluation. The RfD developed for the gamma isomer (lindane) was
used for both isomers. Lindane is considered the most acutely
toxic isomer, while no RfD is available for the alpha isomer.
There is one CPF value available for polycyclic aromatic
hydrocarbons (PAHs) based on the carcinogenicity of
benzo(a)pyrene (Table 2). Not all PAHs are known carcinogens.
PAHs evaluated in Quincy Bay seafood tissues (Table 3), have
varying amounts of evidence that they are carcinogenic in
animals. The individual PAHs have been grouped for evaluation in
the public health assessment as total PAHs. Evaluating all PAHs
as carcinogens is a standard conservative approach, which will
tend to overestimate increased lifetime cancer risk. No RfD for
PAHs was found during the literature search.
Polychlorinated biphenyl (PCBs) contamination was also
evaluated by grouping the data and evaluating total PCBs. EPA
determined that there is positive evidence for carcinogenicity in
animals for Aroclor 1254, Aroclor 1260, and some other PCBs.
Because any PCB mixture can contain appreciable amounts of
carcinogenic PCBs and because of the variability of PCB mixtures,
EPA has recommended that all commercial PCB mixtures be
considered to have a similar carcinogenic potential and
classified all PCB mixtures as Group B2 - Probable Human
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TABLE 3. EVIDENCE OF CARCINOGENICITY IN ANIMALS
Polycyclic-Sufficient LimitedInadequate No
aromatic Hydrocarbons Evidence Evidence Evidence Evidence
Fluorene X
Phenanthrene X
Anthracene X
Fluoranthene X
Pyrene X
Benzo(a)anthracene X
Chrysene X
Benzofluoranthenes X
Benzo(e)pyrene X
Benzo(a)pyrene X
Perylene X
Indeno(l,2,3-cd)pyrene X
Benzo(ghi)perylene X
Dibenz(a,h)anthracene X
Corene X
Source:Clement.1985.
Carcinogens, based on sufficient evidence of carcinogenicity in
animal studies (USPHS. 1987). A CPF of 4.34 (mg/kg/day)'1 has
been used in risk assessments in the recent past as the generally
accepted value. A new CPF of 7.7 (mg/kg/day)'1 based on
carcinogenicity data for Aroclor 1260 has been proposed in a
draft report (USPHS. 1987). Work by the US EPA Exposure
Assessment Group indicates that based on the thirteen congeners
15
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measured in Quincy Bay seafood the mixture more closely resembles
Aroclor 1254 than 1260. (USEPA. 1988b) An upper bound CPF of
2.6 (mg/kg/day)"1 was developed by the EPA Carcinogen Assessment
Group for Aroclor 1254, and is used in this evaluation. Appendix
C documents the development of this CPF. A sensitivity analysis
was performed as part of the results and conclusions in
Section VI to determine the effect on plausible upper bound
increased lifetime cancer risk given the use of different CPF
values for PCBs. The RfD for non-cancer risks for PCBs proposed
in the 1987 USPHS draft document has been used in this public
health evaluation at the suggestion of USEPA-OHEA.
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IV. Exposure Assessment
Portions of the data developed by Gardner and Pruell
(1987) were used to define exposure estimates in risk assessment
scenarios related to the extent of contamination of sediments and
selected biota in Quincy Bay. This section discusses the
selection of the chemical data, consumption data and population
characteristics required for the exposure portion of the Public
Health Assessment below.
A. Species Selection and Characteristics
Based on the guidance provided in the Report 99-731 of the
99th Congress, sampling for this study was conducted in Quincy
Bay during early spring and summer, 1987, for sediments, winter
flounder (Pseudopleuronectes americanus) , soft shelled clams (Wya
arenaria) , and the American lobster (Homarus americanus).
Additionally, 400 oysters (Crassostrea virginica) were suspended
at four locations: three in Quincy Bay and one at the Graves in
Massachusetts Bay.
Surface sediments were collected at 22 locations in Quincy
Bay and core samples were collected at four locations (see
Figure 1). Inorganic contaminant levels were measured in all
samples, and selected organics were measured in the core samples
and 14 of the surface sediment samples. Sediment sampling and
analyses methodologies are discussed in detail by Gardner and
Pruell (1987). Levels of contaminants at many locations
throughout the Bay were elevated beyond the levels generally
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1000
LEGEND
0 SURFACE SEDIMENT SAMPLING LOCATIONS
INORGANIC ANALYSIS ONLY
—, SURFACE SEDIMENT SAMPLING LOCATIONS
9 INORGANIC AND ORGANIC CHEMICAL ANALYSIS
p. CORE SEDIMENT SAMPLING LOCATION
INORGANIC AND ORGANIC CHEMICAL ANALYSIS
FIGURE 1. QUINCY BAY SAMPLING AREA. LOCATION OF SEDIMENT SAMPLING SITES.
SOURCE: US EPA. 1987.
18
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reported by others for Boston Harbor (see the Task I report for
details). Some organics (e.g. PCBs, DDE, PAEs) and some
inorganics were found at the highest levels offshore of Noon Head
and Long Island in the vicinity of sewer system discharges, and
around Peddocks Island and Nut Island near the discharges from
the Nut Island wastewater treatment facility. While contaminants
from these sediments may be released to surrounding water and may
be linked to contaminant levels in organisms, there is no
generally accepted method for directly quantifying the importance
of marine sediment contaminant levels to human health risks from
ingestion of contaminated seafood. Thus, sediment contamination
is not directly included in the computations of the exposure
assessment. Possible implications of the measured sediment
contamination levels are discussed further in the Task V report.
The biological sampling and analyses by* Gardner and Pruell
(1987) included collection and evaluation of histopathological
condition and chemical contamination in three species of
indigenous marine organisms of high -commercial and/or
recreational value in Quincy Bay: Winter flounder, American
lobster, and soft-shelled clam. Oysters brought in from Cotuit
populations were also placed in the Bay to allow in situ
evaluations of contaminant uptake after a 40 day exposure to
Quincy Bay conditions. Since this species is not commercially or
recreationally harvested from Quincy Bay, these results were not
included in the public health assessment of exposure to the
Quincy Bay fishery.
19
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Gardner and Pruell (1987) indicate that histopathological
evaluations provide strong evidence that Quincy Bay populations
of Winter flounder and soft-shelled clams are in poor health (See
Task V Report). At present, theoretical or analytical methods
for correlating the histopathological results with any potential
effects in humans do not exist. The exposure assessment was thus
limited to evaluation of the ingestion by humans of the residues
of chemicals contained in lobster, flounder, and soft-shelled
clams from Quincy Bay.
One hundred Winter flounder were collected by otter trawls
from four transect locations in the Bay (Figure 2). An
additional transect from Moon Head to the eastern end of Long
Island was eliminated due to lack of fishing success. Lobster
collections occurred at nine locations in the Quincy Bay study
area (Figure 3). Collections of specimens later analyzed were
made by traps. Seven sites were chosen for soft-shelled clam
collection, but the organisms were present only at Moonhead and
Moon Islands (Figure 4).
The three species chosen for chemical contamination
evaluation from Quincy Bay are the more commercially/
recreationally significant species harvested from the Bay. In
addition, each was determined to be sufficiently narrow-ranging
to be considered indigenous to the area. Soft-shelled clams are
essentially sedentary as adults. Winter flounder do move.
However, an extensive tagging study conducted in the early 1960's
(Howe & Coates. 1975) suggested that winter flounder in the
20
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Boston Harbor area showed only limited movement from inshore
release areas. Specifically, sexually mature fish moved in to
shallow water to spawn during winter and spring. Many remained
near spawning areas. Some migrated to deeper waters near the
harbor mouth and farther in warmer months. Howe and Coates
(1975) provide sufficient evidence of very high local recapture
rates to allow the simplifying assumption that the flounder
caught by trawl in Quincy Bay during May/ 1987 and those caught
and consumed by fishermen had been in the study area for at least
several months preceding their capture.
A similar simplifying assumption was made concerning
lobster caught in Quincy Bay. Lobster fishermen trap near shore
in the spring when mature lobsters are in shallow water to
spawn. They follow lobster movement to deeper water through
summer and fall months (Jones. 1987). Fishermen believe this
suggests movement of lobster populations that is temperature
related. There is additional evidence, according to the State
Division of Marine Fisheries (Estrella. 1987) that such movement
may occur in older lobsters, with juvenile populations being less
migratory. The DMF also indicates that there is evidence to
suggest that up to 95 percent of the legal size inshore
population is cropped by fishing pressure. At legal size, a
number of captured lobsters may not be sexually mature
(Estrella. 1987). In conclusion, it is possible but not
verifiable that many of the captured Quincy Bay lobsters in the
fishery and for the sampling and analysis in this study may have
24
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spent long enough to have become contaminated (i.e. at least
several months) in the Quincy Bay environment.
B. Contaminant Characterization
As may be seen from the narrow ranges of inorganic and
organic contaminant levels found in soft-shelled clams (Table 1),
residues from both the Moon Head and Moon Island locations were
very similar. Sediment samples were not collected near enough to
the clam collection locations to provide a basis for comparison.
The differences in inorganic and organic levels in lobster
tissues and the lobster hepatopancreas from different sampling
locations were not large (around 2X) and did not follow any clear
geographic pattern. The sample of lobsters was small and it is
likely that movement was sufficient to preclude definitive
conclusions concerning the relationship between lobster and
sediment contamination in this study. The significant difference
between lobster muscle tissue (tail) and hepatopancreas
concentrations, however, requires special consideration in this
assessment. Specifically, consumers of lobster "tomalley"
(hepatopancreas) may have a much higher exposure to the studied
contaminants than would those who only consume lobster meat.
Similarly, while there is some variability among
contaminant residues in individual flounder samples, geographic
patterns of tissue contamination that might be associated with
differing sediment contaminant levels within the Bay can not be
established in view of the lengths of the collection trawl
25
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transects (covering both more and less contaminated sediments)
and the opportunities for either or both the fish and sediments
to move. The Task V Summary Report discusses potential
sediment/organism contaminant relationships on a broader basis,
including discussion of Quincy Bay data versus data from other
locations.
Given the above results, potential consumer exposure was
assumed to be best represented for the analysis by the maximum
and the mean concentrations of contaminants found in each of the
three types of organisms analyzed (Table 1).
C. Estimates of Seafood Consumption
1. Commercial Catch
Seafood consumption estimates for risk assessment include
assumptions about the amount of seafood consumed as well as its
source. EPA guidance (PTI. 1987) recommends against attempting
to quantify the inherently variable commercial catch to consumer
distribution patterns for a risk assessment. As discussed below,
the distribution pattern for commercially harvested seafood from
Quincy Bay is typically irregular, and supports the guidance.
The scope for this study limits the number of species
considered in the consumption estimates to three. On the basis
of available harvest data and interviews with fisheries industry
participants, it is believed that clams, flounder and lobster do
in fact constitute the great majority of the consumed significant
catch from Quincy Bay. Other species seasonally harvested from
26
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the Bay in minor but measurable amounts include bluefish, eel and
smelt, all of which are migrant visitors.
Soft-shelled clams are in theory represented by only a
commercial fishery in Quincy Bay, as only "Master Diggers" are
legally permitted to harvest clams. The clams must then go
through the state's shellfish depuration plant (along with clams
from other areas in the state) prior to distribution.
Reportedly, local clam harvest makes up about 15 percent of the
local demand in metropolitan Boston (Kennedy. 1987). The
remainder of Boston's demand for soft-shelled clams is filled
with imports from areas such as Maryland. It would likely be
impossible to accurately trace Quincy Bay clams through the local
distribution system to ultimate consumers as destinations change
daily and sources are not well tracked (Connerty. 1987).
Additionally, individuals can hold "bait licenses" for clams. It
is believed that some (perhaps many) of these individuals and
others who may or may not hold licenses are clamming for personal
consumption (Ayers. 1987).
Over 12 million pounds of lobster were taken from the
coastal waters of Massachusetts in 1986 (Hoopes. 1986). The
coastal lobster permit reporting area that includes Boston Harbor
and Quincy Bay has been the most productive according to reports
for the last three years (Hoopes. 1985; Hoopes. 1986; Nash.
1984). This reporting system tracks the home port of vessels and
general reporting of areas harvested, but does not provide
overall harvest from an area that corresponds to the geographic
27
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boundaries of Quincy Bay. Lobster fishermen sell to a number of
different distributors. As with clams, it would be practically
impossible to track commercial catch from one area in sufficient
detail to generate a commercial Quincy Bay lobster distribution
to consumption profile over a long period of time.
There is no commercial winter flounder fishery in Quincy
Bay/ although some flounder taken in the Bay are sold by
recreational fishermen (Ayers. 1987; Jones. 1987).
t
2. Recreational Catch
The most recent EPA guidelines for seafood consumption
risk assessment (PTI. 1987) suggest that quantitative
considerations of recreational harvest and distribution to
consumption patterns may be appropriate for risk assessment,
depending on the quality of available data. These were
investigated for the three target species in this assessment.
The recreational flounder fishery in Quincy Bay has been
renowned for many years, with as many as 17,000 estimated annual
boat trips in the mid 1960's (Jerome. 1966). The fishery is
reportedly in decline due to publicity concerning water quality
(Childs. 1987). The state plans an updated recreational survey
but such numbers are not available at this time. Several marinas
rent boats in the area. On a summer day with good weather,
several independent estimates by local fishermen suggest that up
to 1,000 boats may be on the bay. The number of these boats
engaged in fishing is not known. A large number of recreational
28
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flounder fishermen are from out-of-state. Many evidently come a
number of weekends every year, returning home with large amounts
(e.g. 50 pounds or more) of flounder. It is reported that some
of these individuals sell some of their catch dockside and/or
out-of-state. However, it is known that some also keep
considerable amounts for regular personal consumption throughout
the year. Local fishermen can and do also keep enough flounder
for regular consumption, and such has been assumed here and
tabulated later in this chapter. Some local fishermen also fish
for striped bass and bluefish in the summer, and/or smelt in the
winter along with flounder, and fish areas outside Quincy Bay.
The data available at this time limit the consideration of
finfish consumption risks in this study to flounder.
Approximately 250 Quincy residents hold 10 pot
(recreational) licenses for lobster (MDMF. 1987. Unpublished
data). It is assumed that many of these individuals likely fish
Quincy Bay or its environs at least some of the time. In
addition, an unquantified number of license holders from other
nearby areas likely harvest the bay as well. (Reporting of
harvest location is not required for these license holders).
Local commercial lobstermen would anticipate five or six
"keepers" per set per 10 pot string of baited pots (Jones.
1987). Using the unpublished 1987 DMF License data (MDMF.
1987), the average catch per license holder was over 38
lobsters/year. These data provide a basis for judging the
29
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reasonableness of lobster consumption estimates from recreational
catch as tabulated below.
As noted above, soft-shelled clams are only legally
available for consumption by purchase following depuration.
However, some individuals are believed to consume Quincy Bay
clams that are collected illegally or with a bait license. This
is a form of illegal activity for which no records are available
although some estimates have been made below.
3. Consumption Estimates
The above assessment of the fishery makes it clear that
risk should be assessed for several different exposure
assumptions for the Quincy Bay fishery. The data also document
that while it is possible to generate a range of consumption
profiles, the fishery data are not adequate for definitively
assigning the population sizes that fit each profile.
Several levels of Quincy Bay seafood consumption were
developed for the risk assessment. These numbers were derived
using published surveys of a range of seafood consumption, along
with the approach recommended in the risk assessment guidance
manual (PTI. 1987).
According to PTI (1987) the standard value for maximum
consumption estimates in risk assessment is based on the
approximately 0.1 percent of the U.S. population which reportedly
consumes 165 grams/day of seafood. This is a slightly more than
1/3 Ib. serving of fish per day on average. On the basis of
30
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local interviews, it has been assumed that there is a small
percentage of the "local" population of Quincy area residents
that consume this much Quincy Bay fish or seafood, although the
actual population size was not estimated. In the absence of more
definitive, site-specific data, the consumer of this amount of
seafood, in various combinations (see below) is considered the
"Maximally Exposed Individual", (MEI), for this study.
Regional data for seafood consumption by species were
available for New England, so that consumption levels could be
estimated for "typical" consumers without relying on the Guidance
Manual default value.
Several national consumer surveys place New England
residents among the highest consumers of fish and shellfish. The
consumption estimates for "typical Quincy area resident" were
based on the survey data for New England consumers reviewed in a
study for the National Marine Fisheries Service (Penn State.
1985). Data from three surveys were cited in summaries of
regional consumption patterns. One represented a year (1969-70)
of survey results. The other two (1973-74; 1977-78) represented
more recent surveys of greater numbers of individuals, but were
conducted over shorter time periods of 3 days to one month. Each
survey represented a different bias. The differences reported
for average yearly flounder consumption in New England were 0.618
Ibs per capita to 1.005 Ibs per capita, and for average yearly
lobster consumption were 0.601 Ibs per capita to 1.895 Ibs per
capita. The choice was made to rely more heavily on the year
31
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long (1969-70) survey and to modify the averages, slightly up in
the case of flounder and slightly down in the case of lobster,
based on the data from the later surveys. The differences among
the averages in these surveys is small. From these data, then,
it is assumed that the typical local consumers with access to the
recreational fish harvest from Quincy Bay could consume a long-
term average of 1 gram/day (0.8 Ibs/year) of Quincy Bay flounder
and 2.1 grams/day (1.71 Ibs/year) of lobster.
Both of these figures appeared reasonable considering the
apparent recreational lobster and flounder harvest levels and the
exposure that could be associated with commercially distributed
catch. They were used to provide a departure point for
comparison with the "maximally exposed individual". Again the
number of individuals in the consuming population was not
estimated, but there is reason to believe that it could be
relatively large, given the catch volumes.
These above estimates resulted in four separate
consumption profiles, which are discussed below and summarized in
Table 4.
32
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TABLE 4. SUMMARY OF ASSUMED LIFETIME CONSUMPTION LEVELS
Maximally Exposed Individual
Mixed Diet Flounder Only
Typical Local Consumer
xed Diet Mixed Diet
Ty
Mi
Quincy Bay
Clams 16 g/day
Quincy Bay
Flounder
Quincy B
113 g/day
Tissue 30 g/day
Tomalley 6 g/day
165 g/day
1 g/day
1 g/day
2.1 g/day 1.7 g/day
— 0.4 g/day
(a) Breakdown of tomalley versus other edible lobster tissue
based on MDMF, unpublished data.
la. Maximally Exposed Individual, Mixed Seafood Diet from Quincy
Bay.
This represents a potential group of local residents
(likely small) who consume an average of 165 g/day of locally
caught seafood. This group typically would include individuals
who, for economic reasons catch a large amount of seafood for
home consumption. A local individual could catch and consume
this amount of flounder, Quincy Bay clams (illegally dug) and
Quincy Bay lobster as a recreational fisherman in the normal
course of the typical fishing seasons. The whole lobster,
including tomalley, is assumed to be eaten in this diet. it is
assumed to be available and consumed within the practical limits
of reported catch rates (see above) and season imposed by a
10-pot license. The distribution among the three seafood
33
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categories reflect the understanding gained concerning potential
recreational catch levels for all species. Because lobster
tissue residues for the study chemicals were higher than those in
the other two species, the amount of lobster assigned to this
mixed diet was estimated first, based on assumptions of
availability and catch success within the practical limits of the
reported catch rates and season imposed by a 10 pot license (see
above). Next, the assignment of clam consumption levels was
based on some discussion of maximum consumption of this species
with local fishermen and health officials. Finally, the
remainder of the 165 g/day total was assumed to consist of
flounder, on the basis of interviews with local residents
indicating that such a level of consumption likely took place.
Ib. Maximally Exposed Individual, flounder only diet from Quincy
Bay.
This represents a group of individuals (likely small) who
consume an average of the 99.9 percentile value of 165 g/day of
seafood, in this case, of Quincy Bay flounder. This could be
represented by either local or out-of-state flounder fishermen
who keep large enough amounts of caught flounder for year-round
home consumption.
34
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2a. Typical Local Consumer
This group represents those metropolitan Boston area
residents who consume the regional averages of 1 g/day of Quincy
Bay flounder and 2.1 grams/day of lobster meat, in this case
assuming that both came consistently from Quincy Bay. It is
first assumed that these consumers eat lobster without tomalley,
as many individuals do not consume this organ. This typical
local consumer is assumed to have no access to the small number
of Quincy Bay clams that may be available in the area.
2b. Typical Local Consumer
This group would be the same as (2a) above, except these
individuals do consume the lobster tomalley as well.
Clearly any of the above groups could consume fish from
other sources, or other species from the bay. A more detailed
survey of recreational fishing and local consumption patterns
conducted over a full year would allow some estimate of the size
of each of the populations affected and could allow a better
sensitivity analysis based on the more typical consumption
patterns. In the absence of such data, the figures in Section VI
were developed to illustrate sensitivity of some of the risk
estimates to the assumptions about the amounts of seafood
consumed.
35
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V. Public Health Evaluation
To determine whether adverse health effects are likely from
exposure to contaminated Quincy Bay Seafood, exposure scenarios
for maximally exposed individuals and for typical Quincy area
residents, who ingest average amounts of seafood, have been
developed. The method used to calculate dose, hazard index, and
the plausible upper limit of excess cancer risk follows guidance
provided by PTI (1987).
A. Dose Calculation
The human dose of a specific chemical from ingestion of
Quincy Bay seafood is calculated as:
(Cij) (CRj)
= Dij
BW
Where,
Cij = Concentration of contaminant i in species j
(units: pg/gram tissue, wet weight)
CRj = Consumption rate for species j
(units: grams seafood/day)
BW = Average American body weight
(units: kilograms)
Dij = dose of contaminant (i) from ingestion
of species j (units: vg/kg/day)
The concentrations of chemicals in seafood were obtained
from the EPA study of Quincy Bay chemistry results (Gardner and
Pruell. 1987). Maximum and mean contaminant levels detected
were used to calculate dose. Following the EPA Guidance,
36
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(PTI. 1987) the mean concentrations were calculated assuming the
detection limit for undetected observations, and recalculated
assuming a value of zero where the concentration was below the
limit of detection. The results where zeros were used to
calculate means differed very little from those where the
detection limits were used, and are presented in Appendix B
(Tables B-5 through B-8, and B-10).
The consumption rates used to calculate dose are presented
in Section IV. Different consumption rates were assumed for the
maximally exposed and average exposed individuals. The dose
calculations were made utilizing the standard assumptions for an
integrated risk analysis, including exposure over an entire 70-
year lifetime and a 70 kilogram body weight for an average
American adult male. In addition, it was assumed in accordance
with EPA Guidance (PTI. 1987) that the ingested dose is equal to
the absorbed contaminant dose, and that cooking has no effect on
the contaminants.
B. Risk Characterization
To calculate the plausible upper bound to excess lifetime
risk of cancer, the contaminant-specific dose is multiplied by
the carcinogenic potency factor (CPF) for oral exposures. This
equation is considered valid only at low risk levels where it is
assumed that the slope of the dose response curve is linear and
equal to the CPF. To indicate the level of non-carcinogenic
risk, the ratio of calculated contaminant-specific dose to the
37
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reference dose (RfD) is presented. In addition, in some cases, a
sum of the hazard ratios for similarly acting chemicals is
calculated. The CPF and RfD values used in this assessment are
values established by the EPA (Table 2) and described in greater
detail in Section III.
The plausible upperbound excess lifetime risk of cancer
associated with the estimated exposure is expressed as a fraction
(e.g. 1 * 10~6 or 1 in 1,000,000). It represents the estimated
incrementally increased risk in an individual's lifetime of
developing cancer attributable to the exposure. In this
assessment, incremental excess lifetime cancer risks from the
various seafood contaminants were assumed to be numerically
additive in accordance with the Guidance Manual (PTI. 1987).
Chemical-specific cancer risks were thus used to calculate total
plausible, upperbound excess lifetime cancer risks adding across
species and species-specific cancer risks were totalled across
chemicals. Taken together, these provided the basis for
estimating total plausible, upperbound excess lifetime cancer
risks from exposure to Quincy Bay seafood.
The hazard ratio is a ratio of calculated dose to reference
dose. Hazard ratios are summed across similarly acting
t
chemicals. Since the reference dose is defined as the level
unlikely to cause significant adverse health effects associated
with a threshold mechanism of action in humans exposed for a
lifetime, a sum of hazard ratios of less than one indicates that
overall the calculated dose is less than the RfD, and adverse
38
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health effects from this exposure are not likely. A sum of
hazard ratios of greater than one indicates that adverse health
effects may occur from the exposure, however, does not by itself
indicate that adverse effects will occur as there are margins of
safety and/or uncertainty in the derivation of the RfDs upon
which the ratios are based. Margins of safety or safety factors
are generally multiples of 10, each representing a specific area
of uncertainty in the available data. Three types of uncertainty
to which a factor of 10 are often applied are: (1) expected
differences in responsiveness between humans and animals in
prolonged exposure studies, (2) the variability among individuals
within the human population, (3) incomplete data (USEPA. 1986a).
Following the Guidance Manual (PTI. 1987) and generally
accepted practice, chemical-specific hazard ratios were assumed
additive only where the contaminants act on the same target
organ. Species-specific hazard ratios are additive for the same
contaminant, so a hazard ratio for a given chemical in flounder
can be added to a hazard ratio for the same chemical in lobster
to determine the total hazard ratio for one chemical from
ingestion of both flounder and lobster.
C. Maximally Exposed Individual
Exposure scenarios were developed to evaluate the risk from
eating Quincy Bay seafood by two types of maximally exposed
individuals (MEI). The first is a person who consumes an average
of 165 grams of Quincy Bay seafood each day which consists of
39
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flounder, clams, and lobster. This MEI would eat both the
lobster tissue and the hepatopancreas or tomalley. Calculations
of dose, hazard ratio, and plausible upperbound increased
lifetime cancer risk for each of the different seafoods consumed
are presented in Appendix B, Tables B-l, B-2, B-3, and B-4. A
summary of hazard ratio and plausible upper bound increased
cancer risk values (Table 5) documents that the only individual
species- and chemical-specific hazard ratio that exceeds one is
the hazard ratio for PCBs. When the species specific hazard
ratios are summed, the hazard ratio for exposure to the maximum
concentration of chlordane is also larger than one and the hazard
ratio for exposure to maximum and mean concentrations of PCBs are
67 and 43 respectively. Most (79 percent) of the calculated PCB
hazard is associated with exposure to the lobster
hepatopancreas. Since the critical effect for PCBs is reduced
size of offspring (Table 2), and no other indicator chemical in
this study has a similar critical effect, this hazard ratio also
serves as a hazard index.
The largest part (69.2 percent) of the chlordane hazard
(Table 5) comes from exposure to the flounder portion of the
diet. The critical effect for establishing the RfD for
chlordane is liver necrosis. Since some of the other chlorinated
organics also affect the liver, the hazard ratios for chlordane,
DDT, HCB and HCH were added. The RfDs for the metals and PCBs
are not based on adverse effects on the liver. Thus, the hazard
ratios for metals and PCBs are not included in this hazard ratio
40
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total. Summing the maximum ratios across the organic compounds
listed above results in a total hazard ratio sum of 1.92. The
sum of the mean hazard ratios for organic compounds is less than
one. Exposure to seafood contaminated at the maximum level
detected in Quincy Bay samples may result in adverse non-
carcinogenic health effects to the maximally exposed individual
if the dose received not only exceeds the reference dose but
actually reaches a level corresponding to a health effects
threshold. The RfD for PCBs of 1 * 10"^ mg/kg/day is based on a
no observable adverse effects level (NOAEL) in monkeys with an
intake of 0.01 mg/kg/day divided by an uncertainty factor of 100
made up of uncertainty factors for interspecies (10) and
intraspecies (10) extrapolation. The critical effect of smaller
offspring size (Table 2) was seen in monkeys with an intake of
0.4 mg/kg/day. The uncertainty associated with the determination
of this RfD (LOAEL/RFD=4,000) indicates that while exceeding the
RfD by the amount indicated in Table 2 (i.e. by a factor of
between 43 and 67 times) may increase the probability of an
adverse health effect, there is no basis for expectation of a
specific adverse non-carcinogenic response.
The range of estimated total upper limit increased cancer
risk for this maximally exposed individual (Table 5} is 1.5 * 10"
^ to 2.3 * 10~2, based on exposure to mean and maximum
concentrations of contaminants. These numbers are estimates of
the plausible upper bound of lifetime cancer risk and may not
represent the actual risk. The largest increased lifetime cancer
42
-------
risks are primarily (82 percent) associated with consumption of
the tomalley (hepatopancreas) and secondarily the flounder
(14 percent) and lobster meat (3.4 percent). As shown in Table
6, the contaminants contributing the largest portion
(approximately 75 percent) of the excess cancer risk are
polychlorinated biphenyls (PCBs), followed by polycyclic aromatic
hydrocarbons (PAHs) (about 20 to 25 percent).
The hazard ratios and cancer risks calculated for the mean
levels where undetected observations are assumed equal to zero
(Appendix B, Table B-5) are essentially the same as the values
calculated where undetected observations are set equal to the
detection limit (Table 5).
The risk characterization for a maximally exposed individual
consuming 165 grams per day of Quincy Bay flounder and no other
Quincy Bay seafood (Table 7 and Table B-6 in Appendix B)
indicates that both the hazard ratios for ingesting flounder
contaminated with chlordane and PCBs exceed one. Summing the
indices across chlordane, DDT, HCB, and HCH as discussed
previously, results in a hazard ratio total or index of 1.58
associated with adverse health effects on the liver for the
maximum contaminant level and 0.18 for the average contaminant
level. The hazard ratios associated with exposure to PCBs are
17.51 and 6.44 for maximum and mean concentrations respectively,
indicating that adverse noncarcinogenic health effects may occur
from exposure to the level of contamination detected in Quincy
Bay flounder.
43
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For this ME I consumer of flounder only, total plausible
upperbound increased estimated cancer risks range from 1.7 * 10~3
to 4.7 * 10"^ for mean and maximum contaminant levels
respectively. Comparison with the values of Table 5 shows that
the MEI flounder-only diet leads to a projection of between about
10 percent and about 30 percent of the estimated upper bound
cancer risk of the MEI mixed diet.
D. Typical Quincy Area Resident
Risk characterizations are presented for two types of
typical Quincy area residents (Tables 8 and 9 and B-7 and B-8 in
Appendix B). One case was based on the assumption was that the
resident regularly consumes locally caught flounder and lobster
in average amounts (Table 8) without eating the lobster tomalley
(hepatopancreas). The second case (Table 9) was for the resident
who consumes flounder, lobster and tomalley.
None of the hazard ratios associated with typical ingestion
of flounder and lobster without the tomalley (Table 8 or 9) are
larger than 0.22, indicating that non-carcinogenic health effects
are not likely from ingesting seafood at the levels suggested in
the first scenario. The estimated upper bound increased lifetime
cancer risks range from 4.7 * 10~5 to 8.4 * 10~5 for exposure to
mean and maximum levels of contamination for individuals who cc
not eat tomalley.
For the typical Quincy resident who eats flounder, lobster
and tomalley, the hazard ratios (Table 9) associated with maximurr.
46
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TOTAL
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and mean PCB contaminant levels are 3.73 and 2.71 respectively,
indicating non-carcinogenic health effects may occur from
exposure. The largest portion (approximately 95 percent) of the
hazard ratio comes from consuming the tomalley. The plausible
upper bound increased lifetime cancer risk levels are 9.2 * 10~4
to 1.3 * 10~3, with PCBs contributing approximately 74 percent of
the risk.
49
-------
VI. Conclusions and Uncertainties
Risk Comparison
Tabel 10 shows a summary comparison of the total upperbound
estimated cancer risks in this study by consumption scenario and
by type of seafood. A comparison of the estimated upperbound
increased lifetime cancer risks of consuming Quincy Bay seafood
with other estimated lifetime cancer risks from eating and
drinking (Table 11) shows that some of the cases analyzed in this
study result in risk estimates considerably higher than those
estimated by others from other types of activities. In
particular, the estimated incremental risk (plausible upper
bound) for the hypothetical Maximally Exposed Individual eating a
mixed diet of Quincy Bay seafood is about 50-100 times higher
than the estimates for any of the other typical eating and
drinking activities shown on the table. The estimated risk
(plausible upper bound) for just consumption of Quincy Bay Winter
flounder is about 10 times higher than the levels estimated for
the other eating and drinking activities.
50
-------
TABLE 10. UPPER BOUND ESTIMATED LIFE TIME
CANCER RISKS FROM CONSUMPTION OF QUINCY BAY SEAFOOD
Clams
Flounder
Lobster
Meat
Tomalley
Total Risk
Maximally
Mixed Diet
2.1*10~4
(1%)
3.2*10~3
(14%)
8.0*10~4
(3.4%)
1.9*10~2
(82.6%)
2.3*10~2
Exposed Individual
Flounder
4.7*10~3
(100%)
-
-
4.7*10~3
Typical Exposed
Mixed Diet
2.8*10~5
(33%)
5.6*10~5
(67%)
-
8.4*10~5
Individual
Mixed Diet
2.8*10"5
(2.2%)
4.5*10"5
(3.5%)
1.2*10"3
(92.3%)
1 3*10~3
51
-------
TABLE 11. COMPARISON OF ESTIMATED LIFETIME CANCER RISKS
(PLAUSIBLE UPPER LIMIT)
Lifetime Cancer Risks From
Eating and Drinking Activities Estimated Lifetime Risks(a)
• Maximally Exposed Individual - 1.5 to 2.3 * 10~2
mixed diet of Quincy Bay seafood
• Maximally Exposed Individual - 1.7 to 4.7 * 10~3
diet of Quincy Bay winter flounder
• Typical Quincy area resident - 9.2 * 10~4 to
mixed diet of Quincy Bay seafood, 1.3 * 10~3
including lobster tomalley
• Four Tablespoons peanut butter per day 5.6 * 10~4
• One 12 1/2 ounce diet drink per day (6) 7.0 * 10~4
• Average saccharin consumption in the 1.4 * 10~4
United States
• One pint milk per day(b) 1.4 * 10~4
• Typical Quincy area resident - 4.7 * 10"^ to
mixed diet of Quincy Bay seafood 8.4 * 10 5
without lobster tomalley
• Miami or New Orleans drinking water 7.0 * 10
• 1/2 Ib. charcoal broiled steak per week 2.1 * 10~7
(cancer risk only; heart attack and
other risks additional)
(a) Except for Quincy Bay seafood consumption estimates for sub-
populations, all other estimates are averaged over the whole
population of the United States, assuming a 70 year lifetime.
(b) Based on human data for aflatoxin carcinogenicity. Note that it
is assumed that the measured aflatoxins are aflatoxin B, the most
potent. If some corresponds to other aflatoxins, these estirr.atef
risks should be reduced.
Sources: modified from Meta Systems, Inc. 1986.
Note: Meta Systems Inc., (1986), modified the original annual
risk estimates from Crouch and Wilson, (1982), to represent
estimated lifetime risks.
52
-------
The estimated lifetime risk for the hypothetical "typical" local
resident consumer of a mixed Quincy Bay seafood diet including lobster
tomalley is about two to ten times higher than the estimate for the
other eating and drinking activities. Note that without lobster
tomalley, the estimated risks for the hypothetical typical Quincy area
consumer of Quincy Bay seafood drop into the 10"^ range corresponding
to the risks of the other illustrated eating and drinking
activities.
In work done by the Canadian Government, (Environment Canada.
1987), the estimated dietary intake of PCBs was calculated for a
variety of food items. The calculations were based on a mixture of
measured PCB residues for most food items and the assumed presence of
maximum allowed PCB residues in fish. These data are presented in
Table 12 with the data used in this public health evaluation to
provide a comparison of how PCB intake from fish compares with PCB
intake from other food sources. Under any of the consumption
scenarios documented by the Penn State (1985) report, the Canadian
studies, or otherwise assumed in this study, more than half of the
total exposure to PCBs comes from seafood consumption. In the case of
the ME I for this study, estimated PCB exposure from seafood is more
than 20 times higher than that estimated by the Canadian data from all
other dietary sources combined.
53
-------
TABLE 12. SOURCES OF INTAKE OF PCBs
Food
Food
Intake(a)
(g/person-day)
Maximum
Residue
Leve?-v, A
(yg/g)(h)
PCB
Intake
(ug/person-day)
Canadian data (b)
dairy 32.8
meat 48
poultry 3.6
eggs 34
fish 20
Quincy Bay seafood
maximally exposed
individual 165
typical Quincy area
resident 3.1
0.2(d)
0.2(d)
0.5(d)
O.l(e)
2(f)
(9)
(9)
6.6
9.6
1.8
3.4
40
470
26
(a) Based on statistics Canada uses for disappearance of foods frorr.
the marketplace.
(b) Reference for Canadian data: Environment Canada. 1987.
(c) Includes milk, cheese, and butter.
(d) Fat basis.
(e) Whole weight minus shell.
(f) Edible portion, assumed based on maximum residue level allowed
(g) Varies by different kind of seafood, see Table 1.
(h) Based on measured residues for all Canadian data except fish.
Fish value based on maximum allowed.
54
-------
Uncertainties
Extreme caution must be exercised in the interpretation and
use of any risk data due to a variety of uncertainties. Sources
of uncertainty in this risk assessment are discussed individually
below.
1. Representativeness of the measured values for
contaminants in seafood. Comparisons of the PCB values
obtained by Gardner and Pruell (1987) with other data
for the same species in Quincy Bay and other parts of
Boston Harbor (Table 13) suggest that the 1987 EPA
values are representative for Quincy Bay, given the
differences in sample locations and analytical
methodologies of the various studies. Preliminary
results of ongoing studies involving inter-laboratory
calibration of EPA, MDMF and FDA methods of determining
PCB concentrations in various edible portions of lobster
indicate that differences among the agency analytical
techniques are likely not significant.
2. Use of standard risk assessment assumptions. Many of
the assumptions used in this risk assessment are
standard risk assessment assumptions chosen to be
55
-------
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conservative, albeit uncertain. These include the
following assumptions: (1) Probable Human Carcinogens,
(Group B2, where human evidence of carcinogenicity is
limited or inadequate but animal evidence of
carcinogenicity is available), contribute to estimated
increased cancer risks; (2) related compounds such as
PAHs or PCB congeners, can be evaluated by the assumed
toxicity of the more toxic compounds for which data are
available (e.g. B(a)P for the PAHs) and (3) ingested
doses of contaminants are totally absorbed. Since
these assumptions generally apply to the other types of
contaminant risk assessments conducted by the EPA, their
use was considered valid as an initial reference point
for this study. It is acknowledged that not all PAHs
are known carcinogens (see section III). However, all 5
of the PAHs rated as having sufficient evidence of
animal carcinogenicity (Table 3) were detected in Quincy
Bay seafood in this study. These five compounds
comprised up to about ten percent of the total PAHs in
some of the organisms analyzed (Gardner and Pruell.
1987). Given this detection and the lack of importance
of PAHs versus PCBs in the total risk calculation (see
table 6) the effect of treating all PAHs as carcinogens
in this study was minor.
Note that for PCBs there are different CPFs used by
different agencies. Agencies have recognized the need
57
-------
for congener-specific t-i-. ^. for PCBs and are working
towards development of such numbers. Tne CPF used in
this assessment was developed by EPA based on laboratory
experiments using Aroclor 1254. The CPF for Aroclor
1254 was used because the congener mix detected in
Quincy Bay seafood more closely resembled Aroclor 1254
than Aroclor 1260. Appendix C documents this
derivation.
A sensitivity analysis was performed to determine
the effect of varying the CPF for PCBs, from 0.22
(mg/kg/day)"1 to 7.7 (mg/kg/day)~^ on the calculation of
plausible upper bound lifetime cancer risk. The lower
CPF value cited has been used by New York State to
evaluate PCB levels in fish (Bro. 1987). The 7.7 value
is proposed by EPA to replace the current US EPA CPF
value of 4.34 based on the carcinogenicity of Aroclor
1260. Figures 5 and 6 show the effect on the risk
calculations of using different CPF assumptions for
PCBs. For the maximally exposed individual the
plausible upper bound increased lifetime cancer risk is
4.7 x 10"^ using a CPF of 0.22 (mg/kg/day)"1 and average
contaminant levels, and 3.7 x 10~2 for a CPF of 7.7
(mg/kg/day)'1.
3. Assumption that cooking does not change contaminant
levels. This assumption is recommended in EPA guidance
(PTI. 1987) for seafood consumption risk assessments.
The same authors acknowledge that the assumption may not
58
-------
Fig. 5. EFFECT of CPFs for PCBs
3.8E-02
lrdvUual-Uu«d Diet
o
1
I
3.«E-02 -
3.4E-02 -
3.2E-02 -
3.0E-02 -
2JE-O2 -
2.«E-02 -
2.4E-02 -
2-2E-02 -
2.0E-02 -
1JE-02 -
1.«E-Q2 -
1.4E-Q2 -
1-3E-02 -
1.CE-02 -
8.CE-Q3 -
6.0E-O3 -
4.0E-03
2 4
CfT f cr PCB: (mg/kg/dy)- 1
KEY
A. Uted by N.T. State in • stud/ on PCB* In fish (Bre tt «l. 1987).
B. USEPA OHEA developed this value for thU study (USEPA. 1988*).
C. From the Superfmd Public Neslth Evslustion Mnusl
(USEPA.19866).
D. Value developed by USEPA based on the carcinogen*city of
Aroclor 1260 (USPHS. 1987).
59
-------
2.4E-O3
Fig. 6. EFFECT of CPFs for PCBs
•ftp. QUncy NftK-UiMad DM */
1JE-Q3 -
e.QC-O4 H
I
a
T
e
PCfh (mg/'kg/dey)-1
KEY
A. U»ed by N.Y. State In • study on PCi» in ff»h (Iro «t «l. 1987).
B. USEPA ONEA developed this vilue for thU study (USEPA. 1986*).
C. From the Superfund Public Health Evaluation Manual
(USEPA.1986b).
D. Value developed by USEPA bated on the carcinogen!city of
Aroclor 1260 (USPHS. 1987).
60
-------
be valid in all cases, citing, for example, that there
have been discussions of possible decreases in
concentrations of PCBs in cooked versus uncooked samples
of Great Lakes salmonids. However, there is no evidence
to support alternative assumptions in this case. Also
of interest is the possibility that some of the
contaminants in lobster tomalley may be released by
cooking, thereby becoming potentially available to
affect (increase) the concentration in other edible
lobster tissue being cooked in the same vessel. Further
investigation of this uncertainty by sampling and
analysis of uncooked and cooked lobsters would resolve
this uncertanity.
4. Affected population size and consumption patterns. As
noted in Section IV of this report, estimates of the
actual size of affected populations were not made in
this study due to the necessary reliance on a fall-
winter study period.. Some of the data obtained from
earlier consumption surveys (Penn State. 1985) may need
to be checked. For example, the reported average
regional lobster consumption values may be high if the
respondents described their consumption in terms of
whole lobster rather than edible lobster tissue, and if
the researchers failed to adjust the reported values.
Figures 7, 8 and 9 show the sensitivity of the upper-
bound increased cancer risk calculations to the assumed
61
-------
1.7E-03
FIG, 7. FLOUNDER CONSUMPTION EFFECTS
StnrffivOy tna/fyftm 1OS
160
FLOUNDER INGESTION
Cor»unption
Typictl Local Concimer
M«x. Exp. Ind. • Mixed Dltt
M*x. Exp. Ind. - Floinder Only
GriM/Dty
1
113
165
Ri»k
1.0E-05
1.2E-03
1.7E-03
62
-------
FIG. 8. TOMALLEY CONSUMPTION EFFECTS
Stnrifivfty
O.CE+00
Consumption Scenario
HEPATOPANCftEAS INGESTION
6r«M/D«y
Typic»i local Conauner
Maxim*Ily Exposed Individual
0.4
.6
Mean Riak
8.8E-W
1.3E-02
63
-------
6.CE-O4
FIG. 9. LOBSTER CONSUMPTION EFFECTS
Arolyvii
O.OE+00
Meat IngKtian (g/dcy)
LOBSTER NEAT CONSUMPTION
Consumption Scenario Gr«M/D*y
Typicil Local Concuner (With Tonal ley) 1.7
Typical Local Consuner (Without ToMllcy) 2.1
Max. Exp. Ind. • Mixed Diet 30
Mean Rick
3.0E-05
3.7E-05
5.2E-04
64
-------
cornsumption rates for flounder, lobster tissue and
lobster tomalley.
5. Other sources of the same contaminants. The study
results (Tables 5, 7, 8 and 9) indicate that PCB and/or
chlordane residues in Quincy Bay flounder may constitute
a significant fraction of threshold-based, non-
carcinogenic health risks if taken in combination with
other sources of exposure of the same chemicals.
Estimation of total risks due to PCB and chlordane
exposure would require a specially focused
investigation, but is feasible.
In summary, several areas of uncertainty remain, some of
which have been assessed above by sensitivity analysis. The
results of the sensitivity analyses do not change the conclusions
stated earlier regarding human health risks.
65
-------
VII. References
Ayers, Andrew. November, 1987. (Shellfish Constable). Quincy
Health Department. Personal communication.
Boehm, Paul D., William Steinhauer, John Brown. 1984. Organic
Pollutant Biogeochemistry Studies in the Northeast U.S.
Marine Environment. Battelle, MERL for National Oceanic and
Atmospheric Administration. (NA-83-7A-C-00022).
Bro, Kenneth. 1987. Relative Cancer Risks of Chemical
Contaminants in the Great Lakes. Environmental Management
Vol. II, No. 4. pp. 495-505.
Childs, Abigail. November, 1987. Massachusetts Division of
Marine Fisheries. Personal communication.
Clayton and Clayton. 1981. Patty's Industrial Hygiene and
Toxicology. Volume IIB. New York, NY. 2879-3769 pp.
Clement Associates, Inc. 1985. Chemical, Physical and
Biological Properties of Compounds Present at Hazardous Waste
Sites. Prepared for U.S. Environmental Protection Agency.
Connerty, Ray. 1987. (Fisherman). Personal communication.
Crouch, E.A.C. and Richard Wilson. 1982. Risk/Benefit
Analysis. Ballinger Company. Cambridge, MA.
Estrella, Bruce. November, 1987. Massachusetts Division of
Marine Fisheries. Personal communication.
Environment Canada. 1987. Summary of Environmental Criteria for
Polychlorinated Biphenyls (PCBs). Report # EPS 4/HA/l.
25 pp.
Gardner, George R., and Richard J. Pruell. December, 1987.
Quincy Bay Histopathological and Chemical Assessment of
Winter Flounder Lobster and Boston Harbor and an In Situ
Evaluation of Oysters Including Sediment (Surface and Cores)
Chemistry. USEPA ERL. Narragansett, RI. Ill pp.
Guessing, Frank. February, 1988. U.S. Food and Drug
Administration. Personal communication.
Hoopes, Thomas B. 1985. Massachusetts Lobster Fishery
Statistics, Technical Series 20. Massachusetts Division of
Marine Fisheries. Publication # 14530-21-310-8-12-86-C.R.
20 pp.
66
-------
Hoopes, Thomas B. 1986. Massachusetts Lobster Fishery
Statistics, Technical Series 21. Massachusetts Division of
Marine Fisheries. Publication I 14, 996-19-310-8-87-CR.
18 pp.
Howe, Arnold B., and Phillip G. Coates. 1975. Winter Flounder
Movementsf Growth and Mortality of Massachusetts,
Transactions of the American Fisheries Society; Vol. 104,
No. 1. 13-29 pp.
Jerome, William C., Jr., Arthur P. Chessman and Charles 0.
Anderson, Jr. 1966. A Study of the Marine Resources of
Quincy Bay. Monograph Series No. 2. Division of Marine
Fisheries, Department of Natural Resources, Commonwealth of
Massachusetts. 62 pp.
Jones, Chris. December, 1987. (Lobster fisherman). Personal
Communication.
Kennedy, Jeffrey P. November, 1987. Massachusetts Division of
Marine Fisheries. Personal communication.
Klassen. C.D., M.O. Amdur, and J. Doull. 1986. Toxicology: The
Science of Poisons. Third Edition. New York, NY. 974 pp.
Massachusetts Division of Marine Fisheries. 1987. "Special, Non-
Commercial" Lobster License List. Unpublished data.
Merck & Co., Inc. 1983. The Merck Index: An Encyclopedia of
Chemicals, Drugs, and Biologicals. Rahway, New Jersey.
1463 pp.
Meta Systems, Inc. 1986. Qualitative and Quantitative Health
Risk Assessment - Airborne Emissions from the City of Boston
Waste to Energy Facility. 74 pp.
Nash, Gerald M. 1984. 1984 Massachusetts Lobster Fishery
Statistics. Technical Series 19. Massachusetts Division of
Marine Fisheries Publication # 14 181-22-300-9-85-CR. 20 pp.
Pennsylvania State University. September, 1985. Analysis of
Seafood Consumption in the United States 1970, 1974, 1978,
1981. National Marine Fisheries Service. Washington, D.C.
95 pp.
PTI Environmental Services, Inc. September, 1987. Guidance
Manual for Assessing Human Health Risks from Chemically
Contaminated, Fish and Shellfish: Executive Summary. Draft
Report to Battelle New England Marine Research Laboratory for
USEPA Office of Marine and Estuarine Protection. 125 pp.
Sax, N.I., and R.J. Lewis, Sr. 1987. Hazardous Chemicals Desk
Reference. New York, NY. 1084 pp.
67
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Schwartz, Jack P. 1987. PCS Concentrations in Marine Fish and
Shellfish from Boston and Salem HarborSf and Coastal
Massachusetts. Massachusetts Division of Marine Fisheries
CAV Cove Marine Laboratory. Progress Report I 14, 997-36-
110-8-87-CR. 29 pp.
Tetra Tech, Inc. 1986. Guidance Manual for Health Risk
Assessment of Chemically Contaminated Seafood. Prepared for
U.S. Environmental Protection Agency, Region 10 - Office of
Puget Sound. TC-3991-07. 75 pp.
U.S. Environmental Protection Agency (USEPA). 1984. Health
Effects Assessment Documents.
a) Cadmium. PB86-134491
b) Chlordane. PB86-134343
c) Chromium. PB86-134301
d) Copper PB86-134368
e) DDT. PB86-134368
f) Hexachlorobenzene. PB86-134285
g) Lead. PB86-134665
h) Lindane. PB86-134673
i) Mercury PB86-134533
j) Polycyclic Aromatic Hydrocarbons. PB86-134244
k) Polychlorinated Biphenyls. PB86-134512
U.S. Environmental Protection Agency. 1986. Integrated Risk
Information System (IRIS). Chemical Abstracts.
a) Chromium III CAS No.: 16065-83-1
b) Lindane CAS No.: 58-89-9.
U.S. Environmental Protection Agency. 1986c. Superfund Public
Health Evaluation Manual. EPA/540/1-86/060. Washington,
D.C. 175 pp.
U.S. Public Health Service (USPHS). 1987. Toxicological Profile
for Selected PCBs (Aroclors - 1260-1254, -148, -1242, -1232,
-1221, and -1016). Draft for Public Comment. Syracuse
Research Corporation. Contract No. 68-03-3228. 136 pp.
U.S. Environmental Protection Agency (USEPA). 1987(a). Health
Advisories for Legionella and Seven Inorganics. PB87-
235586. Washington, D.C. 125 pp.
U.S. Environmental Protection Agency (USEPA. 1987(b). Health
Advisories for 16 Pesticides. PB87-235586. Washington,
D.C. 262 pp.
68
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U.S. Environmental Protection Agency (USEPA).
Advisories for 25 Organics . PB87-235586.
397 pp.
1987(c). Health
Washington, D.C.
U.S. Environmental Protection Agency (USEPA). 1987. Integrated
Risk Information System (IRIS). Chemical Abstracts.
d) Cadmium
e) Chlordane
f) Chromium VI
g) DDT
h) Lindane
CAS No. :
CAS No.:
CAS No.:
CAS No.:
CAS No.:
7440-43-9
57-74-9
7440-47-3
50-29-3
58-89-9.
U.S. Environmental Protection Agency (USEPA). 1988(a).
Memorandum from Dr. W. Farland, Acting Director, Office of
Health and Environmental Assessment to K. Kipp, Quincy Bay
Coordinator Region I EPA.
U.S. Environmental Protection Agency (USEPA). 1988(b).
Memorandum from S. Braen Norton, Exposure Assessment Group to
Dr. W. Farland Acting Director, Office of Health and
Environmental Assessment.
U.S. Environmental Protection Agency (EPA). 1988(c). Memorandum
from J. Cogliano, Carcinogen Assessment Group, to
K. Garrahan, Exposure Asessment Group.
69
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Appendix A
Toxicity Profiles
-------
TABLE Al. EPA WEIGHT-OF-EVIDENCE
CATEGORIES FOR POTENTIAL CARCINOGENS
EPA Description
Category of Group
Description of Evidence
Group A Human Carcinogen
Group Bl Probable Human
Carcinogen
Group B2 Probable Human
Carcinogen
Group C Possible Human
Carcinogen
Group D Not Classified
Group E No Evidence of
Carcinogenicity
in Humans
Sufficient evidence from
epidemiologic studies to support a
causal association between exposure
and cancer
Limited evidence of carcinogenicity
in humans from epidemiologic studies
Sufficient evidence of carcinogenic
ity in animals, inadequate evidence
of carcinogenicity in humans
Limited evidence of carcinogenicity
in animals
Inadequate evidence of carcinogenic-
ity in animals
No evidence for carcinogenicity in
at least two adequate animal tests
or in both epidemiologic and animal
studies
Source: USEPA. 1986 (a).
A-l
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TABLE A2. RATING CONSTANTS (RVs) FOR NONCARCINOGENS ( a
_ Effect _ Rating (RVs)
Enzyme induction or other biochemical change 1
with no pathologic changes and
no change in organ weights.
Enzyme induction and subcellular proliferation 2
or other changes in organelles but no
other apparent effects.
Hyperplasia, hypertrophy or atrophy, but no 3
change in organ weights.
Hyperplasia, hypertrophy or atrophy with changes 4
in organ weights.
Reversible cellular changes: cloudy swelling, 5
hydropic change, or fatty changes.
Necrosis, or metaplasia with no apparent 6
decrement of organ function. Any neuropathy
without apparent behavioral, sensory, or
physiologic changes.
Necrosis, atrophy, hypertrophy, or metaplasia 7
with a detectable decrement of organ functions.
Any neuropathy with a measurable change in
behavioral, sensory, or physiologic activity.
Necrosis, atrophy, hypertrophy, or metaplasia 8
with definitive organ dysfunction. Any neuropathy
with gross changes in behavior, sensory, or
motor performance, Any decrease in repro-
ductive capacity, any evidence of fetotoxicity .
Pronounced pathologic changes with severe 9
organ dysfunction. Any neuropathy with loss
of behavioral or motor control or loss of
sensory ability. Reproductive dysfunction.
Any teratogenic effect with maternal toxicity.
Death or pronounced life-shortening. Any teratogenic 10
effect without signs of maternal toxicity.
(a) Rating scale identical to that used by EPA in the RQ
adjustment process, as described in US EPA (1983).
Source: USEPA. 1986 (a).
A-2
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TOXICITY PROFILE FOR: CADMIUM
BACKGROUND INFORMATION
Cadmium is a soft metal, and is found naturally occurring in zinc
ores. This element often serves as a constituent of easily
feasible alloys, amalgam in dentistry, electrodes for cadmium
vapor lamp, batteries, color pigment, electroplating and
photometry of ultraviolet rays (Merck. 1983).
Cases of acute industrial cadmium poisoning date as far back as
the 1920's. The first definite reports of chronic effects due to
industrial cadmium exposure date to the late 1940's. It was not
until the 1960's that health effects were noted due to cadmium
associated with environmental pollution, when the Itai-Itai
disease complex in Japan was linked to rice paddy contamination
by smelter wastes (USEPA. 1980a).
The population, in general, is exposed to cadmium through
drinking water and food. For the vast majority of the U.S.
population, ambient air is not a significant source of cadmium
exposure (USEPA. 1980a). A major non-occupational source of
cadmium exposure is derived from cigarettes (Klaassen. 1986).
TRANSPORT & FATE
Cadmium reaches surface water in municipal effluents, and
effluents from pigment, plastics, alloys and other manufacture.
Fallout from air is also a source of cadmium in water (USEPA.
1980a) .
Cadmium is relatively mobile in water, compared with other
metals, and may be transported as hydrated cations or as organic
or inorganic complexes. In saltwater (typical salinity) the
number of probable cadmium species is reduced to a few, with
cadmium chloride complexes likely predominant (USEPA. 1980a).
A-3
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Cadmium is strongly adsorbed to clays/ muds, humic and organic
materials. In polluted waters, cadmium complexing with organic
materials is an important fate/transport process. Sorption
processes account for removal of dissolved cadmium to sediments
(USEPA. 1980a).
Cadmium does bioaccumulate in aquatic organisms and evidently is
eliminated slowly. A high degree of variability exists among the
BCFs reported for saltwater organisms. Fish and most shellfish
bioconcentration factors were generally lower than the uptake
factors for bivalves examined. The latter organisms however, are
noted as not reaching a steady-state with water concentrations
(USEPA. 1980a). The visceral meat of terrestrial organisms
(liver, kidney, pancreas) are noted as organs that bioconcentrate
cadmium (USEPA. 1980a). Lobster hepatopancreas (analogous
structure) in this study had higher residues of cadmium than
muscle tissue.
HEALTH EFFECTS
The major non-occupational routes of human exposure to cadmium
are through food and tobacco smoke. It is estimated that
approximately 5% of cadmium is absorbed by the human gastro-
intestinal tract. This is less efficient than uptake across
pulmonary membranes.
The major effects of long-term oral exposure to cadmium in humans
include: increased proteinuria and renal dysfunction, which
results in kidney stone formation and mineral metabolism
disturbances (USEPA. 1984a). The US EPA (1980) estimated a
Lowest Observed Effect Level (LOEL) of 228 ug Cd/day, based upon
the human dietary intake of contaminated rice from areas of Japan
in which itai-itai disease is prevalent. Since chronic renal
dysfunction occurs approximately at this intake level, the
kidneys are the critically affected organ (USEPA. 1984a).
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This element has been demonstrated to be teratogenic and can
reduce fertility, following intravenous, intraperitoneal, and
subcutaneous administration in rats (USEPA. 1984a).
The calculated reference dose for cadmium is 2.9 x 10~4
(mg/kg/day) (USEPA. 1984a).
Based on exposure to cadmium by inhalation, cadmium has been
classified as a Group Bl, Probable Human Carcinogen. There is no
conclusive evidence that cadmium is carcinogenic following oral
exposures.
A-5
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TOXICITY PROFILE FOR: CHROMIUM
BACKGROUND INFORMATION
Chromium is a metal that exists in four naturally occurring
isotopes (Merck. 1983). It is a relatively rare element which
occurs naturally in the earth's crust.
Among the uses of Chromium VI are the manufacture of chrome-steel
or chrome-nickel-steel alloys (stainless steel). Chromium salts
are also contained in paints and pigments, and are utilized in
the plating and leather tanning industry (USEPA. 1987d).
The adverse effects on skin of high level exposure to chromium in
industrial exposure have been known for a century (USEPA.
1980b). The known harmful effects of chromium have been
predominantly associated with exposure to the hexavelent state
(Chromium VI) of this element (Klaassen. 1986).
TRANSPORT & FATE
Although chromium is widely distributed, it is rarely found in
significant concentrations in natural waters or air in non-urban
areas. Much of the chromium detected in air and water is
presumably derived from industrial processes. Chromium may enter
waterbodies in discharges or as fallout from airborne sources
(USEPA. 1980b).
The trivalent (CrIII) and hexavalent (CrVI) forms of chromium are
the most environmentally and biologically significant forms.
Hexavalent chromium (more widely used in industry) is very
soluble in water as a component of a complex anion. These are
readily reduced to the more insoluble trivalent chromium compound
sulfur dioxide or organic reducing matter (USEPA. 1980b) .
Chromium III only slowly oxidizes to Chromium VI. The hexavalent
form is relatively more stable in neutral or alkaline solutions
A-6
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and traces can be found. Trivalent Chromium has low solubility
in saltwater, and tends to precipitate out, being associated with
the sediments (USEPA. 1980b).
The evidence for bioconcentration of Chromium VI in fish muscle
appears to be at or below 1.0. Bivalves, on the other hand,
apparently bioconcentrate CrIII and/or CrVI. Thus shellfish
consumption may become a source of chromium in human consumption.
HEALTH EFFECTS
Chromium plays a role in human nutrition and is generally
considered essential in small amounts. Chromium levels found to
have adverse effects on humans or other test organisms are
several orders of magnitude higher than those recommended as safe
in consumed sources, including drinking water (USEPA. 1980b).
Hexavalent chromium is more toxic than trivalent chromium and
more readily taken up by cells than trivalent chromium.
Adsorption of chromium from the gut is generally poor. Once
inside cells, chromium VI is likely reduced to the trivalent
state (USEPA. 1980b).
The major acute effects from ingested chromium include renal
tubular necrosis (Klaassen. 1986). Chromium VI (chromic acid
and its salts) have a corrosive action on the skin and mucous
membranes.
Mutagenic effects by chromium have been documented. It has been
suggested that the chromium mutagenesis causative agent is
trivalent chromium bound to genetic material after its reduction
from the hexavalent form (Klaassen. 1986).
There is inadequate evidence of chromium carcinogenicity by oral
exposure and it has not been classified as a carcinogen by this
exposure route. Chromium carcinogenicity has only been shown
A-7
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through the occupational inhalation of chromium VI, where its
effects are observed in the human respiratory passages and in the
lungs (USEPA. 1987d).
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TOXICITY PROFILE FOR: COPPER
BACKGROUND INFORMATION
Copper is a soft heavy metal. Elemented copper is very reactive
with organic or mineral acids that contain or act as oxidizing
agents. Copper has two oxidation states, the cuprous and
cupric. The cuprous state is unstable in aerated water over the
pH range of most natural waters (6 to 8) and will oxidize to the
cupric state (USEPA. 1980c).
Many copper containing compounds are used as fungicides.
Medicinally, copper sulfate is used as an emetic (Klaassen.
1986).
TRANSPORT & FATE
Copper is ubiquitous in rocks and minerals of the earth's crust
and these sources are responsible for background levels of copper
in water typically below 20 vg/1. Higher levels are likely from
corrosion of brass/copper pipe, effluent and fallout from
industry and sewage treatment plant effluents (USEPA. 1980c).
Some copper compounds are highly soluble in water (copper
sulfate, chloride, nitrate), while others may precipitate out of
solution more readily (basic copper carbonate, cupric hydroxide,
oxide, or sulfide). Cupric ions are adsorbed by clays, sediments
and organic particles, or may form complexes with a number of
inorganic compounds (USEPA. 1980c).
The levels of copper in water are dependent upon water chemistry,
including pH, temperature, alkalinity and the concentration of
bicarbonate, sulfide and organic ligands. Acid conditions and
low concentrations of complexing agents favor ionic copper
solubility. Alkalinity and complexing agents reduce levels of
A-9
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cupric ions in water. Many of the various copper complexes and
precipitates appear to be largely non-toxic (USEPA. 1980c).
Copper is an essential element, especially in plant and
crustaceans. Bivalves bioconcentrate copper to the highest
levels but the highest observed are not known harmful to man
(USEPA. 1980c).
HEALTH EFFECTS
Copper is an essential element in humans. There are two
inherited diseases that represent abnormal copper metabolism. In
Menke's disease there is reduced absorption of copper, resulting
in symptoms resembling copper deficiency. In Wilson's disease,
copper accumulates in the liver and brain, resulting in copper
toxicosis (USEPA. 1980c).
Copper has toxic effects at high dose levels and is an essential
element in lower levels. Excessive ingestion of copper salts
(i.e. copper sulfate) may result in acute poisoning and
eventually produce death. Symptoms such as vomiting,
hypotension, coma and jaundice are particular to acute copper
poisoning. The use of copper containing dialysis equipment and
burn treatment with copper compounds has also produced hemolytic
anemia (Klaassen. 1986).
No evidence of human teratogenesis associated with oral exposure
has been reported by the US EPA. Data regarding the
carcinogenicity of copper were not sufficient to rate this
element adequately, therefore, it was categorized by EPA's
Carcinogen Assessment Group as a group D, Not Classified
substance.
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TOXICITY PROFILE FOR: LEAD
BACKGROUND INFORMATION
Lead is a ubiquitous soft gray acid-soluble metal that exists in
three oxidation states. Lead is widely used in industry because
of its high density, softness, resistance to corrosion and
radiation. It has often been used in electroplating, metallurgy,
the manufacture of construction materials, radiation protection
devices, plastics, and electronics equipment, as a gasoline
additive and as a pigment in paint. (USEPA. 1980d).
Unlike many contaminants where exposures may be related to a
specific route or situation, substantial "background" lead
exposure occurs, primarily through food and water. Lead gasoline
combustion has also been a major source of environmental exposure
(USEPA. 1984g).
TRANSPORT & FATE
Lead reaches the aquatic environment through precipitation,
fallout of lead dust, sheet runoff, and both industrial and
municipal waste water discharges. (USEPA. 1980d).
Inorganic lead compounds are most stable in the +2 valence state,
while the organic lead compounds are most stable in the + 4
valence state. Neither metallic lead nor the common lead
minerals are considered soluble in water. They can be
solubilized by some acids. However, some of the lead compounds
produced in industry are considered water soluble. Natural lead
compounds typically become adsorbed by ferric hydroxide or
combined with carbonate or sulfate ions. These are insoluble in
water. The solubility of lead compounds in natural waters
depends heavily on pH. It ranges from 10,000,000 yg/liter at pH
of 5.5 to 1 yg/1 at pH of 9 (USEPA. 1980d).
A-ll
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A few available studies have shown that lead can be
bioaccumulated. The range of bioconcentration factors for
species examined was 17.5 to 2,570. The species were largely
bivalves. No saltwater fish species were examined in these
studies. (USEPA. 1980d).
HEALTH EFFECTS
Approximately 8% of the lead ingested by adults is absorbed by
the gastrointestinal tract (USEPA. 1984g). Age has a major
influence on the extent of lead absorption. It has been observed
that absorption of lead in infant rats was considerably greater
than in adults. Similar results have been seen in humans
(USEPA. 1984g). Lead is a cumulative poison which most directly
affects the blood cells (Merck. 1983). Lead tends to produce a
brittleness within the red blood cells thus causing intensified
fragility of the tissue. This phenomenon results in a faster
destruction of cells, leading to anemic symptoms (USEPA. 1984g).
Neurological effects are most common in those children having
direct contact and exposure to lead contents in paint films.
Chronic exposures to low levels of lead can cause subtle learning
disabilities in children. Among the neurological effects caused
by lead poisoning in children are alterations in cognitive
functioning, inappropriate social behavior and the inability to
focus attention (Klaassen. 1986). IQ decrements and EEC brain
wave pattern alterations were observed among those children
exposed to lead, with an average blood lead level ranging from
30-50 yg/dl (USEPA. 1984g). They also showed weight loss,
weakness and anemia (Merck. 1983).
In a multigenerational study of rats, histological changes in
kidney were noted as a sensitive indicator of liver toxicity.
Data concerning the carcinogenic potential of lead to humans
after oral exposure proved inconclusive (Clement. 1985). There
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is some animal evidence that several lead salts are
carcinogenic. Lead and lead compounds were classified by the US
EPA as a Group C, Possible Human Carcinogen.
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TOXICITY PROFILE FOR: MERCURY
BACKGROUND INFORMATION
Mercury is a silver-white metal that exists in three oxidation
states: elemental, mercurous and mercuric. It can be part of
both organic and inorganic compounds. Mercurous salts are much
less soluble than mercuric salts and are much less toxic than the
mercuric forms. (USEPA. 1980e, 1984i).
Natural degassing of the earth's crust releases mercury, although
mining, smelting and industrial discharge have contributed
greatly to the environmental pollution from mercury (Klaassen.
1986).
Mercury is used in the manufacture of mercury and incandescent
lamps, in amalgams with copper, tin, silver and gold, in
photography, paints and as a fungicide (Merck. 1983).
TRANSPORT & FATE
The atmosphere is the major pathway for distribution of
mercury. The main input is from natural sources, although input
from industry is significant. Mercury is removed from the
atmosphere mainly through precipitation. Mercury is also added
to aquatic systems through runoff and discharges (USEPA. 1980e).
At one time elemental mercury was considered relatively inert,
and was thought to settle to the bottom of a water body and
remain inoccuous. It is now known that elemental mercury can be
oxidized in sediments to divalent mercury. Both aerobic and
anaerobic bacteria can methylate divalent mercury in sediments,
with the reverse reaction occurring very slowly. Evidently, the
slime coat and intestines of fish can methylate mercury. Methyl
mercury is both directly toxic and bioaccumulates. It is more
A-14
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toxic to mammals than inorganic mercury. Uptake of methyl
mercury is extremely rapid. These compounds rapidly cross cell
membranes and bind to ligands in tissue - importantly, in muscle
tissue (the part of fish consumed by man). Depuration by
excretion evidently requires demethylation, a slow process. This
is evidently responsible for mercury's biological half-life of 2-
3 years and high bioconcentration (up to 40,OOOX reported for
oyster) (USEPA. 1980e).
HEALTH EFFECTS
The main sources of human mercury exposure are methylmercury
compounds in the food supply and mercury vapor in the atmosphere
of occupational settings.
Metallic mercury (inorganic form) appears to be poorly absorbed
from the GI tract as demonstrated by a study in which animals who
ingested gram quantities of mercury only absorbed 0.01 percent of
the element. Methylmercury (alkyl form of mercury), however, was
essentially completely absorbed in volunteers who consumed tuna
contaminated with the compound (USEPA. 1984i).
After oral ingestion of inorganic mercury and mercuric salts,
microscopic evaluation of the kidneys in exposed rats was
performed and showed various degrees of damage to the proximal
convulated tubules (PCT) and the glomeruli. Other portions of
the tubule were affected in later stages (USEPA. 1984i).
The acute and chronic effects of methylmercury (an alkyl mercury)
have been observed in the central nervous system in poisoning
incidents, including the well-documented case of seafood
contamination in residents of the area around Minimata Bay,
Japan. Effects such as visual and hearing impairment, ataxia and
loss of sensation in the extremities and around the mouth have
been recorded in man and seem related to cortical neuron
destruction (USEPA. 1984i).
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Data regarding teratogenicity could not be located for inorganic
mercury, however, several investigators have reported embryotoxic
and teratogenic effects in animals treated with methylmercury
(alkyl mercury) (USEPA. 1984i). Neurological defects were the
most common effect noted but an increased frequency of cleft
palate in mice was also documented at doses of 0.1 mg/kg/day of
methylmercury. In humans, brain damage has been reported in
incidents of methylmercury poisoning (DSEPA. 1984i).
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TOXICITY PROFILE FOR: CHLORDANE
BACKGROUND INFORMATION
Chlordane is a complex mixture that includes two isomers of
chlordane, heptachlor, and two isomers of nonachlor (Clement.
1985). This compound has a high chlorine content ranging between
64-67% (Merck. 1983).
Chlordane's solubility ranges from 0.56 to 1.85 mg/liter at 25 Cc
and is miscible in aliphatic and aromatic solvents (Clement.
1985). Although relatively insoluble in water, this compound
loses chlorine content in the presence of alkaline reagents.
With the exception of its use through subsurface ground insertion
(as a pesticide for termite control and dipping of roots or tops
of non-food plants) the DSEPA has cancelled registrations of
pesticides which contain this toxic compound (Merck. 1983). It
previously served as an agricultural home & garden pesticide or
insecticide (USEPA. 1987g).
TRANSPORT & FATE
Atmospheric transport of vapors and contaminated dust particles
from soil application sites can occur.
Chlordane, however, is a compound with a high resistance to
chemical and biological degradation making it very persistent in
the environment. Chlordane is somewhat volatile in clear water,
and this may be a loss process. Adsorption to organic particles
in water is likely. Sorption to sediments is also a likely
important mechanism for removal of chlordane from the water
column. Residue concentrations in sediment are often much higher
than in the water. (Clement. 1985).
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Chlordane degradation to photoisomers, (i.e. photo-cis-chlordane)
occurs under natural environmental conditions. These can be even
more toxic to certain animals and can bioaccumulate to a much
higher degree (USEPA. 1987).
Chlordane accumulates in tissues of aquatic organisms to levels
higher than in the water. Bioconcentration factors thousands of
times greater than water concentrations have been observed in a
wide variety of aquatic organisms. (Clement. 1985).
HEALTH EFFECTS
Chlordane has been found to be poisonous to humans by ingestion,
inhalation, intravenous and percutaneous absorption. Chlordane
has been determined to be a CNS stimulant whose exact mode of
action, although unknown, may involve some microsomal enzyme
stimulation (Sax. 1987).
The fatal Chlordane dose to humans has been estimated to range
from 6 to 60 grains (.2 to 2 ounces) (Sax. 1987). Low oral
Chlordane doses showed severe chronic fatty degeneration of the
liver. This phenomenon is corroborated by the results of
numerous laboratory studies in which Chlordane exposed animals
show degenerative changes in the liver and kidney tubules (Sax.
1987). Chlordane is associated also with reproductive and
metabolic disorders as observed in exposed laboratory mice
(Clement. 1985).
The reference dose for Chlordane has been determined to be 5 x
10~5 mg/kg/day based on a 1983 study, where the LOEL was 1 ppm in
the diet for Chlordane exposed rats. The critical effect was
liver necrosis (USEPA. 1987c).
Several oral cancer bioassays have been conducted. Data indicate
increased incidence in hepatocellular carcinomas in chlordane
A-18
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exposed mice and rats. From these studies, a human carcinogenic
potency risk factor of 1.3 (nig/kg/day)-! was computed (USEPA.
1987c). Chlordane was categorized by EPA's Carcinogen Assessment
Group as a B2 group compound, Probable Human Carcinogen (USEPA.
1986c).
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TOXICITY PROFILE FOR: DDT
BACKGROUND INFORMATION
DDT is a colorless crystal or a white to slightly off-white
powder and is odorless or with a slightly aromatic odor.
Technical DDT (Dichlorodiphenyltrichloroethane) is generally a
mixture of p,p'-DDTf o,p'-DDT, p,p'-DDD, and traces of other
materials. Metabolites of DDT include p,p'-DDE and o/p'-DDD.
DDT isomers and metabolites are often found together and have
similar properties. (Clement. 1985).
DDT is the best known of all the synthetic insecticides. This
compound was synthesized in 1874, albeit it wasn't until 1939
that its insecticidal effectiveness was discovered and later
patented in 1942. During World War II, DDT was directly applied
to humans for the control of lice and other insects. It was one
of the most widely used agricultural insecticides in the United
States and other countries from 1946 to 1972 (Klaassen. 1986).
TRANSPORT & FATE
Due to its high molecular stability, DDT, along with all its
metabolites, is very persistent in the environment. DDT' s
primary transport from application sites was probably
volatilization from soil and water. Isomers of DDT, however, are
most often transported via sorption on sediments and
bioaccumulation (Clement. 1985). This compound's half-life in
water has been determined to range from 56 to 110 years in lake
water, and from 3-15 years in soil (Sax. 1987).
DDT is unusually stable in the environment due to its very low
solubility in water and its resistance to destruction by light
and oxidation. (Merck. 1983).
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Bioaccumulation of DDT is well documented, and is a particularly
important fate process for this compound in aquatic systems.
Analysis of environmental samples indicate that direct uptake,
sorption to biota, and biomagnification in food chains result
from DDT contamination (USEPA. 1984e).
HEALTH EFFECTS
While DDT is classified as a neuropoison, no unequivocal reports
of fatal human poisoning have been recorded despite widespread
use of the substance for 30 to 40 years (Klaassen. 1986). A
dose of 200 mg/kg of DDT has been determined to be highly
dangerous though not fatal to man (Sax. 1987). Chronic
exposures to DDT, DDD and DDE in humans lead to accumulation of
the chemical in fatty tissues. DDT's location of primary toxic
action is the sensory, motor nerve fibers and the motor cortex
(Klaassen. 1986).
Most toxicological data are based on oral exposures. Acute oral
exposures can lead to symptoms of burning or prickling of the
tongue, lips and face, apprehension, irritability, dizziness and
tremors (Klaassen. 1986). Chronic oral exposures resulted in
liver lesions at all doses tested, the lowest of which was 10 ppm
in food or 0.5 mg/kg/day. Additional animal studies showed
increased incidence of tumors and increased mortality of
offspring in a six generation study with an exposure of 100 ppm
(13 mg/kg/day). Oral exposures of 2.5 mg/kg/day of DDT ingested
by pregnant mice proved embryotoxic and fetotoxic (USEPA.
1984e). DDT has consistently caused a decrease in the
reproductive capacity of organisms tested.
DDT and all its metabolites are compounds with a capacity to
bioconcentrate, typically in the adipose issues of most
animals. Toxic doses produce vomiting, muscle weakness,
disturbance of equilibrium, and finally chronic or asphyxial
A-21
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convulsions, followed by death from respiratory failure or
ventribular fibrillation (Clayton. 1981). The RfD of 5.0 x 10~4
mg/kg/day was derived from a study of rats fed commercial grade
DDT, where hepatocellular hypertrophy were observed at some
doses, and a NOEL was shown to be 0.05 mg/kg/day (USEPA. 1987h).
There is evidence of carcinogenicity in animals with exposures to
DDT. Exposures to DDT and its metabolites have lead to liver
tumors in mice (USEPA. 1984e). Exposures to DDT have also shown
to develop hepatomas in rats and lymphomas and lung cancers in
mice. DDT is classified as a Group B2, Probable Human Carcinogen
by the US EPA (USEPA. 1986c). Results from six animal studies
were used to develop a q^*/ carcinogenic potency value of 0.34
(mg/kg/day) ~1 (USEPA. 1984e).
A-22
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TOXICITY PROFILE FOR: HEXACHLOROBENZENE (HCB)
BACKGROUND INFORMATION
HCB is an intermediate in dye manufacture and issued as a wood
preservative. ECB is a very stabler unreactive compound that
when exposed to heat emits highly toxic chlorides (Sax. 1987).
In its physical state, BCB consists of white needles or
monoclinic prisms, and is insoluble in water (Merck. 1983).
Since 1978, ECB is no longer manufactured in the U.S.
(Klaassen. 1986).
TRANSPORT & FATE
Although a half-life value cannot be determined, HCB's detection
in remote areas may suggest that it could be a long one, due to
evidence of long distance transport (USEPA. 1984f). Aerial
dispersion of this compound at HCB manufacturing plants is the
major entry pathway of this compound into the environment
(Clayton. 1981). Rainout and dry deposition are effective
mechanisms for the atmospheric removal of HCB and consequent
entry into the aquatic environment (USEPA. 1984f).
Photodecomposition is extremely slow and rarely observed.
Excessively high temperatures will destroy this compound. An
aromatic hydrocarbon, HCB degrades very slowly and is persistent
in the environment. It is a hydrophilic compound and as such is
expected to bioaccumulate in aquatic organisms. Depuration
occurs over time and HCB levels can decrease in biological
organisms, once removed from the exposure sources.
HEALTH EFFECTS
A classified fungicide, hexachlorobenzene, produced numerous
cases of acquired porphyria cutanea tarda, (PCT), which is
characterized by symptoms such as pigmentary changes, deep
A-23
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scarring, hepatomegaly, permanent loss of hair, skin atrophy and
death. Accidental exposure vas traced to the consumption of feed
grains treated with this compound. In this case, ninety-five
percent of the infants of the mothers that had PCT died within a
year of birth and others acquired the disorder known as "pink
sore" from their HCB affected mother. The presence of HCB in the
mother's milk suggested that "pink sore" was a resulting effect
due to lactation as an exposure route rather than HCB placental
transfer (Klaassen. 1986; USEPA. 1984f). Teratogenic and
reproductive effects, however, have been found to be minimal in
experimental animals (USEPA. 1984f).
Hexachlorobenzene has been demonstrated to be carcinogenic in
rodents (rats, mice, and hamsters), following oral exposure.
Data for humans is not available at this point (USEPA. 1984f).
A carcinogenic potency value of 1.688 (mg/kg/day)-l was derived
by the US EPA in 1980 based on the incidence of hepatomas in male
Syrian Golden hamsters. Hexachlorobenze has been categorized as
a group B2, Probable Human Carcinogen, by the US EPA Carcinogen
Assessment Group (USEPA. 1986c).
A-24
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TOXICITY PROFILE FOR: HEXACHLOROCYCLOHEXANE (HCH)
BACKGROUND INFORMATION
HCH is the common name for the family of isomers of
hexachlorocyclohexane. Technical HCH contains approximately 64%
alpha, 10% beta, 13% gamma, 9% delta and 1% epsilon isomers of
1,2,3,4,5,6-hexachlorocyclohexane (Sax. 1987). HCHs are the
chlorination products of benzene. All the isomers are crystals
with melting points ranging from 112 to 309 degrees Celsius
(Merck. 1983). These compounds exhibit very low volatility and
are slightly soluble in water.
Technical hexachlorocyclohexane is used as an insecticide for the
control of insects on cotton, fruits and vegetables. Lindane,
the gamma isomer, is more often used in insect control on both
livestock and pets (Clayton. 1981). Lindane is presently
imported into the U.S. and according to a 1970 import level
report, less than one million pounds were imported in that year
(USEPA. 1984h).
TRANSPORT & FATE
Adsorption to sediments seems to be a major transport mechanism
in the aquatic environment (USEPA. 1984h). A low mobility in
soil has been recorded for lindane, although surface runoff could
represent a transport mechanism for surface water. Based on the
saturation vapor pressure data, lindane may not be absorbed onto
particulate matter in the air. Nevertheless, in this media,
rainout has been the demonstrated removal mechanism.
HEALTH EFFECTS
The alpha and gamma (lindane) HCH isomers have been recorded as
convulsant poisons, while the beta and gamma isomers are central
A-25
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nervous system depressant. The epsilon isomer appears to have no
observable effects on our system (Klaassen. 1986). Toxicity
studies have been complicated by the fact that each of the
isomers has its own characteristic toxic effect(s) (USEPA.
1984h).
Lindane, along with the other four HCH isomers, has been
associated with aplastic anemia and paramyeloblastic leukemia. A
study in which technical grade HCH was administered through a
diet to Wistar rats demonstrated numerous physiological changes
such as depression, liver increase, fatty accumulation and kidney
degeneration (USEPA. 1984h). Lindane intake affects stimulation
of the central nervous system, causing violent convulsions and is
generally followed by either death or slow recovery. Elevated
body temperatures and pulmonary edema have been reported in
children (USEPA. 1986b).
In a study where rats were administered lindane (99.85%) in a
diet, lindane exposure related effects were not noted on
mortality, hematology, clinical chemistry or urinanalysis
(USEPA. 1986b). Rats receiving 20 and 100 ppm lindane were
observed to have a higher incidence of livehypertrophy,
interstitial nephritis and kidney tubular degeneration. Since
these effects were mild and rare at a level of 4 ppm, this
represents a No Observable Adverse Effect Level (NOAEL) (USEPA.
1986b). An oral reference dose value for a-HCH and g-HCH
(lindane) has been determined to be 3 x 10~4 mg/kg/day.
The teratogenic and other fetotoxic effects on female rats
treated with lindane for four months resulted in: (a) disturbed
estrous cycles, (b) lowered embryonic viability, (c) reduced
fertility and (d) delayed sexual maturation at the 0.5 mg/kg
bw/day level (USEPA. 1984h). These effects were not observed at
a 0.05 mg/kg bw/day level.
A-26
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Lindane appears to fall between a Group B2 and Group C for its
carcinogenic risk category (USEPA. 1986c) while alpha HCH is
considered a Group B2 carcinogen. The carcinogenic potency
factor for the alpha isomer is 6.3 (mg/kg/day)'1 based on
increased incidence of liver tumors in mice and rats, while it is
1.33 for the gamma isomer (USEPA. 1987f).
A-27
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TOXICITY PROFILE FOR: POLYCHLORINATED BIPHENYLS (PCBs)
BACKGROUND INFORMATION
Polychlorinated Biphenyls are a family of the chlorinated
aromatic compounds. The physical, chemical and biological
characteristics of these chemicals vary widely, depending on the
number of chlorine atoms substituted in the aromatic ring(s)
(Klaassen. 1986). The Aroclors are characterized by an
exclusive four digit number. The first two digits indicate
whether the compound contains biphenyls (denoted by a 12)
triphenyls (by a 54) or both compounds (25,44), while the last
two digits state the weight percent of chlorine in the compound
(Merck. 1983). The chlorine content ranges from 12 to 68
percent. In general, all PCBs have very low water solubilities
(0.003-0.6 mg/1) and vapor pressures 10-3 to 10-5 mm Hg at 20° C
(USEPA. 1984k).
Polychlorinated Biphenyls or PCBs were once widely used
industrial chemicals. Their high stability contributed to both
their commercial usefulness and their subsequent long-term
environmental and health effects. PCBs have been commercially
available since 1930. PCBs have been used primarily as
insulating material in electrical capacitors and transformers,
for the insulation of electrical cables and wires, fire
retardants, and in heat transfer systems (Clayton. 1981). The
manufacture and distribution of PCB-containing products has been
banned since 1979 (Klaassen. 1986).
FATE & TRANSPORT
PCB's ubiquitous nature can be attributed to volatilization
mechanism followed by adsorption onto dust and fallout
(Klaassen. 1986). Lighter PCB species, with fewer chlorine
atoms, tend to volatalize more easily.
A-28
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PCBs are relatively inert and therefore persistent. Adsorption
to organic material in sediment is probably a fate mechanism for
at least the more heavily chlorinated PCBs. Slow desorption can
provide continuous low-level contamination. These less heavy
PCBs can be biodegraded by some soil microorganisms. The heavier
PCBs are not measurably biodegraded, but can be photodegraded by
ultraviolet light at a very slow rate. (Clement. 1985). PCBs
are bioaccumulated and biomagnified in the aquatic environment.
HEALTH EFFECTS
In 1968, accidental ingestion of PCBs occurred in Yusho, Japan,
as a result of rice bran oil contamination with Kanechlor-400, a
PCB product used as a heat transfer agent (USEPA. 1984). This
incident known as Yusho poisoning, affected approximately
1,000 persons, altering their dermal and respiratory systems.
Palmar sweating and muscular weakness were also common
complaints. By 1979, 31 Yusho patients had already died
(USEPA. 1984k) from causes such as malignant neoplasms, stomach
and liver cancers, and malignant lymphomas.
Cancer caused by Kanechlor-500 has been demonstrated in
laboratory mice, while Aroclor 1260 has also been shown to be
carcinogenic in rats (USEPA. 1984k). The reference dose of
1x10"^ mg/kg/day is based on a study of rhesus monkeys where
exposures to Aroclor 1016 in the diet during mating and gestation
resulted in smaller offsprings in the study animals than those of
the control group (USEPA. 1987h). Studies have shown an
increased number of different liver cancers such as
adenocarcinomas, trabecular carcinomas and neoplastic nodules in
rats fed PCBs. No significant teratogenic effects were recorded
but fetotoxicity was evident (USEPA. 1984k).
PCBs have been classified as a group B2, Probable Human
Carcinogen compound. A draft document is available from the U.S.
A-29
-------
Public Health Services (Nov. 1987)r which designates a
carcinogenic potency factor of 7.7 (mg/kg/day)"1 for PCBs based
on the carcinogenicity of Aroclor 1260 (USPHS. 1987). Previous
to this newly developed CPF the generally accepted value of 4.34
(mg/kg/day)"1 was used. A congener specific analysis of
Quincy Bay biota samples was conducted by USEPA Exposure
Assessment Group where it was concluded that based on the
thirteen congeners measured, the mixture of PCBs in the
Quincy Bay seafood resembles Aroclor 1254 more closely than
Aroclor 1260 or 1242 (OSEPA. 1988b). Additional work by the US
EPA Carcinogen Assessment Group indicates that the plausible
upper bound cancer potency factor for Aroclor 1254 is 2.6
(mg/kg/day)"1. It is based on a National Cancer Institute study
in which statically significant dose related increases in liver
modules, benign tumors, and malignant tumors, were seen in rats
fed a diet containing Aroclor 1254 (USEPA. 1988c). The CPF of
2.6 (mg/kg/day)"1 was used in the evaluation of risk for this
study. This CPF and the two others mentioned previously are not
considered substantially different due to the uncertainty
associated with the experimental data from which the CPF value of
2.6 (mg/kg/day)"1 was derived (USEPA. 1988c).
A-30
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TOXICITY PROFILE FOR: POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
BACKGROUND INFORMATION
PAHs are chemicals which consist of two or more fused benzene
rings and occur in a variety of commercial products such as soot,
coal, tar, tobacco smoke, cutting oils and petroleum (Klaassen.
1986). These compounds form as a result of breakdown of
hydrocarbon compounds when exposed to ultraviolet radiation or by
incomplete combustion of organic compounds with insufficient
oxygen availability.
TRANSPORT & FATE
Little information is available on the range of compounds that
are classified as PAHs, however, much is inferred from the more
researched benzo(a) pyrene. Atmospheric fallout, surface runoff
are likely existing sources to aquatic environments and
adsorption on to sediments is a probable transport mechanism.
(Clement. 1985).
PAHs are relatively insoluble in water, but the dissolved portion
is believed to undergo direct photolysis. Some may also be
oxidized by chlorine and ozone. (Clement. 1985).
PAHs are bioaccumulated, although rapidly metabolized and
eliminated by most organisms (not shellfish). Biodegradation is
believed to occur more slowly in water than in soil, but to more
significant in systems chronically affected by PAH
contamination. (Clement. 1985).
HEALTH EFFECTS
Due to the high lipophylic nature of PAHs, they are readily
absorbed in the gastrointestinal tract of animals. In a study
A-31
-------
where rats were administered B[a]P contained in a starch
solution, 50% of the compound was absorbed. There is often no
sign of toxicity until the dose is high enough to produce a high
tumor incidence thus carcinogenicity dominates health effect
considerations (Clement. 1985). When benzo [a] pyrene is
administered to the skin of mice quick carcinoma formation
results. Subcutaneous injection produces sarcomas in rats and
mice. Oral administration of some PAHs to rhesus monkeys and
other primates has so far not yielded tumors (Klaassen. 1986).
Benzo [a] pyrene was administered to study mice through diet at
concentrations ranging from 1 to 250 ppm and stomach tumors
(papillomas and carcinomas) were reported. Control mice did not
have similar tumors (USEPA. 1984J). At increased concentrations
ranging from 250 to 1000 ppm, B[a]P produced a higher incidence
of stomach tumors, as well as lung adenoma and leukemia in the
studied mice (USEPA. 1984J).
The US EPA used incidences of stomach tumors in B[a]P exposed
mice in a 1957 study to derive a carcinogenic potency factor of
11.53 (mg/kg/day)"1 for oral intake. This CPF is issued for all
PAHs using the conservative default assumption that with the
absence of sufficent data to the contrary all PAHs are
carcinogenic and equal in potency to B[a]P.
A-32
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REFERENCES FOR TOXICITY PROFILES
Clayton and Clayton. 1981. Patty's Industrial Hygiene and
Toxicology. Volume IIB. New York, NY. 2879-3769 pp.
Clement Associates, Inc. 1985. cheaical, Physical and
Biological Properties of Compounds Present at Hazardous Waste
sites. Prepared for U.S. Environmental Protection Agency.
Klassen. C.D., M.O. Amdur, and J. Doull. 1986. Toxicology: The
science of Poisons. Third Edition. New York, NY. 974 pp.
Merck & CO., Inc. 1983. The Merck Index; An Encyclopedia of
Chemicals, Drugs, and Biologicals. Rahway, New Jersey.
1463 pp.
Sax, N.I., R.J. Lewis, Sr. 1987. Hazardous Chemicals Desk
Reference. New York, NY. 1084 pp.
U.S. Environmental Protection Agency (USEPA). 1980. Ambient
Water Quality Criteria Documents. USEPA Office Of Water
Regulation and Standards. Criteria Division. Washington D.C.
a) Cadmium EPA 440/5-80-025
b) Chromium EPA 440/5-80-035
c) Copper EPA 440/5-80-036
d) Lead EPA 440/5-80-057
e) Mercury EPA 440/5-80-058
U.S. Environmental Protection Agency (USEPA). 1984. Health
Effects Assessment Documents.
a) Cadmium. PB86-134491
b) Chlordane. PB86-134343
c) Chromium. PB86-134301
d) Copper. PB86-134368
e) DDT. PB86-134376
f) Hexachlorobenzene. PB86-134285
g) Lead. PB86-134665
h) Lindane. PB86-134673
i) Mercury. PB86-134533
j) Polycyclic Aromatic Hydrocarbons. PB86-134244
k) Polychlorinated Biphenyls. PB86-134512
U.S. Environmental Protection Agency (USEPA). 1986. integrated
Risk Information System (IRIS). Chemical Abstracts.
a) Chromium. CAS No.: 16065-83-1
b) Lindane. CAS No.: 58-89-9
A-33
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U.S. Environmental Protection Agency (USEPA). 1986c. superfund
Public Health Evaluation Manual. EPA/540/1-86/060.
Washington, D.C. 175 pp.
U.S. Environmental Protection Agency (DSEPA) . 1987a. Health
Advisories for Legionella and Seven Inorganics .
PB87-235586. Washington, D.C. 125 pp.
U.S. Environmental Protection Agency (USEPA). 1987.
Risk Information System (IRIS). Chemical Abstracts.
b) Cadmium.
c) Chlordane.
d) Chromium VI
e) DDT.
f) Lindane.
CAS No.:
CAS No.:
CAS No.:
CAS No.:
CAS No.:
integrated
7440-43-9
57-74-9
7440-47-3
50-29-3
58-89-9
U.S. Environmental Protection Agency (USEPA).
Advisories for 16 Pesticides.
Washington, D.C. 262 pp.
1987g. Health
PB87-235586.
U.S. Environmental Protection Agency (USEPA). 1987h. Health
Advisories for 25 Organics . PB87-235578. Washington, D.C.
397 pp.
U.S. Public Health Services (USPHS) 1987. Toxicological Profile
for selected PCBs (Aroclors - 1260-1254, -1248, -1242, -1232,
-1221, and -1016). Draft for Public Comment. Prepared by
Syracuse Research Corporation under Contract No. 68-03-
3228. 136 pp.
U.S. Environmental Protection Agency (USEPA). 1988a. Memorandum
from Dr. W. Farland, Acting Director, Office of Health and
Environmental Assessment, to K. Kipp, Quincy Bay Coordinator
- Region I EPA.
U.S. Environmental Protection Agency (EPA). 1988 b. Memorandum
from S. Braen Norton, Exposure Assessment Group, to
Dr. W. Farland, Acting Director, Office of Health and
Environmental Assessment.
U.S. Environmental Protection Agency (EPA). 1988c. Memorandum
from J. Cogliano, Carcinogen Assessment Group, to
K. Garrahan, Exposure Assessment Group.
A-34
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Table B-ll. ORGANICS AND METALS INCLUDED IN ANALYTICAL RESULTS PROVIDED
BY US ENVIRONMENTAL PROTECTION AGENCY (a)
ELEMENTS/METALS
Silver
Arsenic
Beryllium
Cadmium
Cobalt
Chromium
Copper
Iron
Mercury
Magnesium
Manganese
Nickel
Lead
Antimony
Selenium
Thallium
Vanadium
Zinc
ORGANIC COMPOUNDS
Bis(2-Ethyl-Hexyl)Phthalate
Chlordane (total)
a-Chlordane
g-Chlordane
Coprostanol (Coprosterol)
PP-DDD
PP-DDE
PP-DDT
Hexachlorobenzene (HCB)
Hexachlorocyclohexane (HCH)
a-HCH
g-HCH (lindane)
Heptachlor
Methylene chloride
Methyl Chloride
Endrin
Toxaphene (chlorocamphene)
Polyarom. Hydrocarbons(PAH)
Fluorene
Phenanthrene
Anthracene
ORGANIC COMPOUNDS(Continued)
C1PA(homologs/Phen-Anthr)
C2PA(homologs/Phen-Anthr)
C3PA(homologs/Phen-Anthr)
C4PA (homologs/Phen-Anthr)
Fluoranthene
Pyrene
Benzo [a] anthracene
Chrysene
Benzofluoranthenes (sum)
Benzo [e] pyrene
Benzo [a] pyrene
Perylene
Indeno [1,2,3-cd] pyrene
Benzo [ghi] perylene
PAHs (Sum of mol. weight 276)
PAHs (Sum of aol. weight 278)
Corene
PAHs (Sum of mol. weight 302)
Total of measured PAHs
PCBs (total) (a)
Aroclor 1242
Aroclor 1254
-CB052 (2,2',5,5'-PCB)
CB047 (2,2',4,4'-PCB)
CB101 (2/2',4,5,5l-PCB)
CB151 (2,2',3,5,5',6-PCB)
CB118 (2,3',4,4',5-PCB)
CB153 (2,2,4,4',5,5(-PCB)
CB138 (2,2',3,4,4',5',-PCB)
CB128 (2,2',3,3',4,4',-PCB)
CB180 (2,2',3,4,4,5,5'-PCB)
CB195 (2,2',3,3',4,4',5,6-PCB)
CB194 (2,2',3,3',4,4',5,5'-PCB)
CB206 (2,2',3,3',4,4'.5,5',6-PCB)
CB209 (CL10-PCB)
(a): Gardner & Pruell. 1987.
B-15
-------
Appendix C
Development of Carcinogenic
Potency Factor
for PCBs
-------
V
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON. D.C. 20460
March 15, 1983
OFFICE OF
RESEARCH AND DEVELOPMENT
MEMORANDUM
SUBJECT: Cancer potency for AroclorR 1254
FROM: Jim Cogliano
Carcinogen Assessment Group (RD-689). *>
<-s
TO: Kevin Garrahan
Exposure Assessment Group (RD-689)
In response to your inquiry about a separate cancer potency
for AroclorR 1254, I have prepared the following analysis.
My preliminary calculations indicate a cancer potency of 2.6
per mg/kg/d continuous lifetime exposure to AroclorR 1254. This
is a plausible upper bound, meaning that the true potency is not
likely to exceed this estimate and may be lower. It is based on
the 1978 National Cancer Institute (NCI) study of AroclorR 1254,
in which statistically significant, dose-related increases in
liver nodules, benign tumors, and malignant tumors combined were
seen in Fischer 344 rats fed a diet containing AroclorR 1254.
Several uncertainties deserve your attention:
1. NCI used only 24 rats per group (50 is considered
standard today), so the potency estimate is rather
imprecise.
2. The NCI study lasted 24 months. Although this is
today's standard, a recent, longer study by Norback anc
Weltman indicates that PCB-fed rats develop many tumors
after 24 months. CAG considers the Norback and Weltman
study superior for estimating the potency. The NCI
study is analogous to the study that was superseded by
the Norback and Weltman study.
c-i
-------
3. NCI^s female rats developed only benign liver tumors and
nodules, so some may argue that there was no cancer.
Norback and Weltman, however, demonstrated that nodules
progress to benign tumors, which in turn progress to
malignant tumors. Under EPA's cancer guidelines it is,
therefore, appropriate to consider benign tumors and
nodules. Furthermore, some male rats did develop
malignant liver tumors.
CAG's current cancer potency for AroclorR 1260, which is
presumed to apply to other PCB mixtures as well, is 7.7 per
mg/kg/d continuous lifetime exposure. CAG's previous estimate
was 4.3 per mg/kg/d. In light of the uncertainties cited above,
these figures are not substantially different from the new figure
for AroclorR 1254. Larger differences are commonly seen between
different sexes and animal strains. For example, a comparison of
the NCI and Norback and Weltman studies suggests that AroclorR
1254 may be more potent in male Fischer 344 rats than AroclorR
1260 is in male Sprague-Dawley rats.
Further investigation, perhaps taking into consideration
potency differences between PCB mixtures for other toxic effects,
is needed before there can be separate cancer potencies for each
PCB mixture. Until then, it appears that the cancer potency of
AroclorR 1254 is either similar to, or slightly less than, that
of AroclorR 1260.
Attached is a summary of the new potency calculation. If
you have any questions, or if I can be of further assistance,
please call me at 382-2575.
Attachment: Summary of potency calculation for AroclorR 1254
cc: Charles Ris
C-2
-------
SUBSTANCE
. Aroclor(R) 125*
DEFERENCE NCI, 1976
SEX. STRAIN, SPECIES Feaale Fischer 344 rats
EXPOSURE ROUTE, VEHICLE Oral, diet
TUMOR SITE, TYPE Liver nodular hyperplasia and adenoaas
NOMINAL DOSE
AVERAGE DAILY DOSE
EQUIVALENT HUMAN DOSE
O 25 50 100 ppn
0 1.25 2.50 5.00 »g/kg/d (5% food factor)
0 1.16 2.32 4.65 «g/kg/d (1O5/113 weeks)
0 0.17 0.33 0.64 Dg/kg/d (surf-area adj)
TUMOR INCIDENCE 0/24 6/24 10/22 19/24
TUMOR PERCENTAGE O* 25* 45% 79*
STATISTICAL SIGNIFICANCE -- 1E-02 2E-O4 1E-08
TREND SIGNIFICANCE <0.001, linearity OK
ANIMAL WEIGHT
EXPOSURE PERIOD
STUDY LENGTH
ANIMAL LIFESPAN
POTENCY (qi«)
220 200 160 f (at end of study)
250
105 wk
113 wk
113 wk (assumed)
2.6 per ng/kg/d
Cogliano 16:14 09-Mar-66
C-3
-------
., ^ UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON. O.C. 20460
.V
161988
OFFICE OF
RESEARCH AND DEVELOPMENT
MEMORANDUM
SUBJECT: Congener-Specific Analysis of Quincy Bay Biota Samples
FROM: Susan Braen Norton, Environmental Scientist •**&* ^^
Exposure Assessment Applications Branch
Exposure Assessment Group (RD-689)
TO: William H. Farland, Ph.D.
Acting Director
Office of Health and Environmental Assessment (RD-689)
THRU: Michael A. Callahan, Director
Exposure Assessment Group (RD-689)
One of the concerns expressed in the March 1 meeting of the
Fish Contamination Committee vas that the mixture of PCBs
measured in seafood from Quincy Bay may be more like Aroclor 1254
than Aroclor 1260. This concern vas raised because the cancer
potency factor for Aroclor 1260 vas used to assess risks
associated with the ingestion of seafood (from Quincy Bay.
To address this issue, I conducted a simple analysis using
the thirteen congeners that vere measured in Quincy Bay seafood
(U.S. EPA 1987). The conclusions of this analysis are that,
based on the 13 congeners measured, the mixture of PCBs in the
seafood resembles Aroclor 1254 more closely than Aroclor 1260 or
Aroclor 1242.
Bar graphs of the 13 congeners measured in flounder, clams,
oysters, lobster flesh, and lobster hepatopancreas are attached.
The congener concentrations in these graphs have been normalized
relative to congener 138 (2,2 ' ,3,4,4 ' ,5-PCB) in order to more
easily distinguish patterns. Also attached are bar graphs of the
normalized congener concentrations present in the commercial
mixtures Aroclors 1254, 1242, and 1260 (as per Rapaport and
C-4
-------
Eisenreich 1984; and Capel et al. 1985). On the basis of these
graphs, Aroclors 1254 and 1260 were selected for further
analysis.
To more quantitatively compare the PCBs in seafood with the
commercial PCB mixtures, I summed the squares of the differences
between each of the normalized congener concentrations in the
seafood and the commercial mixture. The results of the sums of
squares analysis are also attached. As can be seen, the mixture
of PCBs measured in oyster tissue most resembles the commercial
mixture Aroclor 1254 as quantified by Rapaport and Eisenreich,
(1984). Residues measured in flounder, clams, and lobster flesh
and hepatopancreas most closely resemble Aroclor 1254 as reported
by Capel ct al. (1985).
There are several important uncertainties in using the
results of this analysis in risk assessment:
1. The analysis was based on only 13 of 209 possible PCB
congeners. However, the 13 congeners vary greatly with
chlorination; for the purposes of this analysis, they
were considered to sufficiently represent the large
range of possible congeners.
2. No congener-specific data were available on the actual
PCB mixture that was fed to the test animals in the
cancer bioassays. Because the congener concentrations
can vary greatly with batch, the congener
concentrations reported in the literature may differ
from those used in the bioassays.
3. Congener-specific toxicity data are not yet available.
Because it is not known whether the most toxic PCB
congeners were used to compare seafood residues to the
commercial mixtures," the PCB mixture in the seafood may
actually be more or less toxic than Aroclor 1254.
Attachments
C-5
-------
REFERENCES
Capel, P.D., Rapaport, R.A., Eisenreich, S.J., and Looney, 3.B.
1985. PCBQ: Computerized Quantification of Total PCB and
Congeners in Environmental Samples. Chemosphere 14: 439-
450
Rapaport, R.A. and Eisenreich, S.J. 1984. Chromatographic
determination of octanol-water partition coefficients
(Kow's) for 58 polychlorinated biphenyl congeners. Environ.
Sci. Technol. 18: 163-170
U.S. Environmental Protection Agency (USEPA). 1987. A
Histopathological and Chemical Assessment of Winter
Flounder, Lobster, and Soft-Shelled Clam Indigenous to
Quincy Bay, Boston Harbor and an in situ Evaluation of
Oysters Including Sediment (Surface and Cores) Chemistry.
Environmental Research Laboratory Narragansett, Rhode
Island. December 1, 1987
C-6
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C-15
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SUMS OF SQUARES COMPARISON BETWEEN QUINCY BAY SAMPLES
AND LITERATURE VALUES
Sample
Type
Oyster
Clam
Lobster
Flesh
Lobster
Hepato-
pancreas
Sample
Number
75261
75253
75255
75254
75256
75257
75259
75260
75237
75239
75241
75219
75220
75244
75245
75212
75214
75223
75225
75228
75249
75250
75230
75234
75237
75219
75244
75212
75223
75228
75249
75230
1254
(b)
11227 *
7578 *
7114 *
7715 *
6952 *
6539 *
16850
11577
32169
31808
31489
27380
30404
28888
29832
31933
30014
32710
28051
32237
29021
31601
33365
34883
37540
33769
34388
37119
37460
61796
36212
37267
1254
(a)
15053
18044
13699
9603
17593
16249
6563 *
6537 *
14359 *
14633 *
17918 *
11104 *
12880 *
12859 *
13163 *
14221 *
12324 *
13691 *
10356 *
14601 *
12647 *
14819 *
16071 *
15928 *
14927 *
11642 *
14272 *
14517
14677
39223
13482
16343
1260
(a)
48971
60767
54969
44954
63189
60065
24173
31195
27108
28282
36550
25924
26322
28772
25922
25696
24942
24811
23140
27604
30535
29737
30424
26394
18762
16306
20574
17962
18258
43196
18369
22266
* The sample most resembles the denoted mixture.
(a) Capel, P.D., Rapaport, R.A., Eisenreich, S.J., and Looney,
B.B. 1985. PCBQ: Computerized Quantification of Total PCB
and Congeners in Environmental Samples. Chemosphere 14:
439-450
(b) Rapaport, R.A., and Eisenreich, S.J. 1984.
Chromatographic determination of octanol-water partition
coefficients (Kow's) for 58 polychlorinated biphenyl
congeners. Environ. Sci. Technol. 18: 163-170
C-16
-------
SUMS OF SQUARES COMPARISON BETWEEN QUINCY BAY SAMPLES
AND LITERATURE VALUES (continued)
Sample Sample
Type Number
Flounder 75124
75168
75185
75190
75191
75194
75195
75198
75101
75113
75114
75115
75179
75160
75164
75167
75170
75172
75180
75182
75128
75133
75145
75148
75149
1254
(b)
38091
39240
38799
35293
36659
40640
34944
43187
39044
43059
39194
24979
31190
34156
29918
33996
31627
38724
24708
35733
39303
35544
36571
36325
27244
1254
(a)
16492 *
19739 *
20954 *
16045 *
17618 *
17993 *
15859 *
23745 *
18299 *
23844
19690 *
10781 *
17733 *
17567 *
12239 *
15024 *
18337 *
19059 *
12261 *
18046 *
18148 *
15479 *
18824 *
15417 *
14093 *
1260
(a)
20642
23926
25383
19456
19S12
19216
18789
24368
19839
23843 *
21606
21346
28029
24312
20876
22622
27089
22791
26087
23313
18788
20236
22506
19947
23345
* The sample most resembles the denoted mixture.
(a) Capel, P.P., Rapaport, R.A., Eisenreich, S.J., and Looney,
B.B. 1985. PCBQ: Computerized Quantification of Total PCB
and Congeners in Environmental Samples. Chemosphere 14:
439-450
(b) Rapaport, R.A., and Eisenreich, S.J. 1984.
Chromatographic determination of octanol-water partition
coefficients (Kow's) for 58 polychlorinated biphenyl
congeners. Environ. Sci. Technol. 18: 163-170
C-17
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