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NOTICES
This document has been reviewed by the Criteria and Standards Division, Office
of Water Regulations and Standards, U.S. Environmental Protection Agency, and
approved for publication.
Mention of trade names or commercial products does not constitute endorsement
or recommendation for use.
This document is available to the public through the National Technical
Information Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161.
11
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FOREWORD
Section 304(a)(l) of the Clean Water Act requires the Administrator of
the Environmental Protection Agency to publish water quality criteria that
accurately reflect the latest scientific knowledge on the kind and extent of
all identifiable effects on health and welfare that might be expected from the
presence of pollutants in any body of water. Pursuant to that end, this
document proposes water quality criteria for the protection of aquatic life.
These criteria do not involve consideration of effects on human health.
This document is a draft, distributed for public review and comment.
After considering all public comments and making any needed changes, EPA will
issue the criteria in final form, at which time they will replace any
previously published EPA aquatic life criteria for the same pollutant.
The term "water quality criteria" is used in two sections of the Clean
Water Act, section 304(a)(l) and section 303(c)(2). In section 304, the term
represents a non-regulatory, scientific assessment of effects. Criteria
presented in this document are such scientific assessments. If water quality
criteria associated with specific stream uses are adopted by a State as water
quality standards under section 303, then they become maximum acceptable
pollutant concentrations that can be used to derive enforceable permit limits
for discharges to such waters.
Water quality criteria adopted in State water quality standards could
have the same numerical values as criteria developed under section 304.
However, in many situations States might want to adjust water quality criteria
developed under section 304 to reflect local environmental conditions before
incorporation into water quality standards. Guidance is available from EPA to
assist States in the modification of section 304(a)(l) criteria, and in the
development of water quality standards. It is not until their adoption as
part of State water quality standards that the criteria become regulatory.
Martha G. Prothro
Director
Office of Water Regulations and Standards
111
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ACKNOWLEDGMENTS
Daniel J. Call
(freshwater author)
University of Wisconsin-Superior
Superior, Wisconsin
Charles E. Stephan
(document coordinator)
Environmental Research Laboratory
Duluth, Minnesota
IV
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CONTENTS
Notices [[[ i i
Foreword [[[ i i i
Acknowledgments [[[ iv
Tables [[[ vi
Introduction [[[ 1
Acute Toxicity to Aquatic Animal?. ....................................... 2
Chronic Toxicity to Aquatic Animals ...................................... 3
Toxicity to Aquatic Plants ............................................... 3
Bioaccumulat i on [[[ 4
Other Data [[[ 6
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TABLES
Page
1. Acute Toxicity of Hexachlorobenzene to Aquatic Animals 11
2. Chronic Toxicity of Hexachlorobenzene to Aquatic Animals 14
3. Toxicity of Hexachlorobenzene to Aquatic Plants 15
4. Bioaccumulation of Hexachlorobenzene by Aquatic Organisms 16
5. Other Data on Effects of Hexachlorobenzene to Aquatic Organisms 19
VI
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Introduction
Hexachlorobenzene (HCB) has been detected in environmental samples from
around the world in recent years, and is recognized as a global pollutant
(Ballschraiter and Zell 1980; Brevik 1981; Brunn and Manz 1982; Chovelon et
al. 1984; Galassi et al. 1981; Jan and Malnersic 1980; Johnson et al. 1974;
Kaiser 1977; Kuehl et al. 1983,1984; Laska et al. 1976; Niimi 1979;
Paasivirta et al. 1981; Sackmauerova et al. 1977; Schmitt et al. 1985;
Skaftason and Johannesson 1982; Tsui and McCart 1981; Veith et al. 1979b).
Contamination of Lake Ontario with heiachlorobenzene and other chlorobenzenes
has occurred since 1915, but at a much greater rate in the past three decades
(Oliver and Nicol 1982). HCB is soluble in water to about 6 ng/L
(Abernethy et al. 1986; Metcalf et al. 1973), and its concentrations in
rivers and lakes are generally reported in terms of ng/L (Niimi and Cho
1981). It has an octanol/water partition coefficient of 6.18 (Neely et al.
1974). The common occurrence of HCB in tissues of aquatic organisms can be
attributed to its high affinity for lipids and its persistence in the
environment (Callahan et al. 1979). Connor (1984) estimated that HCB
contributed little to increased risk of cancer associated with consumption of
freshwater fish from contaminated water.
HCB is used as a seed grain fungicide, a peptizing agent in the
production of nitroso and styrene rubber for tires, a wood preservative, a
porosity controller in the manufacture of graphite electrodes, a fluxing
agent in aluminum smelting, and in pyrotechnics manufacture (Courtney 1979;
Quinlivan et al. 1975/1977; U.S. EPA 1980; Young et al. 1980). HCB wastes
also originate from the production of certain pesticides (Cleveland et al.
1982), chlorinated solvents, vinyl chloride monomers, and chlorine or sodium
chlorate when produced electrolytically (Quinlivan et al. 1975/77). Some of
the commercial formulations of HCB contain toxic impurities, including
1
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pentachlorobenzene, polychlorinated dibenzo-p-dioxins, and polychlorodi-
benzofurans (Villanueva et al. 1974).
This document does not contain information concerning the effects of HCB
on saltwater species and their uses because adequate data and resources were
not available. An understanding of the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" (Stephan et al. 1985), hereinafter referred to as the Guidelines,
and the response to public comment (U.S. EPA 1985a), is necessary in order to
evaluate the following text, tables, and calculations. Results of such
intermediate calculations as recalculated LCSQs and Species Mean Acute Values
are given to four significant figures to prevent roundoff error in subsequent
calculations, not to reflect the precision of the value. The criteria
presented herein supersede previous information on HCB (U.S. EPA 1980)
because these criteria were derived using additional information. The latest
comprehensive literature search for information for this document was
conducted in February, 1988. Some more recent information was also
included. Data in the files of the U.S. EPA's Office of Pesticide Programs
concerning the effects of hexachlorobenzene on aquatic organisms and their
uses have not been evaluated for possible use in the derivation of aquatic
1i fe criteria.
Acute Toxicity to Aquatic Animals
Data that nay be used, according to the Guidelines, in the derivation of
a freshwater Final Acute Value for HCB are presented in Table 1. The only
experimentally determined acute values were at concentrations that are at
least 1000 times the solubility limit. Although the quantitative value of
these measurements might be suspect, they do indicate that concentrations
near solubility should not cause acute toxicity. Some acute exposures were
2
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continued for longer than 96 hr with no resultant mortalities. For example,
adult crayfish, Procambarus clarki. were unaffected over a period of 20 days
at a mean HCB concentration of 27.3 ng/L, and largemouth bass, Micropterus
salmoides. were unaffected over 10 days at 25.8 ng/L (Laseter et al. 1976;
Laska et al. 1978). Because an acute value is not available for any species
at or below solubility, a freshwater Final Acute Value cannot be determined.
If one could be determined, it would be higher than 6 ng/L, which was
reported by Metcalf et al. (1973) to be the solubility of HCB in water.
Chronic Toxicity to Aquatic Animals
The available data that are useable according to the Guidelines
concerning the chronic toxicity of HCB are presented in Table 2. Rainbow
trout, Salmo gai rdneri . were exposed to HCB in a 90-day early life-stage test
(Spehar 1986). No adverse effects on hatching, survival, or growth were
observed at the highest tested concentration of 3.68 /^g/L (Table 2).
Similarly, fathead minnows, Pimephales promelaa. were not affected at
exposures up to 4.8 jug/L in a 32-day early life-stage test (Ahmad et al.
1984; Carlson and Kosian 1987). Survival at hatch, survival at 32 days, and
wet weight were the same at all exposure concentrations as in the control.
In a 7-day life-cycle test with the cladoceran, Ceriodaphnia dubia. HCB
did not cause any measurable effect upon survival or reproduction at
concentrations as high as 7.0 ng/L (Spehar 1986). Because a chronic value
is not available for any species, no Final Chronic Value can be calculated.
If one could be calculated, it would be higher than 3.68
Toxicity to Aquatic Plants
Geyer et al. (1985) reported that the green alga, Scenedesmus
subspicatus. was not affected in 4 days by HCB at a concentration of
3
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10 ng/L (Table 3). A 4-day exposure of Selenastrum capricornutum to
30 ng/L caused a 12% inhibition of population growth (Calamari et al.
1983). A Final Plant Value, as defined in the Guidelines, cannot be obtained
because no test in which the concentrations of HCB were measured and the
endpoint was biologically important has been conducted with an important
aquatic plant species.
Bioaccumulation
HCB concentrations in tissues of aquatic organisms appear to equilibrate
very slowly with concentrations in the water. Uptake via the food might be
more important than direct uptake from water for top carnivores, particularly
in waters where HCB concentrations are very low (Niimi and Cho 1980). Oliver
and Niimi (1983) reported that equilibration had not yet occurred after a
119-day exposure of rainbow trout to HCB at a concentration of
0.00032 Mg/L. at which time the bioconcentration factor was 12,000 (Table
4). An exposure to 0.008 M8/L resulted in a bioconcentration factor of
20,000 with the trout. HCB residues in fathead minnows continually increased
in 28-day tests to concentrations that were from 32,000 to 39,000 times
greater than the concentrations of HCB in water (Kosian et al. 1981).
Following a rapid uptake during the first week of exposure, HCB residues in
fathead minnows gradually increased through 115 days (Veith et al. 1979a).
HCB residues in the minnows were from 16,200 to 45,700 times greater than
concentrations in water following exposures of 32 to 115 days. Carlson and
Kosian (1987) obtained similar results with a BCF of 22,000 for fathead
minnows exposed for 32 days.
Green sunfish, Lepomis cyanellus. bioconcentrated HCB approximately
22,000 times, whereas fingerling rainbow trout bioconcentrated HCB
considerably lower at 5,500 times (Veith et al. 1979a). In a test with a
4
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mixture of chlorinated benzenes, a BCF of 15,660 was obtained for HCB with
the guppy, Poeci1ia reticulata (Konemann and Van Leeuwen 1980). An apparent
steady-state between HCB residues in guppy tissue and concentration in the
water occurred within 7 days.
Coho salmon (Oncorhvnchus kisutch) and green sunfish accumulated HCB when
fed contaminated food (Leatherland and Sonstegard 1982; Sanborn et al.
1977). Crayfish (Procambarus clarki) exposed at a contaminated field site to
a mean HCB concentration of 74.9 ng/L had a mean whole-body concentration
factor of 1,464 (Laseter et al. 1978). Fox et al. (1983) found a direct
correlation between HCB concentrations in surficial sediments of Lake Ontario
and HCB concentrations in oligochaetes that inhabit the sediments. They also
reported that HCB accumulation was greater at the higher trophic levels.
HCB and other lipophilic, persistent chlorinated organics are passed from
parent to offspring via the yolk and oil of the fish egg. Survival and
growth of larvae might be affected by this route of uptake in some fish
species (Niimi 1983; Westin et al. 1985).
Elimination of HCB from salmonids occurs slowly. Norheim and Roald
(1985) determined the half-life of HCB in rainbow trout that had been
injected intraperitoneally and maintained at a water temperature of 7°C to be
81, 146, and 139 days in liver, fat, and muscle, respectively. Niimi and Cho
(1981) reported the biological half-life of HCB to be 224 to 770 days in
rainbow trout fed HCB and maintained at a water temperature of 15°C. In a
subsequent study (Niimi and Palazzo 1985), HCB-fed rainbow trout held at 12
and 18°C had half-lives of HCB of 173 and 198 days, respectively, whereas
fish held at 4°C did not eliminate HCB. Thus, it appears that HCB is highly
persistent in the tissues of fish inhabiting cold waters.
Crayfish that were exposed to 74.9 jug/L f°r 10 days under field
conditions eliminated about 15% of the HCB after three days in clean water.
5
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Males eliminated 47% and females 64% after 25 days (Laseter et al. 1976).
Largemouth bass eliminated from 73.1 to 91.4% of the whole body HCB residue
during 13 days in clean water (Laseter et al. 1976).
McKim et al. (1985) studied the uptake efficiency of HCB and 13 other
chemicals by rainbow trout gills. HCB was among the chemicals most
efficiently removed from water by gills. However, it was not much more
efficiently removed than several chlorophenols which, by comparison to HCB,
have a reduced capacity for bioaccumulation. The high accumulation levels of
HCB can be explained by a slow rate of conversion to water soluble
metabolites and subsequent elimination, combined with an efficient uptake
from water and food due to its high n-octanol water partition coefficient of
6.18.
No U.S. FDA action level or other maximum acceptable concentration in
tissue for HCB. as defined in the Guidelines, is available for HCB.
Therefore, a Final Residue Value cannot be calculated.
Other Data
Additional data on the lethal and sublethal effects of HCB on freshwater
species are presented in Table 5. A green alga, Anki strodesmus falcatus. was
unaffected in a 4-hr exposure to 3 ng/L (Wong et al. 1984). Daphni a magna
were unaffected after 24 to 48 hr at concentrations that exceeded the
solubility of HCB in water (Calamari et al. 1983; Sugatt et al. 1984).
Exposure of Daphnia magna to HCB concentrations in excess of its solubility
in water for a period of 14 days resulted in an EC50 of 16 /^g/L and a
reduced instantaneous growth rate at 23 ^g/L (Calamari et al. 1983). Due
to insufficient deaths, LCSOs could not be determined for rainbow trout
exposed to 30 ng/L for 24 hr (Calamari et al. 1983) or for guppies exposed
to 730 ng/L for 14 days (Konemann 1979,1981). However, largemouth bass
6
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had liver and kidney damage after 10 days of exposure to 3.5 /ig/L (Laseter
et al. 1976).
Unused Data
Some data on the effects of HCB on aquatic organisms and their uses were
not used because the studies were conducted with species that are not
resident in North America (e.g., Hattori et al. 1984; Sugiura et al. 1984).
Data were not used when HCB was a component of a mixture or wettable powder
(e.g., Gjessing et al. 1984; Mayer and Ellersieck 1986; Poels et al. 1980).
Chiou (1985), Chiou et al. (1977), Davies and Dobbs (1984), Glass et al.
(1977), Kaiser et al. (1984), Klein et al. (1984), Koch (1982a,b), Matsuo
(1980,1981), Oliver (1984a), and Sabljic and Protic (1982) compiled data from
other sources.
Tests were not used when only physiological effects or enzyme activity
were measured (e.g., Gluth and Hanke 1984; Hanke et al: 1983; Law and Addison
1981). Schmidt-Bleek et al. (1982) did not adequately describe the methods
used for the toxicity tests. Toxicity data were not used when the test was
conducted in distilled water (e.g., Abernethy et al. 1986; Bobra et al.
1985).
Studies of HCB residues in aquatic organisms were not used when there
were insufficient tissue residue data or accompanying data on HCB
concentrations in the water to determine a bioconcentration or
bioaccumulation factor (e.g., Atuma and Eigbe 1985; Clark et al. 1984;
DeVault 1985; Evans et al. 1982; Giesy et al. 1986; Glass 1975; Heida 1983;
Jaffe and Hites 1986; Kaiser 1982; Kaiser and Valdmanis 1978; Keck and
Raffenot 1979; Kuehl et al. 1980,1981; Niimi 1979; Norstrom et al. 1978;
Paasivirta et al. 1983a,b; Pennington et al. 1982; Schmitt et al. 1985;
7
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Schuler et al. 1985; Swain 1978; Veith et al. 1977,1981; Wiemeyer et al.
1978; Zitko 1971). Results of studies in which HCB was administered by
gavage or injection were not used (e.g., Ingebrigtsen and Skaare 1983; Safe
et al. 1976). Studies using radiolabeled HCB in which only radioactivity was
measured in the water or in exposed organisms were not used (e.g., Freitag
1987; Freitag et al. 1982,1984,1985; Geyer et al. 1981,1984; Schauerte et al.
1982).
Results of laboratory bioaccumulation tests were not used when the test
was not flow-through or renewal (e.g., Bel luck and Felsot 1981; Korte et al.
1978; Lu and Metcalf 1975; Metcalf et al. 1973; Zitko and Hutzinger 1976) or
when the concentration of HCB in the test solution was not adequately
measured (e.g., Isensee 1976; Isensee et al. 1976). A study of the
bioaccumulation of HCB by organisms inhabiting contaminated sediment was not
used (Oliver 1984b).
Results of bioconcentration tests were not used when the duration of
exposure was not specified (e.g., Neely et al. 1974). Studies on the
dynamics and transport of HCB at the sediment-water interface were not used
(Baker et al. 1985; Karickhoff and Morris 1985).
Summary
Hexachlorobenzene at a concentration of 6 Mg/L has been shown not to
cause acute toxicity to any tested freshwater species and 3.68 ^g/L has
been shown not to cause chronic toxicity to any tested species. Freshwater
plants were .not affected at 10 ng/L. In a 14-day test, 16 ng/L caused
a 50% reduction in the fertility of Daphnia magna. Bioconcentration factors
ranged from 5,500 to 45,700 in tests with freshwater fish.
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National Criteria
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important species
is very sensitive, freshwater aquatic organisms and their uses should not be
affected unacceptably if the four-day average concentration of hexachloro-
benzene does not exceed 3.68 ng/L more than once every three years on the
average and if the one-hour average concentration does not exceed 6.0 pg/L
more than once every three years on the average. The only adverse effects
that have been observed in tests on hexachlorobenzene include a 50% reduction
in fertility of Daphni a magna at 16 ng/L and organ damage to largemouth
bass at 3.5 pg/L. However, histopathology data are not used to develop
criteria and the available data do not justify establishing criteria
different from those given above.
ImplementatI on
As discussed in the Water Quality Standards Regulation (U.S. EPA 1983a)
and the Foreword to this document, a water quality criterion for aquatic life
has regulatory impact only after it has been adopted in a state water quality
standard. Such a standard specifies a criterion for a pollutant that is
consistent with a particular designated use. With the concurrence of the
U.S. EPA, states designate one or more uses for each body of water or segment
thereof and adopt criteria that are consistent with the use(s) (U.S. EPA
1983b,1987). In each standard a state may adopt the national criterion, if
one exists, or, if adequately justified, a site-specific criterion.
Site-specific criteria may include not only site-specific criterion
concentrations (U.S. EPA 1983b), but also site-specific, and possibly
pollutant-specific, durations of averaging periods and frequencies of allowed
excursions (U.S. EPA 1985b). The averaging periods of "one hour" and "four
9
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days" were selected by the U.S. EPA on the basis of data concerning how
rapidly some aquatic species react to increases in the concentrations of some
aquatic pollutants, and "three years" is the Agency's best scientific
judgment of the average amount of time aquatic ecosystems should be provided
between excursions (Stephan et al. 1985; U.S. EPA 1985b). However, various
species and ecosystems react and recover at greatly differing rates.
Therefore, if adequate justification is provided, site-specific and/or
pollutant-specific concentrations, durations, and frequencies may be higher
or lower than those given in national water quality criteria for aquatic
life.
Use of criteria, which have been adopted in state water quality
standards, for developing water quality-based permit limits and for designing
waste treatment facilities requires selection of an appropriate wasteload
allocation model. Although dynamic models are preferred for the application
of these criteria (U.S. EPA 1985b), limited data or other considerations
might require the use of a steady-state model (U.S. EPA 1986). Guidance on
mixing zones and the design of monitoring programs is also available (U.S.
EPA 1985b,1987).
10
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