SEPA
              United States
              Environmental Protection
              Agency
               Office of Water
               Regulations and Standards
               Criteria and Standards Division
               Washington DC 20460
EPA 440/5-80-028
October 1980
                                            C.I
Ambient
Water Quality
Criteria for
Chlorinated Benzenes

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      AMBIENT  WATER  QUALITY CRITERIA FOR

            CHLORINATED  BENZENES
                 Prepared By
    U.S. ENVIRONMENTAL PROTECTION AGENCY

  Office of Water Regulations and Standards
       Criteria and Standards Division
              Washington, D.C.

    Office of Research and Development
Environmental Criteria and Assessment Office
              Cincinnati, Ohio

        Carcinogen Assessment Group
             Washington, D.C.

    Environmental  Research Laboratories
             Corvalis, Oregon
             Duluth, Minnesota
           Gulf Breeze, Florida
        Narragansett,  Rhode Island
                           U.S. Envtronmentat Protection Agency
                           Region 5, Library 
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                              DISCLAIMER
     This  report  has  been reviewed by the  Environmental  Criteria and
Assessment Office,  U.S.  Environmental  Protection  Agency,  and approved
for publication.   Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
                          AVAILABILITY  NOTICE
      This  document is available  to  the public  through  the National
Technical Information Service, (NTIS), Springfield, Virginia  22161.
                                   ii

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                                FOREWORD

     Section  304 (a)(l) of  the  Clean  Water. Act of  1977  (P.L.  95-217),
 requires  the Administrator of  the  Environmental  Protection Agency  to
 publish  criteria for  water quality  accurately  reflecting the  latest
 scientific knowledge on the kind and extent of all  identifiable effects
 on  health  and welfare  which  may  be  expected  from  the  presence  of
 pollutants in any body of  water,  including ground water.  Proposed water
 quality criteria  for  the  65 toxic pollutants  listed under  section 307
 (a)(l)  of the  Clean  Water  Act  were  developed and a  notice  of  their
 availability was published  for  public  comment  on March  15,  1979  (44  FR
 15926), July 25, 1979 (44 FR 43660),  and October  1,  1979  (44 FR 56628).
 This  document  is a revision  of those  proposed  criteria based  upon  a
 consideration  of comments received  from other Federal  Agencies,  State
 agencies,  special   interest  groups,   and  individual scientists.   The
 criteria contained in  this document replace any previously published EPA
 criteria  for  the  65  pollutants.    This  criterion document  is also
 published in satisifaction  of paragraph  11 of  the  Settlement Agreement
 in  Natural  Resources  Defense  Council,  et. al. vs.  Train,  8  ERC 2120
 (D.D.C. 1976), modified,  12 «C  1833  (D.D.C. 1979).	

    The term "water quality criteria"  is  used in  two  sections  of the
 Clean Water Act, section 304 (a)(l) and section  303  (c)(2).  The  term has
 a different  program impact  in  each  section.    In section 304,  the term
 represents a non-regulatory,  scientific assessment  of ecological ef-
 fects. The criteria presented  in this publication  are  such scientific
 assessments.   Such water quality criteria  associated  with   specific
 stream uses when adopted as  State water quality standards under section
 303  become  enforceable maximum  acceptable levels  of  a pollutant   in
 ambient waters.  The water quality criteria adopted in  the State water
 quality standards could have the same numerical limits  as the criteria
 developed under section  304.  However,  in many situations States  may want
 to adjust  water quality criteria developed  under section 304 to reflect
 local  environmental  conditions  and  human  exposure  patterns  before
 incorporation  into  water  quality standards.   It  is not  until  their
 adoption as part of  the  State water quality standards that the criteria
 become regulatory.

    Guidelines to  assist  the States  in the modification  of  criteria
presented   in this  document,  in the  development of  water  quality
standards, and  in other water-related programs of this Agency, are being
developed  by EPA.
                                    STEVEN SCHATZOW
                                    Deputy Assistant Administrator
                                    Office of Water Regulations and Standards
                                  111

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                                ACKNOWLEDGEMENTS
Aquatic Life Toxicology

   William A. Brungs, ERL-Narragansett
   U.S. Environmental Protection Agency

Mammalian Toxicology and Human Health Effects

   Albert Munson (author)
   Medical College of Virginia
   Terence M. Grady  (doc. mgr.), ECAO-Cin
   U.S. Environmental Protection Agency

   Jerry F. Stara  (doc. mgr.), ECAO-Cin
   U.S. Environmental Protection Agency

   John Buccini
   Health and Welfare, Canada

   Herbert Cornish
   University of Michigan

   Larry Fishbein
   National Center for Toxicological Research

   Ronald W. Hart
   Ohio State University

   Steven D. Lutkenhoff  (doc. mgr.), ECAO-Cin
   U.S. Environmental Protection Agency

   Martha Radike
   University of Cincinnati

   Sorrel 1 L. Schwartz
   Georgetown University

   Bonnie Smith, ECAO-Cin
   U.S. Environmental Protection Agency
David J. Hansen, ERL-Gulf Breeze
U.S. Environmental Protection Agency.
Roy E. Albert*
Carcinogen Assessment Group
U.S. Environmental Protection Agency

Donald Barnes
East Carolina University

S.G. Bradley
Medical College of Virginia

Richard A. Carchman
Medical College of Virginia

Patrick Dugan
Ohio State University

George Fuller
University of Rhode  Island

Krystyna Locke
U.S. Environmental Protection Agency

Gordon Newell
National Academy  of  Sciences

Larry  Rosenstein
SRI International

Robert E. McGaughy,  CAG
U.S. Environmental Protection Agency

David  L. West
National  Institute for Occupational
   Safety and Health
 *CAG  Participating  Members:
    Elizabeth  L.  Anderson,  Larry Anderson,  Dolph  Arnicar,  Steven  Bayard,
    David  L. Bayliss,  Chao  W.  Chen,  John  R.  Fowle III,  Bernard  Haberman,
    Charalingayya Hiremath, Chang  S.  Lao, Robert  McGaughy,  Jeffrey Rosen-
    blatt,  Dharm  V.  Singh,  and Todd  W. Thorslund.

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Technical Support Services Staff:  D.J. Reisman, M.A.  Garlough, B.L. Zwayer
P.A. Daunt, K.S. Edwards, T.A. Scandura, A.T. Pressley,  C.A. Cooper,
M.M. Denessen.


Clerical Staff:  C.A. Haynes, S.J. Faehr, L.A. Wade,  D.  Jones,  B.J.  Bordicks,
B.J. Quesnell, T. Highland, R. Rubinstein.

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                         TABLE OF CONTENTS

                                                           Page

Criteria Summary

Introduction                                               A-l

Aquatic Life Toxicology                                    B-l
     Introduction                                          B-l
     Effects                                               B-l
          Acute Toxicity                                   B-l
          Chronic Toxicity                                 B-3
          Plant Effects                                    B-3
          Residues                                         B-4
          Miscellaneous                                    B-6
          Summary                                          8-6
     Criteria                                              B-8
     References                                            B-20

                         Monochlorobenzene                 C-l

Mammalian Toxicology and Human Health Effects              C-l
     Introduction                                          C-l
     Exposure                                              C-l
          Ingestion from Water                             C-l
          Ingestion from Food                              C-2
          Inhalation                                       C-6
          Dermal                                           C-6
          Summary and Conclusions                          C-8
     Pharmacokinetics                                      C-8
          Absorption                                       C-8
          Distribution                                     C-8
          Metabolism                                       C-9
     Effects                                               C-ll
          Acute, Subacute,  and Chronic  Toxicity           C-ll
          Synergism and/or  Antagonism         '             C-l5
          Teratogenicity, Mutagenicity,  and
             Carcinogenicity                               C-l6
     Criterion  Formulation                                 C-17
          Existing Guidelines and  Standards                C-17
          Current Levels of Exposure                       C-17
          Special Groups at Risk                           C-18
          Basis and Derivation of  Criterion                C-19
     References                                            C-22

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                        Trichlorobenzenes                  C-29

Mammalian Toxicology and Human Health Effects              C-29
     Introduction                                          C-29
     Exposure                                              C-29
          Ingestion from Water                             C-29
          Ingestion from Food                              C-29
          Inhalation and Dermal                            C-32
     Pharmacokinetics                                      C-35
          Absorption                                       C-35
          Metabolism                                       C-35
          Excretion                                        C-38
     Effects                                               C-39
          Acute, Subacute, and Chronic Toxicity            C-39
          Synergism and/or Antagonism                      C-42
          Teratogenicity, Mutagenicity, and
             Carcinogenicity                               C-42
     Criterion Formulation                                 C-44
          Existing Guidelines and Standards                C-44
          Current Levels of Exposure                       C-44
          Basis and Derivation of Criterion                C-45
     References                                            C-46

                        Tetrachlorobenzenes                C-51

Mammalian Toxicology and Human Health Effects              C-51
     Introduction                                          C-51
     Exposure                                              C-51
          Ingestion from Water                             C-51
          Ingestion from Food                              C-52
          Inhalation and Dermal                            C-54
     Pharmacokinetics                                      C-54
          Absorption, Distribution, Metabolism, and
             Excretion                                     C-54
     Effects                                               C-60
          Acute, Subacute,  and Chronic Toxicity            C-60
          Synergism and/or Antagonism                      C-62
          Teratogenicity,  Mutagenicity,  and
             Carcinogenicity                               C-62
     Criterion Formulation                                 C-64
          Existing Guidelines and Standards                C-64
          Current Levels of Exposure                       C-64
          Special  Groups at Risk                            C-64
          Basis and Derivaiton of Criterion                C-65
     References                                            C-67

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                        Pentachlorobenzene                 C-71

Mammalian Toxicology and Human Health Effects              C-71
     Introduction                                          C-71
     Exposure                                              C-73
          Ingestion from Water                             C-73
          Ingestion from Food                              C-73
          Inhalation                                       C-75
          Dermal                                           C-76
     Pharmacokinetics                                      C-76
          Absorption, Distribution, Metabolism, Excretion  C-76
     Effects                                               C-80
          Acute, Subacute, and Chronic Toxicity            C-80
          Synergism and/or Antagonism                      C-81
        Carcinogenicity, Mutagenicity, Teratogenicity      C-81
     Criterion Formulation                                 C-34
          Current Levels of Exposure                       C-84
          Special Groups at Risk                           C-84
          Basis and Derivation of Criterion                C-84
     References                                            C-86

                         Hexachlorobenzene                 C-93

Mammalian Toxicology and Human Health Effects              C-93
     Introduction                                          C-93
     Exposure                                              C-94
          Ingestion from Water                             C-94
          Ingestion from Food                              C-99
          Inhalation and Dermal                            C-104
     Pharmacokinetics                                      C-108
          Absorption                                       C-108
          Distribution                                     C-110
          Metabolism                                       C-112
          Excretion                                        C-114
     Effects                                               C-116
          Acute, Subacute, and Chronic Toxicity            C-116
          Synergism and/or Antagonism                      C-121
          Teratogenicity                                   C-122
          Mutagenicity                                     C-124
          Carcinogenicity                                  C-125
     Criterion Formulation                                 C-126
          Existing Guidelines and Standards                C-126
          Current Levels of Exposure                       C-127
          Special Groups at Risk                           C-128
          Basis and Derivation of Criterion                C-132
     References                                            C-135
                               vm

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                  Summary - Criterion Formulation           C-150

     Existing Guidelines and Standards                      C-150
          Monochlorobenzene                                 C-150
          Trichlorobenzenes                                 C-150
          Tetrachlorobenzenes                               C-150
          Pentachlorobenzene                                C-150
          Hexachlorobenzene                                 C-151
     Current Levels of Exposure                             C-152
          Monochlorobenzene                                 C-152
          Trichlorobenzenes                                 C-152
          Tetrachlorobenzenes                               C-153
          Pentachlorobenzenes                               C-153
          Hexachlorobenzene                                 C-154
     Special Groups at Risk                                 C-155
          Monochlorobenzene                                 C-155
          Tetrachlorobenzene                                C-159
          Pentachlorobenzene                                C-159
          Hexachlorobenzene                                 C-159
     Basis and Derivation of Criteria                       C-161
          Monochlorobenzene                                 C-161
          Trichlorobenzene                                  C-163
          Tetrachlorobenzene                                C-163
          Pentachlorobenzene                                C-164
          Hexachlorobenzene                                 C-165

Appendix                                                    C-169
                                  IX

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                              CRITERIA DOCUMENT
                             CHLORINATED  BENZENES
CRITERIA
                                 Aquatic  Life
    The available data for chlorinated benzenes  indicate that acute  toxicity
to freshwater  aauatic  life occurs at  concentrations  as  low as 250  ug/1  and
would  occur  at lower  concentrations  among species that  are more  sensitive
than those tested.  No data are  available  concerning  the chronic  toxicity of
the more  toxic of the chlorinated benzenes  to sensitive freshwater  aauatic
life but  toxicity occurs at concentrations as  low  as  50  ug/1  for  a  fish  spe-
cies exposed for 7.5 days.
    The  available data  for  chlorinated   benzenes  indicate  that  acute  and
chronic toxicity to saltwater aauatic life occur at concentrations as  low as
160  and  129  ug/1,  respectively,  and would  occur at  lower concentrations
among species that are more sensitive than  those  tested.

                                 Human Health
Monoch1orobenzene
    For comparison  purposes,  two  approaches were used  to  derive  criterion
levels  for  monochlorobenzene.  Based  on   available toxicity data,  for  the
protection of  public health, the derived level is  488 ug/1.   Using  available
organoleptic data,  for controlling  undesirable  taste  and  odor  auality  of
ambient water, the estimated  level is 20 ug/1.   It should be  recognized  that
organoleptic data as a basis  for establishing  a  water auality criteria  have
limitations and have no demonstrated  relationship  to  potential adverse human
health effects.

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 Trichlorobenzenes
     Due  to the insufficiency in the available information for  the  trichloro-
 benzenes,  a  criterion cannot  be  derived  at  this  time  using  the  present
 guidelines.
 1,2,4,5-Tetrachlorobenzene
     For  the  protection of human health from the toxic properties of  1,2,4,5-
 tetrachlorobenzene  ingested  through  water  and  contaminated  aouatic orga-
 nisms, the ambient water  criterion  is  determined to be 38 «g/l.
     For  the  protection of human health from the toxic properties of  1,2,4,5-
 tetrachlorobenzene  ingested  through  contaminated  aauatic  organisms alone,
 the  ambient water criterion  is  determined to be 48 ug/1.
 Pentachlorobenzene
     For  the  protection of human  health from the  toxic  properties  of penta-
 chlorobenzene  ingested  through  water and  contaminated  aauatic organisms,  the
 ambient  water criterion is determined  to be 74 yg/1.
     For  the  protection of human health from  the toxic  properties  of penta-
 chlorobenzene  ingested  through contaminated  aouatic  organisms alone,  the
 ambient water criterion is determined  to be 85 wg/1.
 Hexachlorobenzene
    For  the  maximum  protection, of  human  health from the  potential carcino-
 genic effects due to exposure of  hexachlorobenzene  through  ingestion of con-
 taminated  water  and  contaminated aauatic  organisms,  the ambient water  con-
centration should  be  zero  based on the  non-threshold  assumption   for  this
chemical.  However,  zero level  may not be attainable  at the  present time.
therefore,  the  levels   which  may result  in incremental  increase  of  cancer
risk  over  the  lifetime  are  estimated   at   10~5,   10~6,   and  10~7.   The

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corresponding recommended criteria  are 7.2 ng/1, 0.72  ng/1,  and 0.072 ng/1,
respectively.   If  the above  estimates are made  for consumption  of aauatic
organisms  only,  exlcuding  consumption of  water,  the  levels  are  7.4  ng/1,
0.74 ng/1, and 0.074 ng/1, respectively.
                                   xii

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                                  INTRODUCTION

     The  chlorinated   benzenes,    excluding   dichlorobenzenes,   are  mono-
 chlorobenzene     (cgH5cl)»     1,2,3-trichl orobenzene    (£5^013),    1,2,4-
 trichlorobenzene       (C6H3C13)»       1,3,5-trichlorobenzene       (C6H3C13^'
 1,2,3,4-tetrachlorobenzene        (CgH2Cl4),        1,2,3,5-tetrachlorobenzene
 (CgH2Cl4),      1,2,4,5-tetrachlorobenzene     (Cgh^Cl^),     pentachloroben-
 zene   (CgHCl5),   and  hexachlorobenzene   (CgClg).    Based   on   annual  pro-
 duction  in  the U.S.,  139,105  kkg  of monochlorobenzene were produced in 1975;
 12,849  kkg of  1,2,4-trichlorobenzene,  8,182 kkg  of 1,2,4,5-tetrachloroben-
 zene  and 318  kkg  of hexachlorobenzene were produced  in  1973  (West and Ware,
 1977; U.S.  EPA, 1975a).
    The  remaining chlorinated  benzenes  are produced mainly as  by-products
 from  the production  processes for the above four  chemicals.   Production  and
 use  of  chlorinated  benzenes  results in  34,278  kkg of  monochlorobenzene,
 8,182 kkg  of  trichlorobenzenes and  about  1,500 kkg  of  tetra-,  penta-,  and
 hexa-chlorinated  benzenes entering  the aquatic environment  yearly.  Annual-
 ly, 690  kkg of monochlorobenzene  and  1,628 kkg  of  hexachlorobenzene contami-
 nate  solid  wastes.   Yearly estimates  of  atmospheric contamination  of mono-
 chlorobenzene  and tetrachlorobenzenes  are  362  and  909  kkg,   respectively
 (West and Ware,  1977).
    Chlorination   of  benzene yields  12  different  compounds:  monochloroben-
zene  (CgH5Cl),  three  isomers  of dichlorobenzene  (the  subject  of  another
criterion  document),   three  trichlorobenzenes,  three   tetrachlorobenzenes,
pentachlorobenzene and  hexachlorobenzene.
    All   are colorless  liquids or  solids  with  a  pleasant  aroma.   The most
important properties  imparted  by chlorine  to  these compounds  are  solvent
                                     A-l

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power,  viscosity,  and moderate chemical  reactivity.   All  of the  chloroben-
zenes are heat stable (Kirk and Othmer, 1963; Snell, et al.  1969;  Hampel  and
Hawley, 1973; Stecher, 1968; Mardsen and Marr,  1963).
    Viscosity  data are  not  available for  all  the  chlorinated   benzenes.
Nevertheless,  the  trend  is for viscosity  to  increase  from  chlorobenzene  to
the more  highly chlorinated  benzenes.  The  nonflammability of  these com-
pounds  follows the same trend.  Chlorobenzene  is  flammable;  trichlorobenzene
is nonflammable but gives off combustable fumes;  the remaining compounds  are
nonflammable (Kirk and Othmer, 1963; Mardsen  and Marr,  1963).
    Vapor pressures of the  chlorinated benzenes decrease progressively from
monochlorobenzene  to  hexachlorobenzene,  i.e.,   at  60°C,  the vapor  pressures
of monochlorobenzene,  trichlorobenzenes  and  1,2,3,5-tetrachlorobenzene  are
60, 3 to 4.4 and 2 mm of  mercury,  respectively  (Hampel  and  Hawley,  1973).
    Some  physical  properties of  the  chlorinated  benzenes  are   given   in
Table 1 (Weast, 1975).
    Monochlorobenzene, which is the most polar  compound, is  soluble  in  water
to the  extent  of 488 mg/1  at 25°C (Mellan,  1970;  Mardsen  and Marr,  1963).
Solubilities of  the  other chlorobenzenes in water  were  not available.   The
chlorinated benzenes  are  generally good solvents for  fats,  waxes, oils  and
greases.  These  compounds  have  a  high  lipid  solubility  and are expected  to
accumulate in ecosystems  (Mardsen  and  Marr, 1963;  Mellan,  1970).
    Monochlorobenzene  is  used for  the synthesis  of  ortho  and  para  nitro-
chlorobenzenes (50 percent),  as  a solvent (20  percent),  in phenol manufac-
turing  (10  percent)  and  in DOT manufacturing (7.5  percent).   1,2,4-Trichlo-
robenzene is  used  as  a dye  carrier (46 percent),  a  herbicide  intermediate
(28 percent),  a  heat  transfer medium,  a dielectric fluid  in transformers,  a
degreaser,  a  lubricant  and a  potential  insecticide against termites.   The
                                     A-2

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                                                  TABLE 1



                                Physical  Properties of Chlorinated Benzenes*
Compound
Monoch 1 orobenzene
Trichlorobenzene
1,2,3-
1,2,4-
1,3,5-
"T Tetrachlorobenzene
1,2,3,4-
1,2,3,5-
1,2,4,5-
Pentachlorobenzene
Hexachl orobenzene
MW
112.56
181.45
215.90
250.34
284.79
mp('C)
-45.6
52.6
17
63.4
47.5
54.5
138-140
86
230
bp('C)
131-132
218-219
213.5
208
254
246
243-246
277
322
Density
1.107
143
1.454
145
146
1.858
1.858
2.044
Log Octanol
Water
Partition
2.83
4.23
4.93
5.63
6.43
*Source:  Weast, 1975

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other  trichlorobenzene  isomers  are  not  used  in any quantity.   1,2,4,5-Tetra-
chlorobenzene  is  the  only tetrachloro-isomer used in  industrial  quantities.
Fifty-six  percent  of  the annual consumption  of 1,2,4,5-tetrachlorobenzene is
used in the  production  of  the  defoliant,  2,4,5-trichlorophenoxy  acetic  acid,
33  percent  in  the synthesis  of 2,4,5-trichlorophenol and  11  percent  as  a
fungicide.   Pentachlorobenzene  is  used  in  small  quantities as  a  captive
intermediate  in  the  synthesis  of specialty chemicals  (West and  Ware,  1977).
Hexachlorobenzene ..in  1972 was  used  as  a fungicide  (23  percent) to  control
wheat  bunt  and smut  on seed grains.  Other  industrial uses  (77  percent) in-
cluded  dye manufacturing,  an   intermediate  in organic  synthesis,   porosity
controller  in the manufacturing  of electrodes,  a  wood  preservative and  an
additive in pyrotechnic compositions for the  military (U.S.  EPA,  1975a).
    In recent  years,  hexachlorobenzene  has  become of concern because of its
widespread distribution as an environmental contaminant and a contaminant  of
food products  used for human consumption.  Hexachlorobenzene has been  found
in  adipose tissue and  milk  of  cattle  being  raised  in  the  vicinity of  an
industrialized  region  bordering  the  Mississippi  River  between  Baton  Rouge
and New Orleans,  Louisiana.   Hexachlorobenzene residues have been  found  in
adipose tissue  of sheep in western  Texas and eastern California (U.S.  EPA,
1975b).  The  occurrence  and  effects of  hexachlorobenzene have been  reported
in  many organisms,  e.g., birds  (Vos, et  al.  1971; Cromartie, et al.  1975),
rats (Medline, et  al.  1973), man (Cam  and Nigogosyan,  1963)  and fish  (Hoi-
den, 1970; Johnson, et  al. 1974;  Zitko, 1971).  Magnification in  the natural
food chain is  indicated  by Gilbertson  and  Reynolds  (1972)  observation  of
hexachlorobenzene  in  the eggs   of  common  terns,  which  had  apparently  eaten
contaminated fish.  This  compound has  also  been  found in  samples  of  ocean
water,  and  its persistence in   the  environment has been acknowledged  (Selt-
zer, 1975).
                                     A-4

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    Specimens of levee soil taken from along the Mississippi  River,  known  to
be  contaminated  with hexachlorobenzene  waste,  had  levels of  the  compound
ranging from 107.0 to 874.0 ug/kg (wet weight)  (U.S.  EPA,  1976).
    Among seven samples of sediments taken from the  lower  Mississippi River,
only  one had  detectable  amounts  of  hexachlorobenzene.   The concentration
found was 231  ug/1.   This  site was known to be  contaminated by  hexachloro-
benzene in the past (Laska, et al.  1976).
    The National  Organics Reconnaissance Survey (NORS) tested ten water sup-
plies  for  a  variety  of organic chemcials.  Monochlorobenzene  was  detected
but not  Quantified  in three of  the ten drinking  water  supplies.   Drinking
water  supplies from  83  locations in  EPA Region  V  were analyzed  for various
pesticides and organic  chemicals.   Hexachlorobenzene was  detected  in three
locations with concentrations  ranging from 6  to 10  ng/1.
    The National  Organics Reconnaissance Survey tested ten finished  drinking
waters for a variety of  organic chemicals.
                                     A-5

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                                   REFERENCES

 Cam,  C. and  G.  Nigogosyan.   1963.   Acauired toxic  porphyria cutanea  tarda
 due to  hexachlorobenzene.   Jour. Am. Med. Assoc.   183: 88.

 Cromartie,  E.W.  et  al.   1975.   Residues of organochlorine  pesticides and
 polychlorinated  biphenyls  and  autopsy  data  for  bald  eagles,   1971-1972.
 Pestic. Monit. Jour.  9: 11.

 Gilbertson,  M.  and  L.M.  Reynolds.   1972.   Hexachlorobenzene (HCB)  in the
 eggs  of common terns in  Hamilton  Harbour, Ontario.   Bull.  Environ. Contam.
 Toxicol.  7: 371.

 Hampel,  C.  and  G.  Hawley.   1973.   The  Encyclopedia of  Chemistry,  3rd ed.
 Van Nostrand Reinhold Co., New York.

 Holden, A.V.   1970.   International co-operative  Study of organochlorine pes-
 ticide residues  in terrestrial and aouatic wildlife,  1967, 1968,  1970.  Pes-
 tic. Monit. Jour.  4: 117.

 Johnson,  J.L.,  et  al.    1974.  Hexachlorobenzene (HCB)  residues  in  fish.
Bull.  Environ. Contam. Toxicol.  11:  393.

Kirk,  R.E.  and  O.F.  Othmer.   1963.   Kirk-Othmer Encyclopedia of Chemical
Technology, 2nd ed.   John Wiley and Sons,  New York.
                                     A-6

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Laska,  A.L.,  et al.   1976.   Distribution  of hexachlorobenzene and  hexachlo-
robutadiene  in water  soil  and  selected  aauatic  organisms  along  the  lower
Mississippi River,  Louisiana.  Bull. Environ. Contam. Toxicol.  15:  535.

Mardsen,  C.  and S. Marr.   1963.   Solvents  Guide.   Cleaver-Hume  Press  Ltd.,
London.

Medline,  A.,  et al.  1973.   Hexachlorobenzene  and  rat  liver.  Arch.  Pathol.
96: 61.

Mellan,  I.   1970.   Industrial Solvents.   Noyes Data Corp.   Park Ridge, New
Jersey.

Seltzer,  R.J.   1975.   Ocean pollutants pose potential  danger to  man.   Chem.
Engr. News.  53: 19.

Snell,  D., et al.   1969.    Encyclopedia  of  Industrial  Chemical  Analysis.
Vol. 9.   Interscience Publishers, New York.

Stecher,  P.G.  (ed.)   1968.   The Merck  Index.   9th  ed.   Merck and Co.,  Inc.,
Rahway, New Jersey.

U.S. EPA.  1975a.   Survey of  industrial  processing  data:  Task I,  Hexachloro-
benzene and hexachlorobutadiene pollution  from  chlorocarbon processes.  Mid.
Res. Inst.  Off. Toxic Subs.  U.S. EPA, Washington,  D.C.
                                     A-7

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U.S.  EPA.   1975b.   HCB  review report: Fifth 90-day HCB meeting  and  status  of
HCB studies.

U.S.  EPA.   1976.   An  ecological study  of hexachlorobenzene.   EPA-560/6-76-
009.

Vos,  J.G.,  et  al.   1971.   Toxicity  of  hexachlorobenzene  in  Japanese ouail
with  special  reference  to  prophyria, liver damage,  reproduction,  and tissue
residues.  Toxicol. Appl. Pharmacol.  18:  944.

Weast, R.C.  (ed.)   1975.  Handbook  of  Chemistry and  Physics.   The Chemical
Rubber Co., Cleveland, Ohio.

West,   W.I.  and S.A.  Ware.    1977.   Preliminary  Report.   Investigation  of
Selected Potential  Environmental  Contaminants: Halogenated  Benzenes.   Envi-
ron. Prot.  Agency,  Washington, O.C.

Zitko, V.   1971.   Polychlorinated biphenyls  and organochlorine pesticides in
some freshwater and marine  fish.  Bull.  Environ. Contam. Toxicol.  6: 464.
                                     A-8

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 Aquatic Life Toxicology*
                                  INTRODUCTION
      This discussion does not  cover  any dichlorobenzene since tney  are  dis-
 cussed in a separate criterion  document.   There is a diversity  of  toxicolo-
 gical  data with numerous species, and there  is  a  consistent  direct  relation-
 ship between toxicity  and bioconcentration  and degree  of  chlorination  for
 fish,  invertebrate, and plant species.  The  acute  toxicity of  hexachlorooen-
 zene  is  difficult  to   ascertain  since  acute  mortality  for  a  variety  of
 species appears to occur at or  above  solubility.
                                    EFFECTS
 Acute  Toxicity
      The 48-hour  ECgQ  values  reported  for  Daphnia  magna (U.S.  EPA, 1978)
 are  (ug/1):   chlorobenzene,  86,000;  1,2,4-trichlorobenzene, 50,200;  1,2,3,5-
 tetrachlorobenzene,  9,710;  and  pentachlorobenzene,  5,280  (Table  1).    The
 48-hour  ECgQ   value  for  1,2,4,5-tetrachlorobenzene  was  greater  than   the
 highest exposure  concentration,  530,000  ug/1  (Table 5).  The 48-hour  EC5Q
 for  three dichlorobenzenes  and Daphnia  magna  ranged  from 2,440  to 28,100
 ug/1.   For  Daphnia magna  the  toxicity  of  chlorinated  benzenes generally
 tended  to  increase as the degree of chlorination increased.
     No  marked  difference  in  sensitivity  between   fish  and   inverteorate
 species  is  evident from the available data.   Pickering  and  Henderson (1966)
 reported  96-hour  LC5Q  values  for   goldfish,   guppy,   and  oluegill  to   be
 51,620, 45,530,  and  24,000 ug/1, respectively,  for cnlorobenzene  (Table  1).
*The  reader  is  referred  to   the  Guidelines  for  Deriving  Water  Quality
Criteria for the Protection of  Aquatic  Life  and Its Uses  in  order  to better
understand  the  following  discussion  and  recommendation.   The  following
tables contain  the  appropriate  data that were  found in the  literature,  ana
at the  bottom  of each  table  are calculations  for  deriving various measures
of toxicity as described in the Guidelines.
                                     3-1

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 The  96-hour LC5Q  values  for chlorobenzene  and  fathead minnows  were 33,930
 and  29,120  yg/1  in soft water  (20  mg/1)  and 33,930 yg/l  in  hard water (360
 mg/1)  (Table 1).   This  indicates  that hardness does not significantly affect
 the  toxicity  of  chlorobenzene.    U.S.   EPA  (1978)  reported  96-hour  LC5Q
 values   for  bluegill  exposed   to   chlorobenzene,  1,2,4-trichlorobenzene,
 1,2,3,5-tetrachlorobenzene,  1,2,4,5-tetrachlorobenzene  and  pentachloroben-
 zene to  be  15,900, 3,360,  6,420,  1,550 and  250  wg/l,  respectively.  Compar-
 able tests  (U.S.   EPA, 1978) were conducted with  three  dichlorobenzenes and
 the  96-hour LC5Q  values  ranged  from 4,280 to  5,590  pg/1.   Only 1,2,3,5-
 tetrachlorobenzene  is  an apparent anomaly  in  the trend of  increasing toxi-
 city with increasing chlorination.
     Mysid  shrimp,  the  only saltwater invertebrate species  tested, was more
 sensitive to  three of five  chlorinated  benzenes than the  sheepshead minnow
 and  more sensitive to  all  chlorinated benzenes  tested  than the freshwater
 cladoceran, Daphnia  magna  (Table 1).  Chlorobenzene  (96-hour  LC50 = 16,400
 ug/1) was the least toxic  to mysid  shrimp,  while pentachlorobenzene (96-hour
 LC5Q  =  160 ug/1)  was  the most acutely  toxic.   As  with  the  freshwater
 species, and as will  be  seen with the sheepshead minnow,  sensitivity to the
 chlorinated benzenes  (including the  dichlorobenzenes) generally increased  as
 chlorination increased.
     Toxicity  tests  with  the   sheepshead minnow  have   also  been  conducted
 (U.S.  EPA,  1978)   with  five chlorinated  benzenes  (Table  1)'.  As  with  the
mysid  shrimp,  all  tests  were conducted  under  static conditions  and  concen-
trations in water  were  not measured.  Concentrations  acutely  toxic  to  this
saltwater fish species  were  relatively high for the lower chlorinated  ben-
zenes and toxicity generally increased with  increasing chlorination;  96-hour
  50 values for  sheepshead minnows  and  dichlorobenzenes  (7,440  to  9,660
ug/1) were sightly lower than that for chlorobenzene.  The  sheepshead minnow
                                     B-2

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 was generally more acutely sensitive to the chlorinated benzenes,  except  for
 1,2,4-trichlorobenzene  and pentachlorobenzene,  than were the freshwater fish
 species   tested  similarly  (Table  1);   96-hour  LC5Q  values  for  sheepshead
 minnows  and bluegills  differed  by factors of 1.5  to  6.4.   The 96-hour IC™
 values  for sheepshead  minnows  ranged from 21,400  pg/1  for 1,2,4-trichloro-
 benzene  to 830 u9/l  for pentachlorobenzene.
 Chronic  Toxicity
      No   chronic  test  has been  conducted  with  any  invertebrate species.
 However,  five embryo-larval  tests have  been conducted  with chlorinated ben-
 zenes  and the fathead  minnow  and the  sheepshead  minnow (Table 2).  Chronic
 values  for the fathead minnow  and  1,2,4-trichlorobenzene  have  been  deter-
 mined  in  two studies   (U.S.   EPA, 1978,  1980).   These  results  provide  an
 acute-chronic  ratio of  6.4  (geometric  mean of  the ratios  10 and  4.1)   for
 this  species  and  1,2,4-trichlorobenzene (Table 2).  The fathead  minnow is a
 little more sensitive to 1,2,3,4-tetrachlorobenzene when tested  by the same
 investigators  (U.S. EPA,  1980), with  a chronic  value  of  318 Mg/l  and  an
 acute-chronic ratio of  3.4 (Table 2).
     The  chronic values for the sheepshead minnow and 1,2,4-trichlorobenzene
 and  1,2,4,5-tetrachlorobenzene  are  222 and  129  wg/1,  respectively  (Table
 2).   The  LC5Q  values  obtained  by  the same  investigator  (U.S.  EPA,  1978)
 are used  to calculate the  acute-chronic ratios of 96 and 6.5  for 1,2,4-tri-
 chlorobenzene  and  1,2,4,5-tetrachlorobenzene,  respectively  (Table  2).   The
 acute-chronic ratio of  96  for  the sheepshead minnow  and  1,2,4-trichloroben-
 zene  is   atypical   when   compared  to  the narrow  range  of  ratios  for  other
 species and/or chlorinated benzenes of 3.4  to 10.
Plant Effects
     Ninety-six-hour  EC5Q  tests,  using chlorophyll  a_  inhibition and  cell
number production  as  measured  responses,  were conducted  with  the  alga,
                                      3-3

-------
Selenastrum capricornutum (Table 3).  The  effects  of chlorinated benzenes on
this alga  generally increased as  chlorination  increased,  but the  trend was
not smooth.  The  alga  was considerably less  sensitive  than  fish and Daphnia
magna with  96-hour  ECgQ values ranging  from 6,630  ug/1  for pentachloroben-
zene to 232,000 ug/1 for chlorobenzene.
     The  saltwater  alga,  Skeletonema costatum,  was  less  sensitive  to the
chlorinated benzenes than  the mysid  shrimp  or  sheepshead minnow  (Table 3).
Ninety-six-hour EC5Q  values for growth,  based  on  concentrations  of chloro-
phyll £ in culture, were comparable  to 96-hour ECgQ  values calculated from
cell  numbers   and,  except  for chlorobenzene,  ECcn  values   for  Skeletonema
costatum  were   3  to  25  times  lower  than  ECcn  values  for  the  freshwater
alga.   Those ECgn values for  the  saltwater alga based on chlorophyll  a^ and
cell numbers,  respectively,  are:   343,000  and  341,000 wg/1  chlorobenzene;
8,750 and 8,930 wg/l 1,2,4-trichlorobenzene;  830  and 700  ug/1 1,2,3,5-tetra-
chlorobenzene;   7,100  and 7,320  ug/1  1,2,4,5-tetrachlorobenzene;   and  2,230
and 1,980 yg/1  pentachlorobenzene.
     There are no data  reported on effects of chlorinated  benzenes on fresh-
water or saltwater vascular plants.
Residues
     Data which are adequate for  computing  acceptable bioconcentration fac-
tors are  available for several  chlorinated benzenes.   After  28-day  expo-
sures,  the  steady-state bioconcentration  factors  for  bluegill  (whole body)
for  pentachlorobenzene,  1,2,3,5-tetrachlorobenzene,   and   1,2,4-trichloro-
benzene are 3,400,  1,800, and 182,  respectively (Table 4).   The  half-lives
for these compounds were between 2 and  4  days for 1,2,3,5-tetrachlorobenzene
and  1,2,4-trichlorobenzene  and greater  than 7  days  for  pentachlorobenzene
(U.S. EPA,  1978).   Hexachlorobenzene has  also  been tested,  and  the fathead
minnow  (whole  body) bioconcentrated  that compound  22,000  times  (Table 4).
                                      8-4

-------
 For three dichlorobenzenes the bioconcentration  factors  ranged  from 60 to 89
 (U.S.  EPA, 1978); these  results  are discussed in the  criterion  document for
 that group of compounds.
      Bioconcentration  factors  correlate well  with  an  increase  in  chlorine
 content.  The sequence of  measured  bioconcentration factors are  72  (mean of
 dichlorobenzene data), 182 (1,2,4,-trichlorobenzene), 1,800  (1,2,3,5-tetra-
 chlorobenzene),  3,400  (pentachlorobenzene),  and 22,000  (hexachlorobenzene)
 for freshwater species.
      Hexachlorobenzene is  bioconcentrated  from  water  into tissues  of salt-
 water  organisms  (Tables 4 and 5).  Bioconcentration factors range from 1,964
 to 23,000 for fish  and shellfish (Parrish, et al.  1974).  However,  the bio-
 concentration factors for fish and  invertebrate  species  exposed  for only 96
 hours  probably underestimate steady-state  factors for organisms  chronically
 exposed  to hexachlorobenzene.   Bioconcentration  factors  for  grass shrimp,
 pink  shrimp,  and sheepshead  minnows  exposed  to hexachlorobenzene  for  96
 hours  ranged  from  1,964 to 4,116 while  the bioconcentration  factor for  pin-
 fish  was  15,203  (Table  4).   Concentrations  of  hexachlorobenzene  in   these
 whole-body samples were probably  not at  equilibrium  after such a  short  expo-
 sure period;  highly  chlorinated  compounds generally  do not reach  equilibrium
 in  exposed animals in short exposure periods.
     The bioconcentration factor  in  the  flesh  of pinfish  exposed  for 42 days
 to  hexachlorobenzene  was  23,000  (Table  4)  for the  five  exposure concentra-
 tions  tested  (0.06 to  5.2  ug/1).   Analysis  of  the concentrations  of  hexa-
 chlorobenzene in pinfish  indicates that  concentrations  after  7  days of  expo-
 sure were approximately one quarter  of the  total  concentration  after 42 days
of  exposure;  concentrations after  42 days  of exposure  appear  to be  near
equilibrium.    Concentrations  of  hexachlorobenzene  in pinfish  muscle  were
reduced  only  16 percent  after 28  days  of depuration;  this  slow rate  is
                                     B-5

-------
similar to that for DDT  in  fishes  (Parrish,  et al.  1974).   Since hexachloro-
benzene bioconcentrated to high concentrations  in all  tissues  of pinfish  and
depuration  was slow   compared  to  several   other  organochlorine  pesticides
(Parrish, et  al.   1974),  this compound  has   a high  potential   for  transfer
through and retention in aquatic food webs.
Miscellaneous
     A variety  of  data on  other  adverse effects on  freshwater  organisms  is
presented in  Table 5.   Biconcentration  factors derived  from  a model  eco-
system (Isensee, et  al.  1976) ranged from 730 to  9,870 but it  could  not  be
determined if these were steady-state results.
     Birge,   et al.  (1979)  exposed  rainbow   trout  embryos for  16 days  and
goldfish  and largemouth bass  embryos  and  larvae for up to 4 days  post-hatch
to chlorobenzene and  observed total  mortality of  the rainbow trout  embryos
at the lowest  measured  exposure  concentration of 90  ug/1.  Hardness  did  not
affect the  LC5Q values  for the goldfish  (880 and  1,040  ug/1)  or  the  much
more sensitive largemouth bass (50 and 60 wg/1).
     As  mentioned  in  the introduction,  the  acute toxicity of  hexachloroben-
zene is  difficult  to  determine.   Tests  with   a midge,  Tanytarsus dissimilis,
rainbow  trout,  fathead  minnow,  and the  bluegill (U.S  EPA,  1980) produced  no
LCrQ values  at concentrations  of hexachlorobenzene  above what  appeared  to
be its solubility limit.
Summary
     In  general,   the  toxicity  of  the  chlorinated  benzenes  to  freshwater
organisms  increases  with  increasing chlorination.    Chlorobenzene is  least
toxic  with  50  percent  effect  concentrations  for  Daphnia magna,  goldfish,
fathead  minnows,  guppy,  and  bluegill  in  the  range  of concentrations  from
15,900 ug/1 to  36,000  yg/1  with  the  cladoceran being a little  more resistant
tnan the  tested  fish  species.  The  dichlorobenzenes,  discussed  in  detail  in
                                      3-6

-------
 another  document,  are  slightly  more  toxic  than chlorobenzene.   Toxicity
 reaches  its  maximum with acute  effect concentrations  of  pentachlorobenzene
 in the  range of 250 ug/1 for  the  bluegill  to 5,280  ug/1  for  Daphnia magna.
 Embryo-larval tests  have been  conducted with  the fathead  minnow,  and  the
 chronic values  are  286 and  705  ug/1  for 1,2,4-trichlorobenzene  (two tests)
 and 318 ug/1 for  1,2,3,4-tetrachlorobenzene.  Acute-chronic  ratios  from fat-
 head  minnow data were  6.4 for  1,2,4-trichlorobenzene  and 3.4 for 1,2,3,4-te-
 trachlorobenzene.   A  freshwater  algal  species  also  was   more  sensitive  to
 more  highly  chlorinated benzenes  with 96-hour EC™  values for  chlorophyll
 a_ in  the  range  of 232,000  ug/1  for chlorobenzene  to  6,780 ug/l for  penta-
 chlorobenzene.   The bioconcentration  of chlorinated benzenes also  increased
 with  increasing  chlorination.  The  whole body  bioconcentration factors  in-
 creased   from  182  for  1,2,4-trichlorobenzene  to   22,000  for hexachloroben-
 zene.   Acute lethal  effects  in a  midge, rainbow  trout,  fathead minnow,  and
 bluegill  were not observed at concentrations approximating  the  solubility  of
 hexach1orobenzene.
     As  with  the  freshwater  toxicity  tests  with  fish   and   invertebrate
 species,  there was an  increase in  effects  with the more  highly chlorinated
 compounds  with  at  least  a one order  of magnitude  decrease  in  96-hour LC™
 values  between   chlorobenzene  and  pentachlorobenzene  for  the  mysid  shrimp
 (16,400  and  160  ug/1)  and  the  sheepshead  minnow (10,500  and  830  ug/l).
 Chronic values for the  sheepshead minnow were 222  ug/1  for  1,2,4-tricnloro-
 benzene  and  129  ug/l  for  1,2,4,5-tetrachlorobenzene.  A  saltwater  algal
 species   was more  resistant   than  the  fish  and  invertebrate species,  with
 96-hour   EC50 values  for chlorophyll  a_  in  the  range of  343,000  yg/l  for
 chlorobenzene to  2,230  ug/1   for  pentachlorobenzene.  Bioconcentration  fac-
 tors for hexachlorobenzene were as high  as 23,000  for  edible portions  of the
pinfish.
                                      5-7

-------
                                   CRITERIA
     The  available  data  for  chlorinated  benzenes   indicate   that  acute
toxicity to  freshwater  aquatic life occurs  at concentrations as  low as 250
ug/1 and  would occur at  lower concentrations  among species  that are  more
sensitive than  those  tested.  No  data  are available concerning  the  chronic
toxicity  of  the  more   toxic  of  the  chlorinated  benzenes  to  sensitive
freshwater aquatic  life  but toxicity occurs  at concentrations as  low  as 50
ug/1 for fish species exposed for 7.5 days.
     The  available  data  for  chlorinated  benzenes  indicate  that  acute and
chronic toxicity to saltwater  aquatic life occur at  concentrations  as low as
160  and 129  yg/1,  respectively,  and  would  occur  at  lower  concentrations
among species that are more sensitive than those tested.
                                      3-8

-------
Table I.  Acute values for chlorinated benzenes
Species
Cladoceran,
Daphnla magna
C ladoceran,
Daphnla magna
Cladoceran,
Daphnla magna
Cladoceran,
Daphnla magna
Rainbow trout.
Sal mo galrdnerl
Goldfish,
Carasslus auratus
W
1 Fathead minnow,
*° Plmephales promelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Guppy,
Poecl lla retlculata
Blueglll,
Lepomis roacrochlrus
Blueyll I,
Lepomis macrochlrus
Method*
S, U
s, u
S, U
s, u
FT, M
s, u
S, U
s, u
S, U
FT, M
FT, M
S, U
S, U
S, U
Chemical
FRESHWATER
ch lorobenzene
1,2,4-trlchloro-
benzene
1,2,3,5-tetra-
ch lorobenzene
pent ach loro-
benzene
1,2,4-trlchloro-
benzene
ch lorobenzene
ch lorobenzene
ch lorobenzene
ch lorobenzene
1, 2, 4-trlch loro-
benzene
1,2.3,4-tetra-
ch lorobenzene
ch lorobenzene
ch lorobenzene
ch lorobenzene
LC50/EC50
Cug/l)
SPECIES
86,000
50,200
9,710
5,280
1,500
51,620
33,930
29,120
33,930
2,870
1,070
45,530
24,000
15,900
Species Acute
Value (uq/l)
86,000
50,200
9,710
5,280
1,500
51,620
~
32,200
2,870
1,070
45,530
19,500
Reference
U.S. EPA, 1978
U.S. EPA, 1978
U.S. EPA, 1978
U.S. EPA, 1978
U.S. EPA, 1980
Pickering &
Henderson, 1966
Pickering A
Henderson, 1966
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
U.S. EPA, 1980
U.S. EPA, 1980
Pickering &
Henderson, 1966
Pickering &
Henderson, 1966
U.S. EPA, 1978

-------
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09
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                      Table  I.  (Continued)
Species Method*
Sheepshead minnow, S, U
Cyprlnodon yarlegatus
Chemical
pent ach lor o-
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LC50/EC50 Species Acute

-------
                                                    Table 2.   Chronic values for chlorinated  benzenes
CXI
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Species Method*
Fathead minnow, ELS
Plmephales promelas
Fathead minnow, ELS
Plntephales promelas
Fathead minnow, ELS
Pimephales promelas
Chemical
FRESHWATER SP
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1, 2,4- trich loro-
benzene
1,2,3,4-tetra-
ch lorobenzene
Llnlts
(M9/D
ECIES
200-410
499-995
245-412
Chronic
Value
(ug/l)
286
705
318
Reference
U.S. EPA, 1978
U.S. EPA, 1980
U.S. EPA, 1980
SALTWATER SPECIES
Sheepshead winnow, ELS
Cyprlnodon varleqatus
Sheepshead minnow, ELS
Cyprlnodon varleqatus
* ELS = Early life stage
1,2, 4- trlch loro-
benzene
1,2,4,5-tetra-
ch lorobenzene

150-330
92-180

222
129

U.S. EPA, 1978
U.S. EPA, 1978

Acute-Chronic Ratio
Species
Fathead minnow
Plmephales promelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Sheepshead minnow.
Chemical
1, 2, 4-trich loro-
benzene
1, 2, 4-trlch loro-
benzene
1,2,3,4-tetra-
ch lorobenzene
1,2,4-trlchloro-
Acute
Value
(ug/l)
2,870
2,870
1,070
21,400
Chronic
Value
(ug/l)
286
705
318
222
Ratio
10
4.1
3.4
96
                                       Cyprlnodon varlegatus
                                                                    benzene

-------
                        Table 2.  (Continued)
                                                                  Acute-Chronic Ratio


Species
Sheepshead minnow.
Cyprlnodon varlegatus


Chemical
1, 2,4,5- tetra-
chlorobenzene
Acute
Value
(M9/U
640

Chronic
Value
(M9/D
129



Ratio
6.5

W

M
OJ

-------
                                           Table 3.  Plant value* tor chlorinated benzenes (U.S.  EPA.  1978)
O>

Species
Alga,
Se 1 enastrum capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastruro capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastrua capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Selenastrum capr 1 cornutum
Alga,
Skeletonema costatum
Alga,
Cbola+nnaiut rostatuffl

Chemical
FRESHWATER SPECIES
ch lorobenzene
ch lorobenzene
1, 2, 4-trlch loro-
benzene
1, 2, 4-trlch loro-
benzene
1, 2,3,5- tetra-
ch lorobenzene
1,2.3,5-tetra-
ch lorobenzene
1.2,4,5-tetra-
ch lorobenzene
1, 2,4,5- tetra-
ch lorobenzene
pent act) loro-
benzene
pentach loro-
benzene
SALTWATER SPECIES
ch lorobenzene
ch lorobenzene

Effect
Chlorophyll a
96- hr EC50
Cel 1 numbers
96-hr EC50
Chlorophyll a
96-hr EC50
Cell numbers
96-hr EC50
Chlorophyll a
96-hr EC50
Cell numbers
96- hr EC50
Ch lorophy 1 1 a
96-hr EC50
Cell numbers
96-hr EC50
Ch lorophy 1 1 a
96-hr EC50
Cel 1 numbers
96-hr EC50
Chlorophyll a
96-hr EC50
Cel 1 numbers
96-hr EC50
Result
(ug/D
232,000
224.000
35,300
36,700
17,200
17,700
52,900
46,800
6,780
6,630
343,000
341,000

-------
                                        TabU 3.   (ContlMMd)
O)
 I
I-*
01
Species
Skalatonena costatum
Alga,
Skeletoned costatu*
Alga.
Skeletonama costatum
Alga.
Skeletonema costatum
Alga.
Skeletoned costatum
Alga,
SkalatoneM costatuai
Alga,
Skeletonema costatum
Alga.
Skeletonema costatum
Chemical
1,2.4-trlchloro-
banzene
1.2.4-trlchloro-
banzene
1,2,3,5-tetra-
ch lorobanzene
1.2.3,5-tetra-
chlorobenzene
1 2,4,5-tetra-
ch lorobenzene
1.2,4,5-tetra-
ch lorobenzene
peatach loro-
benzene
pent ach loro-
benzene
Effect
Chlorophyll a
96-hr EC50 ~
Cell numbers
96-hr EC50
Chlorophyll a
96-hr EC50 ~
Cell numbers
96-hr EC50
Chlorophyll a
96-hr EC50 ~
Cell numbers
96-hr EC50
Chlorophyll a
96-hr EC50 ~
Cell numbers
96-hr EC50
ftoawlt
8,750
8,930
830
700
7,100
7,320
2.230
1,980

-------
                                    Table 4.  Residues for chlorinated benzenes

                                                                BIoconcentratIon
                                                                     Factor
Duration
 (days)      Reference
03
t
I-1
CT>
species
Bluegill,
Lepomls macroch 1 rus
Bluegill,
Lepomls macrochirus
Bluegill,
Lepomls macrochirus
Fathead minnow,
Plmephales promelas
Plnfish.
Lagodon rhoroboldes

FRESHWATER SPECIES
whole body 1.2,4-trlchloro- 182 28 U.S. EPA, 1978
benzene
whole body 1,2,3,5-tetra- 1.800 28 U.S. EPA, 1978
ch lorobenzene
whole body pentachloro- 3,400 28 U.S. EPA, 1978
benzene
whole body hexachloro- 22,000 30 U.S. EPA, 1980
benzene
SALTWATER SPECIES
edible hexachloro- 23.000* 42 Parrlsh, et al. 1974
portion benzene
• Mean concentration factor In 25 muscle samples.

-------
                                                     Table 5.  Other data for chlorinated benzenes
CD
 I
                   Species
                   Alga.
                   ChioreI la pyrenoldosa

                   Alga,
                   Oedogonlum cardlacum

                   Snail,
                   He 11soma sp

                   Cladoceran,
                   Daphnla magna

                   Cladoceran,
                   Daphnla magna

                   Re'd swamp crayfish,
                   Procambarus clarkl
                   Midge,
                   Tanytarsus dlsslmllls

                   Rainbow trout  530,000
B 1 oconcentrat 1 on
factor = 910
Mortality LC50 not
reached
at 27.3
Non- lethal at 57
approx. saturat Ion
100} mortality 90
LC50 258
Non- lethal at 80
approx. saturat Ion
Estimated steady
state bioconcentra-
tion factor = 7,800
B loconcentrat Ion
factor = 690
Reference
Gelke & Par as her,
I976b
Isensee. et al. 1976
Isensee, et al. 1976
U.S. EPA, 1978
Isensee, et al. 1976
Laska, et al. 1978
U.S. EPA, 1980
Birge, et al. 1979
U.S. EPA, 1980
U.S. EPA, 1980
Neely, et al. 1974
Zltko & Hutzlnger,
1976
ch lorobenzene
8 days     LC50 at 50 mg/l
           hardness
                                                       880
Blrge, et al. 1979

-------
Table 5.  (Continued)
                                 Chemical
Goldfish (embryo-larval),
Carassius auratus
Fathead minnow,
Plmephales promelas
Channel catfish,
Ictalurus punctatus
Mosquitof ish,
Gambusla af finis
Bluegi II,
Lepomis macrochirus
Largemouth bass
(embryo- larval),
,-jj Micropterus salmoides
t-> Largemouth bass
00 (embryo-larval),
Micropterus salmoides
Largemouth bass,
Micropterus salmoides
Protozoan,
Tetrahymena pyrlformls
Grass shrimp,
Pal aemonetes pugio
Pink shrimp,
Penaeus duorarum
Pink shrimp,
Penaeus duorarum
ch lorobenzene
hexach loro-
benzene
hexach loro-
benzene
hexach loro-
benzene
hexach loro-
benzene
ch lorobenzene
ch lorobenzene
hexach loro-
benzene
hexach loro-
benzene
hexach loro-
benzene
hexach loro-
benzene
hexach loro-
benzene
FRESHWATER
8 days
4 days
8 days
3 days
4 days
7.5 days
7.5 days
10 and
15 days
SALTWATER
10 days
96 hrs
96 hrs
96 hrs
SPECIES
LC50 at 200 mg/l
hardness
Non- lethal at
approx. saturation
B ioconcentrat ion
factor = 9,870
B ioconcentrat ion
factor =1,580
Non- lethal at
approx. saturation
LC50 at 50 mg/l
hardness
LC50 at 200 mg/l
hardness
No mortality
SPECIES
Decrease growth
Mean bioconcentra-
tion factor = 4, 1 16
Mean bloconcentra-
tion factor = 1,964
33? mortality
during exposure to
OK . . ~ / 1
                                                                                   Result
                                                                                   (|ig/»     Reference
                                                                                    1,040     Blrge, et al.  1979


                                                                                        4.8   U.S. EPA, 1980  ,


                                                                                              Isensee, et al.  1976


                                                                                              Isensee, et al.  1976


                                                                                       78     U.S. EPA, 1980


                                                                                       50     Blrge, et al.  1979
                                                                                       60
Birge, et al. 1979
                                                                                      10  and   Laska,  et  al.  1978
                                                                                       26
                                                                                         1     Gelke  & Prasher,
                                                                                               1976

                                                                                              Parrlsh, et al.  1974
                                                                                               Parrlsh,  et al.  1974
                                                                                               Parrish,  et al.  1974

-------
                    Table 5.  (Continued)
00
 I
I-1
<£>
                    Species

                    Sheepshead minnow,
                    CyprInodon var i egatus

                    Plnflsh,
                    Lagodon rhomboides
 Chemical        Duration          Effect

hexachloro-       96 hrs      Mean bioconcentra-
  benzene                     tion factor = 2,254

hexachloro-       96 hrs      Mean bioconcentra-
  benzene                     tIon factor - 15,203
Result
(ua/l)
Reference

Parrlsh, et al.  1974
           Parrish, et  al.  1974

-------
                                  REFERENCES

Birge, W.J.,  et  al.  1979.  Toxicity  of organic chemicals  to  embryo-larval
stages of fish.  EPA-560/11-79-007.  U.S. Environ. Prot.  Agency.  69 p.

Geike, F.  and C.D. Parasher.   1976a.   Effect of hexachlorobenzene  (HCB)  on
growth of Tetrahymena pyriformis.  Bull. Environ. Contain. Toxicol.  16: 347.

Geike, F.  and C.D.  Parasher.   1976b.    Effect  of hexachlorobenzene on  some
growth   parameters  of  Chiore 11 a   pyrenoidosa.   Bull.   Environ.   Contam.
Toxicol.  15: 670.

Isensee, A.R.,  et  al.   1976.   Soil  persistence  and  aquatic bioaccumulation
potential of hexachlorobenzene (HCB).  Jour. Agric. Food Chem.  24: 1210.

Laska, A.L.,  et  al.  1978.  Acute  and chronic  effects  of hexachlorobenzene
and  hexachlorobutadiene  in  Red   Swamp  Crayfish  (Procambarus  clarki)  and
selected fish species.  Toxicol. Appl.  Pharmacol.  43: 1.

Neely, W.B.,  et  al.  1974.  Partition  coefficient to measure  bioconcentra-
tion potential of  organic chemicals in fish.  Environ. Sci. Tech.  8: 1113.

                                  •
Parrish, P.R., et  al.   1974.   Hexachlorobenzene:  Effects  on  several  estua-
rine  animls.   In:  Proc. 28th  Annu^  Conf. S^E. Assoc.  Game  Fish Comm.  p.
179.
                                     B-20

-------
Pickering, Q.H.  and  C. Henderson.   1966.   Acute toxicity  of  some important
petrochemicals to fish.  Jour. Water Pollut. Control Fed.  38:  1419.

U.S.  EPA.   1978.  In-depth  studies on health  and environmental  impacts  of
selected  water  pollutants.   U.S.  Environ.   Prot.   Agency,   Contract  No.
68-01-4646.

U.S.  EPA.   1980.   Unpublished  laboratory  data.   Environmental  Research
Laboratory - Duluth.

Zitko,  V.  and 0.  Hutzinger.   1976.  Uptake  of chloro-  and  bromobiphenyls,
hexachloro- and  hexabromobenzene  by fish.   Bull.  Environ. Contam.  Toxicol.
16: 665.
                                     B-21

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                        .MONOCHLOROBENZ ENE
Mammalian Toxicology  and Human Health Effects
                            INTRODUCTION
     Monochlorobenzene  (MCB)  is used industrially  both as a  syn-
thetic  intermediate  and as a solvent.   As  a synthetic  intermedi-
ate,  it is primarily  used in the  production  of phenol,  DDT and
aniline.  Because  it  is noncorrosive,  it has technological use as
a  solvent  for a large  number  of compounds  in  the  manufacture of
adhesives, paints, polishes, waxes, diisocyanates, Pharmaceuticals
and natural rubber.
     Data derived from  U.S. International Trade Commission reports
show that between 1966  and  1975, the U.S. annual production of MCB
decreased by  50  percent from approximately  600  million pounds to
approximately 300 million  pounds  (U.S.  EPA,  1977).   It is, as ex-
pected  from  its  structure, lipophilic and  hydrophobic, its solu-
bility  in water being about 100 parts per million.  The log of the
octanol  to  water partition  coefficient  for MCB  is 2.83.   Mono-
chlorobenzene also has  a relatively high vapor pressure (9 torr at
20°C).  As will  be  seen from the next section,  this  is an impor-
tant consideration  in estimating the  likely retention  of  MCB in
surface waters.
                             EXPOSURE
Ingestion from Water
     Based on the vapor pressure, water  solubility,  and molecular
weight of chlorobenzene, Mackay and Leinonen (1975)  estimated the
half-life of  evaporation  from water  for MCB to  be 5.8  hours as
compared to 4.8 hours for benzene and 73.9 hours for DDT.
                               C-l

-------
     MCB has  been  detected  in ground water,  "uncontaminated"  up-



land water, and in waters contaminated either by industrial, muni-



cipal, or  agricultural waste.   It has been  identified  in textile



plant effluents (Erisman and Goldman, 1975) .   Table  1  consists of



a compilation of data from other EPA reports and shows the results



of various water surveys as related to MCB.  Considering the vola-



tile nature of  MCB,  these  data should be  considered from a point



of view of gross estimate of exposure.  For example,  in the aHaly-



sis  of  the water  for Lawson's  Fork Creek,  South  Carolina,  the



range  indicated is  the  result  of  two  analyses  four  days apart



(U.S. EPA,  1977).    The  presence of  MCB  at other sites  has been



demonstrated  qualitatively  by  volatile  organic analysis.   It  has



been detected in "uncontaminated" upland  water in Seattle, Wash.,



(Erisman  and  Goldman, 1975)  and in  raw water contaminated with



agricultural runoff  in Ottumwa, Iowa  and Grand Falls, North Dakota



(U.S. EPA, 1977).   Some  information  is  available  which might give



insight as  to the source of  contamination.   For example,  it  has



been estimated  that  during  the manufacture of MCB,  800 mg escape



into column water  streams  for  every  kg  manufactured.  Another  4 g



of MCB per kg manufactured is recovered from fractionating  columns



for land disposal  (U.S. EPA, 1977).



Ingestion  from  Food



     Lu and Metcalf  (1975) determined the  ecological magnification



of  MCB in various  aquatic  species.    Their  data  are  shown in



Table 2.   For the  purposes  of  comparison, the ecological  magnifi-



cation  of  aldrin and DDT  in  mosquito fish was 1,312  and  16,960,



respectively.
                               C-2

-------
                                                TABLE 1

                              Examples of Occurrence of Monochlorobenzene
        Location
                                   Source
                                       Concentration
                                           (ug/D
o
i
ui
Miami, PL

Philadelphia, PA


Cincinnati, OH


New *ork, NY


Lawrence, MA


Terrebone Parish, LA


Lawsons Fork Creek, SC

Coosa River, GA
Ground water

Raw water contaminated
with municipal waste

Raw water contaminated
with industrial discharge

"Uncontaminated" upland
water

Raw water contaminated
with industrial discharge

Raw water contaminated
with municipal waste

Industrial discharge

Municipal
   1.0

   0.1


0.1 - 0.5


   4.7


   0.12


   5.6


8.0 - 17.0

  27.0
         Source: U.S. EPA, 1975;  1977,

-------
                           TABLE 2
        Ecological Magnification of  Monochlorobenzene
                 in Various Aquatic  Organisms*
        Species                      Ecological Magnification
                                       *
     Mosquito fish                              645
       Gambusia af finis

     Mosquito larvae                           1292
       Culex quinquifasciatus

     Snails                                    1313
       Physa

     Daphnia                                   2789
       Daphnia magna
     Algae                                     4185
       Oedogonium cardiacum

*Source: Lu and Metcalf, 1975;  U.S.  EPA,  1977.
                             C-4

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      Further  data by Lu and Metcalf  (1975)  indicate that MCB  re-
 sists  biodegradation.   They determined the  biodegradability  index
 (BI)  which  was  defined as  the ratio of polar products of  degrada-
 tion  to the nonpolar products.  For MCB, the BI ranged  from  0.014
 to  0.063  in the  organisms  shown in Table 2.  The low value for BI
 was similar to that  seen for DDT and aldrin.  For example,  in mos-
 quito fish  the BI for  MCB  was  0.014,  for DDT it was 0.012 and  for
 aldrin it was 0.015.
     A bioconcentration factor (BCF)  relates the concentration of
 a chemical  in aquatic animals, to  the  concentration in  the  water
 in  which  they live.   The  steady-state BCFs for  a lipid-soluble
 compound  in the  tissues of various  aquatic  animals  seem  to be pro-
 portional  to the  percent  lipid  in the  tissue.   Thus,  the  per
 capita ingestion  of  a  lipid-soluble chemical can be estimated from
 the per capita  consumption  of fish  and shellfish,  the  weighted
 average  percent  lipids  of  consumed   fish  and  shellfish, and a
 steady-state BCF  for the chemical.
     Data from a  recent  survey on  fish  and shellfish consumption
 in the United States were analyzed by SRI International  (U.S. EPA,
 1980).  These data were used  to estimate that  the  per capita con-
 sumption  of freshwater and  estuarine  fish  and  shellfish  in  the
United  States is  6.5 g/day  (Stephan,  1980).   In  addition,   these
data were used with  data on  the fat content of  the  edible portion
 of the same species  to estimate that  the  weighted  average percent
lipids for consumed freshwater and estuarine fish and shellfish is
3.0  percent.
                               C-5

-------
     No  measured  steady-state  bioconcentration  factor  (BCF)  is



available for chlorobenzene, but the equation "Log BCF =  (0.85 Log



P) - 0.70" can  be  used (Veith,  et al.,  1979)  to estimate the BCF



for  aquatic   organisms  that  contain  about  7.6  percent  lipids



(Veith,  1980)  from the  octanol-water  partition  coefficient  (P).



Based on an average measured  log  P  value of 2.49 (Hansch and Leo,



1979), the steady-state  bioconcentration factor for chlorobenzene



is estimated to  be  26.1.   An  adjustment  factor of 3.0/7.6 =  0.395



can  be  used  to adjust  the estimated  BCF from the  7.6 percent



lipids on  which the equation is  based to the  3.0  percent lipids



that is  the  weighted  average bioconcentration  factor  for chloro-



benzene  and  the edible  portion of  all  freshwater  and  estuarine



aquatic organisms consumed by Americans  is calculated  to  be 26.1 x



0.395 » 10.3.



Inhalation



      No data have  been  found which deal  with  exposure to MCB by



air outside of  the industrial working environment.   The  informa-



tion concerning  the  industrial  exposure of workers  has  come pri-



marily from eastern European  sources and is tabulated  in Table 3.



In addition to  that information, Girard, et al.  (1969)  reported on



a case of  an  elderly  female who was exposed to  a glue, containing



0.07 percent MCB, for a period of six  years (see Special  Groups at




Risk).



De rma1



     Pertinent  data concerning the  dermal exposure of MCB  could



not be located  in  the available literature.
                               C-6

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                               TABLE 3

          Recorded Industrial Exposures to Monochlorobenzene
Plant Activity
Concentration of
   MCB (mg/1)
                   Reference
Manufacture of DDT
Manufacture of monurorf
0.020
0.300

0.001

0.004
average
highest

0.01

0.01
Gabor and Raucher,
  1960

Levina, et al. 1966

Stepanyan, 1966
                               C-7

-------
Summary and Conclusions



     Water  is  a  documented  source  of  environmental  exposure  to



MCB.  Because of  the short half-life  of  MCB  in water,  it would be



relatively  difficult  to  monitor human  exposure  unless  multiple



sampling was done. Compared to substances such as DDT,  the accumu-



lation of MCB within the food chain is limited; however, even this



accumulation tends  to  magnify  the possible  human  exposure  to MCB



via discharge into water.



                      PHARMACOKINETICS



Absorption



     There  is  little question,  based on  human  effects  and  mam-



malian  toxicity  studies, that MCB  is absorbed  through  the lungs



and from  the  gastrointestinal  tract  (U.S.  EPA,  1977).   Based on



what is known  about congeners,  it is also probably absorbed from



the surface of the skin.



Distribution



     Because MCB  is highly lipophilic and hydrophobic, it would be



expected  that  it  would  be  distributed throughout total body water



space,  with body  lipid providing  a deposition  site.   The  data



available  on  the  related halobenzene, bromobenzene,  show this to



be  the  case (Reid, et  al.  1971).  Barring  some abnormal kinetic



pattern,  it would  also  be expected that  redistribution from tissue



sites would reflect plasma decay rates.   Again,  with  bromobenzene



this  was   the  case,   the  plasma   t(^/2)   being   5.8 hours  and




the t(i/2)  ^or ^at  being 6.2 hours.
                               C-8

-------
 Metabolism




      Metabolism  of  MCB has  been studied  in  a  number  of  labora-



 tories.  Hydroxylation  occurs  para  to the  chloride  via an NADPH-



 cytochrome P-448 dependent microsomal  enzyme  system.   Further hy-



 droxylation then  occurs to  form the  corresponding  catechol  com-



 pound.   The diphenolic derivative is a predominant form, quantita-



 tively, in comparison to the monophenolic compounds.  Various con-



 jugates of  these phenolic derivatives  are the  primary excretory



 products (Lu,  et al.  1974).   The conjugates are  formed  by micro-



 somal enzymes, in this case, the NADPH-cytochrome P-450 dependent



 system.  However, it  would  appear  that the rate-limiting  step  in



 metabolism  of  MCB is the initial hydroxylation of the ring.  There



 are  some differences  in the nature  of the conjugates,  depending



 upon  the animal  species  studied.  Williams,  et al.  (1975)  found



 that  among 13  species of nonhuman  mammals, 21  to  65 percent  of



 excreted  radioactivity  from the administration  of  14C-MCB was



 present in  the urine as p-chlorophenylmercapturic acid.  The  out-



 put  of  this conjugate in man was  only 16  percent  of the  admin-



 istered dose.  Williams  (1959) also  reported that  about  27  percent



 of MCB  administered  to the rabbit was expired  unchanged  in  the air



 over a  1 to  2  day period; 47 percent of  the dose was excreted as



 glucuronic acid or sulfate conjugate and 25  percent as mercapturic



 acid conjugate.   This  accounts  for  the  total dose and would imply



 that very little  is  excreted unchanged.  This  would  be expected,



as the  lipophilic nature of  MCB would predict  that  it  would be
                               C-9

-------
almost totally reabsotbed by the renal tubules such that  its decay
from the plasma  would  rely totally on metabolism  and on ventila-
tory excretion.
     The ease with which MCB is eliminated via the lungs  or metab-
olized would  predict that its  bioaccumulation  potential is some-
what limited.   Varshavskaya (1968) found that when MCB was admin-
istered  to  rats  at 0.001 mg/day for  nine  months,  the coefficient
of accumulation  was  1.25.   This  would mean  that  accumulation  is
somewhat less than if the exposure  level is kept  constant.    For
example, if  a single dose were taken  every  24  hours and this  re-
sulted  in   a  total  body accumulation of  1.25  x  the  dose,   the
t(l/2)   would  be  calculated  to   be  approximately  11  hours.
This  would  suggest that  in  the rat,  upon exposure  to a constant
dose,  the maximum body concentration  is  reached  in  about  two  days.
The  same numbers cannot be  applied  to man because of  differences
in organ clearance rates,  but relatively speaking it would be  ex-
pected  that equiblibrium would be reached in a  short  time  from an
environmental point of  view and that prolonged exposure  to  con-
stant  levels  in  the  environment would not be expected to result in
continuous  accumulation.
      Evidence has been  accumulating  which implies that the metabo-
lism of halogenated benzene compounds results in  the  formation of
toxic  intermediates.    Brodie,  et al.  (1971)   pretreated  animals
with phenobarbital to stimulate the  activity of drug metabolizing
enzymes in the  liver.   This  treatment  potentiated  liver necrosis
 induced by halogenated  aromatic  compounds  (of  which monobromoben-
 zene was the primary example).   This is  apparently related to  the
                                C-10

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 formation of metabolites capable of forming  complexes  with cellu-



 lar ligands.  The covalent binding  of  the  metabolites  of halogen-



 ated benzene derivatives with protein  has been correlated with the



 ability of  these  compounds to  induce  hepatic necrosis  (Reid,  et



 al. 1971, 1973;  Reid and  Krishna,  1973).    Oesch,  et al.  (1973)



 have reported that  rats  pretreated  with 3-methylcholanthrene  are



 protected  from MCB-evoked  hepatotoxicity.    This  was ascribed  to



 the modification  of  a coupled  monooxygenase  epoxidehydrase system



 (Oesch,  et al. 1973).   Carlson and Tardiff  (1976)  reported  that



 the oral administration of 10 to 40 mg/day of MCB to rats  for  14



 days induced  a  variety  of microsomal enzymes  which metabolize



 foreign  organic compounds  including  benzpyrenehydroxylase.   Cellu-



 lar toxicity,  including carcinogenic and mutagenic  activity, may



 be  related  to  the  formation of highly  active metabolic intermedi-



 ates  such  as epoxides.  In this connection,  Kohli,  et al.  (1976)



 have  suggested  that  the metabolism of  MCB  occurs  via arene oxide



 intermediates as shown  in Figure 1.



                             EFFECTS



 Acute, Subacute, and Chronic Toxicity




     The acute toxic effects of MCB were quantitatively similar  in



 some cases to chlorinated  hydrocarbons  such as carbon tetrachlor-



 ide.   The  oral  LD50  of   monochlorobenzene  in  the  rat  is  ap-



proximately 3 g/kg.  When  administered  by subcutaneous injection,



the  LD50  increases  by  about  25  percent.   Von Oettingen   (1955)



found that large  doses of MCB  (7  to 8  g/kg  subcutaneously) were



fatal in a few hours as a result of CNS  depression.  When  the dose



utilized was  4  to  5  g/kg, death  occurred  after  a  few  days  and
                               C-ll

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                 C!
                                          OH
                        ,OH
                       OH
                FIGURE 1

Proposed Route for Biotransformation of
   Monochlorobenzene Via Arene Oxides
      Source: Kohli, et al. 1976
                  C-12

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 resulted  from hepatic  and/or renal  necrosis.   Vecerek,  et al.
 (1976)  found  the oral  LD50  of MCB  in  rats to  be  3.4 g/kg.   At
 this  dose,  the  animals died  after  about  seven days  and showed
 signs of  a number  of  metabolic  disturbances  including  elevated
 levels of SCOT, lactate dehydrogenase, alkaline phosphatase, blood
 urea nitrogen, and decreased  levels  of  glycogen phosphorylase and
 blood sugars.    Yang  and Peterson  (1977)  administered  MCB  at  5
 mmol/kg  (about 563 mg/kg) intraperitoneally to male rats and found
 an increase in the flow of bile duct pancreatic fluid.
      Data on the subchronic  and chronic toxicity of MCB are sparse
 and  somewhat contradictory.   Lecca-Radu (1959)  administered MCB by
 inhalation to  rats and  guinea  pigs  for periods up to  one  year in
 doses which did not affect the liver or  the kidney but did modify
 erythrocyte  carbonic  anhydrase and leukocyte  indolephenol  oxidase
 activities.  Knapp,  et al. (1971)  administered MCB orally  by  cap-
 sule  to  dogs in  doses  of 27.2,  54.5,  and  272.5  mg/kg/day  five  days
 a  week over  a  90-day  period.   Four of eight of  the animals  in  the
 high  dose  group  died after 14  to 21 daily doses.   Clinical  studies
 prior  to death  revealed an  increase  in  immature leukocytes,  low
 blood  sugar, elevated  SGPT and alkaline  phosphatase  and,   in  some
 dogs,  increases  in  total bilirubin and total cholesterol.    "Gross
 and/or microscopic  pathological changes" were  seen  in the  liver,
kidneys, gastrointestinal  mucosa,  and  hematopoietic tissue of  the
dogs  which died  and,  less extensively,   in  the dogs  which were
sacrificed after 65 or 66 daily doses.  No consistent signs of  MCB
toxicity were  seen in  dogs  in the  intermediate and  low  levels.
                               C-13

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MCB was given to rats by diet  at  doses  of 12.5,  50 and 250 mg/kg/
day for a  period  of 93 to 99  days.   Growth was  retarded  in male
rats in the  high  dose  group.   There was  an  increase  in liver and
kidney weight for  rats  in  the  high and  intermediate levels.  This
was not accompanied by any "histopathological" findings (Knapp, et
al. 1971).
     The toxicity  of  MCB following exposure by  inhalation and by
oral administration has  been  studied by  the  Dow Chemical  Company
(Irish, 1963).   Rats,  rabbits and guinea  pigs  were exposed  seven
hours a day,  five  days  a week,  for a total of 32 exposures over  a
period  of  44 days  at  concentrations  of 200, 475,  and 1,000 ppm.
The response of  the animals  in the high dose group was character-
ized  by "histopathological changes" in  the  lungs, liver  and kid-
neys.   In the middle  dose group, there  was  an  increase  in  liver
weight  and  a slight  liver   "histopathology".    In the  low dose
group,  no  apparent effects were  observed.   In  none of the  groups
was  a hematological change seen.   MCB  was administered orally  to
rats  five  days  a week for a  total of  137 doses over  192  days,  in
dose  groups  of 14.4,  144  and 228 mg/kg.   In  the middle  and high
dose  groups  there  were  significant  increases  in  liver and kidney
weight  and  some  "histopathological  changes"  in  the liver.   Blood
and  bone  marrow  were  normal  in all animals (Irish, 1963).
      Rimington  and Ziegler (1963), citing  the  widespread  outbreak
 of human  cutaneous porphyria  in  Turkey  in  1959  apparently caused
 by wheat  treated  with  hexachlorobenzene  fungicide,  examined  a
 series of chlorinated benzene compounds in rats with regard  to ex-
 perimental  porphyria.   MCB at  an oral  dose of  1140 mg/kg   for five
                                C-14

-------
 days increased the excretion  of  utinary coproporphyrin, porphoro-
 bilinogen, and delta-aminolevulinic acid.  Some hair loss was also
 observed due to follicular hyperkeratosis.
      A study by Varshavskaya  (1968) describes  the  central nervous
 system (CNS), liver and hematopoietic system changes in seven male
 rats per  group  which received oral  doses of  0.1  mg/kg  to  0.001
 mg/kg MCB for a period of nine months.   This report indicates that
 doses of  0.001  mg/kg MCB  for seven  months affected  the CNS  of
 rats and that similar effects  resulted  from similar o-dichloroben-
 zene dosages.  However,  these results  are somewhat  unexpected  in
 light of other studies in the literature.  For  example,  Hollings-
 worth,  et al. (1956)  reported similar  results  from  an  experiment
 with o-dichlorobenzene which differed  by over  three orders of mag-
 nitude  from  those  of  the  Varshavskaya  (1968) study.  This  discrep-
 ancy in o-dichlorobenzene results  leaves  the  MCB  results of  the
 Varshavskaya  study  open to question.
 Synergism  and/or Antagonism
      In  general,  the halogenated benzenes  appear  to increase  the
 activity  of  microsomal   NADPH-cytochrome  P-450 dependent enzyme
 systems.   Induction of microsomal enzyme  activity  has  been  shown
 to enhance the metabolism of a wide  variety of drugs, pesticides,
 and  other  xenobiotics.    Exposure  to  monochlorobenzene  could,
 therefore, result   in  decreased  pharmacologic  and/or toxicologic
                                                  \
activity of  numerous  compounds.   Frequently,  chemical  agents are
metabolized to more active or  toxic  "reactive"  intermediates.  in
this event, exposure to monochlorobenzene would result in enhanced
activity and/or toxicity of these agents.
                               C-15

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Teratogenicity, Mutagenicity, and Carcinogenicity
     There have been  no  studies  conducted  to evaluate the terato-
genic, mutagenic or carcinogenic potentials of MCB.
                                016

-------
                       CRITERION  FORMULATION
 Existing  Guidelines  and  Standards
      The  Threshold Limit  Value  (TLV) for  MCB as  adopted  by  the
 American  Conference  of Governmental Industrial Hygienists  (ACGIH,
 1971)  is  75  ppm  (350  mg/m3).    The American Industrial  Hygiene
 Association  Guide (1964) considered  75  ppm  to be  too  high.    The
 recommended maximal  allowable concentrations in air  in  other  coun-
 tries  are:   Soviet Union,  10   ppm;  Czechoslovakia, 43  ppm;  and
 Romania,  0.05 mg/1.   The latter  value for Romania was  reported by
 Gabor and  Raucher  (1960) and  is  equivalent to  10 ppm.
 Current Levels of Exposure
      MCB  has been detected  in water monitoring surveys  of various
 U.S.  cities  (U.S. EPA,  1975; 1977)  as  was presented in  Table 1.
 Levels reported were:  ground water - 1.0 ug/1; raw water contami-
 nated by  various  discharges  - 0.1  to 5.6 ug/1; upland water -  4.7
 ug/1; industrial discharge - 8.0 to 17.0 ug/1; and municipal water
 - 27 ug/1.  These data show a gross estimate of possible human  ex-
 posure to MCB through the water  route.
     Evidence  of  possible  exposure  from  food  ingestion  is   in-
 direct.  MCB is stable  in  water  and thus can be bioaccumulated by
 edible fish species.
     The only data concerning exposure to MCB via air are from  the
 industrial working environment.   Reported  industrial exposures to
MCB are  0.02  mg/1  (average  value)  and  0.3  mg/1  (highest  value)
 (Gabor and Raucher,  1960);  0.001   to 0.01 mg/1  (Levina,  et  al.
1966); and 0.004  to 0.01 mg/1 (Stepanyan, 1966).
                               C-17

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Special Groups at Risk
     The major group  at  risk of MCB  intoxication are individuals
exposed to MCB  in  the workplace.  Girard,  et  al. (1969) reported
the case  of  an elderly  female  exposed to a glue containing 0.07
percent MCB for a period  of  six  years.  She had  symptoms of head-
ache,  irritation of the  eyes  and the  upper respiratory tract, and
was  diagnosed to  have medullary  aplasia.   Smirnova  and  Granik
(1970)  reported  on three  adults who  developed  numbness,  loss of
consciousness, hyperemia  of  the conjunctiva and the pharynx  fol-
lowing  exposure  to  "high" levels of  MCB.   Information  concerning
the ultimate  course of these  individuals  is not  available.  Gabor,
et al.  (1962) reported on  individuals  who  were exposed to benzene,
chlorobenzene,  and  vinyl  chloride.    Eighty-two workers examined
for certain biochemical  indices  showed a  decreased catalase activ-
ity  in the blood and  an increase in  peroxidase, indophenol  ox i-
dase,  and  glutathione noted  levels.   Dunaeveskii (1972) reported
on the occupational  exposure of workers  exposed to  the chemicals
involved  in  the  manufacture  of  chlorobenzene  at limits below  the
allowable  levels.  After over three  years, cardiovascular effects
were  noted as pain  in the area of the heart, bradycardia,  irregu-
lar  variations in electrocardiogram,  decreased  contractile  func-
tion  of myocardium,  and  disorders in adaptation  to  physical  load-
ing.   Filatova,  et al.  (1973)  reported  on the  prolonged exposure
of individuals  involved  in the  production of diisocyanates  to  fac-
tory  air  which contained  MCB  as  well  as  other  chemicals.  Diseases
noted include bronchitis, sinus arrhythmia, tachycardia, arterial
dystrophy, and  anemia tendencies.   Petrova and  Vishnevskii  (1972)
                                C-18

-------
 studied the course'of pregnancy and deliveries in women exposed to
 air  in  a varnish  manufacturing factory  where  the  air  contained
 three times the maximum permissible level of MCB but also  included
 toluene, ethyl chloride, butanol,  ethyl  bromide,  and orthosilisic
 acid ester.  The only  reported  significant  adverse  effect  of this
 mixed exposure was toxemia during pregnancy.
 Basis and Derivation of Criterion
      There is  no  information  in  the  literature which  indicates
 that monochlorobenzene  is,  or  is  not,  carcinogenic.   There  is
 enough  evidence to suggest that MCB causes dose-related target or-
 gan toxicity,  although the data are lacking  an  acceptable  chronic
 toxicity study.   There  is little,  if  any,  usable  human  exposure
 data primarily because  the  exposure was  not only  to  MCB but  to
 other compounds of known  toxicity.
      A  no-observable-adverse effect  level (NOAEL)  for  derivation
 of  the  water  quality criterion can be extracted from  the  informa-
 tion in  the  studies by  Knapp,  et  al.  (1971)  and  Irish  (1963).
 These are 27.25 mg/kg/day for  the  dog  (the  next highest dose  was
 54.5 mg/kg .and showed  an  effect);  12.5 mg/kg/rat  from the  Knapp
 study (the next highest dose was  50 mg/kg and showed an effect);
 and  14.5 mg/kg/rat  from the Irish  study  (the  next  highest dose  was
 144  mg/kg  and showed an effect).  When toxic  effects  were observed
 at  higher  doses,  the  dog  was  judged to be somewhat more sensitive
 than  rats.  The  duration of  the  study  by  Irish  (1963)  was  six
months, which was  twice as  long as the Knapp study of two  species
 (rat, dog).   Since  the Knapp  and Irish  studies  appear  to give
similar  results  and since there  are  no  chronic  toxicity  data on
                               C-19

-------
which to  rely,  the NOAEL level,  14.4  mg/kg for  six  months, from
                                *

the  longest  term  study  (Irish,  1963)  is  used,  to  calculate  the


acceptable daily intake (ADI).


     Considering that  there  are relatively little  human exposure


data, that there are no long-term animal data, and that some theo-


retical questions, at least, can be raised on the possible effects


of chlorobenzene on blood-forming tissue  an uncertainty factor of


1,000 is  used.   From  this  (ADI)  can  then be calculated  as fol-


lows:


                 70 kg x 14.4 mg/kg   .. nnn    ,,
           ADI =	? 1,000—^"^ = 1-008 m9/day





     The  average daily consumption of  water  was taken  to  be two


liters and the  consumption  of fish and shellfish to  be 0.0065 kg


daily.  A bioconcentration  factor of 10.3 was utilized.   This is


the value  reported  by  the Duluth EPA  Laboratories  (see Ingestion


from Food  section).   The following calculation  results in  a cri-


terion based on the available toxicologic data:



                 2 + (10*.3 x 0.006S) =  488 ug/1


     Varshavskya  (1968)  has reported  the  threshold concentration


for  odor   and  taste of  MCB  in  reservoir  water.    The  specific


methods whereby  the organoleptic data  were obtained  are  not de-


tailed in  this report.  The only statement made was that different


methods provided  similar  estimates of  threshold concentrations.


The reported olfactory and  gustatory  threshold  was  found to be 10


to  20  ug/1.   A  value of  20 ug/1  is about  4.5 percent  of the
                               C-20

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possible  criterion calculated  above.    It is,  however,  approxi-



mately  17 times greater  than  the  highest concentrations  of MCB



measured  in survey  sites  (see Table 1).   Since water of disagree-



able taste  and  odor has  significant  influence on  the  quality of



life and,  thus, is related  to health,  it would appear  that the



organoleptic level of 20 ug/1 should be the recommended criterion.
                              C-21

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American Conference of Governmental  Industrial  Hygienists.   1971.



Documentation  of the  Threshold Limit  Values  for  Substances  in



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American Industrial  Hygiene Association.   1964.   Chlorobenzene.



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Brodie, B.B., et  al.   1971.   Possible mechanism of liver necrosis



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Carlson, G.P. and R.G. Tardiff.  1976.  Effect of chlorinated ben-



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Dunaeveskii, G.A.   1972.   Functional condition  of circulatory or-



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Gig. Tr. Prof. Zabol.  16: 48.







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                               C-22

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Filatova, V.S., et al.  1973.  Industrial hygiene and pathology in



production and  use of diisocyanates.   Profil.  Trarmatizma Prof.



Zabol., Lech. Traum. 219.







Gabor, S. and K.  Raucher.   1960.   Studies on the determination of



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chlorine derivatives  of  benzene.    (A  report of 7 cases.)   Jour.



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Hansch, C. and J.  Leo.   1979.   Substituent Constants for Correla-



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Hollingsworth, R.L.,  et al.   1956.  Toxicity of paradichloroben-



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AMA Arc. Ind. Health.   14: 138.
                               C-23

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 Knapp,  W.K.  Jr.,  et al.   1971.   Subacute  oral  toxicity of mono-
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 Kohli,  I.,  et  al.   1976.   The  metabolism of  higher chlorinated
 benzene isomers.  Can. Jour. Biochem.  54:  203.

 Lecca-Radu, M.   1959.   Modifications of  blood  carbonic  anhydrase
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 Levina, M.M., et  al.    1966.   Issues concerning  sanitary working
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Lu, A.Y.H.,  et  al.   1974.   Liver microsomal  electron   transport
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Lu, P.Y. and  R.L.  Metcalf.   1975.   Environmental  fate  and biode-
gradability of  benzene  derivatives  as studied in  a  model aquatic
ecosystem.   Environ. Health Perspect.   10: 269.
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Mackay, D. and P.J. Leinonen.   1975.   Rate  of evaporation of low-



solubility contaminants from water bodies to atmosphere.  Environ.



Sci. Technol.  9: 1178.







Oesch, P., et  al.   1973.   Induction activation  and  inhibition of



epoxide  hydrase.    Anomalous  prevention  of chlorobenzene-induced



hepatotoxicity by  an  inhibitor of  epoxide  hydrase.   Chem.  Biol.



Interact.  6: 189.







Petrova, N.L.  and  A.A.  Vishnevskii.   1972.   Course  of pregnancy



and deliveries in  women  working in the  organosilicon varnish and



enamel industries.  Nauch Jr. Inrutsk. Med.  Inst.  115: 102.







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related to covalent binding of aromatic hydrocarbons.   Exp. Molec.



Pathol.  18: 80.







Reid,  W.D.,  et  al.   1971.  Bromobenzene  metabolism  and  hepatic



necrosis.  Pharmacol.   6: 41.







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necrosis.  Am. Rev. Resp. Dis.  107: 539.
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 Rimington,  C.  and G.  Ziegler.   1963.   Experimental porphyria  in
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 Smirnova,  N.A.  and N.P.  Granik.   1970.    Remote  consequences  of
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 U.S.  EPA.   1977.   Investigation  of  selected  potential environ-
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U.S.  International  Trade Commission.    1975.   Synthetic  organic



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                               028

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                         TRICHLOROBEN Z EN ES
 Mammalian Toxicology and Human Health Effects
                            INTRODUCTION
      There are  three isomers of trichlorobenzene (TCB): 1,2,3-tri-
 chlorobenzene,  1,2,4-trichlorobenzene  and  1,3,5-trichlorobenzene.
 Of  the three, 1,2,4-TCB  is  the most economically  important  (U.S.
 EPA,  1977).  It  is used as  a dye  carrier  in the  application  of
 dyes  to polyester materials,  as  an  intermediate in  the  synthesis
 of  herbicides, as  a  flame  retardant,  and  for  other  functional
 uses.   The U.S. production  of 1,2,4-trichlorobenzene  in  1973 was
 over   28  million  pounds  (U.S.  International  Trade  Commission,
 1975).   A mixture  of  the three  isomers  is  used  as a solvent,  a
 lubricant,  and  as  a dielectric  fluid.   The  1,2,3 and  1,3,5-TCB
 isomers  as individual compounds  are primarily used as  intermedi-
 ates  in  chemical  synthesis.   TCBs are most probably intermediates
                       t
 in  the mammalian metabolism  of lindane  (Kujawa,  et  al.  1977).
                              EXPOSURE
 Ingestion  from Water
     Table  1  shows  data  from  monitoring  the  various water sites.
 These  data suggest  the  possibility  of  TCB  contamination of the
 drinking water.   In a  report (U.S. EPA, 1975) in which the sample
 site  was not  identified,  the  highest reported  concentration  of
 trichlorobenzene in drinking  water was  1.0 ug/1.
 Ingestion from Food
     Whereas  the  bioaccumulation  of  some of  the  other members  of
 the chlorinated  benzene series has  been  studied  with  regard  to
model  aquatic ecosystems,  apparently  such  has not  been  the  case
                               C-29

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                                                 TABLE  1
                                       Occurrence  of TCBs  in Water*
         Compound
       Location
      Source
Concentration
   (ug/D
o
i
CO
o
         1,2,3-TCB
         1,2,4-TCB
         1,3,5-TCB
Catawba Creek, NC
Catawba Creek, NC
Chattanooga Creek, TN
Joint Water Pollution
Control Plant (JWPCP)
Hyperion Sewage Treatment
Works, LA (HSTW)
HSTW

Orange County Sewage
Department (OCSD)
Port Loma Sewage Treat-
ment Plant (PLSTP)
Oxnard, CA Sewage
Treatment Plant (OSTP)
Los Angeles River
Holston River, TN
JWPCP
HSTW

HSTW

OCSD
PLSTP
OSTP
Los Angeles River
Municipal discharge
Industrial discharge
Industrial discharge
Municipal waste water

5 mile effluent, municipal
waste water
7 mile effluent, municipal
waste water
Municipal waste water

Municipal waste water

Municipal waste water

Surface run off
Industrial discharge
Municipal waste water
5 mile effluent, municipal
waste water
7 mile effluent, municipal
waste water
Municipal waste water
Municipal waste water
Municipal waste water
Surface run off
  21-46a
  500b
  6.0; 1.8a

  6.7; 3.1C

  275; 130C

  0.30a

  0.23; <0.01C

  0.9; 0.25C

  0.007d
  26
  0.2; 0.8C
  <0.01; <0.01C

  0.9; <0.2C

  0.2
  0.02; <0.01C
  0.4; <0.0ic
  0.006d
         *Source:  U.S.  EPA,  1977.
         aSummer;  bSpring;  cSummer,  Fall; ^Winter; eFall

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with the TCBs.  The accumulation of TCBs in the food chain depends



upon  their  concentrations  in  aquatic  organisms.    Haas,  et  al.



(1974) have found  that  40 percent of  the  1,2,4-TCB in wastewater



was absorbed  by  microorganisms,  and the suggestion has  been made



by EPA that the material concentrates  in the cell wall.  This type



of information  indicates  that  TCBs will persist in a  water envi-



ronment and are  available for  incorporation  into  fish.   TCB has



been detected  in  trout  taken from Lake  Superior  and turbot taken



from Lake Huron  (U.S. EPA, 1977).



     A bioconcentration factor BCF  relates  the concentration of  a



chemical  in aquatic  animals to the concentration  in the water  in



which  they  live.   The  steady-state  BCF for  a lipid-soluble com-



pound  in  the  tissues of various aquatic animals  seems to be pro-



portional  to   the  percent lipid  in the  tissue.    Thus,  the per



capita ingestion of a lipid-soluble chemical can be estimated from



the  per  capita  consumption of fish  and shellfish,  the weighted



average  percent  lipids  of  consumed  fish  and  shellfish,  and   a



steady-state BCF for the  chemical.



     Data from a recent survey on  fish and shellfish consumption



in the United  States were analyzed  by  SRI International  (U.S. EPA,



1980).  These  data were used to estimate that the  per capita con-



sumption  of freshwater and estuarine  fish  and shellfish  in the



United States is  6.5 g/day  (Stephan,  1980).   In   addition,  these



data were used with  data  on the fat content of the  edible  portion



of the same species  to  estimate that  the weighted  average  percent



lipids for consumed  freshwater and  estuarine  fish  and  shellfish  is



3.0 percent.
                                C-31

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     A measured  steady-state bioconcentration  factor of  182 was



obtained  for  1,2,4-trichlorobenzene  using  bluegills  (U.S.  EPA,



1978).   Similar  bluegills  contained  an average  of  4.8  percent



lipids (Johnson,  1980).   An adjustment factor of  3.0/4.8  =  0.625



can be used to adjust the measured BCF from the 4.8 percent lipids



of  the  bluegill  to  the  3.0 percent  lipids  that  is  the weighted



average for consumed fish and shellfish.  Thus, the weighted  aver-



age  bioconcentration factor  for  1,2,4-trichlorobenzene  and the



edible portion of all freshwater  and  estuarine  aquatic organisms



consumed by Americans is calculated to be 14.



     There  is  some  information  on studies on  biochemical oxygen



demand (BOD) in waste water  containing  microorganisms from treat-



ment plants.   This  information  has been  compiled previously  (U.S.



EPA, 1977) and is presented in Table 2.  This table summarizes the



20-day BOD  for  1,2,4-TCB.   As can be seen,  the  results vary from



no biodegradation to complete biodegradation of the 1,2,4-TCB.



     Simmons,  et  al. (1976) also  noted  a lack  of degradation of



1,2,4-TCB based  on BOD determinations.   However,  direct chemical



analysis indicated a 14 percent  reduction of TCB concentrations in



industrial wastewater after 24 hours, a 36 percent reduction  in 72



hours and 43 percent reduction at  seven days.  This would  indicate



that  the  limitation in change  of  BOD is due  primarily to incom-



pletely oxidized metabolites.



Inhalation and Dermal



     Vapor  pressures for TCBs  are shown in  Table 3.   These are



relatively  low compared  to  mono- and  dichlorobenzenes.  Neverthe-



less, TCBs have been detected in particulates .from aerial  fallout.
                               C-32

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                             TABLE 2

            Effects of 1,2,4-Trichlorobenzene on BOD*
Source of Organisms
BOD2Q (percent of
theoretical value)
References
Microorganisms from
industrial waste treat-
ment plant

Microorganisms from
industrial waste treat-
ment plant

Mixture of microorganisms
from 4 different textile
treatment plants

Microorganisms from "typi-
cal" treatment plant
        78
       100
        50
     (2 days)
Hintz, 1962
Alexander, 1972
Porter and Snider,
1974
                       Haas, et al. 1974
*Source: U.S. EPA, 1977.
                               C-33

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                        TABLE 3

        Vapor Pressures of Trichlorobenzenes*
 TCB-             Vapor Pressure            Temperature
isomer               (ram Hg)(«c)
1/2,3                 0.07                     25
                      1.0                      40
1/2,4                 0.29                     25
                      1-0                      38.4
1/3/5                 0.15                     25
                      1.0                      78

*Source: U.S. EPA/ 1977;  Sax, 1975.
                          C-34

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 In a  study  of aerial  fallout  in  southern  California  (spring,
 1976) /  five B sampling sites showed median levels of "less than 11
 ng/m2/day"   for  1,2,4-TCB   and   "less  than   6   ng/m2/day"   for
 1,3,5-TCB (U.S. EPA,  1977).
     There  have been no reports of exposures of  humans  to TCB via
 inhalation  that resulted in  toxicity.  The amount of TCB necessary
 to induce a  toxic  reaction  via application to  the skin  is  quite
 high  and thus exposure  to TCB  via water  on  the skin is  not  con-
 sidered  to  be  a significant factor  in  the determination of  cri-
 teria standards (Brown,  et al.  1969).
        
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1,3,5-TCB,  the  two metabolites were 2,3,^-TCP  and  2,4,6-TCP (1.5
and 3.0 percents, respectively).  These authors proposed a pathway
for metabolism  which goes through arene  oxide steps as  shown  in
Figure  1.   Parke  and  Williams  (1960)  have also  described  small
quantitites of monochlorobenzene and parachlorophenol in the urine
of rabbits  following  the  administration of 1,3,5-TCB.   It  can  be
assumed that the TCB is transformed  by  the NADPH-cytochrome P-450
microsomal  enzyme  system.   Although the  evidence  suggests this
metabolic mechanism, the  experiments designed  to demonstrate this
point specifically have not  been conducted.   Egyankor  and  Frank-
lin (1977) incubated TCB ispmers with rat hepatic microsomal cyto-
chrome  P-450.   They  found that the order  of affinity of the iso-
mers  for  cytochrome  P-450  was  1,2,3-TCB   <1,2,4-TCB <1,3,5-TCB.
Interestingly, this is the same order which has been found for the
metabolism  of  TCB   isomers  to  phenol.    They also noted  that
1,3,5-TCB inhibits the  hepatic microsomal  mixed  function  oxidase
system  while  1,2,3-TCB  and 1,2,4-TCB enhanced it.   Ariyoshi,  et
al. (1975a,b,c) reported  on  the microsomal enzyme  systems  in in-
tact  rats.    They  found  that  1,3,5-TCB increased  the  amount  of
microsomal protein, phospholipids, and cytochrome P-450 as well  as
stimulating  the activities  of  aminopyrine demethylase,  aniline
hydroxylase, and delta  aminolevulinic  acid  synthetase  (Ariyoshi,
et al. 1975a).  Similar results were obtained for 1,2,4-trichloro-
benzene.  Increases  were  observed in cytochrome P-450  content  of
the liver, enhanced delta  aminolevulinic acid synthetase activity,
                               C-36

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n
i
                     Proposed Pathways for the Biotransformation of  Trichlorobenzene

                                Isomers Through Arene Oxide Intermediates      \j

                           ,'  -.   ,    ,   Source: Kohli, et al. 1976

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aminopyrine demethylase  activity, microsomal  protein,  microsomal
phosphate, liver weight, and aniline hydroxylase (Ariyoshi, et al.
1975b).
     Carlson and  Tardiff  (1976) reported  that  1,2,4-TCB  caused  a
decrease  in  hexobarbital  sleeping  time  and an  increase  in  the
activities of  cytochrome-c  reductase,  cytochrome P-450 glucuronyl
transferase, benzpyrene  hydroxylase,  and azoreductase.   Carlson
(1978),  investigating  the  effect of 1,2,4-TCB  on  metabolism sys-
tems in  the liver, concluded  that  the  compound  induces xenobiotic
metabolism  of  the phenobarbital type  rather  than  the 3-methyl-
cholanthrene type.
     There is a paucity of kinetic data concerning TCBs.  However,
based on data  from Williams  (1959) and Parke and Williams (1960),
some estimates can be  made as to the  biological  half-life  of the
isomers.   From these  data,  it was estimated  that  the approximate
half-lives of the isomers  are:   1,2,3-TCB,  2 days;  1,2,4-TCB, 5.5
days; and  1,3,5-TCB,  8.5 days.   This  is a consideration  in the
evaluation of  toxicity studies  for all  species,  especially those
which are considered subchronic.
Excretion
     Williams  (1959)  reported  that five  days after  oral  adminis-
tration  of 1,2,3-TCB,  1,2,4-TCB or 1,3,5-TCB to  rabbits,  78, 42,
or 9 percent,  respectively, of  the administered  dose was  excreted
as monophenols.   There was no evidence  for  the  existence  of sig-
nificant alternative metabolic pathways implying that the elimina-
tion  of 1,3,5-TCB  is  significantly   slower  than  the  other  two
                               C-38

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 isomers.   This is related to  the ease'of  oxidation of the various
 isomers and  is reflected in  the monophenol metabolites excreted.
                              EFFECTS
 Acute/  Subacute,  and Chronic Toxicity
      There  is  a limited amount of  relevant data on the toxicity of
 1,2,4-TCB and  essentially no data  on the toxicity of the other two
 isomers.   Cameron, et  al.  (1937)  first described  hepatotoxic ef-
 fects of trichlorobenzene,  finding  it to  be less  than  that  of
 monochlorobenzene or orthodichlorobenzene.   Brown, et  al.  (1969)
 reported  the  single  dose   acute  oral  LD$Q  in   rats   to  be  756
 mg/kg (556  to  939 mg/kg, 95  percent confidence limits).   In  mice,
 the   single  dose  acute  oral  LD50  was   766  mg/kg  (601  to  979
 mg/kg,  95  percent  confidence limits).    With   the  rats,  deaths
 occurred  within  five  days  of  exposure  and  in mice within  three
 days  of exposure.  For  both species,  intoxication  was  manifested
 as depression  of  activity at low doses and predeath extensor  con-
 vulsions  at  lethal  doses.    They also  determined  a  single  dose
 acute percutaneous toxicity  in  rats.  This  was  6139 mg/kg  (4299  to
 9056  mg/kg,  95 percent confidenc'e limits).   From the same  study,
 data  on skin irritation  were reported.  The authors concluded  that
 1,2,4-TCB was  not very  irritating, although fissuring  typical  of  a
defatting action  was observed  after prolonged  contact  in  rabbits
and guinea pigs.   Spongiosis,  acanthosis,  and parakeratosis  were
noted in both  species  along  with some  inflammation of the super-
ficial  dermis  in  rabbits exposed daily for  three weeks.    Some
                               C-39

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guinea pigs exposed to 0.5 ml/day  for  5  days/week for three weeks
died following extensor convulsions..  The  livers  of these animals
were found to have necrotic lesions.
     Coate, et al.  (1977)  reported on a  chronic  inhalation expo-
sure  of  rats  (30 animals  per  group),  rabbits  (16  animals  per
group) and monkeys  (9  animals per group) to  1,2,4-TCB at 25,  50,
and 100  ppm  for  periods of  up to 26  weeks.   No exposure-related
ophthalmologic changes were detected  in  rabbits  and monkeys after
26  weeks of  exposure  (rats  were  not examined).   Similarly,  no
exposure-related  changes  were detected  in BUN,  total bilirubin,
SCOT, SGPT, alkaline phosphatase and LDH when determined  at 4, 13,
and 26 weeks  of  exposure.   Hematological values  were also normal
when  examined  at 4,  13, and  26  weeks.   Pulmonary  function tests
were  conducted  on the monkeys.    No  treatment-associated changes
were noted in  static compliance,  carbon  monoxide diffusion capac-
ity,  distribution of  ventilation,  transpulmonary  pressure,  or  a
battery  of lung  volume determinations.  Histological  changes were
noted in the  livers and  kidneys  of rats  necropsied after 4 and 13
weeks of exposure.   These changes were  noted in  animals from all
treatment  groups and  were manifested  as an  increase  in size and
vacuolation of hepatocytes.   However,  after 26 weeks,  no  compound-
related  histopathological  changes were  noted in  rabbits or mon-
keys.
     Rowe  (1975)  reported  that persons exposed to  1,2,4-TCB vapor
at  3  to 5 ppm  experienced minor  eye  and respiratory  irritation.
                               C-40

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 The odor was  described as easily  noticeable at  these  concentra-
 tions.   There was  a  detectable odor at  concentrations  up  to  2.4
 ppm,  but no eye  irritation was evident.   No odor was  noted at con-
 centrations up to 0.88  ppm.
      Smith,  et  al.  (1978)  conducted a  90-day,  daily  oral  dose
 study of 1,2,4-TCB in  rhesus  monkeys  (four animals per group)  at
 concentrations of 1, 5, 25,  90,  125,  and  174  mg/kg.   Their report,
 which is an abstract,  states  that single  oral  daily doses  of  25
 mg/kg  or less were nontoxic  whereas  doses of  90  mg/kg  or  higher
 were  toxic  and doses of 173.6 mg/kg  were  lethal  within 20  to  30
 days.   There were  no deaths  observed in the 1,  5,  and 25  mg/kg
 groups;  one death occurred in each of the  90 mg/kg and 125  mg/kg
 groups and  two deaths occurred in  the 174 mg/kg  group.   Animals  on
 the  highest dose exhibited  severe weight  loss  and predeath fine
 tremors.   All of the animals  in the  highest  dose group had ele-
 vated  BUN,   Na+,  K+,  CPK,  SCOT,  SGPT,   LDH,  and  alkaline  phos-
 phatase  as  well  as hypercalcemia  and  hyperphosphatemia  from  30
 days  on.  Smith,  et al.   (1978) have been  using  the  urinary  pattern
 of  chlorguanide  metabolites  as an  indication  of cytochrome  P-450
 dependent drug metabolism.   The  abstract states  that  at the high
 doses, monkeys showed   evidence  of  the  hepatic  induction  as well
 as increased clearance  of  intravenous doses of labeled TCB.  Fur-
 ther information  on the  study  (Smith,  personal communication) gave
 evidence of  liver enzyme induction in the  90,  125, and  174 mg/kg
animals.   There were some pathological changes noted in  the livers
 of  the  high dose groups, primarily  a fatty  infiltration.   The
point at which there was  no effect  related  to the compound was at
                               C-41

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the 5 mg/kg level.  Since only an abstract of this study is avail-
able and since the  interpretation  of  this  study  is complicated by
the use of other drugs and weight losses in the control animals, a
valid no-observed-effect level (NOEL) cannot be derived from these
data.
     Rimington and  Ziegler  (1963)  were able to  induce an experi-
mental porphyria in rats with 1,2,4-TCB which was marked by an  in-
creased urinary coproporphyrin excretion and an increased porphor-
obilinogen excretion  in  urine.   This  porphyria  could be reversed
by glutathione.   They also noted a  hair  loss  due to  hyperkerato-
sis.  This study  cannot  be  used  for criterion formulation because
the compound was given only at one  (maximum tolerated) dose.
Synergism and/or Antagonism
      In  general,  the halogenated  benzenes  appear to  increase  the
activity  of  microsomal  NADPH-cytochrome  P-450  dependent enzyme
systems.   Induction of microsomal  enzyme  activity has been  shown
to enhance  the  metabolism  of a wide variety of drugs,  pesticides,
and  other  xenobiotics.  Exposure  to TCB could,  therefore, result
in decreased  pharmacologic  and/or  toxicologic  activity of  numerous
compounds.   Frequently, chemical  agents are  metabolized  to  more
active  or  toxic  "reactive"  intermediates.   In  this event,  exposure
to TCB  would  result in enhanced activity and/or  toxicity  of  these
agents.
Teratogenicity,  Mutagenicity and Carcinogenicity
      Studies  have not been conducted primarily for  the  purpose of
determining  the teratogenic or  mutagenic properties  of trichloro-
 benzene isomers.   Gotto,   et  al.   (1972),  in a  study to  examine
                                C-42

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hepatomas caused  by  hexachlorocyclohexane,  administered 1,2,4-TCB
at a dose of 600 ppm  by  inhalation daily for  six  months  to mice
and  reported  no  incidence  of  hepatomas.    There  are no  other
studies which have been designed for  the  purpose of studying car-
cinogenicity of TCB;  nor  have there been any  other reports indi-
cating such activity.
                              C-43

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                      CRITERION FORMULATION
Existing Guideline and Standards
     A proposed American Conference of Governmental and Industrial
Hygienists Threshold Limit Value  (TLV)  for 1,2,4-trichlorobenzene
is 5  ppm  (40 mg/m3) as  a  ceiling value  (ACGIH,  1979).   Sax,  et
al. (1975) recommends a maximum allowable  concentration of 50 ppm
in air  for  commercial TCB, a  mixture  of  isomers.   Coate,  et al.
(1977), citing  their  studies,  recommended that the  TLV should  be
set below  25 ppm,  preferably  at  5 ppm  (40  mg/m3).   Gurfein and
Parlova  (1962)   indicate  that  in the  Soviet  Union  the  maximum
allowable concentration for TCB in water  is  30  ug/1/ which is in-
tended  to prevent  organoleptic  effects.   They  also report  that in
a  study of  40  rats and 8  rabbits  administered  TCB  in drinking
water  at   60 ug/1  for  7 to 8  months,  no effects  were observed.
This  information was obtained from an abstract only, as evaluation
of the  study was not possible.
Current Levels  of  Exposure
      Possible  human exposure to  TCBs  might occur  from municipal
and   industrial  wastewater  and  from  surface  runoff   (U.S.  EPA,
1977).   Municipal  and  industrial  discharges  contained  from 0.1
ug/1  to 500  ug/1.   Surface runoff has been found to contain  0.006
to 0.007 ug/1.
      In the National  Organics Reconaissance  Survey  (NORS)  con-
ducted  by  EPA  in  1975,  trichlorobenzene was  found  in  drinking
water at a  level  of 1.0  ug/1.
                                C-44

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Basis and Derivation of Criterion




     Reliable toxicologic data on which to base a defensible water



quality  criterion  do not  exist  for  the  trichlorobenzenes.   The



studies  by  Smith,  et al.  (1978) and  Coate,  et al.  (1977)  do not



give sufficient detail or suffer from inherent problems in experi-



mental design.   Therefore, according  to  the guidelines  for  cri-



terion development, a criterion cannot be  recommended for any tri-



chlorobenzene isomer.   For future  derivation of  a human  health



criterion, sound data must be developed describing  the  effects  of



trichlorobenzenes on humans and experimental  animals.
                              C-45

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                            REFERENCES







Alexander,  M.    1972.    Pollution  characteristics  of  1,2,4-tri-



chlorobenzene.  Dow Chemical Co., Midland, Michigan.   (Unpub.)







American Conference of Governmental  Industrial  Hygienists.   1979.



Threshold Limit Values for Chemical Substances and Physical Agents



in Workroom Environment with Intended Changes for 1979.







Ariyoshi, T., et al.   1975a.   Relation  between  chemical structure



and activity.  I. Effects of the number of chlorine atoms in chlo-



rinated  benzenes  on the  components  of  drug  metabolizing  systems



and hepatic constitutents.  Chem. Pharm. Bull.  23: 817.







Ariyoshi, T., et al.   1975b.   Relation  between  chemical structure



and activity.   II.  Influences  of isomers  of dichlorobenzene, tri-



chlorobenzene  and  tetrchlorobenzene on  the activities   of  drug



metabolizing enzymes.  Chem. Pharm.  Bull.  23: 824.







Ariyoshi, T., et al.   1975c.   Relation  between  chemical structure



and activity.  III. Dose  response on  tissue course of induction of



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zene.  Chem. Pharm. Bull.   23: 831.







Brown,  V.K.H.,  et  al.    1969.   Acute toxicity  and  skin  irritant



properties  of 1,2,4-trichlorobenzene.  Ann. Occup. Hyg.  12: 209.
                                C-46

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 Cameron,  G.R., et  al.   1937.   The  toxicity  of certain  chlorine
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 Carlson,  G.P.   1978.  Induction  of  cytochrome  P-450  by halogenated
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 Carlson,  G.P.  and R.G. Tardiff.   1976.  Effect of chlorinated ben-
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 Coate,  W.B.,  et al.  1977.  Chronic inhalation  exposure  of rats,
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 Egyankor, K.B.  and  C.S.  Franklin.   1977.   Interaction of  the tri-
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 Gotto, M.,  et  al.   1972.   Hepatoma  formation  in  mice after admin-
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Gurfein,  L.W.  and  Z.K.  Parlova.   1962.   The  Limit  of Allowable
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                               C-47

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Haas, J.M., et al.  1974.  Environmental considerations concerning
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Hintz, M.  1962.  Pollution characteristics  of 1,2,4-trichloroben-
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Johnson,  K.     1980.     Memorandum  to  D.W.  Kuehl.    U.S.   EPA.
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Kohli,  I., et al.   1976.   The metabolism  of higher  chlorinated
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Parke,  D.V. and R.T.  Williams.   1960.   Studies in  detoxication.
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Further  observations  on 1,3,5-trichlorobenzene.   Biochem.  Jour.
74: 1.

Porter, E.M.  and  J.J.  Snider.   1974.    30-day biodegradability of
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                               C-48

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 Rimington,  C. and G.  Ziegler.    1963.   Experimental porphyria  in
 rats  induced  by  chlorinated   benzenes.    Biochem.   Pharmacol.
 12:  1387.

 Rowe.   1975.   Written  communication.   April.

 Sax, N.I.    1975.   Dangerous Properties  of Industrial  Materials.
 4th  ed.  Van  Nostrand  Reinhold, New York.

 Simmons, P.,  et  al.   1976.  1,2,4-Trichlorobenzene: Biodegradable
 or  not?  Can.  Assoc.  Textile  Colourists  Chem.  Int.  Tech.  Conf.
 Quebec.  October 13-15.

 Smith, C.C.   1979.  Personal communication.  April 10.

 Smith, C.C.,  et  al.  1978.  Subacute  toxicity of 1,2,4-trichloro-
 benzene (TCB) in subhuman  primates.  Fed. Proc.   37: 248.

 Stephan, C.E.  1980.  Memorandum  to J. Stara.  U.S. EPA.  July 3.

 U.S. EPA.   1975.  Preliminary assessment  of suspected  carcinogens
 in drinking water.  Rep.  Cong.  No. PB-250961.

U.S. EPA.   1977.   Investigation  of  selected potential  environ-
mental contaminants:  Halogenated benzenes.   EPA  560/2-77-004.
                               C-49

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U.S. EPA.   1978.   In  depth studies on  health  and environmental



impacts of selected water pollutants.   U.S.  Environ. Prot. Agency.



Contract No. 68-01-4646.







U.S. EPA.   1980.   Seafood consumption data  analysis.   Stanford



Research Institute  International, Menlo  Park,  Calif.   Final rep.



Task II, Contract No. 68-01-3887.







U.S.  International  Trade  Commission.   1975.   Synthetic  organic



chemicals:   U.S.  production and sales.   U.S.  Govt. Print.  Off.,



Washington, D.C.







Williams,  R.T.    1959.    The  Metabolism  of Halogenated Aromatic



Hydrocarbons.   In;  Detoxication Mechanisms.   2nd  ed.   John  Wiley



and Sons, New York.  p. 237.
                               C-50

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                         TETRACHLOROBEN Z ENE
 Mammalian Toxicology and Human Health Effects
                            INTRODUCTION
      Tetrachlorobenzene   (TeCB)   exists   as   three   isomers-1,2,
 3,4-TeCB, 1,2,3,5-TeCB  and  1,2,4,5-TeCB.   Of  these,  1,2,4,5-TeCB
 is the most  widely used.    1,2,4,5-TeCB  is  used primarily  in  the
 manufacture  of  2,4,5-trichlorophenoxyacetic  acid  (2,4,5-T)  and
 2,4,5-trichlorophenol (2,4,5-TCP).  In 1973,  an estimated ten mil-
 lion  pounds  of  1,2,4,5-TeCB were  utilized  in the manufacture  of
 2,4,5-T while six million pounds were utilized  in  the  manufacture
 of 2,4,5-TCP  (U.S. EPA, 1977).  In the Soviet  Union,  1,2,4,5-TeCB
 is used as a  soil and grain pesticide (Fomenko, 1965).   It  is  not
 used  for  this purpose in the United States.
      Tetrachlorobenzene (TeCB)  has been  found to be  among  the
 metabolites of hexachlorobenzene (Mehendale,  et al. 1975; Rozman,
 et al.  1975),  lindane,  pentachlorocyclohexane,  pentachlorobenzene,
 and pentachlorophenol  (Engst,  et al. 1976a,b).
      1,2,4,5-TeCB  has an extremely low vapor  pressure, less than
 0.1 mm  Hg  at  25°C  (Sax, 1975).  The log of the octanol/water par-
 tition  coefficient  for  TeCB  is  4.93.
                             EXPOSURE
 Ingestion  from Water
     No literature was  found which  identified TeCB in water in the
United  States.   However, contamination of  runoff  as a  result  of
its  industrial use  is  certainly  feasible and  may  in part,  be
responsible  for   the   contamination   of   the   aquatic  organisms
described below.    Soil  microorganisms  are  capable  of  metabolizing
                               C-51

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lindane to tetrachlorobenzene, among others (Tu, 1976;  Mathur  and
Sana, 1977).  TeCB  derived  in this manner is available  from soil
runoff.
Ingestion from Food
     There are  some data to  show that  TeCB  will concentrate  in
fish  exposed  to  industrial effluent  discharge.   Kaiser  (1977)
identified two isomers of TeCB in  three  species  of  fish caught at
various distances  from  a pulp and  paper mill.    Similarly,  Lunde
and Ofstad (1976)  identified  tetrachlorobenzene  in  sprat (a  small
herring) from different locations in southeastern Norway.
     Qualitatively,  tetrachlorobenzenes  have  been  identified  in
the  food  chain  as a  result of  the biotransformation  of lindane.
Saha and Burrage  (1976)  administered  lindane to  hen  pheasants and
identified tetrachlorobenzene as part  of the array  of metabolites
found  in  eggs  and chicks as  well as  in the  body tissues  of  the
hens.     Balba   and  Saha  (1974)  followed   the   metabolism  of
14C-lindane  in  wheat plants  grown from treated seeds  and  iden-
tified  two  and  possibly  three of the  isomers of TeCB.   Kohli, et
al.  (1976  b,c)  in  laboratory studies  identified TeCB  as  a  minor
metabolite of lindane in lettuce and endives.
     Tetrachlorobenzenes have also been  identified  as metabolites
of gamma pentachlorocyclohexane in corn  and pea  seedlings.   Penta-
chlorobenzenes  have also been identified in the essential oil of
marsh  grass  (Miles,  et al.  1973).
      There  is legitimate doubt  as to whether  exposure to TeCBs as
breakdown  products  of lindane  and other substances  represents  a
                                C-52

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 significant  exposure, especially  considering that  concentrations
 of  the more  toxic  parent  compounds  are  higher.
     A bioconcentration  factor  (BCF)  relates the concentration  of
 a chemical in  aquatic animals to  the  concentration  in  the  water  in
 which  they  live.   The steady-state BCFs  for a lipid-soluble  com-
 pound  in  the tissues  of various aquatic animals seem to  be propor-
 tional to the  percent lipid in the tissue.   Thus,   the per  capita
 ingestion of a lipid-soluble chemical  can  be  estimated from the
 per capita consumption of fish and  shellfish, the weighted average
 percent lipids  of  consumed  fish  and shellfish,  and   a  steady-state
 BCF for the chemical.
     Data from  a  recent  survey on  fish and shellfish consumption
 in  the United States  were analyzed  by SRI International  (U.S.  EPA,
 1980).  These  data were  used  to  estimate  that the per capita  con-
 sumption  of  freshwater and  estuarine  fish  and shellfish  in the
 United States  is  6.5 g/day  (Stephan,  1980).   In  addition,   these
 data were used  with data  on  the  fat content of the   edible portion
 of  the same  species to estimate  that  the  weighted average percent
 lipids for consumed freshwater and estuarine fish and  shellfish is
 3.0 percent.
     A measured steady-state bioconcentration  factor of  1,800 was
 obtained   for 1,2,3,5-tetrachlorobenzene using bluegills (U.S.  EPA,
1978).     Similar bluegills  contained  an  average  of  4.8  percent
lipids (Johnson, 1980).   An  adjustment factor of 3.0/4.8  =  0.625
can be used to adjust the measured BCF from the 4.8  percent lipids
of  the bluegill to the 3.0  percent lipids  that  is the  weighted
average for  consumed  fish  and  shellfish.   Thus,  the  weighted
                               C-53

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average bioconcentration factor for 1,2,3,5-tetrachlorobenzene and



the edible portion  of  all  freshwater and estuarine  aquatic orga-



nisms consumed by Americans is calculated to be 1,125.



     No measured steady-state bioconcentration factor is available



for 1,2,4,5-tetrachlorobenzene.  However, the weighted average BCF



of 1,125 obtained for  the  very similar 1,2,3,5-tetrachlorobenzene



can also be used for this compound.



Inhalation and Dermal



     No reliable information has been located dealing with inhala-



tion or dermal exposure to TeCB.



                         PHARMACOKINETICS



Absorption, Distribution, Metabolism, Excretion



     Jondorf, et al. (1958) administered each of the three isomers



of TeCB to three rabbits  at oral doses  of 0.5  g/kg.   The animals



were followed for six days after dosing.  The percentage of admin-



istered dose recovered in the feces over this time for the respec-



tive  compounds  was:   1,2,3,4-TeCB,  5  percent;  1,2,3,5-TeCB,  14



percent; and  1,2,4,5-TeCB,  16  percent.   Considering  that  this is



over  a  six-day period  and that  some  of  the  fecal  TeCB  content



could possibly have  been  a result of biliary  excretion,  it would



appear that the gastrointestinal absorption of TeCBs is relatively



efficient.



     Table 1  shows the distribution  of unchanged TeCB  in rabbit



tissues six days after dosing.  Comparative distribution among the



three isomers shows a relative degree of consistency.  The one ex-



ception is in the  gut contents where 12 percent  of  the total re-



maining compound is present for 1,2, 4,5-TeCB which is about twice
                               C-54

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                              TABLE 1

          Unchanged Tetrachlorobenzene in Rabbit Tissues,
              Six Days After Dosing (0.5 g/kg orally)*
Percentage
TeCB
1,2,3,4
1,2,3,5
1,2,4,5
Liver
0.1
<0 .5
0.1
Brain Skin
2
<0.2 5
<0.1 10
Depot
Fat
5
11
25
of Dose
Gut
Contents
0.5
1'.4
6.2

Rest of
Body
2.0
5.2
6.4

Total
10
23
48
*Source:  Jondorf,  et al.  1958.
                                C-55

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that for the  other  isomers.   This  could  reflect lesser absorption
of 1,2,4,5-TeCB or, possibly, biliary excretion.
     Table 2 shows the extent of elimination of the isomers in ex-
pired air.
     Table 3  shows  the  urinary excretory pattern  observed  in the
three isomers.  The 1,2,3,4-TeCB isomer is more freely metabolized
than the  other two isomers,  and  1,2,4,5-TeCB  is  metabolized the
least.
     Kohli, et al.  (1976a) studied  the metabolism  of  TeCB isomers
in rabbits and  identified  the nature of TCP metabolites.   A dose
of 60 to 705 mg/kg was  administered to rabbits by intraperitoneal
injection  and  the urine and  feces  were  collected for  ten  days.
The metabolism  of both  1,2,3,4-TeCB and  1,2,3,5-TeCB  yielded two
common metabolites, 2,3,4,5- and 2 ,3 ,4 ,6-tetrachlorophenol (TeCP).
Another metabolite of 1,2,3,5-TeCB was 2,3,5,6-TeCP.  This metabo-
lite 2,3,5,6-TeCP was also the  only metabolite identified follow-
ing the administration  of  1,2,4,5-TeCB.  The  relationships  among
the various isomers were  strikingly similar to the data reported
by Jondorf, et al. (1958).
     Kohli, et  al.  (1976a) proposed  the  formation of  the  phenol
metabolites through corresponding arene oxides.   The  authors sug-
gested the involvement  of an  "  NIH  shift"  of  the  chlorine atom in
the formation  of  the  metabolites  (except  for  the formation  of
2,3,5,6-TeCP   from   1,2,3,5-TeCB    which   can   be  derived   from
2,3,5,6-TeCB and  oxide  without an  NIH shift  of  chlorine).   The
scheme proposed by Kohli is shown in Figure 1.
                               C-56

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                             TABLE 2

           Elimination of Unchanged Tetrachlorobenzenes
         in Expired Air of Rabbits Following Oral Dosing*
TeCB

1,2,3,4
1,2,3,5
1,2,4,5
Dose
(gAg)

0.5
0.3
0.5
0.3
0.5
Percentage of Dose in Expired Air
Days after Dosing
±
1.9
0.8
2.1
0.9
1.2
2_
2.2
1.7
2,1
3.2
0.2
115
1.6 0.2
6.7
1.2 2.9 2.6
9.8
0.2
Total

5.9
9.2
10.9
13.9
1.6
*Source:  Jondorf,  et.  al.  1958.
                               C-57

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                                        TABLE 3

                Urinary Excretion of Metabolites of Tetrachlorobenzenes
                    in Rabbits Following Oral Dosing (0.5 g/kg/)*






n
i
U1
CO



TeCB Glucuronide
1,2,3,4 30(22-36)
(5)
1,2,3,4 6(2-10)
(9)


1,2,4,5 4(1-8)
(ID
Percentage of
Ethereal
Sulfates
3(1-8)

2(1-6)
(9)


K
-------
n
i
(ji
V£>
                                                 FIGURE  1


                     Proposed Routes for the Biotransformation  of  Tetrachlorobenzene

                                         Isomers Via Arene Oxides

                                       Source: Kohli, et al.  1976a

-------
     From the above  information,  it is reasonable  to expect that
the metabolism of the TeCB  is  via  liver  microsomal  enzymes.  Ari-
yoshi, et al. (1975) reported  an  increase  in  cytochrome P-450 in-
duced by all three isomers in the rat liver as well as an increase
in delta  aminolevulinic  acid synthetase activity.   Rimington and
Ziegler (1963) showed that urinary porphyria and porphyria precur-
sors were increased  in rats  by  administration  of 1,2,3,4-TeCB but
not by 1,2,4,5-TeCB.  This  effect  was  correlated with an increase
in  porphyrins,  porphorobilinogen  and  catalase  activity  in rats
treated with 1,2,3,4-TeCB but not the 1,2,4,5 isomer.
                             EFFECTS
Acute, Subacute, and Chronic Toxicity
     Most -of the  information  on  tetrachlorobenzene comes  from
studies   done   in   the   Soviet   Union  and   is   concerned  with
1,2,4,5-TeCB.   The   LD$Q  values  for white  mice were  reported  to
be 1,035  mg/kg when  the  compound was administered  orally  in sun-
flower' oil  and  2,650 mg/kg  given orally as a  suspension in a 1.5
percent  starch  solution.    In rats  and  rabbits,   the  LDsg  was
reported  to be  1,500 mg/kg  when the compound  was  administered  in
sunflower oil  (Fomenko,  1965).  The apparent  cumulative activity
of this isomer of TeCB was demonstrated by Fomenko (1965).  A dose
of  300  mg/kg, 20 percent of the  LD5Q,  was administered  to rats
daily; 50 percent of the animals  died  when a  dose  equivalent  to
the  LDso  was  obtained.    The  same  investigator  administered
1,2,3,5-TeCB  in   oral  doses of  75  mg/kg  daily  for  two  months.
While there  were  presumptive changes  in  liver  function, prothrom-
bin index,  blood  cholesterol,  and  number  of reticulocytes, histo-
                               C-60

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 pathological  examination showed no  significant  change that would
 alter  liver function.   Adrenal  hypertrophy and decreased  content
 of  ascorbic  acid  in adrenals  were reported.   Histopathological
 examinations  did  not reveal appreciable  differences between con-
 trol and experimental groups.
     Further  experiments are  described  in. the  foregoing report
 (Fomenko,  1965)  from the Soviet Union in  which 1,2,4,5-TeCB was
 administered  in oral doses of 0.001, 0.005, and  0.05 mg/kg  to rats
 and rabbits over  an  8-month  period.   The  report states that doses
 of  0.005  mg/kg and  "especially" 0.05 mg/kg disrupted the condi-
 tioned reflexes.   It  is  stated  that "formation of a positive con-
 ditioned reflex became  slower  but the  latent  period remained the
 same."  It  is  also stated  that  rabbits treated with doses  of 0.05
 mg/kg "began  to display  disorders  in glycogen-forming function in
                             *•
 the  liver  only after  six experimental months."   No hematologic
 changes were   noted  in   the  animals.   At the  end  of  the dosing
 period, liver  weights were increased in animals  receiving doses of
 0.005  and  0.05 mg/kg.   The conclusion was  that  the  two higher
 doses were active and that the lower dose was not.
     The data from  the  above  studies (Fomenko/  1965)   are  only
 partially presented  and  the bulk  of  the  report consists  of  the
 conclusions of the author.  The studies of conditioned reflexes in
 rats were  conducted  on  a control  group  of five animals,  low  and
middle dose groups of seven animals each,  and a high dose group of
 six animals.   It  is  not  clear  from  the report  whether these  ani-
mals represented the total number of animals in each group.
                               C-61

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     Braun, et al. (1978)  administered 1,2,4,5-TeCB in the diet to



beagles at  5  mg/kg/day for  two years.   No  changes  in  clinical



chemistry parameters were  noted after  18 months.   At 24  months



there was a slight elevation of serum alkaline  phosphatase  activ-



ity and bilirubin levels.   The animals  were  then allowed  to re-



cover.  After three months the  serum  chemistry  changes noted were



no longer evident.  Gross and histopathological  studies were done



20  months  after  cessation  of  exposure.   No   treatment  related



changes were noted.



Synergism and/or Antagonism



     Since TeCBs  can  increase  cytochrome P-450 levels,  it,  like



other  halogenated benzenes,  appears  to induce  metabolic  enzymes



(Ariyoshi, et  al. 1975).   In  general,  the halogenated  benzenes



appear  to  increase  the  activity  of microsomal  NADPH-cytochrome



P-450-dependent enzyme  systems.   Induction of  microsomal  enzyme



activity has been shown to enhance  the  metabolism  of a wide vari-



ety of drugs, pesticides,  and other xenobiotics.  Exposure to TeCB



could,  therefore,   result   in  decreased  pharmacologic  and/or



toxicologic activity of numerous compounds.   Frequently,  chemical



agents  are  metabolized  to  more   active  or   toxic  "reactive"



intermediates.   In this  event, exposure  to TeCB would result in



enhanced activity and/or toxicity of these agents.



Teratogenicity, Mutagenicity, and Carcinogenicity



     No studies have been identified  which  directly  or indirectly



address  the  teratogenicity or  Carcinogenicity   of TeCB.    An ab-



stract  of  a  stucjy by Kiraly,  et  al. (1976) describes  a  study of



chromatid disorders  among  workers  involved in  the  manufacture of
                               C-62

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an organophosphorus compound.  Disorders  were  said  to be signifi-



cantly higher in this group  than  in  a  group involved in the manu-



facture of TeCB.  However, the abstract  concludes,  "The mutagenic



properties  of  tetrachlorobenzene were  confirmed."   This  Is  the



only reference seen referring to mutagenic activity of TeCBs.
                               C-63

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                      CRITERION FORMULATION
Existing Guidelines and Standards
      The  maximal  permissible  concentration  of  TeCB  in  water
established by the Soviet Union is 0.02 mg/1  (U.S. EPA, 1977).
Current Levels of Exposure
     No data  are  available on current  levels of exposure.   How-
ever, the report by Morita, et al. (1975) gives some indication of
exposure.  Morita,  et  al.  (1975) examined  adipose  tissue  samples
obtained  at  general hospitals and  medical  examiners  offices  in
central Tokyo.   Samples from  15 individuals were  examined?  this
represented 5 males and 10 females between  the  ages  of 13  and 78.
The tissues were examined  for  1,2,4,5-TeCB as well  as  for  1,4-di-
chlorobenzene and hexachlorobenzene.   The TeCB  content of  the fat
ranged from  0.006  to 0.039 mg/kg of tissue;  the mean  was  0.019
mg/kg.  The  mean  concentrations  of  1,4-dichlorobenzene  and  hexa-
chlorobenzene  were  1.7  mg/kg   and  0.21   mg/kg,   respectively.
Neither age nor sex correlated with  the  level of  any of the  chlo-
rinated hydrocarbons in adipose tissue.
Special Groups at Risk
     The  primary  groups at  risk from  the  exposure  to TeCB  are
those who deal with it in the workplace.  Since it is a metabolite
of certain insecticides, it might be expected  that  certain  indi-
viduals exposed to those agents might experience  more  exposure  to
TeCB, especially  since  its elimination  rate might  be relatively
slow in man.   Individuals  consuming  large quantities  of fish may
also  be  at risk  due  to the  proven  bioconcentration  of TeCB  in
fish.  The bioconcentration factor for  1,2,4,5-TeCB is  1,125.
                               C-64

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Basis and Derivation of Criterion
     The  dose  of  5  mg/kg/day 1,2,4,5-TeCB  reported by  Braun,  et
al.  (1978)  for  beagles caused  no  changes  in  clinical  chemistry
parameters  after 18 months  of exposure  via the  diet.    After  24
months,  however, slight  elevations were  noted in  serum alkaline
phosphatase activities  and bilirubin  levels.    These  changes were
reversible  three months after the  last exposure.   Whether  histo-
pathological changes  related  to treatment  with  1,2,4,5-TeCB  oc-
curred is unclear, as tissue studies were not begun until 20  months
after cessation  of  exposure.   Because no effects  were  observed  at
the dose  level  used  by  Braun, et al.  (1978)  until after  18  months
of exposure, and since  those changes were transient and not clearly
related to any  functional  impairment or  pathological lesions which
would adversly  affect   the performance  of the  animal,  5  mg/kg/day
can  be  considered a  no-observed-adverse-effect level  (NOAEL)  for
calculation of an acceptable daily  intake  (ADI).  Based  on a 70  kg
man, the ADI can be calculated from  the NOAEL using a safety  factor
of  1,000.   This safety factor is  required  by   the  guidelines  for
criteria derivation because:  (1)  the study by Braun,  et  al.   (1978)
was performed  on only  four  animals, (2) gross  and  histopathology
were not done  until 20  months after  the last exposure; and (3) sup-
portive epidemiologic or  subchronic data  are  not available.   For
1,2,4,5-TeCB,  the ADI  can be calculated as follows:
                   70  kg x 5 mg/kg   n oc    ,,
            ADI = 	1/000      =0.35 mg/day

     For the purpose of establishing a water  quality criterion,  it
is assumed that  on the  average, a person  ingests 2 liters of water
                               C-65

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and 6.5  grams  of  fish daily.  Since  fish  may bioconcentrate this
compound,  a  bioconcentration factor  (F)  is used  in  the calcula-
tion.
     The equation for calculating an  acceptable  amount  of TeCB in
water is:
Criterion = 2 1 + (lxOOSS) = 37.6 ug/1 or   38 ug/1
where:
          21=2 liters of drinking water consumed
   0.0065 kg  = amount of fish consumed daily
        1,125 = bioconcentration factor
          ADI « Acceptable Daily Intake (mg/kg) for a 70 kg/person)
     Thus, the recommended criterion for  1,2,4,5-TeCB  in water is
38 ug/1.  Due  to the lack of data  describing  toxicologic effects
of the other TeCB isomers and the predominant  use  of 1,2,4,5-TeCB
by  industry,  no  criteria are  recommended  for  the 1,2,3,4-  or
1,2,3,5-TeCB isomers.  This  criterion  for 1,2,4 ,5-tetrachloroben-
zene  can alternately  be expressed as  48  ug/1   if  exposure  is
assumed to be from the consumption of fish and shellfish alone.
                               C-66

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                             REFERENCES

 Ariyoshi,  T.,  et al.   1975.   Relation between chemical  structure
 and activity.   II.  Influences of isomers in dichlorobenzene,  tri-
 chlorobenzene  and tetrachlorobenzene  on the  activities  of drug-
 metabolizing enzymes.   Chem.  Pharm.  Bull.   23: 824.

 Balba,  M.H. and  J.G.  Saha.    1974.    Metabolism of  lindane  14c
 by  wheat plants grown  from  treated seeds.   Environ. Let.  7: 181.

 Braun,  W.H.,  et  al.    1978.    Pharmacokinetics  and  toxicological
 evaluation  of dogs  fed 1,2,4,5-tetrachlorobenzene in  the diet for
 two years.  Jour. Environ.  Pathol. Toxicol.  2: 225.

 Engst,  R.,  et al.   1976a.   The metabolism  of   hexachlorobenzene
 (HCB) in rats.  Bull.  Environ. Contam. Toxicol.   16:  248.

 Engst,  R.,  et  al.    1976b.   The  metabolism  of  lindane and  its
metabolites  gamma-2,3,4,5,6-pentachlorocyclohexene,   pentachloro-
 benzene and pentachlorophenol  in rats and  the  pathways  of lindane
metabolism.  Jour. Environ. Sci. Health Bull:  95.

Fomenko, V.N.   1965.   Determination of  the   maximum  permissible
concentration of tetrachlorobenzene in water basins.   Gig.  Sanit.
30:  8.
                               C-67

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Johnson,  K.     1980.    Memorandum  to  D.W.  Kuehl.    U.S.   EPA.


March 10.




Jondorf, W.R., et al.   1958.  Studies  in detoxication.   The metab-


olism  of halogenobenzenes  1,2,3,4-,  1,2,3,5- and  1,2,4,5-tetra-

                  i
chlorobenzenes.  Jour. Biol. Chem.  69:  189.




Kaiser, K.L.E.  1977.   Organic contaminant residues in  fishes from


Nipigon Bay Lake Superior.  Jour.  Fish.  Res. Board Can.   34:  850.




Kiraly, J., et  al.   1976.  Chromosome studies in  workers  exposed


to organophosphorus insecticides.   Mankavedelem.   22: 27.




Kohli, J., et al.   1976a.   The  metabolism of higher  chlorinated


benzene isomers.  Can. Jour. Biochem.   54: 203.




Kohli, J., et al.    1976b.   Balance  of  conversion of   [^4C]  lin-


dane in  lettuce in hydroponic culture.   Pestic.  Biochem.  Physiol.


6: 91.




Kohli, J., et al.  1976c.   Contributions  to ecological  chemistry.


CVII.   Fate  of 14C-lindane  in  lettuce,  endives  and   soil  under


outdoor conditions.  Jour. Environ. Sci. Health  Bll: 23.




Lunde, G.  and E.B.  Ofstad.   1976.   Determination  of fat  soluble

chlorinated compounds in fish.  Jour.  Anal. Chem.  282:  395.
                               C-68

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 Mathur,  S.P. and  J.G.  Saha.   1977.   Degradation of  lindane-14C
 in a mineral soil and in an organic soil.  Bull.  Environ.  Contam.
 Toxicol.   17: 424.

 Mehendale,  H.M.,  et  al.   1975.   Metabolism  and  effect  of hexa-
 chlorobenzene  on hepatic  microsomal enzymes  in  the rat.   Jour.
 Agric. Food  Chem.   23: 261.

 Miles, D.H., et al.  1973.  Constituents  of  marsh grass.  Survey
 of  the  essential  oils   in  Juncus  roemerians.    Phytochemistry
 12:  1399.

 Morita, M.,  et  al.   1975.   A  systematic  determination  of  chlori-
 nated  benzenes   in  human   adipose   tissue.     Environ.    Pollut.
 9: 175.

 Rimington, C. and G.  Ziegler.    1963.   Experimental  porphyria in
 rats  induced   by   chlorinated   benzenes.     Biochem.  Pharmacol.
 12: 1387.

 Rozman, K.,  et  al.   1975.    Separation,  body  distribution  and
metabolism of hexachlorobenzene after oral  administration to  rats
and rhesus monkeys.   Chemosphere.  4: 289.
                               C-69

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Saha, J.G. and R.H.  Burrage.   1976.   Residues of lindane  and  its
metabolites  in  eggs, chicks  and  body  tissues  of  hen  pheasants
after  ingestion  of lindane carbon-^4  via treated  wheat seed  or
gelatin capsules.  Jour. Environ. Sci.  Health Bll:  67.

Sax/ N.I.   1975.   Dangerous  properties of  industrial  materials.
4th ed.  Van Nostrand Reinhold,  New York.   p. 1145.

Stephan, C.E.  1980.  Memorandum to J.  Stara.  U.S.  EPA.  July  3.

Tu, C.M.   1976.    Utilization and  degradation of lindane  by  soil
microorganisms.  Arch. Microbiol.  108:  259.

U.S. EPA.   1977.   Investigation  of  selected potential  environ-
mental  contaminants:  halogenated benzenes.   560/2-77-004.   U.S.
Environ. Prot. Agency.

U.S. EPA.  1978.   In-depth studies on health  and  environmental  im-
pacts of  selected  water pollutants.   U.S. Environ. Prot.  Agency.
Contract NO. 68-01-4646.

U.S. EPA.   1980.   Seafood consumption data analysis.   Stanford
Research Institute International, Menlo Park, Calif.   Final rep.,
Task II, Contract No. 68-01-3887.
                               C-70

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                         PENTACHLOROBENZ ENE
 Mammalian Toxicology and Human Health Effects
                            INTRODUCTION
      Pentachlorobenzene (QCB*) is used primarily as a precursor in
 the  synthesis  of  the   fungicide,  pentachloronitrobenzene  (PCNB,
 Quintozene,  Terraclor), and  as a  flame  retardant.   It has  been
 suggested as an  intermediate  in the production  of  thermoplastics
 (Kwiatkowski,  et  al.  1976).   QCB  is a  white  solid  crystalline
 material  at  room  temperature  and,  like other  halogenated benzenes,
 is  both lipophilic and  hydrophobic.   Approximately  1.4 x 106  kg
 of  Pentachlorobenzene was  produced  in 1972  and  it  is  estimated
 that  16.6 x 103  kg of  the  material was  discharged  into  ambient
 water sources.   Much of the  exposure of  the  population  to QCB  is
 derived  from exposure  to  lindane,   hexachlorobenzene  (HCB),   and
 PCNB.   The metabolism of lindane  to  QCB  is  well established,  and
 it  has been demonstrated  in  humans  (Engst,  et  al.  1976a) ,   rats
 (Engst,  et  al.  1976b,c;  Seidler,  et  al.  1975;  Kujawa,  et   al.
 1977), and rabbits  (Karapally, et al. 1973).  Biotransformation  of
 lindane to QCB can  occur earlier in the food chain.  Engst, et  al.
 (1977) identified QCB as a product of the metabolism of  lindane  by
mold grown spontaneously on grated carrots.  Tu (1976)  identified
*QCB (for quintochlorobenzene) rather than PCB will be used as the
abbreviation for Pentachlorobenzene  to  avoid  confusion with poly-
chlorinated biphenyls.
                               C-71

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71 soil microorganisms  which would biodegrade lindane.   Thirteen
of these were  examined  further and were  found  to produce QCB  as
one of the metabolites of the insecticide.  Mathur and Sana (1977)
have also reported QCB as a soil degradation product of lindane.
     QCB  has  been  identified  as  a  metabolite   of  HCB  in  rats
(Mehendale, et al.  1975;  Engst,  et al. 1976c) and  rhesus  monkeys
(Rozman, et al, 1977, 1978; Yang, et al. 1975, 1978).
     Tetrachloronitrobenzene (TCNB) occurs as a residue in techni-
cal grade  PCNB.   Borzelleca, et al.  (1971)  detected  TCNB  storage
in tissue  of rats,  dogs,  and cows following  feeding  studies  with
PCNB.   Rautapaa,  et al.  (1977)  examined  soil samples  in  Finland
from areas  that  have been treated  with PCNB and found  a  maximum
PCNB level  of  27  mg/kg  of soil and the highest  QCB  level  of  0.09
mg/kg of soil.
     Igarashi,  et al. (1975)  identified QCB  as a further degrada-
tion product of pentachlorothioanisole  in soil.
     The  importance  of  QCB  as a  contaminant  of PCNB  in  treated
soil is demonstrated by the  study of Beck and Hansen  (1974).   They
studied 22  soil  samples  from fields where technical PCNB had  been
used regularly during the foregoing 11 years.   The  concentration
range for  PCNB in the samples was from  0.01  to 25.25 mg/kg of soil
and  for  QCB was  from 0.003  to 0.84  mg/kg  of soil.   The  samples
were studied for  a  period of  600  days.  The half-life  of QCB  in
two  separate determinations  was  194 and 345  days.  The calculated
log  octanol/water partition  coefficient for  QCB = 5.63.
                               C-72

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                              EXPOSURE
 Ingestion from Water
      The following discussion concerning the ingestion of QCB from
 food, especially as related  to its presence in  marine  organisms,
 also relates to the presence of the compound in water.  Burlingame
 (1977)  has identified QCB in effluent from  a wastewater  treatment
 plant in southern California.  Access to water  by QBC can occur by
 a number of means including  industrial discharge or as a  breakdown
 product or contaminant of widely used organochlorine compounds.
 Ingestion from Food
      From the available information, it appears  that  the  presence
 of  QCB in soil and  its  persistence there can result  in  accumula-
 tion  within the  food chain.   This  also holds true for  its  ecologi-
 cal precursors.   For example, Balba and Sana (1974) treated  wheat
 seed  with  isotopically  labeled  lindane and  observed  a  number  of
 metabolites,  including QCB,   in  the seedlings  and mature plants.
 Kohli,  et  al.  (1976a)  found  that  isotopically  labeled lindane
 added  to the  nutrient  medium  for lettuce was metabolized  to a  num-
 ber of  products  including QCB.   Dejonckheere,  et al.  (1975, 1976)
 examined  samples  from  soil which had been used  to grow lettuce  and
 witloof-chicory.  The  soil had been  treated with  PCNB  for  a 6-year
 period.   Average  QCB concentrations ranged  from 0.25 to 0.85 ppm.
 Lunde  (1976)  has examined fish  from southeastern  Norway  for  the
 presence  of polychlorinated  aromatic  hydrocarbons.   QCB  was among
 a  number  of  compounds  identified  in  extracts  of plaice,  eel,
 sprat, whiting, and cod.   Lunde  and Ofstad  (1976) quantitated  the
amount  of chlorinated  hydrocarbons  in  sprat   oil.    Six  samples
                               C-73

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taken from different locations and/or at different times contained
QCB at 0.7 to 3.8 ppm.  Ten Berge and Hillebrand (1974) identified
the presence  of a  number  of organochlorine  compounds,  including
QCB, in plankton, shrimp, mussels, and fish from the North Sea and
the Dutch Wadden Sea.  The compounds were present at part per bil-
lion levels.
     Stijve (1971)  detected  QCB  in  chicken  fat  which was ascribed
to residues of HCB.  Kazama, et al.  (1972) administered QCB by in-
tramuscular  injection to hens  and  recovered 7.3  percent  of  the
dose  in  the yolk of  the  egg.   No  material  was found  in  the  egg
white.   Saha  and Burrage  (1976)  administered isotopically labeled
lindane  to  hen pheasants via  treated  wheat seed  or  gelatin cap-
sules and recovered QCB as one of  the  metabolites  in the body of
the  hen, in  the  eggs  and  in the  chicks.    Dejonckheere,  et  al.
(1974) reported on  the  presence of QCB  in animal fat  and suggested
that  it  was derived from pesticide residues of  HCB and lindane in
feed.  Greve  (1973) identified QCB and HCB in wheat products used
for animal  feed and detected QCB  in the  fat of animals utilizing
that  feed.
      A bioconcentration factor  (BCF) relates the concentration of
a  chemical  in aquatic animals  to  the concentration  in  the water in
which they  live.   The  steady-state BCF  for  a  lipid-soluble com-
pound in the  tissues of various  aquatic  animals  seems to be pro-
portional  to  the  percent  lipid  in  the  tissue.   Thus,  the  per
capita ingestion  of a lipid-soluble chemical  can  be  estimated  from
                                C-74

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 the per  capita consumption  of  fish and  shellfish,  the  weighted
 average  percent  lipids  of  consumed  fish  and  shellfish,  and  a
 steady-state BCF  for the chemical.
      Data from a recent  survey  on  fish and  shellfish consumption
 in the United States were analyzed  by  SRI  International (U.S.  EPA,
 1980).  These data were used to estimate  that  the  per capita  con-
 sumption of  freshwater  and  estuarine  fish and  shellfish in  the
 United States  is 6.5  g/day  (Stepahn,  1980).   In  addition,  these
 data  were used with data on  the fat content  of  the edible portion
 of the same species to estimate that the  weighted  average percent
 lipids for  consumed freshwater  and  estuarine  fish and  shellfish is
 3.0 percent.
     A measured steady-state bioconcentration factor  of 3,400  was
 obtained  for  pentachlorobenzene  using  bluegills  (U.S,  EPA, 1978).
 Similar   bluegills  contained  an average  of  4.8  percent lipids
 (Johnson, 1980).   An adjustment factor  of 3.0/4.8  = 0.625 can  be
 used to adjust  the  measured  BCF  from the 4.8 percent lipids of  the
 bluegill  to the 3.0 percent lipids that  is  the weighted  average
 for consumed  fish and  shellfish.  Thus,  the weighted  average bio-
 concentration factor for pentachlorobenzene and  the  edible portion
 of  all  freshwater  and  estuarine   aquatic  organisms  consumed   by
Americans is calculated to be 2,125.
jnhalation
     There is very  little information concerning  atmospheric expo-
sure  to QCB.   The  primary  site  for such exposure could  be   the
workplace in industries utilizing and/or producing QCB.
                               C-75

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Dermal
     No information was obtained which concerns dermal exposure to
pentachlorobenzene.
                         PHARMACOKINETICS
Absorption, Distribution, Metabolism, Excretion
     Table 1 presents  data  from Parke and Williams  (1960)  on the
metabolism of pentachlorobenzene by  rabbits.   It  can be  seen that
a substantial portion  of the oral dose  was  recovered in  the gut
contents three to  four days after dosing.  Except  for  the possi-
bility of biliary  secretion, which appears unlikely from the data
obtained after a  parenterally  administered dose,  it would appear
that pentachlorobenzene is very poorly absorbed from the gastroin-
testinal tract.   It is also evident  that distribution favors depo-
sition in  the fat. Engst, et al.  (1976c)  administered  QCB orally
to  rats  at  a dose  of  8 mg/kg  for 19  days.    They  identified
2,3,4,5-tetrachlorophenol and pentachlorophenol  as the  major uri-
nary  metabolites.    They also  detected  2,3,4,6-tetrachlorophenol
"and/or"  2,3,5,6-tetrachlorophenol  and  unchanged  QCB.    They re-
ported  the  presence  of  1,3,5-trichlorobenzene   in  the  liver.
Kohli,  et  al.   (1976b)   described   2,3,4,5-tetrachlorophenol  and
pentachlorophenol  as  urinary metabolites of  QCB  in  the rabbit.
They were detected at  yields of 1 percent each of  the administered
dose.   The authors suggest  that  the dechlorination-hydroxylation
step  to the tetrachlorophenol derivative proceeds  through  an  arene
oxide  step.   Koss and  Koransky (1977)  reported pentachlorophenol
and  2,3,4,5-tetrachlorophenol  as metabolites  of  QCB in  the rat.
                                C-76

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                                                              TABLE 1

                                        Disposition of Pentachlorobenzene in the Rabbit as
                                                 Percentage  of  Administered Doset




o
1
-J

Dose/Route
mg/kg
0.5 p.o.
0.5 p.o.
0.5 s.c.
Time
After
Dose
(Days)
3
4
10

Urine
Tri- or Penta-
Chlorophenol
0.2
0.2
0.7

Other
Phenol
1
1
1

Gut
Feces Contents
5 45
5 31
1.5 0.5

Pelt
1
5
47*

Depot
Fat
15
9
22*

Rest
of
Body
6
5.5
10

Un-
changed
0
0
0

Other
Hydro-
carbons
9
21
L2

Total
Accounted
For
82
78
85
*l.ocated mainly at site of injection.

-------
However, they  stated  that the amount  of  pentachlorophenol recov-
ered in the urine represented about  9  percent  of  the administered
dose.    Quantitively,   this   is  substantially  greater  than  the
amounts of pentachlorophenol reported by Kohli, et al. (1976b) for
the rabbit.  Parke and Williams (1960) reported that less than 0.2
percent of  the  dose  was recovered as  pentachlorophenol  in rabbit
urine,  also a  substantial  difference from  that  observed  in the
rat.   Rozman,  et al.  (1979)  found  that  biological  half-life for
QCB in  rhesus  monkeys  to be two to  three months.   After 40 days,
10  percent  of  the  total dose was  excreted  in  the  urine; of  this,
58  percent was  pentachlorophenol.  After the same period, about 40
percent of  the  dose was  excreted in  the feces, 99 percent  of which
was unchanged QCB.  These authors  explained  this as  unabsorbed QCB
that  was  secreted in  bile  into the  GI  tract.  Ariyoshi,  et al.
(1975)  reported that,  in female  Wistar rats intubated with QCB at
250 mg/kg for  three days, the compound increased the liver  content
of  cytochrome  P450  and  increased the activities  of aminopyrine
demethylase and aniline  hydroxylase.   Microsomal protein and  phos-
pholipids  were also  increased  as  was the  activity of  delta-ami-
nolevulinic acid synthetase.
      Further  information on  the  biotransformation  and  accumulation
properties  of QCB can  be obtained from a  study  reported   by  Vil-
leneuve and Khera  (1975) who  studied the  placental  transfer  of
halogenated benzene  in rats.  They administered oral doses  of QCB
to pregnant rats on  days 6  through  15  of  gestation.   It can  be
seen in Table 2 that the accumulation in the  organs is  dispropor-
tionate to  the increasing dose,  implying   that  at doses  between
                                C-78

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                                                              TABLE 2

                                    Tissue Distribution of Pentachlorobenzene  (pp* wet  tissue)
                                         Following Oral Administration to Pregnant Rats*
Dose
9A9)
50
0
^ 100
200

Fat3
470+106
824+116
3350^331


Liver3
13
18
91
.9+5.1
.1+2.0
.1+6.6

Brain3
6.9+ 1.2
12.0+ 1.7
62.5+10.2

Heart3
6.2+1.0
12.6+2.0
57.5+9.6

Kidney3
6.00.1
10.6+1.5
43.5+2.6

Spleen3
4 . 5+^1 . 1
8.3+^1.3
46.2+8.1
Whole3 «b
Fetus
9.65+1.3
21.2 ^2.1
55.1 +6.7
Fetalc
Liver
4.37+0.69
10.4 j+1.31
40.4 +6.02
Fetalc
Brain
3.08+0
5.31+^0
20.5 +2


.55
.60
.64
aKepresents the mean of 5 animals + S.E.M.
bRepresents the mean of two fetuses from 15 litters j+ S.E.M.
cHei>resents the mean of five fetuses each from a different litter + S.E.M.

-------
100 and 200 mg/kg,  elimination approaches zero  order  kinetic be-
havior.  The ease of accumulation of the compound within the fetus
is also evident.  This will be discussed further below.
                             EFFECTS
Acute, Subacute, and Chronic Toxicity
     Goerz, et al.  (1978), in a study of the comparative abilities
of QCB  and HCB to  induce  porphyria,  administered a diet  of 0.05
percent QCB  to female adult  rats  for a period  of 60  days.   The
treatment  resulted  in an increased urinary excretion of porphyrins
by the HCB treatment, but  none with  the QCB treatment.  It is un-
certain from  these  experiments whether  the  dosage levels  for QCB
are  adequate.   Induction of experimental  porphyria  can be accom-
plished with  all  of the other  chlorinated  benzenes,  and it  would
appear  that  a more  detailed  examination  of   pentachlorobenzene
should  be  done before  any final conclusions  are made  concerning
its  ability  to induce porphyria.  A  survey of the literature has
revealed   no  other  published  data  on  the  acute,  subchronic  or
chronic  toxicity of QCB.   The  only  exceptions  to  this are  data
which  have been gathered  in  association with pharmacokinetic and
teratologic  studies,  but  on  the basis  of  the  number of  animals
utilized and  the  time  of administration, these  are not particular-
ly useful  for calculating  criteria.   For example, Khera and Ville-
neuve  (1975)  administered QCB  in doses  of 50, 100, and 200  mg/kg
orally to pregnant rats during days  6 to  15  of gestation.  The
adult rats (20 in each group) did  not display any "overt"  signs  of
 toxicity,  though it is not certain  whether the word "overt" refers
 to any particularly informative toxicological examination.
                                C-80

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     There are  no  other studies which describe  the chronic toxi-
city of pentachlororbenzene.
     Koss  and  Koransky  (1977)  have suggested  that a  major con-
sideration in the  toxicity  of  pentachlorobenzene is its biotrans-
formation to pentachlorophenol.  Considering  that the findings by
Rozman, et al.  (1979)  showing  the  half-life of pentachlorobenzene
to  be  two to three  months, and the  urinary excretion  of penta-
chlorophenol  to be  6  percent of  the  administered  dose,  it  is
doubtful  that over  a period of 40 days a  substantial  quantity of
pentachlorophenol would be made available  to the  system.
Synergism and/or Antagonism
     The  interaction of QCB with microsomal  enzyme systems might
result in effects  on biotransformation and toxicity  of  drugs and
other chemicals.   However,  there  are no available  data  on syner-
gistic or antagonistic effects.
Carcinogenicity, Mutagenicity, Teratogenicity
     There is  one  report that alludes  to  the  carcinogenicity of
pentachlorobenzene  in mice  and  the absence  of   this  activity  in
rats and  dogs  (Preussman, 1975).   This paper has not  been evalu-
ated due to difficulties in locating the source.  When made avail-
able it  will be evaluated  as  a  possible  basis  for  a  criterion
standard.
     Teratogenicity  studies with QCB have  been  reported  by Khera
and Villeneuve  (1975).   As  indicated  above,  QCB at  50,  100,  or
200 mg/kg in corn oil was administered by stomach tube to pregnant
rats on days 6  to 15  of gestation.   The  authors  did not interpret
                               C-81

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these data  to demonstrate  the  teratogenicity of  QCB.    However,
extra ribs are considered abnormal in  fetal development.   Table 3
represents findings resulting  from Cesarean sections done  on  day
22 of pregnancy.  The high dose of QCB produced an increased inci-
dence of  uni-  or  bilateral  extra rib, as well as  sternal defects
consisting of  unossified or  nonaligned  sternabrae with  cartila-
genous precursors present.  The  authors  considered  that  the ster-
nal  defects  suggested  a retarded sternal  development,   and  that
these were  related  to  a decreased mean fetal weight.    At lower
doses the sternal defects  were  not  noted, but  there  was  an  in-
creased incidence of  extra  ribs.  The number  of  litters  with  one
or more  litter mates  showing  an  anomalous  rib  number  (14th  and
15th combined), versus  numbers  of litters examined for  each  dose
group, was  3/19 for  0  mg/kg,  14/19  for  50  mg/kg, 11/19  for  100
mg/kg, and 15/19  for  200 mg/kg,  showing  an  apparent  dose-related
incidence.
     No data have been  found concerning the mutagenicity of QCB.
                               C-82

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                           TABLE 3

        Prenatal Data on Rats Dosed on Days 6 to 15 of
              Gestation with Pentachlorobenzene*

No. of rats pregnant at term
No. of live fetuses, mean
% fetal death,
(dead + deciduomas) 100
total implants
Fetal weight, g., mean
No. of fetuses examined
for skeletal anomalies
Anomalies, type and incidence
Extra ribs:
uni
bilateral
Fused ribs
Wavy ribs
Sternal defects
No. of fetuses examined
for visceral defects
Runts
Cleft Palate
Other defects

0
19
12.1

1.3
4.8

127

2
2

5
5

67
1


Dose
50
18
12.5

4.2
4.9

129

18
10

2
4

69
2
1

(mg/kg)
100
19
11.5

3.1
4.8

122

10
11



67




200
17
10.7

3.2
4.4

100

17
46
2
31

52
2

2
*Source: Khera and Villeneuve, 1975
                               C-83

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                      CRITERION FORMULATION
Current Levels of Exposure
     Morita, et al.  (1975) examined  levels  of  QCB  in adipose tis-
sue samples obtained from general hospitals and medical examiners'
offices in central Tokyo.  The samples were collected from a total
of 15 people.  By gas  chromatography the  authors found the resid-
ual level  of QCB  to range from  0.004  ug/g  to 0.020  v.g/g,  with a
mean value of  0.09  ug/g  of  fat.   Lunde  and  Bjorseth  (1977)  ex-
amined blood  samples from workers  with occupational  exposure  to
pentachlorobenzene  and  found that  their  blood  samples  contained
higher levels of this  compound  than  a  comparable group of workers
not exposed to chlorobenzenes.
Special Groups at Risk
     A group at  increased  risk  would appear to  be  those individ-
uals exposed occupationally.   Due to the persistence  of the com-
pound  in  the food  chain,  an increase  in  the  body  burden  of  QCB
might be expected in individuals  on  high  fish  diets or diets high
in agricultural products containing residues of PCNB spraying.
Basis and  Derivation of Criterion
     A survey of the QCB literature  revealed  no  acute, subchronic
or chronic toxicity  data with the exception of the  study by Khera
and Villeneuve (1975).   These authors found an  adverse  effect on
the fetal  development  of embryos  exposed  in utero to pentachloro-
benzene administered to  the  dams at 50 mg/kg  on days 6  to 15  of
gestation.   This dose constitutes  a  low-observed-adverse-effect
level  (LOAEL).    According  to  current guidelines,  extrapolation
from such  data requires application of a  safety  factor  of from
                               C-84

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1 to 10 .  Since the observed effect was only suggestive of terato-
genicity of QCB,  a  safety factor of 3  is  applied.   Because long-
term toxicity  data  on humans  are  not  available  and  the existing
animal data are  sparse,  an  additional  safety factor  of  1,000 is
applied to the calculation  of  an acceptable  daily intake (ADI) as
follows:
                        70 kg  «O
The average  daily  consumption of water  was  taken to  be  2 liters
and the consumption of fish to be 0.0065 kg daily.  The bioconcen-
tration factor for QCB is 2,125.
     Therefore:
Recommended Criterion * 2 + (2 125 x 0 0065)  * °-074 mg/1  (or — »— 74ug/l)

    The recommended water quality criterion for pentachlorobenzene
is 74  ug/1.   The criterion can  alternatively be  expressed  as 85
ug/1 if  exposure is assumed  to be from consumption  of  fish and
shellfish alone.
                               C-85

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Balba,  M.H.  and  J.G.   Saha.    1974.    Metabolism  of  Lindane-14C
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Beck,  J. and  K.E.  Hansen.    1974.    Degradation  of  quintozene,
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Borzelleca, J.F., et al.   1971.  Toxicologic and metabolic studies
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Dejonckheere,  W., et al.   1974.   Hexachlorobenzene  (HCB) and  other
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                                C-86

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Dejonckheere, W., et al.   1976.  Residues  of quintozene,  its con-
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Engst,  R.,  et al.   1976b.   The metabolism  of  hexachlorobenzene
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Engst,  R.,  et  al.    1976c.   The  metabolism  of lindane  and its
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Goerz,  G.,  et  al.   1978.  Hexachlorobenzene (HCB)  induced porphy-
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                                C-87

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Greve,  P.A.   1973.   Pentachlorobenzene  as a contaminant of animal
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Igarashi, H., et  al.   1975.   Studies on the metabolic degradation
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Karapally,  J.C.,  et  al.   1973.   Metabolism  of  lindane-14c  in
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Kazama, M.,  et  al.    1972.   Chemical hygiene studies  on  organic
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Khera, K.S. and D.C.  Villeneuve.   1975.   Teratogenicity studies  on
halogenated benzenes  (pentachloro-,  pentachloronitro-, and  hexa-
bromo-} in rats.  Toxicology.   5: 117.

Kohli,  J.,  et al.   1976a.    Balance of  conversion  of carbon-14
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                               O88

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Kohli, J., et  al.   1976b.   The  metabolism of higher  chlorinated



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Koss, G. and W. Koransky.   1977.  Pentachlorophenol  in  Different



Species  of  Vertebrates after  Administratin of  Hexachlorobenzene



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Plenum Press,  New York,  p.  131.







Kujawa, M.,  et al.  1977.   On the Metabolism of  Lindane.   In;  S.H.



Zaidi  (ed.), Environmental Pollution  and  Human Health,  Proc.  1st



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Kwiatkowski,  G.T.,  et al.   1976.   Chloroaromatic  ether  amines.



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Lunde, G.  1976.   Persistent and  non-persistent fat soluble  chlo-



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sekr., Publ.,  Org. Miljoegifter  Vatten Word.  Symp.  Vattenforsk,



12th.  p. 337.







Lunde, G. and  A.  Bjorseth.   1977.   Human  blood samples  as  indi-



cators of occupational exposure  to persistent chlorinated  hydro-



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Lunde, G. and  E.B. Ofstad.   1976.   Determination of  fat-soluble



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Mathur, S.P.  and J.G. Saha.   1977.   Degradation of  lindane-14C
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Toxicol.  17: 424.

Mehendale, H.,  et al.   1975.   Metabolism and  effects of  hexa-
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Agric. Food Chem.  23: 261.

Morita, M., et  al.   1975.   A systematic determination  of  chlori-
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Parke, D.V. and  R.T.  Williams.   1960.   Studies  in  detoxification
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Preussmann. R.   1975.  Chemical carcinogens in  the  human environ-
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Rautapaa,  J., et al.   1977.   Quintozene in some soils  and plants
in Finland.  Ann. Agric.  Fenn.  16:  277.

Rozman, K., et  al.   1977.   Longterm feeding study  of  hexachloro-
benzene in rhesus monkeys.   Chemosphere.  6: 81.
                               C-90

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Rozman, K.,  et al.   1978.   Chronic  low  dose exposure  of  rhesus



monkeys to hexachlorobenzene.  Chemosphere.   7:  177.







Rozman, K.,  et al.    1979.   Metabolism  and pharmacokinetics  of



pentachlorobenzene  in rhesus  monkeys.    Bull.  Environ.  Contam.



Toxicol.  22: 190.







Sana, J.G. and R.H.  Burrage.   1976.  Residues of  lindane  and  its



metabolites  in  eggs, chicks, and  body tissues  of hen  pheasants



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Seidler, H., et al.   1975.   Studies on the  metabolism of  certain



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                               C-91
                                  \

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             f
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 U.S. EPA.  1978.   In-depth studies on health and environmental im-

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                               C-92

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                        HEXACHLOROBENZENE
Mammalian Toxicology and Human Health Effects
                           INTRODUCTION
     Hexachlorobenzene  (HCB)  is a crystalline  substance  which is
virtually insoluble in water.   It is used  to control  fungal dis-
eases in cereals,  and it is used in  a number  of organic syntheses.
HCB should not be  confused  with  the  more commonly used insecticide
benzene hexachloride (hexachlorocyclohexane).
                 Cl
          Ci
                                              0
rt
  H^'ri

                 Cl
                                                  Cl
                 HCB                       benzene hexachloride
     The main agricultural  use  of  HCB is  on wheat  seed  which is
intended solely for planting.   For  this purpose/ HCB  is mixed with
a blue dye, giving  the treated wheat a distinct blue color.  This
coloration is intended as a warning that  the  seed has been  treated
with a poison and must not be  used  for  stock  or human consumption.
In 1971, about 6,800  kg  were  used  in  the United  States as a seed
fungicide, its only registered  use (Isensee, et al.  1976).   De-
spite advice and  regulations,  treated  seed grain  has been fed to
animals  intended  for, or  whose  products  are  intended  for human
consumption.   HCB does not degrade  easily under normal  conditions.
Trace amounts have  been  found in areas and ecological  systems far
                               C-93

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removed from  the original area, of application.   HCBs  impact  on
agriculture as a result of environmental contamination may be much
larger than its utility as a fungicide to control smut diseases in
cereal grains.   Foodstuffs  such  as eggs,  milk,  and  meat  become
contaminated  with  HCB  as  a  result of  ingestion of  HCB-treated
cereals by livestock.
     Commercial  production  of  HCB  in  the  United States  was  dis-
continued in  1976 (Chem. Econ. Hdbk.,  1977).   However,  even prior
to 1976,  most HCB  was  produced as  a  waste by-product  during  the
manufacture of perchloroethylene,  carbon tetrachloride, trichloro-
ethylene,  and  other  chlorinated hydrocarbons.   This  is  still  the
major source  of  HCB  in  the  U.S.   In 1972,  an  estimated  2.2 x  106
kg of HCB were produced from these industrial processes (Mumma and
Lawless, 1975).  Its generation as  a  by-product remains unabated.
HCB found in  Louisiana  was  apparently related  to airborne  indus-
trial emissions, while residues in sheep from Texas and California
were traced to pesticide  contaminated  with HCB.   Until recently,
HCB was a major impurity  in the  herbicide dimethyl  tetrachloro-
terephthalate and the fungicide pentachloronitrobenzene.   HCB  has
been  found   in  polyethylene  plastic  bottles  from  one  source
(Rourke, et al.  1977).   HCB is used in  industry  as  a plasticizer
for polyvinyl chloride as well as  a flame retardant.
                             EXPOSURE
Ingestion from Water
     Very little is known regarding potential exposure to HCB as a
result of ingestion  of  contaminated water.  HCB has  been detected
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in specific  bodies  of water,  particularly  near points  of  indus-
trial discharge.  Except for  such  source-directed  sampling, there
is little  information of  HCB  concentrations  in  surface  waters.
HCB has  been found  in river  water and soil samples  collected in
the vicinity of an industrialized region bordering the Mississippi
River between Baton Rouge and New  Orleans,  Louisiana.   The levels
of HCB  in  the Mississippi  River  water samples  were  low,   usually
below 2  ug/kg.   Maximum  concentrations of HCB  were  found  in  sam-
ples  of  levee soil  collected  near Plaquemine   (400  ugAg)  and of
ditch mud  collected  near Darrow  (874  ug/kg).    Soil  on  the river
side  of  the  levee accumulated HCB from the  load carried in solu-
tion  and in suspension  in  the river water  (Laska,  et al.  1976).
High  concentrations  of HCB  were sporadically found in a newly dug
pond  near  a  landfill where wastes containing HCB  were buried and
in a  small stream carrying runoff water  from  a field adjacent to
an industrial plant.  The  HCB levels in the  landfill pond water
varied  from  4.8 to 74.9 ug/kg  and from 10,500  to 53,130 ug/kg in
mud  samples.  The HCB levels  in  the stream water varied from 0.1
to 72.8  ug/kg and from 2,520  to 13,800 ug/kg in mud  samples (Lase-
ter,  et  al.  1976) .
      Water samples  from western Lake  Superior  contained HCB;  the
exact concentration  was  not quantitively  measured.  Lake Superior
is  one of the  largest  and cleanest  oligotrophic  bodies of  fresh
water in the world.   The total population density around the  lake
is  low  and  the  concentrations of  trace elements  have  remained
relatively small  compared  to those in  other  Great  Lakes  (Veith, et
al.  1977).  HCB was  detected in  drinking water supplies at  three
                                C-95

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locations, at concentrations ranging  from  6  to 10  ng/kg.   HCB was



detected  in  finished  drinking  water at  two  locations,  at concen-



trations ranging from 4 to 6 ng/kg  (U.S. EPA, 1975).



     HCB has considerable potential to  bioaccumulate  in the aqua-



tic environment and is very  persistent.   The combination  of these



two attributes makes  HCB  a  potentially  hazardous  compound  in the



environment.  Soil contaminated with HCB would retain HCB for many



years.   If  contaminated  soil  finds its  way into  the  aquatic en-



vironment, it will become available to aquatic organisms.



     HCB  enters  the  environment  in  the waste  streams  from  the



manufacture of chlorinated hydrocarbons  and  from its  agricultural



use as  a  pre-emergence fungicide for  small  grains.   HCB becomes



redistributed throughout  the  environment as a  consequence  of its



leaching from industrial  waste dumps  and  its  volatilization from



industrial sources and contaminated impoundments.  HCB adsorbed to



soil may be transported long distances in streams and  rivers.  HCB



is now distributed throughout the world.   The solubility of HCB in



water is low, however, its concentration in water rarely exceeding



2 ug/kg.



     HCB is sufficently volatile  so that one air  drying  of moist



soil or biological samples  causes a 10  to 20 percent  loss  of HCB



(vapor  pressure  1.089xlO~5  mm Hg  at 20°C).    The half-life  of



HCB in  soil  (incorporated at 10 kg/ha)  stored  in  plastic-covered



plastic pots is about  4.2 years (Beck and Hansen,  1974).   HCB is



not lost from soil 2 to 4 cm beneath the surface during 19 months,



but 55  percent is  lost from the  surface 2 cm of  soil  within two



weeks  (Beall,  1976).    Clearly,  volatilization  is a  significant
                               C-96

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factor in  the  loss of HCB  from soil and  for  its entry  into the
atmosphere.   No HCB  is  lost  from  soil treated  with 0.1  to 100
mg/kg of HCB and stored under aerobic (sterile and nonsterile) and
anaerobic nonsterile conditions for one year in covered containers
(Isensee, et al. 1976).  Degradation products of HCB have not been
found in plants and soil.   Hexachlorobenzene  is  relatively resis-
tant  to  photochemical degradation  in water.   Photolysis  of HCB
occurs slowly  in methanol,  62 percent being degraded  in  15 days.
It  is not  known  whether   organic  matter  in  natural waters  or
natural photosensitizers in the environment can  enhance  the rate
of degradation  of HCB  (Plimmer  and  Klingebiel, 1976) .   HCB may be
even more stable than DDT or dieldrin in the environment (Freitag,
et al. 1974).  HCB has been singled out as the only organic chemi-
cal  contaminant present  in  the ocean  at  levels likely  to  cause
serious problems (National Academy of Sciences (NAS), 1975).
     HCB, adsorbed  to soil  or  sand, is  released into  water and
taken  up  by  aquatic  organisms  such as  algae,   snails,  daphnids
(Isensee, et al. 1976), and fish (Zitko and Hutzinger, 1976).  The
alga, Chara,  collected  from the lower  Mississippi  River (Louisi-
ana) contained HCB at 563 ug/kg wet weight.  An undefined plankton
sample contained 561 ug HCB/kg  (Laska, et al. 1976).
     The  aquatic  plants Najas  and  Ellocharios  contained  147  ug
HCB/kg and "423 ug HCB/kg wet weight, respectively (Laseter, et al.
1976).  Three aquatic invertebrate genera:  snail, Physa, crayfish
Procambarus, and dragonfly larvae,  Anisoptera,  also collected from
the  lower Mississippi River,  contained  294  ug/kg,  48.67  ug/g, and
4.7  ug/g, respectively (Laseter, et  al.  1976).   The  HCB levels in
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inland fish from  the United States  ranged  from "none detected" to
62 mg HCB/kg.  The  high  mean  level of HCB  in  carp (16 mg/kg) was
attributed  to runoff  from an  industrial  chemical  storage area.
The mean HCB  concentration  in  seven other  inland  fish ranged from
<1 to  130  ug/kg   (Johnson, et  al.  1974).   The HCB  level  in fish
collected  from the  contaminated lower  Mississippi  River  ranged
from 3.3 to 82.9  mg/kg for fish.   The HCB  levels in mosquitofish
collected some distance from the site of the HCB industrial source
on the  lower  Mississippi River  ranged  from 71.8  to 379.8  ug/kg,
about 100-fold lower than the HCB content in fish near the site of
industrial contamination (La.seter, et al. 1976).
     Marine invertebrates  collected  from   the  central  North Sea
contained substantially less HCB than  invertebrates  from the cen-
tral contaminated lower Mississippi River (Schaefer, et  al. 1976).
Residues of HCB were determined in 104 samples of marine organisms
collected at various sites off the Atlantic Coast of Canada during
1971 and 1972.  The results indicated a widespread, low-level dis-
tribution of HCB  (<1 to  20  ug  HCB/kg).   The highest levels of HCB
were in  fatty samples  (1  ug/kg  in whole cod  vs  39  ug/kg  in cod
liver;  none  detected  in  whole  lobster vs 54  ug/kg  in  lobster
hepatopancreas).   Herring contained the greatest whole body burden
of HCB  (20  ug/kg)  (Sims,  et al. 1977).  The HCB  levels in marine
fish from  the central  North Sea ranged  from 0.2  to 2.9 ug/kg for
muscle and  from  2.9 to 10 ug/kg for liver.   The  organ  concentra-
tions of HCB  increased with increasing  lipid  content of the  organ
(Schaefer, et  al. 1976).
                               C-98

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     HCB  has  been detected  in  a  number of water  and  land  birds.
Carcasses of  immature  ducks  contained  HCB  ranging from >60  to 240
ug/kg  (White and Kaiser, 1976).  The HCB levels ranged from  110 to
500 ug/kg in  carcasses of 4 of 37  bald eagles (Cromartie,  et al.
1975).   The HCB  levels  in  the eggs  of the  common  tern, Sterna,
ranged  from 1.35 to  14.7  mg/kg dry  weight  (Gilbertson  and Rey-
nolds,  1972).   Eggs of  double-crested cormorants, Phalacrocorax,
from the Bay  of  Fundy  were  monitored  from  1973 to 1975.   The eggs
contained 15 to  17 ug HCB/kg wet weight (Zitko, 1976).
     Foxes  and  wild boars,  which  feed on  small  animals  such as
mice and invertebrates, accumulated large amounts  of HCB.  Because
predators  and  scavengers   contain  higher  residues  of   HCB than
herbivores,  it  would  seem that biomagnification  through  the food
chain is occurring  (Koss and Manz, 1976).
Ingestion from Food
     Ingestion of excessive  amounts of HCB  has been a consequence
of carelessness, lack of concern,  and  ignorance.  There is a ten-
dency to dispose of excess wheat seed  by feeding it to stock with-
out due recognition of the  toxic properties of the compounds con-
cerned.  In the mid-1960's, a shipment of Australian powdered eggs
was rejected  for importation into  the United States  by  the Food
and Drug Administration on  the  grounds of  contamination  with HCB.
The New South Wales Egg Marketing  Board tests samples of eggs that
it handles  and  will  not  accept for  distribution any eggs  which
contain significant amounts of HCB.
                               C-99

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     Food materials were collected at retail and department stores
in Tokyo, Japan,  and  were  weighed in the amounts  consumed  a day.
The food materials were classified into four categories:  cereals,
vegetal  products  (vegetables,  vegetal  oils,  seasoning, and  sea-
weed), marine animal products,  and terrestrial animal products in-
cluding dairy products and eggs.  The dietary intake of HCB ranged
from 0.3  ug/day  to 0.8 ug/day.   Contributions from  cereals were
low (<0.05 ug/day).  The contribution from vegetal products ranged
from <0.05 ug/day  to  0.4  ug/day;  that for  marine  animal  products
from <0.05 ug/day  to  0.3  ug/day;  and that  for  terrestrial  animal
products from 0.3 ug/day to-0.4 ug/day (Ushio and Doguchi,  1977).
     Herds of  cattle  in  Louisiana  were  condemned  by the  State
Department of Agriculture in 1972 for excessive HCB residues, that
is, they exceeded 0.3 mg HCB/kg in fat.   Levels as high as  1.52 mg
HCB/kg were reported.   Of  555  animals tested among  157  herds,  29
percent  of  the  cattle  sampled  contained <0.5  mg HCB/kg  in fat.
HCB residues apparently did not arise  from agricultural  applica-
tion of  HCB  fungicide  but  from contamination  of air,  soil,  and
grass  by industrial sources (U.S.  EPA,  1976).   In a  total diet
study conducted in Italy between 1969 and 1974, the average intake
was estimated  to be  4.2  ug/person/day  (Leoni  and  D'Arca,  1976).
HCB contents  of  various foods  can be  found  in Table  1.    In  an
effort to reduce  the  amount of HCB entering  the  environment,  the
Federal  Republic  of Germany no longer allows  application  of HCB-
containing pesticides  (Geike and  Parashar,  1976).   The  New South
Wales  Department  of Health  (Australia)  has recommended that  the
concentration of  HCB  in eggs must  not  exceed  0.1  mg/kg  (Siyali,
                               0100

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                           TABLE 1

        Hexachlorobenzene Content of Food  (y.g HCB/kg)*


                     (Italy: 1969 - 1974)
Food
Bread
Noodles
Maize flour
Rice
Preserved legumes
Dry legumes
Fresh legumes
Fresh vegetables
and artichokes
Tomatoes
Potatoes
Onions
Carrots and other
root vegetables
Fresh fruit
Dried fruit
Exotic fruit
Citrus fruit
Bovine meat

Mutton, game
and rabbits
Giblets

Pork meat

Chicken

Eggs
Fresh fish
Preserved fish
Whole milk
Butter
Cheese

Olive oil
Seed oil
Lard

Wine
Beer
Sugar
Coffee
Mean
1.1
0.7
n.d.
0.8
1.1
2.4
n.d.
0.5

n.d.
n.d.
0.6
n.d.

n.d.
n.d.
n.d.
n.d.
0.7
(33.6)
1.0
(25.4)
0.7
(27.0)
25.0
(96.3)
5.7
(49.0)
4.7
0.7
n.d.
4.1
133.0
12.6
(63.0)
13.1
4.7
46.2
63.4
0.1
n.d.
0.2
n.d.

n.d.
0.2

0.3
n.d.
0.2

n.d.



0.6






n.d.

n.d.

n.d.

9.1
(74.3)
n.d.

1.7
n.d.

0.2

n.d.

n.d.
n.d.


n.d.

n.d.

Range
(a) _ 2.9
2.9

1.1
3.1
5.1

1.8

-
—
0.6
-

-
—
-
—
1.4
- (78.4)
2.6
- (51.3)
1.3
- (53.9)
- 40.9
-(118.3)
- 11.5
- (75.0)
7.5
1.8

- 17.2
—
- 25.1
-(126.0)
- 53.8
- 27.9
—

0.6
-
0.6
"•
Values in parentheses are for extracted fat.
(a)n.d. — not detected
*Source: Adapted from Leoni and D'Arca, 1976.


                             C-101

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1973).  The National  Health and Medical Research  Council  (NHMRC)
(Australia) has set the  tolerance for  cows' milk  at  0.3  mg HCB/kg
in fat (Miller and Fox,  1973).   The  Louisiana  Department of  Agri-
culture has  set  the  tolerance  for  meat -at 0.3  mg HCB/kg  in fat
(U.S. EPA, 1976).
     There is a  substantial body of information on HCB  levels in
human milk for a number of countries.  In the  United States,  human
milk  contained  a mean concentration  of 78  ppb  (Savage,  1976).
Milk  from  45  women living  in a  metropolitan area  (Sydney,  Aus-
tralia) was found to  contain HCB.  The  mean HCB  concentration in
human milk was 15.6 ug/kg,  and  7  percent of  the  samples  contained
51  to 100 ug  HCB/kg.   In  addition,  49  human milk  samples  from
France and 50  from  the Netherlands  contained  HCB, but  no  concen-
trations were reported.  Human milk samples from Germany contained
153 ug HCB/kg of whole milk and those from Sweden 1 ug/kg (Siyali,
1973).  HCB was also detected in all of 40 human milk samples from
Brisbane,  Australia,  and a  rural area  (Mareeba  on  the Atherton
Tablelands).  The excretion of HCB  into human milk was  higher in
Brisbane  samples than  in Mareeba  samples  (2.22  versus  1.23  mg
HCB/kg in milk fat).   The higher levels  of  HCB in Brisbane donors
may  be  related to the close proximity  to  a  major  grain  growing
area,  the  Darling  Downs.   The  daily intake of HCB  by  infants in
Brisbane was  estimated to  be  39.5  ug per  day per  4 kg  of  body
weight and in Mareeba  to  be 14 ug per day per 4 kg of body weight.
The calculated average daily intake of HCB by breast-fed babies in
both  areas exceeded the  acceptable  daily intakes  of  2.4  ug/kg/day
recommended by the Food  and Agriculture  Organization/World Health
                                C-102

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Organization  (FAO/WHO)  (1974).  The HCB content  of  human milk  also
exceeded  the  Australian NHMRC tolerance  for cows' milk (0.3 mg/kg
in milk  fat).  The dietary  intake  by young adults  (15-to-18-year
old males) was  estimated to be 35 ug  HCB per person per day (Mil-
ler and Fox,  1973).  Similarly, HCB was found  in  all of 50  samples
of human  breast milk  collected in Norway.  The mean HCB level was
9.7 ug/kg, with a maximum value of 60.5 ug/kg.  The HCB content  of
colostrum  (7.7  ug/kg)  was within the range  of  that for milk 1  to
16 weeks  after  birth  (5.9  to 10.0 ug/kg).  The HCB content of the
human milk samples  in  this  survey exceeded the maximum concentra-
tion  of  20 ug/kg  for  cows'  milk  approved by FAO/WHO.   The milk
sample with the  highest  HCB  level exceeded  this standard by three-
fold  (Bakken  and Seip, 1976) .
     A bioconcentration  factor  (BCF)  relates the concentration  of
a chemical in aquatic animals to  the  concentration  in the water  in
which they live.  The  steady-state  BCF   for  a  lipid-soluble com-
pound in  the  tissues  of various aquatic  animals  seems  to  be pro-
portional  to  the  percent lipid  in   the  tissue.    Thus,   the  per
capita ingestion of a lipid-soluble chemical can  be estimated from
the per  capita  consumption  of fish  and  shellfish,  the  weighted
average  percent lipids  of  consumed  fish  and  shellfish,  and  a
steady-state  BCF for the chemical.
     Data from  a recent survey on fish  and shellfish  consumption
in the United States was analyzed  by  SRI  International  (U.S.  EPA,
1980a).   These data were used to estimate that the per capita con-
sumption  of  freshwater  and  estuarine fish  and  shellfish  in  the
United States  is 6.5  g/day  (Stephan, 1980).   In addition,  these
                               C-103

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data were used with data on  the  fat  content  of  the edible portion
of the same species to estimate  that  the  weighted  average percent
lipids for consumed freshwater and estuarine fish and shellfish is
3.0 percent.
     A measured steady-state bioconcentration factor of 22,000 was
obtained  for  hexachlorobenzene  using  fathead minnows  (U.S.  EPA,
1980b).   These  fathead  minnows probably  contained  about  7.6 per-
cent  lipids (Veith,  1980).    An adjustment  factor  of  3.0/7.6 -
0.395 can be used  to  adjust  the  measured  BCF from the 4.8 percent
lipids  of  the  bluegill  to  the  3.0  percent lipids  that  is the
weighted  average bioconcentration factor  for hexachlorobenzene and
the edible  portion of all freshwater  and estuarine aquatic  orga-
nisms  consumed  by  Americans is  calculated  to be 22,000 x 0.395 •
8,690.
Inhalation  and  Dermal
      HCB  enters the air  by various mechanisms such  as  release from
stacks  and  vents  of  industrial  plants,  volatilization from  waste
dumps and impoundments,  intentional  spraying and dusting, and un-
intentional dispersion  of HCB-laden  dust  from manufacturing  sites,
during transport of finished material or wastes, and  by  wind from
sites where HCB has been applied.  Plasma HCB concentrations of  86
 individuals living  in  Louisiana  adjacent  to   a  plant  producing
chlorinated solvents, but not occupationally exposed,  averaged 3.6
ug/kg with a maximum of 23  ug/kg.   Plasma HCB  concentrations were
 higher in  males  than in  females (4.71  ug/kg  compared  with 2.79
 ug/kg,  respectively),   but  there was no  significant  difference
                                C-104

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 between age groups.  There was  no  evidence  of  cutaneous porphyria
 in this population, but persons with high plasma concentrations of
 HCB  showed  elevated   coproporphyrin   and   lactic  dehydrogenase
 levels.  Only two of 48 household meals sampled contained signifi-
 cant quantities  of HCB,  but  there was  some  correlation  between
 concentration in plasma and the concentration  of HCB  in household
 dust.    Some  household  dust   contained  as  much  as  3.0  mg/kg.
 Affected households were  on the  route  of a  truck  which regularly
 carried residues containing HCB from a  factory  to a dump.   Workers
 in the adjacent plant  engaged  in manufacturing  carbon  tetrachlo-
 ride and perchloroethylene had  plasma  HCB  concentrations  from  14
 to 233  ug/kg  (Burns and Miller,  1975).
     Pest  control  operators  in  their  day-to-day  work  handle  a
 variety of  toxic  chemicals,   including chlorinated  hydrocarbon
 pesticides.   Pesticides may enter the body  by  inhalation of  spray
 mist which  exists  in confined  spaces.   The  levels  of HCB in  blood
 of  pest control  operators  in  New  South Wales,  Australia,  were
 found to be elevated in a 1970-1971 study  (1 to 226 ug/kg).  The
 pest control  operators  seldom  used  respirators,  and  those in use
 appeared to be  ineffective due to poor service maintenance.   The
 respiratory exposure values were many-fold higher than the  accept-
 able daily  intake  as applied to  food  by WHO  (0.1  ug/kg/day or 7
 ug/day  intake for a 70 kg man)   (Simpson and Shandar, 1972).
     HCB may enter  the  body by  absorption through the  intact skin
as a result of skin contamination.  Workers  involved in  the appli-
cation  or manufacture  of HCB-containing  products  are,   therefore,
at greater risk.
                               C-105

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     HCB enters the body  as  a result of ingestion  and  presumably
by inhalation  and  absorption  through  the  skin.   HCB  remains  in
the blood  for  only a  short  period before  it  is  translocated  to
fatty tissues  or  is excreted.   HCB  blood  levels  reflect  either
recent exposure or mobilization of HCB  from  body  fat  depots.   HCB
finds its  way  into air, water, and  food as a result  of  uninten-
tional escape  from industrial sites, intended application  of HCB
containing products, volatilization from waste disposal sites and
impoundments,  and  Unintentional dispersion  during transport and
storage.   The  result  has been the worldwide dissemination  of HCB
and ubiquity in man's food, at least in  low levels.
     All blood samples  taken  from  children  (1  to 18 years old)  in
upper Bavaria  in 1975  contained  HCB at 2.6  to 77.9 ug/kg.   The
study  included 90  males and  96  females.   HCB  levels  in   blood
showed a positive,  hyperbolic correlation with  age, tending  to an
upper limit  of 22 ug/kg  for boys and 17  ug/kg for girls.  The rate
of  increase  in HCB concentration was  inversely  proportional  to  a
function of  age.  A substantial accumulation of HCB became evident
9  to 10 months after  birth  (Richter and Schmid,  1976).   HCB was
found  in  all  of  a  series  of human  fat  samples  collected  from
autopsy  material  throughout  Germany.   The  highest  levels  of HCB
were  in specimens from  Munster  (22  mg  HCB/kg in  fat)  and Munich
 (21  mg  HCB/kg  in fat)  (Acker and Schulte,  1974).   The  presence  of
HCB  in  Japanese  adipose tissue obtained at autopsy was determined
 for  a  total  of  241  samples from  Aichi  Cancer  Center  Research
                                C-106

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 Institute, Chikusa-Ka Nagoya, Japan.  The  concentration  of HCB in
 these fat samples was  90  ug/kg  + 6 ug/kg  standard  error  (Curley,
 et al. 1973).
      HCB was  found in all  of 75  specimens of Australian human body
 fat (1.25 mg/kg).  Perirenal fat was.taken  at  autopsy from a ran-
 dom selection  of bodies  at  the City  Morgue,  Sydney,  Australia.
 All ages  and  both sexes  were  included  in the  study (Brady  and
 Siyali,  1972).   The  incidence  (63  percent  of samples  tested)  and
 concentration  of HCB  (0.26  mg/kg)  in 38 specimens  of human  body
 fat from  Papua  and  New  Guinea were  lower than  the Australian
 values.   The concentration of HCB  in  whole  blood  of  185 people  who
 had some  occupational  exposure to  organochlorine  compounds   in
 their  working  conditions and of 52 who  had no known  exposure  was
 determined.  None of  the subjects displayed apparent  signs  of  in-
 toxication.   Over  95 percent of  the subjects  had HCB  in  their
 blood.   The HCB  blood level  in the  exposed population  was 55.5
 ug/kg, with 9  percent having more  than  100  ug/kg.   The HCB blood
 level  in  the population with  no  known exposure was 22 ug/kg, with
 none having as much as  100 ug/kg.  Levels of 50 to 100  ug/kg whole
 blood  indicate either recent  exposure over and above that normally
 assimilated from the environment or the mobilization of fat depots
 associated with a  loss  in  total  body weight.   The  mean HCB level
 in 81  samples of human  body  fat  was 1.31  mg/kg,  with a maximum of
 8.2  mg/kg.   All  81  human   fat  samples contained  HCB  (Siyali,
1972).
                               C-107

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     The HCB levels  in  adipose  tissue of Canadians,  collected in
1972 by  Burns  and Miller  (1975),  were  determined.   The  regional
distribution of the  samples  was  as follows:  16  from the eastern
region  (Newfoundland,  Prince Edward  Island,  Nova Scotia  and  New
Brunswick), 50  from  Quebec,  57  from Ontario, 22  from the central
region  (Manitoba  and Saskatchewan) and  27  from  the western region
(Alberta and British Columbia).   All of the  adipose  samples  con-
tained  HCB, with  an overall mean  value of 62 ug/kg.   HCB values
were lowest in  the samples from the eastern (25 ug/kg) and central
(15 ug/kg) regions and  highest in Quebec (107 ug/kg).  The Ontario
samples  averaged  60  ug  HCB/kg and  those from  the western  region 43
ug/kg.   The HCB content of adipose tissue  from females (82 ug/kg)
was  greater  than  that  for males  (52 ug/kg).  The HCB content of
human  adipose  tissue did not show an age-related  trend:   0 to 25
years,  76  ug/kg;  26  to  50  years, 45  ug/kg;  and 51+ years,  70 ug/kg
 (Mes,  et al.  1977).   In the study of Richter and  Schmid,  the  age-
related accumulation of  HCB was  marked  only for  the first  five
years  of life  (Richter and Schmid, 1976).   Plasma HCB levels  in  a
Louisiana population exposed through the transport and disposal  of
chemical waste containing HCB averaged  3.6 ug/kg  in  a study of  86
 subjects.   The highest  level  was 345 ug/kg in  a  sample from  a
waste  disposal worker, while the highest level in a  sample  from  a
 member of the  general  population  was 23  ug/kg (Burns and Miller,
 1975).
                          PHARMACOKINETICS
 Absorption
      To date,  only  absorption of HCB  from  the  gut  has  been ex-
 amined  in detail.   Fish  fed  HCB-contaminated  food take  up the

                                C-108

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 material  in a reasonably direct relationship to  the  concentration
 in the food (Sanborn,  et al. 1977).  Intestinal  absorption  of  HCB
 from an aqueous  suspension was poor  in both  rabbits (Parke  and
 Williams,  1960)  and  rats (Koss  and  Koransky,  1975).   The  amount of
 HCB  left  in the  intestinal contents 24 hours after administration
 was  small.  Intestinal absorption  of HCB  by rats was  substantial
 when the  chemical was  given in cotton seed oil (Albro  and Thomas,
 1974)  or  olive  oil (Koss and Koransky,  1975).  Between 70 percent
 and  80  percent of doses of HCB ranging from 12 mg/kg  to  180 mg/kg
 were absorbed.   The  fact that HCB  is well absorbed when  dissolved
 in oil  is  of particular relevance  for man.   HCB  in  food products
 will selectively  partition  into the  lipid  portion, and  HCB   in
 lipids will  be absorbed  far more efficiently  than that  in an aque-
 ous  media.   This  is consistent with the observation that  the high-
 est  HCB levels ever  observed have  been'  in  tissues  of  carnivorous
 animals (Acker and Schulte,  .1971;  Koeman, 1972).    HCB is readily
 absorbed from the  abdominal cavity after intraperitoneal  injection
 of the chemical dissolved in oil.
     Data  of toxicological experiments  should  take  into account
 how  HCB was  administered.   Relatively little HCB was absorbed  by
 the  walls  of  the  stomach and  duodenum  of  rats one hour after oral
administration of HCB suspended in aqueous methylcellulose.   After
 three hours,  the  ingested  HCB  reached the jejunum  and  ileum,  re-
sulting in increasing  concentrations  in  the walls  of these  parts
of the intestine.  Liver  and kidney contained  some HCB;  however,
the  concentrations in  lymph  nodes  and adipose  tissue were  much
                              C-109

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higher.   During  the  remaining  45  hours,  the concentrations  in
liver and kidney decreased, whereas those  in  lymph  nodes  and adi-
pose tissue remained relatively constant or rose slightly.  Portal
venous  transport  to the liver  seemed to  be  a minor  pathway be-
cause,  in  spite of its slow  metabolism,  HCB  never  achieved high
concentrations  in the liver.  The majority of  the ingested HCB was
absorbed by the lymphatic system in the region of the duodenum and
jejuno-ileum, and deposited in  fat, bypassing the systemic circu-
lation  and  excretory  organs.   There  appears  to  be  an equilibrium
between lymph nodes and fat (latropoulos,  et  al. 1975).
Distribution
      It  is  well known that HCB has a low  solubility in  water  (6
ug/kg)  (Lu  and  Metcalf, 1975) and a high solubility  in fat  (calcu-
lated log  partition  coefficient  in  octanol/H20=6.43).    Accord-
ingly,  the  highest  concentrations  of  HCB are  in fat  tissue  (Lu  and
Metcalf,  1975).  The  concentration  of HCB  in  fish fed contaminated
food  (100 mg/kg)  for three days was  4.99  mg/kg  in  liver and  1.53
mg/kg in muscle (Sanborn,  et al. 1977).   The  concentration  of  HCB
in Japanese quail fed contaminated food (5 mg/kg) for 90 days  was
6.88  mg HCB/kg in  liver  and 0.99 mg/kg  in brain of  female  birds
and 8.56 mg/kg in liver  and  1.44 mg/kg  in  brain  of male  birds
 (Vos, et al.  1971).  As  noted above,  HCB  accumulated in  fatty tis-
 sues.   After  prolonged  feeding of  a constant  level of  HCB,  the
 concentration  of  compound in  the fat  of laying  hens   reached  a
 plateau.   This indicates that  an  equilibrium between uptake  and
 excretion can  be  achieved.   This  phenomenon  allows  one  to calcu-
 late the ratio of  the  concentration  of HCB  in fat  to the  concen-
                                 C-11C

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 tration in the  feed.   This  accumulation or  storage  ratio apparent-
 ly is  independent  of  HCB concentration  in the feed  over a  wide
 range.   The accumulation ratio for HCB in  laying hens  is  about  20
 (Kan  and Tuinstra,  1976).
      The distribution  of HCB in  rat  tissues  was similar  for  ani-
 mals  given a single oral dose  or a single intraperitoneal  injec-
 tion  of  HCB dissolved  in  olive  oil.    Adipose tissue  contained
 about  120-fold,  liver,  4-fold;   brain,   2.5-fold;  and   kidney,
 1.5-fold more HCB  than muscle.   The HCB content of adrenals,  ova-
 ries  and the Harderian  gland  was  essentially  the same as skin,
 whereas that for heart, lungs,  and  intestinal  wall  corresponded  to
 the  level in liver.   The thymus  content was  similar  to that  of
 brain (Koss  and Koransky, 1975).
      The distribution  of  HCB  in mice  fed a diet containing 167  mg
 HCB/kg  was determined  after three and six weeks.  The HCB  level  in
 the serum was  23 mg/kg after  three weeks  and  12 mg/kg after six
 weeks;  for liver, 68.9 mg/kg after  three weeks  and  56 mg/kg at six
 weeks;  for  spleen,  20.9  mg/kg at three weeks  and 47  mg/kg at six
 weeks;  for  lung,  85.1  mg/kg  at three  weeks and 269  mg/kg at six
 weeks;  and for the thymus, 48.6 mg/kg at three weeks and 152 mg/kg
 at six  weeks.   The  only histological alterations seen  in tissues
 of mice  fed  HCB for six weeks was a centrilobular  and pericentral
 hepatic   parenchymal   cell  hypertrophy;   hepatic   Kupffer  cells
 appeared  normal in number and morphology (Loose, et  al.  1978).
     Adipose tissue serves as a reservoir  for  HCB, and depletion
 of fat depots results  in mobilization and redistribution of stored
pesticide.   For  example,  food  restriction  caused mobilization  of
                               C-lll

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HCB stored within  the  fat depots  of  rats  that had been  fed  HCB-
contaminated food  for  14  days.    Although  HCB was  redistributed
into the plasma  and other tissues  of the body,  food  restriction
did not increase the excretion of HCB; therefore,  the  total  body
burden was not reduced.   Rats receiving 100 mg HCB/kg/day orally
for 14 days developed tremors, lost appetite, and  some died during
subsequent  food  restriction.   Weight  loss  from whatever  cause
results in redistribution  of  HCB  contained  in  adipose  tissue,  and
if the initial level of  the  pesticide is  sufficiently  high,  toxic
manifestations may develop (Villeneuve, 1975).
Metabolism
     Although HCB appears  to  be relatively  stable in  the  soil, it
is metabolized by a variety  of animal species.  About half of HCB
taken  into the body of  fish  fed contaminated food  is converted
into pentachlorophenol  (Sanborn,  et  al.  1977).   The  rabbit  does
not  appear to  oxidize  HCB  to pentachlorophenol  (Kohli, et  al.
1976).  In rats  given HCB  intraperitoneally  on two or three  occa-
sions  (total dose 260 to 390 mg HCB/kg), pentachlorophenol, tetra-
chlorohydroquinone,  and   pentachlorothiophenol   were   the  major
metabolites  in  urine.   More  than 90 percent  of  the  radiolabeled
HCB material  in  the urine had been  metabolized,  whereas  only 30
percent of the starting radiolabeled  HCB material in the feces was
metabolized.   Of the HCB  administered  intraperitoneally,  65  per-
cent was  in  the  animal body  (almost  all  as  HCB), 6.5 percent was
excreted in the urine (mostly  as  metabolites) and 27.2 percent was
excreted in  the  feces  (about 70 percent as HCB).   The metabolites
                                C-112

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 in feces  were  (in  decreasing  order)  pentachlorophenol  >  penta-
 chlorothiophenol  >  an  unidentified  substance  (Koss,  et  al.  1976).
      In  organs  of  rats  given  8 mg  HCB/kg  dissolved in  sunflower
 oil by gavage,  only HCB,  pentachlorobenzene,  and  pentachlorophenol
 could  be  identified.  The metabolites  were  present  in small  con-
 centrations.   The  HCB  level in fat was 83  mg/kg,  in muscle,  17
 mg/kg; in  liver,  125 ug total;  in kidneys, 21 ug each; in  spleen,
 9  ug  total;  in  heart,  1.5 ug total  and in adrenals, 0.5 ug each.
 In urine,  the  main metabolites of  orally administered  HCB  were
 pentachlorophenol,  tetrachlorophenol,  trichlorophenol,  and  penta-
 chlorobenzene.  Small  amounts  of trichlorophenol and tetrachloro-
 phenol were  present as  glucuronide conjugates.   The  feces  con-
 tained  a  little  pentachlorobenzene,  but  mostly  the  parent  HCB
 (Engst, et al. 1976) .
     HCB in  corn  oil given  orally  to  rats at a  dose  of  20 mg/kg
 for 14 days  caused  an  elevation of  the levels of cytochrome P-450
 and NADPH-cytochrome c  reductase activity.  HCB  appears  to be an
 inducer of the hepatic microsomal system of the phenobarbital  type
 (Carlson, 1978).  In a  separate study,  the cytochrome P-450 level
 was elevated  in rats (Porton strain) fed HCB mixed  into  the diet
 (dose about  19  mg/kg)  for 14 days,   but  not  in  rats  (Agus strain)
 fed food containing HCB  for  90  days.   In both HCB-exposed groups,
 benzo(a)pyrene  hydroxylation activity  was  elevated,   but  amino-
pyrine N-demethylatse activity was  not significantly  enhanced.   It
has been proposed that  KCB  is  an  inducer of  hepatic  microsomal
enzyme activity having  properties  of both the  phenobarbital type
and the  3-methylcholanthrene  type   (Stonard,   1975;  Stonard  and
                               C-113

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Greig,  1976).    Although  HCB  is   a  well-documented  inducer  of



hepatic  microsomal  enzyme   activity,   the  hexobarbital  sleeping



times of rats fed 2,000 mg HCB/kg/day for 14 days were' the same as



unexposed  control   rats.    The  duration of  hexobarbital-induced



sleep decreased  14  days  after eliminating HCB from  the  diet.   In



rats fed 500 mg HCB/kg/day for 14 days, hepatic glucose-6-phospha-



tase  activity  was  decreased  and  serum  isocitrate  dehydrogenase



activity remained undetectable.  In rats fed 10 mg HCB/kg/ day for



14 days,  the liver  was enlarged; the cytochrome  P-450  level,  de-



toxification  of EPN,  0-ethyl 0-(p-nitrophenyl)  phenylphosphono-



thioate, benzpyrene hydroxylase  activity and azoreductase activity



were  increased, whereas  cytochrome  c  reductase and  glucuronyl



transferase activities were  unaltered.



Excretion



     As  described  in  earlier sections,  HCB  is  excreted  mainly in



the  feces  and  to some extent  in the  urine  in  the form of several



metabolites  that are more polar  than the parent HCB.   Usually  a



plateau  is reached  in  most tissues  when  the dose  is  held  relative-



ly  constant.    If  exposure  increases  or decreases,  however,  the



body concentration  will increase or decrease, accordingly.



      Fish  fed  HCB  contaminated  food  (100  mg/kg) for  three  days



have relatively high  levels  of HCB and pentachlorophenol in their



stomach  (27.16  mg/kg  and  19.14 mg/kg, respectively)  and  intestine



 (26.82  mg/kg and  15.94  mg/kg,  respectively)  by the  fourth  day.



The  half-life  of HCB  in  the stomach,  intestine,  and muscle was  8



 to  8.5  days,  in the carcass  10  days,  and in the liver  19.6 days.



During  the initial elimination  period,  the  clearance  of HCB  from
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 the  intestine  and muscle lagged  behind  that for  the  stomach and
 liver,  and  may indicate biliary  excretion  with enterohepatic re-
 circulation  (Sanborn,  et al.  1977),  which has  been described  in
 dogs  (Sundlof, et al. 1976).
     HCB accumulates  in the eggs of  laying  hens fed contaminated
 food.   The  accumulation ratio (level  of  HCB in whole egg/level  in
 the  feed) was  1.3.    The  actual HCB concentration  in  eggs  was  20
 ug/kg for hens fed 10 ug HCB/kg of feed and 140 ug/kg for hens fed
 100  ug  HCB/kg.   Although  the concentration  of  HCB  in eggs  is
 usually  viewed from  the  perspective  of  accumulation  in a human
 food, it can also be regarded  as an excretion process.  Whereas  10
 percent of  the daily  HCB  intake is  excreted in the feces, 35 per-
 cent  is excreted in  the  eggs  of laying  hens  (Kan  and  Tuinstra,
 1976).  The rate  of elimination of  HCB  from swine was greatest  48
 to  72  hours after a  single  intravenous  injection  of drug.   The
 rate of release  of  HCB  from  fat was  the  rate  limiting  factor for
 excretion at later  times.   Half  of  the  starting HCB  material  in
 the feces was unmetabolized HCB.  All of the HCB material excreted
 in the urine were metabolites  of HCB.   Excretion of HCB from swine
 was 5-fold  to  10-fold slower  than excretion  from dogs  (Wilson and
 Hansen,  1976) .
     Clearance  of HCB from  brain  of rats  given  a single  injection
 intraperitoneally occurs in two steps:  a  slow  phase on  days 1  to
 14, and a very slow phase thereafter.  The  half-life for the slow
phase was 10 days and that  for the very  slow phase was  57 days.
Similarly,   the  half-life of  HCB  in  testes  was  15  days for  the
 initial slow clearance and  62  days  for the  later  very  slow  phase.
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The initial clearance rates  (half-lives)  for the heart,  lung and



kidney were 15, 13, and 16 days  respectively.   In contrast to the



pattern for individual organs, the clearance of HCB from the whole



body proceeded  as  a single step  process,  with  a half-life  of  60



days.   The initial  clearance  of  HGB  from  individual  organs there-



fore reflects a redistribution  of the chemical  among  the tissues



of  the  body  (Morita  and Oishi,  1975).   Clearance  of  HCB  from



organs of  rats  given  a  single dose of HCB dissolved  in olive oil



by gavage  also  occurred in two  stages:  a very  slow  phase between



days two  and  five, or  eight, and a  slow phase  thereafter.   The



overall half-life  of HCB  for  fat,  skin,  liver, brain,  kidney,



blood, and muscle was 8 to 10 days.  The administered chemical was



retained  in  the  tissue as  unaltered  HCB.    During  a  two  week



period, 5  percent  of  the administered  HCB was  excreted  in the



urine; essentially  all  as  metabolites  of  HCB,  and 34  percent was



excreted in the feces, mostly as  unaltered HCB.   The  fecal excre-



tion of a  fairly  high amount of  unmetabolized HCB is  presumed  to



be due to  biliary  secretion.   Unchanged HCB has  been  detected  in



bile of rats after  intraperitoneal administration of  the chemical



(Koss and  Koransky, 1975).



     No radioactivity  was detected  in  the  expired  air  of  rats



administered radiolabeled HCB (Koss and Koransky, 1975).



                             EFFECTS



Acute, Subacute, and Chronic Toxicity



     Japanese quail are  among the most sensitive  species to HCB.



Japanese  quail  fed a  diet  containing  5  mg HCB/kg  for  90  days



developed  enlarged  livers,  had  slight  liver damage  and excreted
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 increased  amounts  of coproporphyrin in  the  feces.   Increased  ex-
 cretion  of coproporphyrin was  noticeable after  10  days (Vos,  et
 al.  1971).
     The acute  toxicity  of HCB  for vertebrates  is low:   500 mg/ kg
 intrapertioneally  is not lethal in rats;  the  lethal oral dose  in
 guinea pigs  is greater  than  3  g/kg;  and  the  lethal oral dose  in
 Japanese  quail is  greater  than 1  g/kg  (Vos,  et al.  1971) .    In
 acute studies,  HCB was more toxic for  guinea  pigs  than rats,  but
 accumulated  to a  lesser degree  in  the  guinea  pig.    Male  rats
 appeared  to  be more  susceptible  to HCB  than  females  (Villeneuve
 and  Newsome,  1975).  HCB is  able  to induce  rat  microsomal liver
 enzymes; HCB was more effective in stimulating aniline  hydroxylase
 than aminopyrine demethylase  or hexobarbital  oxidase.   HCB is  not
 a particularly  effective  inducer of  these microsomal enzymes (den
 Tonkelaar and van Esch, 1974).  Although HCB has a low  acute toxi-
 city for most  species  (>1,000 mg/kg),  it has  a wide range of bio-
 logical effects at prolonged moderate exposure.
     Subacute  toxic effects  of HCB  were  examined   in  rats  after
 feeding with  HCB  for  15 weeks.   Histopathological  changes  were
 confined to the liver and spleen.  In  the liver,  there was an  in-
 crease in  the  severity  of  centrilobular liver  lesions with  as
 little as 2 mg HCB/kg/day in the food.   In contrast  to  the results
 of others, females  were  more susceptible  to  HCB than  male  rats.
 It would appear that  0.5  mg  HCB/kg  of  body weight per  day  is  the
 no-effect level in  the rat (Kuiper-Goodman, et  al. 1977).   Unlike
 in the rat, it  was  not possible to induce porphyria in dogs  with
HCB (Gralla,  et al.  1977).   Swine are  more susceptible to HCB  in
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subacute studies than rats.  Liver microsomal enzymes were induced
in swine and excretion  of  coproporphyrin was increased by  0.5  mg
HCB/kg/day after  13  and 8  weeks,  respectively.   It  would  appear
that 0.05 mg HCB/kg/day in the diet is  the  "no-effect"  level  for
swine (den Tonkelaar, et al. 1978).
     In rats given 50 mg HCB/kg every other day  for  53  weeks,  an
equilibrium between intake  and elimination was achieved after nine
weeks.  In general,  the changes  observed in  the  long term studies
resembled those described for short term studies.  When the admin-
istration of HCB  was discontinued, elimination  of  the xenobiotic
continued slowly for many months (Koss, et al. 1978).
     HCB caused a serious  outbreak  of  hepatic  porphyria  in Turkey
involving cutanea  tarda lesions and porphyrinuria  (Cam  and Nigo-
gosyan, 1963).  This has been  confirmed  in  a number of laboratory
animals including rats  (San Martin de Viale, et al.  1976), rabbits
(Ivanov, et  al.  1976),  Japanese quail (Vos, et  al.  1971),  guinea
pigs  (Strik,  1973),  swine  (den  Tonkelaar,  et  al.   1978),  mice
(Strik, 1973) and Rhesus monkeys (latropoulos, et al. 1976).  Rats
given 50 mg HCB/kg orally for 30 days showed enlarged livers, ele-
vated  liver  porphyrin   and elevated  urine  porphyrin   (Carlson,
1977).  In  both  rabbits and rats,  HCB produced an increase in the
excretion  of uroporphyrin  and  coproporphyrin.    The  mechanism  of
action  of  HCB  is   not known,  but  it  elicits an   increase  in
O -aminolevulinic  acid  synthetase,  which   is   the   rate-limiting
enzyme  in the  biosynthesis  of  porphyrins  (Timme,  et  al.  1974).
The  development  of HCB-induced  porphyria is  accompanied  by a pro-
gressive  fall  in hepatic uroporphyrinogen decarboxylase  activity.
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This change may be causally related  to  the  disease  (Elder,  et al.
1976).  The mitochondrial membrane may  also be  a  factor  in  limit-
ing the rate of porphyrin biosynthesis since some critical enzymes
are intramitochondrial  and  others are  cytoplasmic.   It  has  been
proposed that  HCB  may damage the mitochondrial  membrane,  thereby
facilitating the flow of porphyrin  intermediates through  it  (Si-
mon, et al. 1976).  Consistent  with  this  proposal is the observa-
tion that HCB  causes  marked enlargement  of  rat  hepatocytes,  pro-
liferation  of  smooth  endoplasmic reticulum, formation  of  eosino-
philic  bodies,  generation  of  large  lipid vesicles,  and  mitochon-
drial swelling  (Mollenhauer, et al.  1975).
     It  should be noted that  the   principal  metabolite of  HCB,
pentachlorophenol,  is not porphyrinogenic in the  rat,  so the  for-
mation  of  this metabolite  is unlikely  to play a role  in  HCB-in-
duced porphyria (Lui, et al.  1976).   Nevertheless,  it  is conceiv-
able that  other metabolites of  HCB,  particularly as a  result  of
microsomal  enzyme  induction,  might   be  the actual  porphyrogenic
agent (Lissner, et al. 1975).
     An epidemic of  HCB-induced cutanea  tarda  porphyria occurred
in  Turkey  during  the period  1955   to  1959 (Cam and  Nigogosyan,
1963).   More  than 600  patients were  observed  during  a  5-year
period, and  it was estimated  that   a total of 3,000  people  were
affected.   The  outbreak  was  traced to the consumption  of wheat as
food  after  it  had been prepared for planting by  treatment  with
hexachlorobenzene.   The syndrome involves  blistering  and  epider-
molysis of  the exposed parts  of the body,  particularly  the  face
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and hands.  It was  estimated  that  the  subjects  ingested 50 to 200



mg HCB/day for a relatively long period before the skin manifesta-



tions became apparent.   The symptoms were seen  mostly  during the



summer months, having  been exacerbated by intense  sunlight.   The



disease subsided and symptoms disappeared 20 to 30 days after dis-



continuation of  intake of HCB-contaminated bread.   Relapses were



often seen, either because the subjects were eating HCB-containing



wheat again,  or  because of redistribution  of HCB  stored  in body



fat.



     A disorder called pembe yara was described in infants of Tur-



kish  mothers  who either  had  HCB-induced porphyria or  had  eaten



HCB-contaminated bread  (Cam,  1960).   The maternal  milk contained



HCB.  At least 95 percent  of  these infants  died within a year and



in many villages, there were  no  children  left between the ages of



two  and  five  during  the  period  1955-1960.    With  human  tissue



levels  of  HCB  increasing  measurably  throughout  the  world,  the



effect of  low chronic  doses of  this  pesticide must  be considered.



HCB  is  stored in  the  body fat  and  transmitted  through maternal



milk.  It  is not known whether HCB is responsible for genetic dam-



age to the progeny  (Peters, 1976).



     There was no evidence  of cutaneous porphyria in 86 Louisiana



residents  having an  average plasma HCB  level  of 3.6 ug/kg, with a



maximum level of 345 ug HCB/kg.   There  was  a possible correlation



between plasma HCB  levels  and urinary  coproporphyrin excretion or



plasma lactate dehydrogenase  activity but none  with urinary uro-



porphyrin  excretion  (Burns  and Miller,  1975).   It should be noted



that  the  people  in Turkey  showing symptoms  of  porphyria  had in-
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gested 1 to 4 mg  HCB/kg/day  for  a relatively long period (Cam and



Nigogosyan, 1963).   It is speculated  that some  of  the Louisiana



workers had taken in several mg HCB per kg of body weight per day,



at least sporadically.



Synergism and/or Antagonism



     HCB at doses  far  below  those causing mortality enhances the



capability of animals to metabolize foreign organic  compounds (see



Metabolism section).   This type  of  interaction may  be  of  impor-



tance in determining the effects of other concurrently encountered



xenobiotics on  the animal (Carlson  and Tardiff,  1976).   An in-



crease in paraoxon dealkylation  activity was  a  more  sensitive in-



dicator  of  induction  of  microsomal  enzyme   activity  in a  liver



fraction from rats fed a diet containing 2 mg HCB/kg for two weeks



than cytochrome P-450 content  or  N-demethylase  activity (Iverson,



1976).



     HCB elicits significant and  rather selective changes  in lin-



dane  metabolism  in  rats  (Chadwick,   et 'al.   1977).    Rats  admin-



istered 7.5 mg HCB/kg/day orally for  seven days had  increased cap-



ability to metabolize  and eliminate  1,2,3,4,5,6-hexachlorocyclo-



hexane (lindane).  As  noted  before,  HCB caused  liver enlargement



and enhanced EPN metabolism.   Rats fed HCB also had  significantly



increased ability  to  metabolize p-nitroanisole,  but  not  methyl



orange.  HCB-treated rats excreted 35 percent of  the administered



lindane in their  feces  and 13.7  percent in their  urine  within  24



hours, in  contrast to 12.7  percent  in feces  and 5.0  percent  in



urine of unexposed rats.  The  amount  of lindane  in fat and  liver,



24 hours  after  administering  12.5  mg of  lindane/kg  orally,  was
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less  in HCB-treated rats  than in  unexposed  controls (117  versus



60.7 mg/kg  in  fat  and  9.57  versus 5.24 mg/kg in liver).  The  lin-



dane  content of  the  kidney was  not  significantly  reduced  (6.91



versus  5.94 mg/kg  for  HCB-treated versus  unexposed  rats).   Rats



pretreated  with  HCB excreted  a significantly higher proportion  of



free chlorophenols, with a  corresponding decrease  in polar  metabo-



lites as compared  to unexposed rats.



     Prior  exposure to  HCB  may alter the response of an animal  to



any of  a variety of challenges.  Mice  fed a diet containing 167  mg



HCB/kg  have  altered  susceptibility  to  Salmonella  typhosa  0901



1ipopolysaccharide  (endotoxin).    The LD50  for exposed  mice was



about 40  mg endotoxin/kg,  for  mice  fed  HCB  for  three  weeks 7.4



mg/kg,  and  for  mice fed HCB for  six weeks, 1.4 mg/kg.   Mice fed



HCB were  also somewhat more  susceptible to  the  malaria parasite



Plasmodium  than unexposed mice (Loose, et al. 1978).



Teratogenicity



     The effect of HCB on reproduction has received limited atten-



tion.   Dietary  HCB adversely  affected reproduction  in  the   rat  by



decreasing  the  number  of  litters  whelped  and  the  number   of  pups



surviving to weaning (Grant, et al. 1977).  The fertility (numbers



of litters  whelped/number   of  females exposed  to  mating)   of  rats



fed a diet  containing  320  mg  HCB/kg  was  decreased.   This  concen-



tration of HCB in  the food  led to  cumulative toxicity resulting  in



convulsions and  death  in  some  of  the  animals.   The proportion  of



pups surviving five days was reduced when the parents had been fed



a diet  containing  160 mg HCB/kg  and when the rats  had  been fed a



diet of 80  mg HCB/kg for  three  generations.    Birth  weights  were
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reduced in  rats  fed a diet  containing  320 mg HCB/kg  and in rats



fed a  diet containing  160  mg  HCB/kg  for  two generations.   The



weights of  5-day-old pups were  markedly  less  when the parents had



been fed a diet containing 80 mg HCB/kg.  The tissue of 21-day-old



pups whose dam had been fed graded dietary levels of HCB contained



progressively more  drug.   For  example,  the  level of  HCB  in body



fat was about  250  mg/kg  when the dietary  level  was  10 mg/kg? 500



mg/kg in fat for 20 mg/kg  in diet;  800 mg/kg  in  fat for  40 mg/kg



in diet; 1,900 mg/kg in fat  for 80  mg/kg  in diet; and 2,700 mg/kg



in fat for  160 mg/kg in diet.   The  highest HCB levels were in the



body fat;  for pups whose dam  had  been  fed  a diet containing 10 mg



HCB/kg, the body fat contained  250 mg HCB/kg; liver, 9 mg/kg; kid-



ney and brain,  4  mg/kg; and  plasma,  1.3 mg/kg.   HCB  crossed the



placenta of rats  and  accumulated in  the fetus  in  a dose-related



manner.  HCB  fed  to pregnant mice  and rats was  deposited in the



tissues in  a  dose-related  manner.   The HCB content of placentas



was greater than  that  of the  corresponding fetuses  in  both rats



and mice,  and equivalent to that in the yolk sac.  The fetuses and



placentas   of  rats  had  proportionally  greater  deposition  of HCB



than those of mice at the same dose levels.  Upon multiple dosing,



the deposition  of  HCB  increased  in both  fetuses  and  placentas



(Andrews and Courtney, 1976).   HCB  does not appear  to be terato-



genic for the rat  even  though  the chemical  is  reaching  the fetus



(Khera, 1974).



     Pregnant CD-I  mice  given  50 mg HCB/kg/day  orally  on gesta-



tional days 7  to  11 showed  essentially the  same  tissue distribu-



tion of drug  on day 12  as similarly-treated,  nonpregnant female
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mice.   The  HCB levels (mg/kg) were  as follows:   fat,  about 500;
thymus and skin, about 200;  skeletal  muscle,  about 100;  liver and
brain,  about  25;  and spleen,  kidney, uterus,  and  ovaries,  about
12.  The fetus contained 1.2 mg HCB/kg and the placenta 1.6 mg/kg.
CD-I  mice  administered  100  mg HCB/kg/day  orally  on  gestational
days  7  to  16  had increased  maternal  liver  weight  to  body weight
ratios and decreased fetal body weights.  In addition,  there was a
small  increase  in  the incidence  of  abnormal fetuses  per  litter.
These  abnormalities  included  cleft   palate,  small kidneys,  club
foot,  and  enlarged  renal  pelvis  in   both  unexposed  and  exposed
groups (Courtney, et al. 1976).
Mutagenicity
     The capability  of HCB to  induce  dominant lethal mutations in
rats was tested after administering up to 60  mg HCB/kg/day orally
for  10  days.   There were  no significant differences  between the
exposed  and  unexposed groups  with respect  to the incidence  of
pregnancies, corpora lutea,  liver  implants,  or  deciduomas  (Khera,
1974).
     HCB injected intraperitoneally into rats at 10 mg/kg elicited
a marked induction  of  the  hepatic cytochrome P-450 system.   This
liver  microsomal  fraction mediated   the  metabolic activation  of
2,4-diaminoanisole  to  a mutagen  (as  measured  by  the   Ames  test)
(Dybing and Aune, 1977).  The mutagenic activities of several aro-
matic  and  polycyclic  hydrocarbons are  not  associated with  the
parent compound but with metabolically activated products  that re-
act covalently with nucleic acid.  As noted previously, HCB stimu-
lates  the  hepatic  cytochrome  P-450  system  and, thereby,  has the
potential to enhance the mutagenicity  of other  chemicals.
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Carcinogenicity
     Two studies  have  been conducted which indicate that HCB  is  a
carcinogen.   The  carcinogenic activity of HCB  in hamsters  fed  4,
8,  or  16  mg/kg/day for  life  was assessed  (Cabral,  et al.  1977).
HCB appears  to have multipotential  carcinogenic activity; the  in-
cidence of hepatomas, haemangioendotheliomas, and  thyroid adenomas
was significantly  increased.   Whereas 10  percent of the  unexposed
hamsters developed tumors, 92 percent  of the  hamsters  fed  16  mg
HCB/kg/ day developed tumors.  The  incidence  of  tumor-bearing  ani-
mals was dose-related:   56 percent  for hamsters fed 4 mg HCB/kg/
day and 75 percent for  8 mg/kg/day.  Thyroid tumors, hepatomas  or
liver  haemangioendotheliomas  were  not  detected  in  the  unexposed
group.  An intake  of 4 to  16 mg HCB/kg/day in hamsters  is near  the
exposure range estimated  for  Turkish people  who accidentally  con-
sumed HCB-contaminated grain  (Cabral, et  al. 1977).
     The carcinogenic activity of HCB in mice  fed  6.5,  13,  or  26
mg/kg/day for life was  assessed.   The  incidence of hepatomas was
increased significantly  in mice fed  13 or 26 mg  HCB/kg/ day.  None
of the hepatomas occurred  or metastasized in the  untreated control
groups.  The  results  presented in  the abstract  of  Cabral,  et al.
(1978) confirm their earlier  conclusion  that HCB is carcinogenic.
However, the  incidence   of  lung  tumors in  strain A mice treated
three times a week for  a total of  24  injections of 40 mg/kg each
was not significantly greater than  the incidence  in control mice
(Theiss, et al.  1977).   Moreover,  HCB did not induce hepatocellu-
lar carcinomas in  ICR  mice fed HCB at  1.5 or  7  mg/kg/day  for  24
weeks (Shirai, et al.  1978).
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                      CRITERION FORMULATION
Existing Guidelines and Standards
     As  far  as  can be  determined,  the  Occupational  Safety and
Health Administration  (OSHA)  has  not set a  standard  for occupa-
tional exposure  of  HCB.   HCB  has  been approved  for  use as a pre-
emergence fungicide applied  to seed  grain.   The Federal Republic
of  Germany  no  longer  allows  the  application  of  HCB-containing
pesticides (Geike and  Parasher,  1976).   The  government of Turkey
discontinued the use of  HCB-treated  seed wheat  in  1959 after its
link to acquired toxic porphyria  cutanea  tarda was  reported  (Cam,
1959).  Commercial  production  of  HCB  in  the  United  Staes was dis-
continued in 1976 (Chem.  Econ. Hdbk., 1977).   The Louisiana  State
Department of  Agriculture has  set  the  tolerated level  of  HCB  in
meat fat at 0.3 mg/kg (U.S. EPA, 1976).  The NHMRC (Australia) has
used this same value for  the tolerated level  of HCB in cows' milk
(Miller and Fox,  1973).  WHO has set  the tolerated level of HCB  in
cows' milk at 20 ug/kg  in whole milk  (Bakken and Seip, 1976).  The
New  South Wales  Department of  Health (Australia)  has recommended
that the  concentration  of HCB in  eggs must  not  exceed  0.1 mg/kg
(Siyali,  1973).  The value of.0.6  ug  HCB/kg/day was  suggested  by
FAO/WHO in 1974  as  a reasonable  upper limit for HCB  residues  in
food for  human consumption (FAO/WHO, 1974).    The  FAO/WHO recom-
mendations for residues  in foodstuffs were  0.5 mg/kg  in  fat for
milk and eggs, and  1  mg/kg in  fat for meat and  poultry.   Russia
and Yugoslavia have set  the maximum tolerated  level  of HCB in air
at 0.9 mg/m3 (Int.  Labor Off.  1977).
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Current Levels of Exposure
     HCB appears to  be  distributed  worldwide,  with high levels  of
contamination  found  in  agricultural areas  devoted  to  wheat  and
related cereal  grains  and in  industrial  areas.   HCB is manufac-
tured  and  formulated  for application  to  seed wheat  to  prevent
bunt; however, most  of  the HCB in  the  environment comes from  in-
dustrial processes.   HCB  is  used  as a starting  material   for  the
production of  pentachlorophenol  which is marketed  as a wood pre-
servative.  HCB is one  of  the main  substances  in  the  tarry  residue
which  results  from  the  production of  chlorinated  hydrocarbons.
HCB is formed as a by-product in the production of  chlorine gas  by
the  electrolysis  of  sodium  chloride  using  a mercury electrode
(Gilbertson and Reynolds,  1972).
     People in the United States  are exposed to HCB  in  air,  water
and food.  HCB is disseminated in the air as dust  particles and  as
a result of volatilization from sites having a  high HCB-concentra-
tion.   Airborne HCB-laden  dust  particles  appear to  have  been  a
major cause of  increased blood  concentrations  of  HCB in the gen-
eral public living near an industrial site  in Louisiana  (Burns  and
Miller, 1975).  HCB  is  found  in  river water near  industrial  sites
in quantities of as  high  as  2 ug/kg (Laska, et al. 1976) and even
in finished drinking water at 5 ng/kg  (U.S. EPA,  1975).   HCB  oc-
curs in a wide variety  of  foods, in particular, terrestrial animal
products, including dairy  products  and eggs  (U.S.  EPA, 1976).   The
dietary intake of HCB has  been estimated to  be  0.5  ug/day in  Japan
(Ushio and Doguchi,  1977) and 35 ug/day  in Australia (Miller  and
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Fox, 1973).   Breast-fed infants in Australia  and  Norway may  con-
sume 40 ug HCB/day (Miller and Fox, 1973; Bakken and Seip,  1976).
     Table 2  lists HCB  concentrations  found in human adipose  tis-
sue  collected throughout the  world.    The  maximum  HCB  level  re-
ported was 22 mg/kg  (Acker and Schulte, 1974).
     The HCB  content  of  human  blood  samples collected in Bavaria,
Australia, and Louisiana is shown in Table  3.  The maxium HCB  con-
centration reported was 0.345 mg/kg in  the  sample  from a Louisiana
waste disposal worker (Burns and Miller, 1975).
     The levels of HCB  in  body  fat  of  swine and sheep were 6-fold
and  8-fold greater,  respectively than  the  dietary level (Hansen,
et al. 1977).  If these comparisons are valid when applied  to  man,
it would appear  that some  adult humans have  been  exposed  to  sev-
eral mg HCB/kg/day.   A  similar conclusion  is  reached  by extrapo-
lating  the values for  human  blood.    The  HCB levels  in  blood  of
rats are about one tenth the dietary level  (Kuiper-Goodman, et al.
1977) .
     Current  evidence would  indicate  that  food  intake  may  be  the
primary source of the body burden of  HCB  for  the  general  popula-
tion although inhalation and dermal exposure may be more important
in selected groups, e.g., industrial workers.
Special Groups at Risk
     Several  groups appear to be at increased risk.  These  include
workers engaged  directly in:   (1)  the manufacture  of HCB  or  in
processes  in  which HCB  is  a by-product,   (2)  the  formulation  of
HCB-containing  products,   (3)   the   disposal  of  HCB-containing
                                C-128

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                           TABLE 2



Hexachlorobenzene Content of Human Adipose Tissues at Autopsy
Source No. Samples
Australia 75
81
Papua and
New Guinea 38
Japan 241
Canada 3
16
? " 50
to
vo H 57
22
27
Germany 56
54
54
59
59
93
Mean Values
(mg/kg in
Human Fat)
1.25
1.31
0.26
0.08
0.09
0.025
0.107
0.060
0.015
0.043
2.9
8.2
5.9
4.8
6.4
4.8
Reference
Brady and Siyali,
Siyali, 1972
Brady and Siyali,

1972

1972
Curley, et al . 1973
Mes and Campbell,
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schultr
1976





, 1974
, 1974
, 1974
, 1974
, 1974
, 1974

-------
                TABLE 3




The HCB Content of Human Blood Samples
Mean Values
(mg/kg
Source
Bavaria
0
1 "
U)
0 Australia
"
n
Louisiana
NO.
98

96

185
52
76
86
Samples
boys

girls

exposed
unexposed


in
0

0

0
0
0
0
Blood)
.022

.017

.055
.022
.058
.0036

Richter

Richter

Siyali,
Siyali ,
Siyali
Referemce
and Schmid,

and Schmid,

1972
1972
and Ouw, 1973

1976

1976




Burns and Miller, 1975

-------
wastes; and  (4) the'application  of HCB-containing  products.   Other
groups  at  risk are  the  general public  living  near  industrial
sites,  populations  consuming large amounts  of contaminated  fish,
pregnant  women, fetuses,  and  breast-fed  infants.    Two lines  of
evidence  indicate that  infants may be at risk.  It  has  been  demon-
strated that human  milk contains HCB,  and  some infants  may  be  ex-
posed  to  relatively high  concentrations  of HCB  from that  source
alone  (Miller  and  Fox, 1973;  Bakken  and Seip,  1976).   Moreover,
some  infants  of  Turkish  mothers  who consumed  HCB-contaminated
bread developed a fatal disorder  called pembe  yara.   In  some  Turk-
ish villages in the region most  affected by HCB-poisoning, few  in-
fants survived  during  the period  1955-1960  (Cam, 1960).
     Occupational exposure  is associated  with an  increased  body
burden  of HCB.   Plant  workers   in  Louisiana  have  about  200  ug
HCB/kg  in blood (Burns  and Miller, 1975) .  The HCB  content of body
fat exceeded 1 mg/kg in many parts of the world where HCB contami-
nation  of the  environment  is  extensive (Brady and Siyali,   1972;
Acker and Schulte, 1974) .
     The  massive  episode  of human poisoning  resulting  from  the
consumption of  bread prepared  from HCB-treated seed wh^at brought
to  light  the  misuse  of  HCB-treated  grain  (Cam  and Nigogosyan,
1963).  In spite of warnings,  regulations,  and attempts at public
education, HCB-treated  grain apparently still  finds  its way  into
the food  chain,  for example,  in  fish  food (Hansen,  et  al.   1976;
Laska, et al.  1976).   The  difficulty  in tracing the source of  HCB
cor -amination  in  a  diet for  laboratory  animals   emphasizes  the
                                C-131

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difficulties encountered  in tracing  the source  of  HCB  in  food-
stuffs for human consumption (Yang, et al. 1976).
     As noted previously,  adipose  tissue acts as  a  reservoir for
HCB.   Depletion  of  fat depots  can result in  mobilization and re-
distribution of stored HCB.  Weight loss for any reason may result
in a dramatic redistribution  of  HCB contained  in  adipose tissue;
if the stored levels of HCB are high, adverse effects might ensue.
Many humans  restrict  their dietary intake  voluntarily  or because
of  illness.   In  these instances,  the  redistribution of  the HCB
body burden  becomes  a potential added  health hazard (Villeneuve,
1975).
Basis and Derivation of Criterion
     Among the studies reviewed by this  document,  only  two appear
suitable  for  use in  the  risk  assessment:    the  mouse  study  of
Cabral, et  al.   (1978)  and the hamster  study  of  Cabral, et al.
1977.  These two studies are described in detail in Appendix I.
     Under the Consent Decree  in NRDC  v.  Train,  criteria  are to
state  "recommended  maximum permissible  concentrations  (including
where appropriate, zero) consistent with the protection of aquatic
organisms, human health,  and  recreational   activities".   HCB  is
           *
suspected of being a human carcinogen.  Because there is no recog-
nized  safe concentration  for  a human  carcinogen,  the recommended
concentration of HCB  in  water  for  maximum  protection  of  human
health is zero.
     Because attaining  a  zero  concentration  level may  be unfeas-
ible in some cases,  and in order to  assist  the  Agency  and States
in  the  possible  future development of  water  quality regulations,
                               C-132

-------
the  concentrations  of  HCB corresponding  to  several  incremental

lifetime cancer  risk  levels have been  estimated.   A  cancer risk

level provides an estimate of the additional  incidence  of cancer

that may  be expected  in  an exposed population.   A risk  of 10~5

example, indicates a probability  of  one  additional  case  of cancer

for  every  100,000  people  exposed,  a risk  of 10~^  indicates one

additional case of cancer for every million people exposed, and so

forth.

     In the Federal Register  notice  of  availability  of  draft am-

bient water quality  criteria, EPA stated  that it  is  considering

setting  criteria  at  an  interim  target  risk  level  of  10~5,

10~6, or 10~7  as shown in the table below:

                                         Risk Levels
Exposure Assumption      	and Corresponding Criteria (1)
     (per day)
                              IP"7         IP"6         10~5

2 liters of drinking       0.072 ng/1    0.72 ng/1    7.2 ng/1
water and consumption
of 6.5  grams fish
and shellfish. (2)

Consumption of fish        0.074 ng/1    0.74 ng/1    7.4 ng/1
and shellfish  only.
(1)  Calculated from the  linearized  multistage  model descri-

     bed in  the  Human Health  Methodology  Appendices  to the

     October 1980  Federal Register  notice,  which  announced

     the availability of this document.  Appropriate bioassay

     data used in the calculation of  the model  are  presented

     in Appendix I.  Since the extrapolation model  is  linear

     at low  doses,  the  additional lifetime risk  is directly
                               C-133

-------
     proportional  to the  water  concentration.    Therefore,
     water concentrations corresponding  to other risk levels
     can  be  derived by  multiplying or  dividing one  of  the
     risk levels and corresponding water concentrations shown
     in the  table  by factors such as 10,  100, 1,000,  and so
     forth.
(2)  Ninety-seven  percent  of the  HCB  exposure  results  from
     the  consumption of aquatic  organisms which  exhibit an
     average bioconcentration potential  of 8,690-fold.   The
     remaining 3 percent of  HCB  exposure results from drink-
     ing water.
     Concentration  levels were  derived  assuming a  lifetime expo-
sure to various amounts of HCB,  (1) occurring from the consumption
of both drinking water and aquatic life grown in waters containing
the corresponding HCB concentrations and (2) occurring solely from
consumption  of  aquatic life  grown in  the waters  containing  the
corresponding HCB  concentrations.   Because data indicating other
sources of HCB exposure and  their contributions  to total body bur-
den are  inadequate  for  quantitative use,  the  figures  reflect the
incremental risks associated with the indicated  routes only.
                                C-134

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Siyali,  D.S.   1973.   Polychlorinated biphenyls, hexachlorobenzene



and  other  organochlorine pesticides  in human  milk.    Med.  Jour.




Australia.   2: 815.







Siyali,  D.S.  and  K.H. Ouw.   1973.   Chlorinated hydrocarbon pesti-



cides  in human  blood  - Wee  Waa survey.   Med.  Jour.  Australia.




2: 908.







Stephan, C.E.   1980.  Memorandum to J.  Stara.   U.S.  EPA.  July 3.








Stonard, M.D.   1975.   Mixed type hepatic microsomal enzyme  induc-



tion by hexachlorobenzene.  Biochem.  Pharmacol.   24: 1959.
                                C-146

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 Stonard,  M.D.  and J.B.  Greig.    1976.    Different  patterns   of



 hepatic  microsomal enzyme activity produced  by administration  of



 pure  hexachlorobiphenyl  isomers  and  hexachlorobenzene.    Chem.



 Biol. Interact.   15: 365.








 Strik, J.J.T.W.A.   1973.  Species differences  in experimental por-



 phyria  caused  by  polyhalogenated  aromatic  compounds.    Enzyme.



 16: 224.








 Sundlof,  S.F.,  et al.   1976.   Pharmacokinetics of hexachloroben-



 zene in male  laboratory beagles.  Pharmacologist.  18: 149.








 Theiss,  J.C.,  et  al.   1977.  Test  for  carcinogenicity of organic



 contaminants  of United States drinking waters  by  pulmonary tumor



 response  in strain A mice.  Cancer Res.  37: 2717.








 Timme,  A.H.,  et  al.    1974.    Symptomatic porphyria.    Part  II.



 Hepatic  changes  with  hexachlorobenzene.    S.  African Med.  Jour.



 48: 1833.








 U.S. EPA.  1975.   Preliminary assessment  of suspected carcinogens



 in drinking water.  Report  to Congress.   EPA  560/4-75- 003.   U.S.



 Environ. Prot. Agency,  Washington, D.C.








U.S. EPA.  1976.   Environmental  contamination  from hexachloroben-



 zene.   EPA 560/6-76-014.  Off.  Tox.  Subst.  1-27.
                               C-147

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U.S. EPA.   1980a.   Seafood  consumption  data analysis.   Stanford



Research Institute International, Menlo Park, Calif.   Final rep.,



Task II, Contract No.  68-01-3887.







U.S. EPA.   1980b.   Laboratory data.  Environ. Res.  Lab.   Duluth,



Minnesota.  (Unpub.)







Ushio, F. and M. Doguchi.   1977.   Dietary intakes of some chlori-



nated hydrocarbons  and  heavy metals estimated on the experiment-



ally prepared diets.  Bull. Environ. Contam. Toxicol.  17: 707.







Veith,  G.D.,  et  al.    1977.    Residues   of  PCBs  and  DDT  in  the



western Lake Superior ecosystem.   Arch.  Environ.  Contam.  Toxicol.



5: 487.







Veith,  G.D.,    1980.    Memorandum  to C.E.  Stephan.   U.S.  EPA.



April 14.







Villeneuve,  D.C.   1975.   The effect of  food restriction  on  the



redistribution  of  hexachlorobenzene in  the  rat.    Toxicol. Appl.



Pharmacol.  31: 313.







Villeneuve,  D.C.  and  W.H.  Newsome.  1975.   Toxicity  and tissue



levels  in the rat and guinea  pig following acute  hexachlorobenzene



administration.  Bull.  Environ.  Contam. Toxicol.  14: 297.
                                C-148

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 Vos,  J.G.,  et  al.    1971.    Toxicity  of  hexachlorobenzene   in
 Japanese  quail  with  special  reference to porphyria, liver damage,
 reproduction,  and  tissue residues.    Toxicol.   Appl.  Pharmacol.
 18: 944.

 White,  D.E.  and T.E. Kaiser.  1976.   Residues of  organochlorines
 and  heavy  metals  in  ruddy  ducks   from  Delaware River,  1973.
 Pestic. Monitor. Jour.  9: 155.

 Wilson, D.W.  and L.G. Hansen.   1976.   Pharmacokinetics  of  hexa-
 chlorobenzene in growing swine.  Pharmacologist.  18: 196.

 Yang, R.S.H.,  et al.   1976.  Hexachlorobenzene  contamination  in
 laboratory monkey chow.  Jour. Agric. Food Chem.  24: 563.

 Zitko,  V.  1976.    Levels  of chlorinated hydrocarbons  in  eggs  of
 double-crested cormorants from 1971  to  1975.   Bull. Environ.  Con-
 tarn. Toxicol.  16: 399.

 Zitko, V. and 0.  Hutzinger.   1976.  Uptake  of chloro-  and bromo-
 biphenyls,  hexachloro-  and   hexabromobenzene  by  fish.     Bull.
Environ. Contam. Toxicol.   16: 665.
                               C-149

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                  SUMMARY-CRITERION FORMULATION



Existing Guidelines and Standards



Monochlorobenzene



     The Threshold  Limit  Value  (TLV)  for MCB  as  adopted  by  the



American Conference of Governmental  Industrial Hygienists (ACGIH)



(1971)  is  75  ppm  (350  mg/m3).   The American  Industrial Hygiene



Association Guide  (1964)  considered  75 ppm  to be  too high.   The



recommended maximal allowable concentrations in air in other coun-



tries  are:    Soviet Union,  10  ppm;   Czechoslovakia,  43  ppm;  and



Romania, 0.05 mg/1.  The  latter  value for Romania  was reported by



Gabor and Raucher (1960)  and is equivalent to 10 ppm.



Trichlorobenzene



     A  proposed ACGIH Threshold  Limit Value   (TLV)  standard  for



TCBs  is  5  ppm (40 mg/m3)  as  a ceiling value  (ACGIH,  1977).   Sax



(1975)  recommends a maximum allowable concentration  of  50  ppm in



air  for commercial TCB,  a  mixture   of  isomers.    Coate,  et  al.



(1977),  citing  their  studies, recommends that the TLV  should be



set  below  25  ppm,  preferably  at 5  ppm  (40  mg/m3).   Gurfein  and



Parlova  (1962)   indicate   that  in the  Soviet  Union  the maximum



allowable  concentration  for TCB  in water  is 30 ug/1  which  is  in-



tended  to prevent organoleptic effects.



Tetrachlorobenzene



     The maximal permissible concentration of TeCB in water estab-



lished  by  the Soviet Union  is 0.02 mg/1 (U.S. EPA, 1977).



Pentachlorobenzene



     No guidelines  or standards  for pentachlorobenzene  could be



located  in  the  available  literature.
                               C-150

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Hexachlorobenzene
     As  far as  can  be  determined,  the  Occupational  Safety and
Health Administration has  not  set  a standard for occupational ex-
posure to  HCB.   HCB  has been  approved  for  use as a pre-emergence
fungicide  applied  to  seed  grain.   The Federal Republic of Germany
no  longer   allows   the  application  of  HCB-containing  pesticides
(Geike and Parasher,  1976).  The government  of Turkey discontinued
the  use  of  HCB-treated seed  wheat  in  1959  after  its   link to
acquired toxic  porphyria cutanea tarda was  reported (Cam, 1959).
Commercial production  of HCB in the United States was discontinued
in 1976 (Chem. Econ.  Hdbk., 1977).  The Louisiana State Department
of Agriculture  has  set the tolerated level  of HCB  in  meat fat at
0.3 mg/kg  (U.S.  EPA,  1976).   The NHMRC (Australia)  has used  this
same value  for  the tolerated  level  of  HCB   in cows'  milk (Miller
and Fox,  1973).  WHO  has set the tolerated  level  of HCB  in cows'
milk at 20  ug/kg in whole  milk (Bakken and  Seip,  1976).   The New
South Wales Department of  Health  (Australia)  has recommended  that
the  concentration   of HCB  in  eggs  must  not  exceed  0.1  mg/kg
(Siyali,  1973).  The  value of  0.6 ug HCB/kg/day was suggested by
FAO/WHO in  1974  as a  reasonable upper  limit for HCB  residues in
food for human  consumption  (FAO/WHO,  1974).  The  FAO/WHO recom-
mendations  for  residues  in foodstuffs were  0.5  mg/kg  in  fat for
milk and eggs,  and  1  mg/kg in  fat  for  meat and  poultry.   Russia
and Yugoslavia have set  the maximum  tolerated  level  of  HCB in air
at 0.9  mg/m3 (Int.  Labor Off.,  1977).
                               C-151

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Current Levels of Exposure
Monochlorobenzene
     MCB has been detected  in water  monitoring  surveys  of various
U.S. cities (U.S. EPA, 1975; 1977) as was presented in the text in
Table  1.   Levels  reported were:   ground water  - 1.0  ug/1;  raw
water contaminated by various discharges - 0.1 to 5.6 ug/1; upland
water  -  4.7  ug/1;  industrial  discharge  -  8.0  to 17.0  ug/1  and
municipal water  -  27 ug/1.  These  data show a  gross  estimate of
possible human exposure to MCB through the water route.
     Evidence  of possible  exposure   from  food  ingestion is  in-
direct.  MCB  is  stable in water and  thus could  be bioaccumulated
by edible fish species.
     The only data concerning exposure to MCB via air are from the
industrial working environment.   Reported  industrial  exposures to
MCB  are  0.02  mg/1  (average value)   and  0.3  mg/1  (highest value)
(Gabor  and  Raucher,  1960);  0.001 to 0.01 mg/1  (Levina, et  al.
1966); and 0.004 to 0.01 mg/1 (Stepanyen, 1966).
Trichlorobenzene
     Possible  human  exposure to TCBs  might  occur  from municipal
and  industrial  wastewater   and  from  surface  runoff   (U.S.  EPA,
1977).   Municipal  and  industrial discharges  contained  from  0.1
ug/1  to 500 ug/1.   Suface runoff has  been found  to contain 0.006
to 0.007 ug/1.
     In the  National  Organics  Reconaissance  Survey (NORS)  con-
ducted  by  EPA  in  1975,  trichlorobenzene was  found in  drinking
water at a level of 1.0 ug/1.
                               C-152

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Tetrachlorobenzene



     No data  are  available on  current  levels of  exposure.    How-



ever, the report by Morita, et  al.  (1975) gives some  indication  of



exposure.  Morita,  et al. (1975)  examined  adipose tissue samples



obtained at  general hospitals  and medical  examiners'  offices  in



central Tokyo.   Samples  from 15  individuals  were examined;  this



represented 5 males and  10  females between the ages of 13 and 78.



The tissues were examined  for 1,2,4,5-TeCB as well as for 1,4-di-



chlorobenzene and hexachlorobenzene.  The  TeCB content of the fat



ranged  from  0.006  to  0.039  mg/kg  of tissue;  the  mean  was   0.019



mg/kg.   The  mean concentrations  of l,4-dichlorobenze;ne  and  hexa-



chlorobenzene  were   1.7   mg/kg  and  0.21  mg/kg,   respectively.



Neither age nor sex correlated  with the level of any of the  chlo-



rinated hydrocarbons  in adipose  tissue,.



Pentachlorobenzene



     Morita,  et al. (1975) examined levels  of QCB in adipose  tis-



sue samples obtained  from general  hospitals  and medical examiners'



offices in central Tokyo.  The  samples were  collected from a  total



of 15 people.   By gas  chromatography,  the  authors  found  the re-



sidual  level  of  QCB range from 0.004  ug/g  to 0.020  ug/g/  with  a



mean value  of 0.09 ug/g  of   fat.   Lunde and  Bjorseth  (1977) ex-



amined  blood  samples  from workers with occupational  exposure  to



pentachlorobenzene  and  found that their blood  samples  contained



higher  levels of this  compound  than a comparable group of workers



not exposed to chlorobenzenes.
                                C-153

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Hexachlorobenzene
     HCB appears  to  be  distributed  worldwide,  with high levels  of
contamination  found  in  agricultural areas  devoted  to  wheat  and
related  cereal  grains  and in  industrial  areas.   HCB is manufac-
tured  and  formulated  for application  to  seed wheat  to  prevent
bunt;  however,  most  of  the HCB in  the  environment comes from  in-
dustrial processes.   HCB  is  used  as a starting  material   for  the
production of pentachlorophenol  which is marketed  as a wood pre-
servative.  HCB is one  of  the main  substances  in  the  tarry  residue
which  results   from  the  production of  chlorinated  hydrocarbons.
HCB is formed as a by-product in the production of  chlorine gas  by
the  electrolysis  of  sodium  chloride  using  a mercury electrode
(Gilbertson and Reynolds,  1972) .
     People in  the United  States are exposed to HCB in air,  water,
and food.  HCB  is disseminated in the air as dust  particles and  as
a result of volatilization from sites having a  high HCB-concentra-
tion.   Airborne HCB-laden  dust  particles  appear to have  been  a
major  cause of  increased  blood concentrations  of  HCB in the gen-
eral public living near an industrial site  in Louisiana  (Burns  and
Miller,  1975).  HCB  is  found  in  river water near   industrial  sites
in quantities of as  high  as  2  ug/kg (Laska, et al. 1976) and even
in finished drinking water at 5 ng/kg  (U.S. EPA,  1975).   HCB  oc-
curs in  a wide variety  of  foods, in particular, terrestrial animal
products, including dairy  products  and eggs (U.S.  EPA, 1976).   The
dietary  intake  of HCB has  been estimated to be 0.5  ug/day in  Japan
(Ushio and Doguchi,  1977)  and 35 ug/day  in Australia (Miller  and
                                C-154

-------
 Fox,  1973).   Breast-fed infants in Australia  and Norway may  con-
 sume  40  ug  HCB/day (Miller and Fox,  1973; Bakken and Seip,  1976).
 Table  1  lists HCB  concentrations  found  in  human adipose  tissues
 collected  throughout the  world.   The maximum  HCB  level reported
 was 22 mg/kg  (Acker and Schulte,  1974) .   The HCB content of human
 blood samples  is  given  in  Table 2.   The maximum HCB concentration
 reported  was  0.345 mg/kg  which  was  found  in  a  sample  from a
 Louisiana waste disposal worker.
     The levels of  HCB  in  body  fat of swine and sheep were  sixfold
 and eightfold  greater,  respectively,  than the dietary level (Han-
 sen, et al. 1977).   If  these comparisons  are valid when applied to
 man,  it  would  appear that some adult humans  have been  exposed to
 several mg HCB/kg/day.   A  similar  conclusion  is reached by extra-
 polating the values for human blood.   The HCB  levels  in blood of
 rats  are about  one tenth  less than  the dietary  level  (Kuiper-
 Goodman, et al. 1977).
     Current evidence would  indicate  that food intake  may  be  the
 primary source of  the  body burden of HCB  for  the general  popula-
 tion although inhalation and dermal exposure may be more important
 in selected groups, e.g.,  industrial workers.
 Special Groups at Risk
 Monochlorobenzene
     The major group at risk of MCB  intoxication are  individuals
 exposed to MCB in  the  workplace.   Girard, et  al.  (1969)  reported
 the case of  an elderly female  exposed  to a glue  containing 0.07
percent MCB for a period of  six years.   She  had symptoms  of head-
 ache,  irritation of  the eyes and the  upper  respiratory  tract,  and
                               C-155

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                           TABLE 1




Hexachlorobenzene Content of Human Adipose Tissues  at Autopsy
Sourcee No. Samples
Australia 75
81
Papua and
New Guinea 38
Japan 241
Canada 3
16
0
K " 50
a\
57
22
27
Germany 56
54
54
59
59
93
Mean Values
(mg/kg in
Human Fat)
1.25
1.31
0.26
0.08
0.09
0.025
0.107
0.060
0.015
0.043
2.9
8.2
5.9
4.8
6.4
4.8
Reference
Brady and Siyali,
Siyali, 1972
Brady and Siyali,

1972

1972
Curley, et al. 1973
Mes and Campbell,
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Mes, et al. 1977
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schulte
Acker and Schultr
1976




, 1974
, 1974
, 1974
, 1974
, 1974
, 1974

-------
o
I
                                                TABLE 2



                                The HCB Content of Human Blood Samples
Source
Bavaria
H
Australia
H
H
Louisiana
No. Samples
98 boys
96 girls
185 exposed
52 unexposed
76
86
Mean Values
( mg/kg
in Blood)
0
0
0
0
0
0
.022
.017
.055
.022
.058
.0036
Richter
Richter
Siyali,
Siyali ,
Siyali
Reference
and Schmid,
and Schmid,
1972
1972
and Ouw, 1973

1976
1976



Burns and Miller, 1975

-------
was  diagnosed to  have medullary  aplasia.   Smirnova  and  Granik



(1970)  reported  on three  adults  who developed  numbness,  loss of



consciousness, and  hyperemia of  the  conjunctiva and  the  pharynx



following exposure  to  "high"  levels  of  MCB.   Information concern-



ing  the ultimate  course  of  these  individuals   is  not available.



Gabor,  et  al. (1962)  described  toxic effects  on  individuals who



were  exposed  to   benzene,   chlorobenzene,  and  vinyl  chloride.



Eighty-two workers  examined for certain biochemical indices showed



a  decreased  catalase  activity  in  the  blood and  an  increase in



peroxidase,  indophenol oxidase,  and  glutathione levels.   Dunae-



veskii  (1972)  reported  on the  occupational  exposure  of  workers



exposed to the chemicals involved in the manufacture of chloroben-



zene at limits below the allowable  levels.  After  more than three



years,  cardiovascular  effects were  noted  as  pain  in  the  area of



the heart, bradycardia, irregular variations in electrocardiogram,



decreased  contractile  function  of  myocardium,  and disorders in



adaptation to physical loading.  Filatova, et al.  (1973)  reported



on the  prolonged  exposure  of individuals  involved  in  the  produc-



tion of diisocyanates  to  factory air which contained  MCB  as well



as other chemicals.  Diseases noted  include  asthmatic bronchitis,



sinus  arrhythmia,  tachycardia,  arterial  dystrophy,  and  anemic



tendencies.  Petrova and Vishnevskii  (1972) studied  the course of



pregnancy  and  deliveries  in  women exposed  to  air  in a  varnish



manufacturing  factory  where  the air  contained  three times  the



maximum permissible level  of  MCB but  also  included  toluene, ethyl



chloride,   butanol,  ethyl  bromide,  and  orthosilisic  acid -ester.
                                C-158

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The only  reported  significant adverse effect  of  this mixed expo-



sure was  toxemia during pregnancy.



Tetrachlorobenzene



     The  primary  groups  at risk  from the  exposure to  TeCB  are



those who deal with it in the workplace.  Since it is a metabolite



of certain  insecticides,  it might be  expected  that  certain indi-



viduals exposed to  those  agents  might experience  more exposure  to



TeCB, especially  since its  elimination  rate might  be  relatively



slow in man.   Individuals consuming  large quantities  of  fish may



also be  at risk  due  to  the proven  bioconcentration of  TeCB  in



fish.  The bioconcentration  factor for 1,2,4,5-TeCB  is 1,125.



Pentachlorobenzene



     A group  at  increased risk would appear  to  be  those  individ-



uals exposed  occupationally.  Due to the persistence  of  the  com-



pound in  the  food  chain,  an increase in  the body  burden of  QCB



might be  expected in  individuals  on  high  fish diets  or  diets  high



in agricultural products containing QCB residues of PCNB sprays.



Hexachlorobenzene



     Several groups appear  to  be  at  increased risk;  these  include



workers engaged  directly  in:   (1)  the manufacture  of  HCB  or  in



processes  in  which HCB  is  a  byproduct;   (2)  the formulation  of



HCB-containing  products;   (3)   the   disposal  of  HCB-containing



wastes;  and (4) the application  of HCB-containing  products.   They



also include  the  general public  living  near  industrial  sites,



pregnant  women,  fetuses,  and  breast-fed  infants and  populations



consuming large amounts of  contaminated  fish.  Two  lines  of  evi-



dence indicate that infants may  be  at  risk.  It has  been demon-
                                C-159

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strated that human milk contains HCB,  and  some  infants  may be ex-

posed  to  relatively  high concentrations  of  HCB from  that source
                   •
alone  (Miller  and  Fox,  1973; Bakken  and  Seip,  1976).   Moreover,

some  infants   of  Turkish  mothers  who consumed  HCB-contaminated

bread  developed  a  fatal  disorder  called pembe  yara.    In  some

Turkish villages in the region most affected by HCB-poisoning, few

infants survived during the period 1955-1960 (Cam,  1960).

     Occupational  exposure  is associated  with  an  increased  body

burden  of HCB.   Plant workers  in  Louisiana  have  about  200  ug

HCB/kg in blood (Burns and Miller,  1975).   The HCB content  of body

fat exceeds 1 mg/kg in many.parts  of  the  world  where HCB contami-

nation  of the  environment  is  extensive (Brady and  Siyali, 1972;

Acker and Schulte,  1974).

     The  massive  episode  of human poisoning  resulting  from the

consumption of bread prepared from HCB-treated  seed wheat  brought

to  light  the  misuse  of  HCB-treated   grain  (Cam and  Nigogosyan,

1963).  In spite of warnings, regulations,  and  attempts at public

education, HCB-treated grain apparently still finds  its way into

the food  chain,  for  example, in fish food (Hansen, et  al. 1976;

Laska, et al.  1976).  The  difficulty  in tracing  the source of HCB

contamination  in   a  diet  for  laboratory  animals   emphasizes  the

difficulties encountered  in  tracing  the  source of HCB  in food-
                                                     ^
stuffs for human consumption (Yang, et al. 1976) .

     As noted  previously,  adipose  tissue  acts as  a reservoir for

HCB.   Depletion  of fat depots can result  in mobilization and re-

distribution of stored HCB.  Weight loss for any reason  may result

in  a  dramatic  redistribution of HCB  contained  in  adipose  tissue;
                               C-160

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if the stored levels of HCB are high, adverse effects might  ensue.
          *

Many  humans  restrict their dietary  intake  voluntarily or because


of  illness.   In  these instances,  the  redistribution  of  the HCB


body  burden  becomes a potential  added  health hazard  (Villeneuve,


1975).


Basis and Derivation of Criteria


Monochlorobenzene


     There  is  no  information in  the literature  which indicates


that  monochlorobenzene is,  or is  not,   carcinogenic.    There  is


enough  evidence  to  suggest  that  MCB causes  dose-related  target


organ  toxicity,  although the  data  are  lacking  for  an acceptable


chronic  toxicity  study.    There  is  little,  if  any,  usable   human


exposure data primarily  because  the exposure was  not  only  to MCB


but to other compounds of known toxicity.


     A no-observed-adverse-effect  level  (NOAEL)  for derivation  of


the water quality criterion can be derived from  the information  in


the studies  by Knapp,  et al.  (1971) and  Irish  (1963).   These are


27.25 mg/kg/day for  the dog  (the  next highest dose was 54.5  mg/kg


and showed  an  effect), 12.5  mg/kg/rat  from the Knapp  study  (the


next  highest  dose  was   50   mg/kg   and   showed  an  effect) ,  and


14.5 mg/kg/rat from the Irish  study  (the  next highest dose was 144


mg/kg and showed an  effect).   When  toxic  effects were  observed  at


higher doses,  the  dog was  judged  to  be  somewhat  more sensitive


than  rats.   The  duration of  the  study  by  Irish  (1963)  was six


months which was  twice as long as  the Knapp  study  of  two species


(rat,  dog).   Since  the  Knapp and  Irish studies  appear  to give


similar  results and  since there are  no   chronic toxicity  data  on
                               C-161

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which to  rely,  the NOAEL level,  14.4  mg/kg for  six  months, from
the  longest  term  study  (Irish,  1963)  is  used   to  calculate  the
acceptable daily intake (ADI).
     Considering that  there  are relatively  little  human exposure
data, that there are no long-term animal data, and that  some  theo-
retical questions, at least, can be raised on the possible effects
of chlorobenzene on blood-forming tissue, an uncertainty factor of
1,000 is used.  From this the ADI can be calculated as follows:
           ADI . 7° k? *4       = 1.008 mg/day
     The average  daily consumption of  water  was  taken  to be  two
liters and the consumption  of  fish to be 0.0065 kg daily.  A  bio-
concentration factor of 10.3 was utilized.  The  following  calcula-
tion  results  in  a  criterion  based  on  the  available toxicologic
data:

                  2 +  (10.3  x  0 .0065)    488 ug//1
     Varshavskya  (1968),  the  only report  available,  has  reported
the  threshold concentration  for  odor  and taste  of  MCB in  reservoir
water  as  being  20 ug/1.   This value  is  about  4.5 percent of  the
possible standard calculated above.   It is, however,  approximately
17 times greater  than  the  highest  concentration of MCB measured in
survey sites.   Since  water  of disagreeable  taste  and  odor is  of
significant  influence  on  the quality  of life  and,  thus,  related to
health,  it would appear  that  the organoleptic level  of  20  ug/1
should be  the recommended  criterion.
                                C-162

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 Trichlorobenzene



      Reliable toxicologic data on which to base a defensible water



 quality criteria  do not  exist  for  the  trichlorobenzenes.   The



 studies by Smith, et al.  (1978), and Coate, et al.  (1977)  do not



 give sufficient detail  or suffer from inherent problems in experi-



 mental design.   Therefore,  according  to  the guidelines  for  cri-



 teria derivation, a criterion  cannot be recommended  for  any  tri-



 chlorobenzene isomer.   For future  derivation of  a  human  health



 criterion,  sound data must be developed describing  the  effects  of



 these compounds on humans and experimental animals.   It  should  be



 emphasized  that this is  a criterion  based  on aesthetic rather  than



 on  health  effects.  Data  on  human health  effects   must  be  devel-



 oped as a  more substantial basis for deriving a criterion  for  the



 protection  of human  health.



 Tetrachlorobenzene



      The  dose  of  5 mg/kg/day  1,2,4,5-TeCB  reported  for  beagles



 (Braun, 1978)  was  utilized as  the NOAEL for  criterion  derivation.



 An  acceptable daily  intake (ADI)  can be calculated  from  the NOAEL



 by  using a  safety  factor  of 1,000  based  on a  70 kg/man:



              ADT - 70 kg  x 5 mq/kq    _  ,,_
              ADI	1,000 ^—* =0.35 mg/day








     For the  purpose of establishing a water  quality criterion,  it



 is assumed  that on the average, a person ingests 2 liters of water



and  6.5 grams of fish.    Since  fish  may bioconcentrate  this com-



pound, a bioconcentration factor  (F)  is used  in the  calculation.
                               C-163

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     The equation for calculating  an  acceptable  amount of TeCB  in

water is:
 Criterion = 2 1 + (IIZS    .0065) = 37'6 u^/1 or   38


where :

          21=2 liters of drinking water consumed

    0.0065 kg = amount of fish consumed daily

        1,125 = b i oc once ntr a ti.an— factor

          ADI = Acceptable Daily Intake (mg/kg for a 70 kg/person)

     Thus/ the recommended criterion  for  1,2,4,5-TeCB  in water  is

38 ug/1.  The criterion  canf alternatively be expressed as 48 ug/1
                                                         *
if exposure is assumed to be from the consumption of fish and

shellfish alone.

Pentachlorobenzene

     A survey of the  QCB literature  revealed no acute, subchronic

or chronic toxicity data with  the  exception  of the study by Khera

and Villeneuve  (1975).   These  authors found  an  adverse effect  on

the fetal development  of embryos exposed  in  utero to pentachloro-

benzene administered  to  the  dams at  50 mg/kg on days  6  to 15  of

gestation.   This dose constitutes  a low-observed-adverse-effect-

level  (LOAEL) .    According  to  current guidelines,  extrapolation

from such data  requires  application of a  safety  factor of  from 1

to 10.   Since  the  observed effect was  only  suggestive of terato-

genicity of QCB, a  safety  factor of  3  is applied.   Because long-

term toxicity  data  on  humans  is  not available  and  the existing

animal data  is  sparse,  an additional  safety factor  of  1,000  is

applied to the  calculation  of  an acceptable  daily intake (ADI)  as

f ol 1 ow s :
                               C-164

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                        70 kg  x 50 mg/kg    , ,.,
                  ADI	(3) (1,000)    =  ia7 m*

The average  daily consumption of water  was taken to  be 2  liters
and the consumption of fish to be 0.0065 kg daily.  The  bioconcen-
tration factor for QCB is 2,125.
     Therefore:
Recommended Criterion = 2 + (2,125 'x?6.0065) = 74 ^/l

    The recommended water quality criterion for pentachlorobenzene
is 74  ug/1.   The criterion  can  alternatively be expressed  as 85
ug/1 if exposure  is assumed to be from the  consumption of fish and
shellfish alone.
Hexachlorobenzene
     Among the studies reviewed by  this  document,  only  two  appear
suitable  for  use in  the risk  assessment:  the mouse   study  of
Cabral, et  al.  (1978)  and  the hamster  study  of  Cabral,  et al.
(1977).  These two studies are described in detail in Appendix I.
     Under the Consent  Decree in NRDC  v.  Train, criteria  are to
state  "recommended  maximum  permissible  concentrations  (including
where appropriate, zero)  consistent with the protection  of aquatic
organisms, human health, and recreational activities".   HCB  is
suspected of being a human carcinogen.  Because there is no  recog-
nized  safe concentration  for   a human  carcinogen,  the  recommended
concentration  of HCB in  water for  maximum  protection  of  human
health is zero.
     Because attaining a  zero concentration level may be unfeas-
ible in some  cases,  and  in  order  to assist the  Agency and  states
                               C-165

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in the possible  future development of  water  quality regulations,

the  concentrations  of  HCB corresponding  to  several  incremental

lifetime cancer  risk  levels have  been  estimated.    A  cancer risk

level provides an estimate of the additional  incidence  of cancer

that  may  be  expected  in  an  exposed  population.    A   risk  of

10~5,  for   example,   indicates  a  probability  of  one  additional

case  of  cancer  for  every  100,000   people  exposed,  a   risk  of

10~6  indicates  one  additional  case  of  cancer for  every  million

people exposed,  and so forth.

     In the Federal Register  notice  of  availability of  draft am-

bient water quality criteria, EPA stated  that  it  is  considering

setting  criteria  at  an  interim  target  risk  level  of  10~5,

1006, or 10~7  as shown in the following  table:

                                         Risk Levels
Exposure Assumption      	and  Corresponding Criteria (1)
     (per day)
                              10"7         10~6         1Q-5

2 liters of drinking       0.072 ng/1    0.72 ng/1    7.2  ng/1
water and consumption
of 6.5 grams fish
and shellfish. (2)

Consumption of fish        0.074 ng/1    0.74 ng/1    7.4  ng/1
and shellfish  only.


(1)  Calculated from  the  linearized  multistage model descri-

     bed in the Human Health  Methodology  Appendices  to  the

     October  1980  Federal  Register  notice,  which   announced

     the availability of this document.  Appropriate bioassay

     data used in the  calculation  of  the model are  presented

     in Appendix I.   Since the  extrapolation  model   is linear

     at low doses,  the additional lifetime  risk is  directly
                               C-166

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      propdrtional  to  the  water  concentration.    Therefore,
      water  concentrations corresponding to other  risk  levels
      can  be derived  by multiplying  or dividing  one  of  the
      risk levels and  corresponding water concentrations  shown
      in the table  by factors such as  10,  100,  1,000,  and  so
      forth.
 (2)   Ninety-seven  percent  of  the HCB  exposure  results  from
      the  consumption of  aquatic  organisms which  exhibit  an
      average  bioconcentration potential of  8,690-fold.    The
      remaining 3 percent  of HCB  exposure  results  from drink-
      ing water.
      Concentration levels  were  derived assuming  a lifetime expo-
sure  to various amounts  of  HCB  (1)  occurring  from the consumption
of both drinking water and  aquatic life grown in waters  containing
the corresponding HCB concentrations and (2) occurring solely from
consumption  of  aquatic  life  grown in  the  waters  containing  the
corresponding HCB  concentrations.   Because data  indicating other
sources of HCB exposure and their contributions to total  body bur-
den are inadequate for  quantitative  use,  the figures  reflect  the
incremental risks associated with the indicated routes only.
                                C-167

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                      Summary of  Recommended  Criteria for Chlorinated Benzenes

                Substance                       Criterion        Basis for Criterion

            Monochlorobenzene1                   20  ug/1         organoleptic effects

            Trichlorobenzene2                    none            organoleptic effects

            1,2,4,5-Tetrachlorobenzene           38  ug/1         toxicity study

            Pentachlorobenzene                   74  ug/1         toxicity study

o           Hexachlorobenzene3                    7.2 ng/1       carcinogenicity
en
CO
            *A  toxicological  evaluation of  monochlorobenzene  resulted in  a  level  of
             of 488 ug/1;  however,  organoleptic  effects have been reported  at  20 ug/1.

            2Insufficient data to derive criterion.
            3The value 7.2 ng/1 is at a risk level of 1 in 100,000.

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                          APPENDIX I
               Summary  and  Conclusions  Regarding the
              Carcinogenicity  of  Chlorinated Benzene*
     Monochlorobenzene (MCB)  is  used industrially as a solvent, and
 as  a synthetic  intermediate  primarily for  production of  phenol,
 DDT, and  aniline.   MCB has been detected  in water  contaminated by
 industrial or  agricultural waste,  and  human exposure is  mainly via
 water.  There  are no studies  available concerning  the  mutagenic or
 carcinogenic potential of MCB, so that it is not possible to calcu-
 late a water quality criterion on the basis of an oncogenic effect.
     There are three isomers  of  trichlorobenzene  (TCB).   1,2,4-TCB
 is used as a carrier of dyes,  as a  flame retardant,  and in the syn-
 thesis of  herbicides.   1,2,3-TCB  and  1,3,5-TCB are used as  syn-
 thetic intermediates,  while a mixture  of the three  isomers  is used
 as a solvent  or  lubricant.  TCBs are likely intermediates  in  mam-
 malian metabolism  of  lindane,  and TCBs metabolize to  trichloro-
 phenols (TCP), e.g., 1,3,5-TCB produces 2,4,6-TCP.   TCB  is  present
 in drinking  water,  but there are  no  studies concerning  the  muta-
 genicity or carcinogenicity of  these compounds  and, hence, a  cri-
 terion cannot be calculated on this basis.
     Tetrachlorobenzene  (TeCB)  exists  as  three isomers.   Two  of
 these,  1,2,4,5-TeCB and  1,2,3,6-TeCB,  are  used in  the manufacture
of  2,4,5-trichlorophenoxyacetic  acid  (2,4,5-T)   and  2,4,5-tri-
chlorophenol (2,4,5-TCP).   TeCB is one of the metabolites of  hexa-
chlorobenzene and lindane.   TeCB has not been identified  in water


*This summary  has  been prepared and  approved  by  the  Carcinogens
 Assessment Group of EPA.
                              C-169

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in the  United States.   However,  industrial effluent  may contain
TeCB which causes contamination of aquatic organisms.  Soil micro-
organisms can  metabolize lindane  to TeCB,  which may further con-
taminate water due  to soil runoff.   There  are no carcinogenicity
studies available for TeCBs so that a water quality criterion can-
not be derived on this basis.
     Pentachlorobenzene  (QCB) is used mainly as  a precursor  in  the
synthesis of the fungicide pentachloronitrobenzene, and as a flame
retardant.  Lindane metabolizes in humans to QCB.  QCB has entered
water  from  industrial  discharge,  or  as a  breakdown  product  of
organochlorine compounds.   There  are no data available concerning
the  mutagenicity  of QCB.   There  is a  translated abstract  of  an
article by  Preussman  (1975)  which states that PCB is carcinogenic
in mice, but not  in  rats  and dogs.   The  abstract does  not report  the
data and, since the article has been difficult  to obtain,  the study
is not yet  available  to  evaluate for a water quality criterion.
     Hexachlorobenzene   (HCB)   is  used  as  a fungicide  and  indus-
trially for  the  synthesis of  chlorinated hydrocarbons,  as a plas-
ticizer,  and  as a flame  retardant.  HCB  has been detected in water
near  sites  of  industrial discharge  and leaches  from industrial
waste  dumps.   HCB  is very  stable  in the environment and bioaccu-
mulates,  so that it is present in  many  food  sources,  e.g., cereals,
vegetables,  fish, meat,  and dairy products.  It  is stored in human
adipose  tissue and is  present  in  human milk.   There  is  only  one
mutagenicity  study  reported for HCB which is  negative  for  the  in-
duction of  dominant lethal mutations in  rats.
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      Studies by Cabral,  et al.  (1977,  1978)  indicated  that oral
 administration  of  HCB  induced  hepatomas  and liver  hemangioendo-
 theliomas  in male  and female Syrian Golden hamsters,  and hepatomas
 in male and female Swiss mice.   The  data from the hamster study were
 reported in  detail  for evaluation,  whereas the mouse study was only
 described  in an abstract.   In the  hamster study,  there  was  a sta-
 tistically significant incidence of hepatomas in males fed 50, 100,
 and 200 ppm  (p = 7.5 X 10"7, 2.45 X 10'15,  and 1.30 X 10~19,  respec-
 tively) , and of liver hemangioendtheoliomas  in males fed 100  and
 200 ppm (p = 4.5 X 10~3 and 4.0 X 10~6,  respectively).   There was a
 statistically significant  incidence of hepatomas in females fed 50,
 100,  and 200 ppm (p  = 7.5  X  10~7, 2.0 X 10~8 and  3.05  X  10~19,  re-
 spectively) , and of  liver  hemangioendotheliomas in  females fed  200
 ppm (p = 0.026).
     The water quality criterion for HCB is based on the  induction
 of hepatomas in male Syrian Golden  hamsters given daily oral  doses
of 50, 100, or 200  ppm (Cabral, et al. 1977).  The concentration  of
HCB in  drinking water calculated  to limit human  lifetime cancer
risk from HCB to less than 10~5 is 7.2 ng/1.
                              C-171

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                    SUMMARY  OF  PERTINENT DATA


     The water quality  criterion for  HCB is based on the  induction

of hepatomas in male Syrian Golden hamsters given a daily  oral dose

of 4, 8, or 16 mg/kh for 80 weeks  (Cabral,  et al. 1977).   The hepa-

toma incidences, in the treated  and control groups are  shown  in  the

table below.  The criterion was  calculated  from  the following para-

meters:

              Dose                     Incidence
          mg/kg/day)           (no. reporting/no, tested)

                0                         0/40

                4                        14/30

                8                        26/30

               16                        49/57



      le =  560  days             w = 0.100 kg

      Le =  560  days             R « 8,960 I/kg

      L   =  560  days


      With   these  parameters  the  carcinogenic   potency factor  for

 humans, qL*,  is 1.688  (mg/kg/day)'1.   The resulting water  concen-

 tration of HCB  calculated to  keep  the individual lifetime  cancer

 risk below 10    is 7.2 ng/1.
                                C-172      «O& GOVERNMENT PRINTING OFFICE: 19 80 720-016/4373 1-3

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