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.
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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.
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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.
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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
-------
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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Chemical
1.2,4-trichloro-
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1 =
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09
I
Table I. (Continued)
Species Method*
Sheepshead minnow, S, U
Cyprlnodon yarlegatus
Chemical
pent ach lor o-
benzene
LC50/EC50 Species Acute
-------
Table 2. Chronic values for chlorinated benzenes
CXI
I
h-1
NJ
Species Method*
Fathead minnow, ELS
Plmephales promelas
Fathead minnow, ELS
Plntephales promelas
Fathead minnow, ELS
Pimephales promelas
Chemical
FRESHWATER SP
1, 2,4- trlch loro-
benzene
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.
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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.
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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.
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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,
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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.
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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.
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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.
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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
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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
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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
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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
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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.
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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
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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).
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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
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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
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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 (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.
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REFERENCES
American Conference of Governmental Industrial Hygienists. 1971.
Documentation of the Threshold Limit Values for Substances in
Workroom Air. 3rd ed.
American Industrial Hygiene Association. 1964. Chlorobenzene.
Am. Ind. Hyg. Assoc. Jour. 25: 97.
Brodie, B.B., et al. 1971. Possible mechanism of liver necrosis
caused by aromatic organic compounds. Proc. Natl. Acad. Sci.
68: 160.
Carlson, G.P. and R.G. Tardiff. 1976. Effect of chlorinated ben-
zenes on the metabolism of foreign organic compounds. Toxicol.
Appl. Pharmacol. 36: 383.
Dunaeveskii, G.A. 1972. Functional condition of circulatory or-
gans in workers employed in the production of organic compounds.
Gig. Tr. Prof. Zabol. 16: 48.
Erisman, H. and M. Goldman. 1975. Identification of organic com-
pounds in textile plant effluents. Presented at the First Chemi-
cal Congress of the North American Continent, Mexico City, Mexico,
November 30 - December 5.
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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
the maximum permissible concentrations of benzene and monochloro-
benzene. Jour. Hyg. Epidemiol. Microbiol. Immunol. 4: 223.
Gabor, S., et al. 1962. Certain biochemical indexes of the blood
in workers exposed to toxic substances (benzene, chlorobenzene,
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Girard, R., et al. 1969. Serious blood disorders and exposure to
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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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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
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C-24
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C-25
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C-26
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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
-------
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.
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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.
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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).
-------
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
-------
n
i
Proposed Pathways for the Biotransformation of Trichlorobenzene
Isomers Through Arene Oxide Intermediates \j
,' -. , , Source: Kohli, et al. 1976
-------
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
-------
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
-------
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
-------
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
-------
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
-------
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
-------
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-
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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
microsomal enzymes following treatment with 1,2,4-trichloroben-
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
derivatives of benzene with special reference to o-dichloroben-
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Carlson, G.P. 1978. Induction of cytochrome P-450 by halogenated
benzenes. Biochem. Pharmacol. 27: 361.
Carlson, G.P. and R.G. Tardiff. 1976. Effect of chlorinated ben-
zenes on the metabolism of foreign organic compounds. Toxicol.
Appl. Pharmacol. 36: 383.
Coate, W.B., et al. 1977. Chronic inhalation exposure of rats,
rabbits and monkeys to 1,2,4-trichlorobenzene. Arch. Environ.
Health 32: 249.
Egyankor, K.B. and C.S. Franklin. 1977. Interaction of the tri-
chlorobenzenes with cytochrome P-450. Biochem. Soc. Trans.
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Gotto, M., et al. 1972. Hepatoma formation in mice after admin-
istration of high doses of hexachlorocyclohexane isomers. Chem-
osphere. 1: 279.
Gurfein, L.W. and Z.K. Parlova. 1962. The Limit of Allowable
Concentration of Chlorobenzenes in Water Basins. in; B.S. Levine
(ed.), USSR Literature on Water Pollution. Dept. Commer., Wash-
ington, D.C.
C-47
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Haas, J.M., et al. 1974. Environmental considerations concerning
the selection of dye carrier solvents. Presented at the 1974 Am.
Assoc. Textile Chem. Colourists Natl. Tech. Conf. October 9-11.
Hintz, M. 1962. Pollution characteristics of 1,2,4-trichloroben-
zene. Dow Chemical Co., Midland, Michigan. (Unpub.)
Johnson, K. 1980. Memorandum to D.W. Kuehl. U.S. EPA.
March 10.
Kohli, I., et al. 1976. The metabolism of higher chlorinated
benzene isomers. Can. Jour. Biochem. 54: 203.
Kujawa, M., et al. 1977. On the Metabolism of Lindane. Proc.
1st Int. Symp. Environ. Pollut. Human Health.
Parke, D.V. and R.T. Williams. 1960. Studies in detoxication.
Metabolism of halobenzenes: (a) Penta- and hexachlorobenzene: (b)
Further observations on 1,3,5-trichlorobenzene. Biochem. Jour.
74: 1.
Porter, E.M. and J.J. Snider. 1974. 30-day biodegradability of
textile chemicals and dyes. Presented at 1974 Am. Assoc. Textile
Chem. Colourists Natl. Tech. Conf. October 9-11.
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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.
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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.
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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
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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.
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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.
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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.
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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.
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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.
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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.
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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.
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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.
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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.
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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.
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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
-------
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
-------
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
-------
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
-------
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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halogenated benzenes (pentachloro-, pentachloronitro-, and hexa-
bromo-) in rats. Environ. Physiol. Biochem. 5: 328.
Yang, R.S.H., et al. 1975. Chromatographic methods for the an-
alysis of hexachlorobenzene and possible metabolites in monkey
fecal samples. Jour. Assoc. Off. Anal. Chem. 56: 1197.
Yang, R.S.H., et al. 1978. Pharmacokinetics and metabolism of
hexachlorobenzene in the rat. Jour. Agric. Food Chem. 26: 1076.
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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
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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).
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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.
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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.
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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
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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
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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.
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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
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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).
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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
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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
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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
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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
C-117
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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
C-127
-------
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
-------
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
-------
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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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.
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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
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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 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
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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.
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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.
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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.
-------
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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