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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)
                               11

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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
                                14

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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.
                                16

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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
                              17

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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

-------
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

-------
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

-------
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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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

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                 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

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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-Uud 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.  Ued by N.Y. State In  study on PCi in ffh (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
           Corunption
           Typictl Local Concimer
           Mx. Exp. Ind.  Mixed Dltt
           M*x. Exp. Ind. - Floinder Only
GriM/Dty

   1
  113
  165
   Rik
1.0E-05
1.2E-03
1.7E-03
                                 62

-------
        FIG.  8.   TOMALLEY  CONSUMPTION  EFFECTS
                             Stnrifivfty
O.CE+00
          Consumption Scenario
HEPATOPANCftEAS INGESTION

             6rM/Dy
           Typici 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                 GrM/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

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         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

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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

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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).
                               A-4

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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).
                               A-8

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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.
                              A-10

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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

                              A-12

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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.
                              A-13

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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).
                              A-15

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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).
                              A-16

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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).
                              A-17

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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).
                              A-19

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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).
                              A-20

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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

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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

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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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   Appendix B





Risk Calculations

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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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-------
      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

-------