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NOTICES
This document has been reviewed by the Environmental Research
Laboratories, Duluth, MN and Narragansett, RI, Office of Research and
Development and the Health and Ecological Criteria Division, Office of Science
and Technology, 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) (1) of the Clean Water Act of 1977 (P.L. 95-217) 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,
including ground water. This document is a revision of proposed criteria
based upon consideration of comments received from other federal agencies,
state agencies, special interest groups, and individual scientists. Criteria
contained in this document replace any previously published EPA aquatic life
criteria for the same pollutant(s).
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 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, they represent maximum acceptable pollutant
concentrations in ambient waters within that state that are enforced through
issuance of discharge limitations in NPDES permits. 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 modify water quality criteria developed under section 304 to
reflect local environmental conditions and human exposure patterns.
Alternatively, states may use different data and assumptions than EPA in
deriving numeric criteria that are scientifically defensible and protective of
designated uses. It is not until their adoption as part of state water
quality standards that criteria become regulatory. Guidelines to assist the
states and Indian tribes in modifying the criteria presented in this document
are contained in the Water Quality Standards Handbook (December 1983). This
handbook and additional guidance on the development of water quality standards
and other water-related programs of this Agency have been developed by the
Office of Water.
This document, if finalized, would be guidance only. It would not
establish or affect legal rights or obligations. It would not establish a
binding norm and would not be finally determinative of the issues addressed.
Agency decisions in any particular situation will be made by applying the
Clean Water Act and EPA regulations on the basis of specific facts presented
and scientific information then available.
Tudor T. Davies
Director
Office of Science and Technology
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ACKNOWLEDGMENTS
Daniel J. Call
(freshwater author)
University of Wisconsin-Superior
Superior, Wisconsin
David J. Hansen
(saltwater author)
Environmental Research Laboratory
Narragansett, Rhode Island
Robert L. Spehar
(document coordinator)
Environmental Research Laboratory
Duluth, Minnesota
Suzanne M. Lussier
(saltwater coordinator)
Environmental Research Laboratory
Narragansett, Rhode Island
IV
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CONTENTS
Page
Notices 11
Foreword , 0.11.
Acknowledgments i-v
Tables vi-
Introduction 1
Acute Toxicity to Aquatic Animals 3
Chronic Toxicity to Aquatic Animals 4
Toxicity to Aquatic Plants ... 6
Bioaccumulation 6
Other Data 7
Unused Data 8
Summary 8
National Criteria 9
Implementation 10
References
24
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1.
2.
3.
4.
5.
6.
TABLES
Ranked Genus Mean Acute Values with Species Mean Acute-Chronic
Other Data on Effects of 2 , 4-Dimethylphenol on Aquatic Organisms .
Page
12
15
17
20
21
22
vi
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Introduction '
2,4-Dimethylphenol (2,4-DMP) is a naturally occurring, substituted
phenol derived from the cresol fraction of petroleum or coal tars by
fractional distillation and extraction with aqueous alkaline solutions (Gruse'
and Stevens 1942; Lowry 1963; Rudolfs 1953; U.S. EPA 1976). 2,4-DMP, also
known as'l-hydroxy-2,4-dimethylbenzene, m-xylenol, 2,4-xylenol or m-4-xylenol,
has the empirical formula C8H,0O (Weast 1972). 2,4-DMP is used commercially as
an important chemical feedstock or constituent for the manufacture of a wide
range of commercial products for industry and agriculture. It is also used in
the manufacture of phenolic antioxidants, disinfectants, solvents,
Pharmaceuticals, insecticides, fungicides, plasticizers, rubber chemicals,
polyphenylene oxide, wetting agents, and dyestuffs; and is an additive or
constituent of lubricants, gasolines, and cresylic acid. No direct commercial
application for 2,4-DMP appears to exist at present.
Five other positional isomers of dimethylphenol or xylenol exist and
include 2,3-, 2,5-, 2,6-, 3,4-, and 3,5-dimethylphenol. Since these isomers
result from the different positioning of the two methyl groups on the phenol
ring, they are referred to as positional isomers. As would be expected, there
are variations in their physical, chemical, and biological properties.
2,4-DMP has a molecular weight of 122.17 and in its normal state exists
as a colorless, crystalline solid (Bennet 1974; Weast 1972). It has a melting
point of 27 to 28°C, a boiling point of 210°C (760 mm Hg), a vapor pressure of
1 mm Hg at 52.8°C, and a density of 0.9650 at 20°C (Bennet 1974; Jordan 1954;
Weast 1972). 2,4-DMP is slightly soluble in water and, as a weak acid (pK. of
10.6), is also soluble in alkaline solutions (Sober 1970). 2,4-DMP readily
dissolves in organic solvents such as alcohol and ether (Weast 1972).
A large number of products utilize 2,4-DMP as a feedstock or
constituent. Hence, disposal of chemical and industrial process wastes and
distribution from normal product applications represent feasible routes for
entry of 2,4-DMP into the environment. Examples of the latter route include
pesticide applications, asphalt and roadway runoff, and the washing of dyeJ
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materials (U.S. EPA 1975).
Information regarding the concentration, persistence, fate and effects
of 2,4-DMP in the environment is limited. However, its presence in petroleum
distillate fractions and coal tars, together with its use as a chemical
feedstock or constituent for the manufacture of many products, clearly
indicate the potential for both point and nonpoint source water contamination.
2,4-DMP has been detected in the effluent from coal gasification plants and in
finished drinking water (Shackelford and Keith 1976). The concentration of
2,4-DMP in sediments collected near the Los Angeles County Sanitation
District's sewage outfall located off of Palos Verdes, California, was 40
/jg/kg (Schwartz et al. 1985). It was below detection limits at six other
stations located further away from the outfall (Schwartz et al. 1985).
It is inferred that 2,4-dimethylphenol will undergo some photolysis in
well-aerated surface waters in spite of its apparent persistence (Callahan et
al. 1979). Richards and Shieh (1986) rank it as a persistent, volatile and
accumulative chemical. Callahan et al. (1979), on the other hand, indicate
that there should be little tendency for it to volatilize from water. The
complete biodegradation of 2,4-DMP has been reported to occur in approximately
two months, although the conditions were not stated (Rodd 1952).
2,4-DMP can be oxidized to form pseudoquinone (Rodd 1952). However, the
conditions required for this reaction generally are not found in the
environment. 2,4-DMP reacts with aqueous alkaline solutions to form the
corresponding salt. Such salts are readily soluble in water, provided that an
alkaline pH is maintained. The free position on the aromatic ring, ortho to
the hydroxyl group, may be alkylated (Kirk and Othmer 1964) or halogenated
(Rodd 1952). However, such reactions have not been reported to occur under
normal environmental conditions.
2,4-DMP causes a detectable odor in water when present at relatively low
concentrations (Buikema et al. 1979). Hoak (1957) reported an odor threshold
of 55.5 jug/L.
All concentrations reported herein are expressed as 2,4-DMP. The
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criteria presented herein supersede previous aquatic life water quality
criteria for 2,4-DMP (U.S. EPA 1980) because these new criteria were derived
using improved procedures and additional information. Whenever adequately
justified, a national criterion may be replaced by a site-specific criterion
(U.S. EPA 1983a) that may include not only site-specific concentrations (U.S.
EPA 1983b) but also site-specific frequencies of allowed excursions (U.S. EPA
1985).
A comprehension of the "Guidelines for Deriving Numerical National Water
Quality Criteria for the Protection of Aquatic Organisms and Their Uses"
(Stephan et al. 1985), hereafter referred to as the Guidelines, and the
response to public comment (U.S. EPA 1985) is necessary to evaluate the
following text, tables, and calculations.
The latest comprehensive literature search for information for this
document was conducted in September, 1992. Some more recent information is
included.
Acute Toxicitv to Aquatic Animals
The data that are available according to the Guidelines concerning the
acute toxicity of 2,4-DMP are presented in Table 1. Freshwater Species Mean
Acute Values were calculated according to the Guidelines as geometric means of
the available acute values. Of the 12 freshwater genera for' which mean acute
values are available, the most sensitive genus, Ceriodaphnia, is about 20
times more sensitive than the most resistant, Lumbriculus. Both the most
sensitive and most resistant genera are invertebrates. Fish were intermediate
in sensitivity with a range in Genus Mean Acute Values from 6,300 ^g/L to
19,300 /^g/L. The freshwater Final Acute Value for 2,4-DMP was calculated to
be 2,670 /jg/L using the procedure described in the Guidelines and the Genus
Mean Acute Values in Table 3. The Final Acute Value is lower than the lowest
available freshwater Species Mean Acute Value.
The acute toxicity of 2,4-DMP to resident North American saltwater
animals has been determined with six species of invertebrates and three
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species of fish (Table 1). The acute toxicity of 2,4-DMP differs by a factor
of 42 for saltwater animals, with acute values based on 96-hr LCSOs ranging
from 1,320 pg/L for juvenile inland silversides, Menidia bervllina, to 55,900
pg/L for adult archiannelid worms, Dinophilus gyrociliatus (Thursby and Berry
1987a). Mortality increased with increased duration during 96-hr tests with
six of seven species for which daily survival data are available. The
saltwater Final Acute Value, based on nine Genus Mean Acute Values, is 548.8
fjg/L. The Final Acute Value is lower than the lowest available saltwater
Species Mean Acute Value.
Chronic Toxicitv to Aquatic Animals
The data that are available according to the Guidelines concerning the
chronic toxicity of 2,4-DMP are presented in Table 2. The freshwater
cladoceran, Ceriodaphnia dubia, was tested in a 7-day life-cycle chronic
exposure (Spehar 1987). Mean 2,4-DMP concentrations were 210, 470, 810, 1,870
and 3,410 /^g/L. Survival was slightly reduced, but not significantly at these
concentrations. However, young production was significantly lower (p <0.05)
than controls at the two highest concentrations, with reductions of about 64
and 90 percent in young produced at 1,870 and 3,410 jug/L, respectively. The
chronic limits for this test were between 810 and 1,870 pg/L, which results in
a chronic value of 1,230 pg/L (Table 2). Division of the companion acute
value for Ceriodaphnia of 3,340 pg/L by the chronic value results in an
acute-chronic ratio of 2.715.
Fathead minnows (Pimephales promelas) were exposed to 2,4-DMP in a
32-day early life-stage test at concentrations of 900, 1,360, 1,970, 3,110,
and 5,130 pig/L (Holcombe et al. 1982). The percentage of normal appear ir.v-
larvae at hatch was similar for each exposure as in the control. Survival of
juvenile fish was reduced at the highest exposure and weight was reduced
(S15.6%) at the two highest exposures. The control fish at the end of the
study averaged 72.6 mg in wet weight which is low for fathead minnows of this
age in a toxicity test. Based upon growth, the chronic limits were 1,97C *:: i
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3,110 /jg/L. The chronic value is 2,475 pg/L. Division of the companion acute
value of 17,000 /jg/L by this value results in an acute-chronic ratio of 6.869.
LeBlanc (1984) published an early life-stage study in which fathead
minnows were exposed to 750, 1,500, 3,200, 7,400, and 15,000 A/g/L 2,4-DMP. No
fish survived at 15,000 pg/L, and only 12 percent survived at 7,400 pg/L.
Length and weight were significantly less than controls at 7,400 and 3,200
Hg/L. Based upon growth, the chronic limits were 1,500 and 3,200 ^g/L. The
chronic value is 2,200 A/g/L. No corresponding acute value is available to
determine an acute-chronic ratio.
Fathead minnows were exposed to 2,4-DMP in a third 32-day life-stage
test at concentrations of 398, 605, 966, 1,573, 2,580, and 4,052 /jg/L (Russom
1993). The study was conducted in the same laboratory as the Holcombe et al.
(1982) study, but ten years later. Significant negative effects were observed
for percentage of normal appearing larvae at hatch, and in survival at the end
of the study at the highest exposure concentration of 4,052 A605 i^gfL. Wet weight and total length were
reduced 10.4% and 4.8%, respectively, at the exposure concentration of 605
pg/L. The mean wet weight of the control fish was 144 mg. Based upon growth,
the chronic limits were 398 and 605 /^g/L. The chronic value is 491 ^iqfL. A
corresponding acute value for this test was not measured; therefore, an acute
chronic ratio cannot be calculated.
The chronic toxicity of 2,4-DMP has been determined in an early
life-stage toxicity test with the saltwater inland silverside, Menidia
bervllina (Thursby and Berry 1987b). Ninety percent of the embryos exposed to
722 pg/L diea prior to hatch; all hatched fish died. Survival of fish hatched
in 296 A
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companion acute value of 1,320 /ug/L results in an acute-chronic ratio of
6.704.
The Final Acute-Chronic Ratio of 5.000 is the geometric mean of the
acute-chronic ratio of 2.715 for the freshwater cladoceran, Ceriodaphnia
dubia; 6.869 for the freshwater fathead minnow, Pimephales promelas; and 6.704
for the saltwater inland silverside, Menidia bervllina. Division of the Final
Acute Value of 2,670 pg/L for freshwater species by the ratio of 5.000 results
in a Final Chronic Value of 534.0 pg/L for freshwater aquatic life. The value
of 534.0 /jg/L is a factor of 2.3 less than the chronic value for the
life-cycle test with Ceriodaphnia dubia and is slightly greater than the
lowest chronic value of 491 pg/L reported for the fathead minnow.
Division of the Final Acute Value of 548.8 pg/L for saltwater species by
the ratio of 5.000 results in a Final Chronic Value of 109.8 /jg/L for
saltwater aquatic life. The value of 109.8 pg/L is a factor of 1.8 less than
the chronic value of 196.9 /^g/L determined from the early life-stage test with
inland silversides.
Toxicitv to Aquatic Plants
Results of a test with one species of freshwater algae and 2,4-DMP is
shown in Table 4. A four-day exposure with the alga, Scenedesmus guadricauda,
indicated that 2,4-DMP concentrations of 40,000 /jg/L and above inhibited
growth (Bringman and Kuhn 1959a,b) . No acceptable saltwater plant data with
2,4-DMP were found in the literature. A Final Plant Value, as defined in the
Guidelines, cannot be calculated for 2,4-DMP.
Bioaccumulation
A study to determine the bioconcentration of 2,4-DMP with one freshwater
species is shown in Table 5. UC radiolabelled 2,4-DMP bioconcentrated
150-fold in the whole body of the bluegill, Lepomis macrochirus (Barrows et
al. 1980; Veith et al. 1980) (Table 5). A BCF determined on the basis of
radiolabelling may contain some radiolabelled metabolites; therefore, the BCF
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of 150 may be greater than that for parent 2,4-DMP. 2,4-DMP has a measured
partition coefficient (log n-octanol/water) of 2.42, and the BCF of 150
appears to be a reasonable estimate when compared to other chemicals (Veith et
al. 1980).
No U.S. FDA action level or other maximum acceptable concentration in
tissue, as defined in the Guidelines, is available for 2,4-DMP. Therefore, no
Final Residue Value can be calculated.
Other Data
The incipient inhibition concentration for the bacterium, Escherichia
coli, was in excess of 100,000 pg/L (Bringman and Kuhn 1959a) (Table 6).
Exposure of the alga, Chlorella pyrenoidosa, to 100,000 pg/L for 72 hr
resulted in a 52 percent reduction of chlorophyll a (Huang and Gloyna
1967,1968). A 28-hr exposure of the protozoan, Microregma heterostoma,
produced an incipient inhibition concentration of 70,000 pg/L (Bringman and
Kuhn 1959b), while a 60-hr EC50 of 130,510 pg/L (based on cell number) was
obtained with the protozoan, Tetrahvmena pyriformis (Schultz and Riggin 1985).
Spehar (1987) and Norberg-King (1987) reported 48-hr LCSOs ranging from
3,100 /jg/L to 6,300 pg/L for Ceriodaphnia dubia in eight separate tests in
which daphnids were fed. Bringman and Kuhn (1959a,b) reported immobilization
of Daphnia maqna at 24,000 pg/L.
Rainbow trout, Oncorhvnchus mvkiss, were exposed to acutely lethal
concentrations of 2,4-dimethylphenol to determine the symptomology of
poisoning (Bradbury et al. 1989). At an exposure concentration of 9,040 pg/L,
the trout had a mean survival time of 6 hr. They exhibited a significant
increase from pre-exposure measurements in cough frequency, and significant
decreases in gill oxygen uptake efficiency, total blood carbon dioxide
(arterial) and hematocrit. These responses were consistent with a toxic mode
of action referred to as type- II (polar) narcosis. The mean LC50 for fathead
minnows after 8 days was 13,500 pg/L (Phipps et al. 1981).
The number of sporophytes was reduced in brown kelp, Laminaria
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saccharina, exposed for two days to 12,000 pg/L of 2,4-DMP in two tests which
began with either five- or seven-day-old plants (Thursby and Steele 1987;
Table 6). Reproduction of kelp in 7,200 pg/L was not reduced.
Unused Data
A screening study by Applegate et al. (1957) was not used because not
enough fish were tested per concentration. High control mortalities occurred
in some tests reported by Thursby and Berry (1987a,b), and these results were
not included in the data tables. 2,4-DMP toxicity was reported in cell
cultures only by Babich and Borenfreund (1987). Methods were not adequately
described in some studies (e.g., Curtis et al. 1982; Grushko et al. 1975).
Data were not used when 2,4-DMP was a component in a mixture (e.g., Giddings
and Franco 1985; Swartz et al. 1985) or effluent (Horning et al. 1984;
Pickering 1983). Studies were not used if the exposure duration was not
specified (e.g., Blum and Speece 1991).
Reports of 2,4-DMP toxicity were not used when the data had been
compiled from other sources (e.g., Alexander et al. 1983; Enslein 1987; Hall
and Kier 1984a,b; Kenaga 1982; Sabljic 1987; Schultz et al. 1986; Veith and
Broderius 1987). Similarly, reviews on bioconcentration (Davies and Dobbs
1984) and taste or odor (Persson 1984) were not used.
Summary
The acute toxicity of 2,4-DMP has been determined for 12 species of
freshwater animals. Acute values ranged between 3,340 pg/L and
67,600 /jg/L. Of the eight invertebrate and four fish species tested, two
cladocerans, Ceriodaphnia dubia and Daphnia maqna, were the most sensitive.
Acute values for freshwater fish ranged from 6,300 pg/L to 19,300 fjg/L. The
bluegill, Lepomis macrochirus, was the most sensitive freshwater species.
The chronic value for Ceriodaphnia dubia was 1,230 pg/L. In three tests
with the fathead minnow, Pimephales promelas, chronic values of 2,475, 2,200
and 491 ftg/L were obtained. Acute-chronic ratios were 2.715 and 6.869 for
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Ceriodaphnia and Pimephales, respectively.
The acute toxicity of 2,4-DMP has been determined for nine species of
saltwater animals. Acute values ranged from 1,320 /ug/L for juvenile inland
silversides, Menidia bervllina, to 55,900 /jg/L for archiannelid worms,
Dinophilus gyrociliatus. Of the six invertebrate and three fish species
tested, no taxonomic group appeared particularly sensitive.
Chronic toxicity data for saltwater organisms are available from an
early life-stage test with the inland silverside, Menidia beryllina. Survival
of hatched fish was reduced in 296 /^g/L of 2,4-DMP. No effects on survival or
growth were observed at 131 ^g/L. The acute-chronic ratio for this species is
6.704.
Limited plant data indicate that concentrations of 40,000 /ug/L or more
result in reduced growth of freshwater algae. No acceptable saltwater plant
data were found in the literature.
One test showed that the BCF for 2,4-DMP was 150 based on data for the
bluegill, Lepomis macrochirus. No acceptable saltwater BCFs were found in the
literature.
The freshwater Final Acute Value and Final Chronic Value for 2,4-DMP are
2,670 and 534 /jg/L, respectively. The value of 534 pg/L is slightly greater
than the lowest chronic value of 491 pg/L reported for the fathead minnow,
indicating that this species might not be adequately protected if ambient
water concentrations exceed this concentration for long periods of time. The
saltwater Final Acute Value and Chronic Value are 548.8 and 1*09.8 pgfL,
respectively. Chronic adverse effects to the only saltwater species exposed
to 2,4-DMP occurred at concentrations that are higher than the Final Chronic
Value which should be protective of saltwater species.
National Criteria
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms and
e*
Their Uses" indicate that, except possibly where a locally important species
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is very sensitive, freshwater aquatic organisms and their uses should not be
affected unacceptably if the four-day average concentration of 2,4-DMP does
not exceed 530 A/g/L more than once every three years on the average and if the
one-hour average concentration does not exceed 1,300 ^g/L more than once every
three years on the average.
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, saltwater aquatic organisms and their uses should not be
affected unacceptably if the four-day average concentration of 2,4-DMP does
not exceed 110 ljg/l> more than once every three years on the average and if the
one-hour average concentration does not exceed 270 ^g/L more than once every
three years on the average.
Implementation
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 when 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). Water quality criteria adopted in state water quality standards
could have the same numerical values as criteria developed under Section 304,
of the Clean Water Act. However, in many situations states might want to
adjust water quality criteria developed under Section 304 to reflect local
environmental conditions and human exposure patterns. Alternatively, states
may use different data and assumptions than EPA in deriving numeric criteria
that are scientifically defensible and protective of designated uses. State
water quality standards include both numeric a'nd narrative criteria. A state
may adopt a numeric criterion within its water quality standards and apply it
10
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either state-wide to all waters designated for the use the criterion is
designed to protect or to a specific site. A state may use an indicator
parameter or the national criterion, supplemented with other relevant
information, to interpret its narrative criteria within its water quality
standards when developing NPDES effluent limitations under 40 CFR
122.44(d)(1)(vi).2
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 1991). The averaging periods of "one hour" and "four
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 1991). 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 1991), 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 1987, 1991).
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DRAFT
9/22/93
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DRAFT
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March 2.
Holcombe, G.W., G.L. Phipps and J.T. Fiandt. 1982. Effects of phenol,
2,4-dimethylphenol, 2,4-dichlorophenol, and pentachlorophenol on embryo,
26
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DRAFT
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larval, and early - juvenile fathead minnows (Pimephales promelas1) . Arch.
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LeBlanc, G.A. 1984. Comparative structure-toxicity relationships between acute
and chronic effects to aquatic organisms. In: QSAR in environmental
27
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toxicology. Kaiser, K.L. (Ed.)- D. Reidel Publ. Co., Dordrecht, pp. 235-260.
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and substituted phenols to the fathead minnow. Bull. Environ. Contam. Toxicol.
26:585-593.
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of wastewater from a conventional wastewater treatment system receiving
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substances by wet oxidation. J. Water Pollut. Control Fed. 52:2117-2130.
Redmond, M.S. and K.J. Scott. 1987. Acute toxicity tests with dimethylphenol.
(Memorandum to D.J. Hansen, U.S. EPA, Narragansett, RI), April 17.
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pollutants in the aquatic environment. Water Res. 20:1077-1090.
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York, NY.
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pollutants from the molecular connectivity model. Z. gesam'te Hyg. 33:493-496.
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29
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Stephan, C.E., D.I. Mount, D.J. Hansen, J.FK Gentile, G.A. Chapman and W.A.
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*
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31
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32
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-R-33-02.1-
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AMBIENT AQUATIC LIFE WATER QUALITY CRITERIA FOR
ANILINE
(CAS Registry Number 62-53-3)
SEPTEMBER 1993
U.S. ENVIRONMENTAL PROTECTION AGENCY
OFFICE OF WATER
OFFICE OF SCIENCE AND TECHNOLOGY
HEALTH AND ECOLOGICAL CRITERIA DIVISION
WASHINGTON, D.C.
OFFICE OF RESEARCH AND DEVELOPMENT
ENVIRONMENTAL RESEARCH LABORATORIES
DULUTH, MINNESOTA
NARRAGANSETT, RHODE ISLAND
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NOTICES
This document has been reviewed by the Environmental Research
Laboratories, Duluth, MN and Narragansett, RI, Office of Research and
Development and the Health and Ecological Criteria Division, Office of Science
and Technology, 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.
ii
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FOREWORD
Section 304(a) (1) of the Clean Water Act of 1977 (P.L. 95-217) 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,
including ground water. This document is a revision of proposed criteria
based upon consideration of comments received from other federal agencies,
state agencies, special interest groups, and individual scientists. Criteria
contained in this document replace any previously published EPA aquatic life
criteria for the same pollutant(s).
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 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, they represent maximum acceptable pollutant
concentrations in ambient waters within that state that are enforced through
issuance of discharge limitations in NPDES permits. 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 modify water quality criteria developed under section 304 to
reflect local environmental conditions and human exposure patterns. {
Alternatively, states may use different data and assumptions than EPA in
deriving numeric criteria that are scientifically defensible and protective of
designated uses. It is not until their adoption as part of state water
quality standards that criteria become regulatory. Guidelines to assist the
states and Indian tribes in modifying the criteria presented in this document
are contained in the Water Quality Standards Handbook (December 1983). This
handbook and additional guidance on the development of water quality standards
and other water-related programs of this Agency have been developed by the
Office of Water.
This document, if finalized, would be guidance only. It would not
establish or affect legal rights or obligations. It would not establish a
binding norm and would not be finally determinative of the issues addressed.
Agency decisions in any particular situation will be made by applying the
Clean Water Act and EPA regulations on the basis of specific facts presented
and scientific information then available.
Tudor T. Davies
Director
Office of Science and Technology
iii
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ACKNOWLEDGMENTS
Larry T. Brooke
(freshwater author)
University of Wisconsin-Superior
Superior, Wisconsin
David J. Hansen
(saltwater author)
Environmental Research Laboratory
Narragansett, Rhode Island
Robert L. Spehar
(document coordinator)
Environmental Research Laboratory
Duluth, Minnesota
Suzanne M. Lussier
(saltwater coordinator)
Environmental Research Laboratory
Narragansett, Rhode Island
iv
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CONTENTS
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Page
Notices ............ . ...............
Foreword . .' . " ................ .........
Acknowledgments ............................. iv
Tables ............ .... ................. vi
Introduction .............................. l
Acute toxicity to Aquatic Animals ......... ........... 2
Chronic Toxicity to Aquatic Animals ................... 4
Toxicity to Aquatic Plants ....................... 6
Bioaccumulation ............................. 7
Other Data ............................... 7
Unused Data ........... ....... ............. XO
Summary .................................
National Criteria ............................ 13
Implementation ............................. 13
References
33
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1.
2.
3.
4.
5.
TABLES
Page
Acute Toxicity of Aniline to Aquatic Animals. 15
Chronic Toxicity of Aniline to Aquatic Animals .. 19
Ranked Genus Mean Acute Values with Species Mean Acute-Chronic
Ratios.
21
Toxicity of Aniline to Aquatic Plants. 24
Other Data on Effects of Aniline on Aquatic Organisms. 26
vi
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introduction
Aniline (aminobenzene, benzenamine, phenylaraine) is the simplest of the
aromatic amines (C^KSKH2). It occurs naturally in coal-tars (Shelford 1917)
and is manufactured by the catalytic reduction of nitrobenzene, amination of
chlorobenzene and ammonolysis of phenol.
The major users of aniline are the polymer, rubber, agricultural and dye
industries. Demand for aniline by the dye industry was high prior to the
1970's but decreased markedly in the United States thereafter because of the
increased use of synthetic fabrics. Aniline is used today primarily by the
polymer industry to manufacture products such as polyurethanes. The rubber
industry uses large amounts of aniline to manufacture antioxidants,
antidegradants and vulcanization accelerators. The pharmaceutical industry
uses aniline in the manufacture of sulfa drugs and other products. Important
agricultural uses for aniline derivatives include herbicides, fungicides,
insecticides, repellents and defoliants. Aniline has also been used as an
antiknock compound in gasolines (Kirk-Othmer 1982).
Aniline is soluble in water up to 34,000,000 A/9/L (Verschueren 1977).
The Iog10 of the octanol-water partition coefficient for aniline is 0.90 (Chiou
1985a). Through direct disposal, such as industrial discharges and non-point
sources associatedwith agricultural uses, it enters the aquatic environment.
It is removed from the aquatic environment by several mechanisms. The major
pathway of removal from water is by microbial decomposition (Lyons et al.
1984, 1985). Several minor pathways have been identified including
evaporation, binding to humic substances and autoxidation.
Additions to the aniline molecule of certain functional groups have been
found to increase toxicity (Brooke et al. 1984; Geiger et al. 1986, 1987).
Tests with the fathead minnow fPimephales promelas) have demonstrated that
substitutions with halogens, (chlorine, fluorine, and bromine) increased
toxicity. The addition of alkyl groups also increased toxicity; the toxicity
increases in proportion to the increase in chain length. Twenty-four
substitutions were tested and all except EiEE additions of methyl and nitro
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DRAFT
9/22/93
groups increased the toxicity to the fathead minnow.
All concentrations reported herein are expressed as aniline. Results of
such intermediate calculations as recalculated LCSO's and Species Mean Acute
Values are given to four significant figures to prevent round-off error in
subsequent calculations, not to reflect the precision of the value. Whenever
adequately justified, a national criterion may oe replaced by a site-specific
criterion (U.S. EPA 1983a) that may include not only site-specific
concentrations (U.S. EPA 1983b) but also site-specific frequencies of allowed
excursion (U.S. EPA 1985).
A comprehension 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 1985), is necessary to understand the
following text, tables, and calculations. The latest comprehensive literature
search for information for this document was conducted in September 1992; sime
more recent information is included.
Acute toxicitv to Aquatic Animals
The data that are available according to the Guidelines concerning the
acute toxicity of aniline are presented in Table 1. Cladocera were the most
sensitive group of the 19 species tested. Several species of larval midges
and embryos and larvae of the clawed toad, Xenopus laevis. were the most
resistant to aniline in acute exposures. Fish tended to be in the mid-range
of sensitivity for aquatic organisms.
Forty-eight-hour ECSOs for the cladocerans Ceriodaphnia dubia and
Daphnia maona were 44 nq/l. and 530 pg/L, respectively. Several independent
exposures conducted with both species showed consistency among the tests
(Table 1). However, there appears to be a large increase in tolerance of
aniline between cladocerans and other aquatic species. The 96-hr LC50 for the
next most sensitive species, a planarian, Duaesia tiorina. was 31,600 j/g/L.
Ninety-six-hour LCSOs for fish ranged from 10,600 to 187,000 ^g/L. The
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rainbow trout fOncorhvncus mvkisa) was the most sensitive species of fish
tested, with 96-hr LCSOs ranging from 10,6«0 to 41,000 ftg/l>. The bluegill
(Lepomia macrochirusl was slightly more tolerant of aniline with a 96-hr LC50
of 49,000 pg/L. Fathead minnows, Pimephales promelas, and goldfish, Carassius
auratua. were the most tolerant of aniline of the fish species tested.
Ninety-six-hour LCSOs for tests with fathead minnows ranged from 32,000 to
134,000 pg/L. A 96-hr LC50 for the goldfish was 187,000 /jg/L.
Franco et al. (1984) exposed four species of midge larvae to aniline and
found them to be the most tolerant of aniline of all species tested. The
midge, clinotanvpus pinquis. was the most tolerant of the four species tested;
a 48-hr LC50 of 477,900 A«g/L was calculated for this species. LCBOs for other
midge species tested by Franco et al. (1984), ranged downward to 272,100 /ng/L.
Holcombe et al. (1987) tested another species of midge (Tanvtarsus disslmilis)
and reported a 48-hr LC50 >219,000 ^g/L.
The African clawed frog, Xenopus laevis, was relatively tolerant of »
aniline. In a series of three tests, Davis et al. (1981) found that embryos
of African clawed frogs were more tolerant than the larvae. The 96-hr LCSOs
for embryos and tailbud embryos were 550,000 and 940,000 A«g/L, respectively,
compared to 150,000 pg/L for the larvae.
Genus Mean Acute Values (GMAVs) are ranked from most sensitive to most
resistant for the nineteen freshwater genera tested (Table 3). The freshwater
Final Acute Value (FAV) of 56.97 pg/L was calculated using the GMAVs for the
four most sensitive genera, Ceriodaphnia, paphnia. Duqesia, and Oncorhvnchus
which differ from one another within a factor of 251. The Final Acute Value
is 2.2 times less than the acute value for the most sensitive freshwater
species.
The acute toxicity of aniline to resident North American saltwater
animals has been determined with five species of invertebrates and three
species of fish (Thursby and Berry 1987a, I987b; Redmond and Scott 1987;
Table 1). Grass shrimp, tested as larvae, was the most sensitive' species
based on an acute value of 610 figfL. Crustaceans comprised the three most
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sensitive species tested; acute values ranged from 610 to 16,500 pg/L. Acute
values for three fishes, a mollusc and an echinoderm ranged from 17,400 to
>333,000 pg/L. Mortalities in acute tests with mysids, grass shrimp,
sheepshead minnows and inland silversides increased during 96-hr tests. GMAVs
are ranked from the most sensitive to the most resistant (Table 3) for the
eight saltwater genera tested. The Final Acute Value for saltwater species is
153.4 pg/L which is four times less than the acute value for the most
sensitive saltwater species tested.
Chronic Toxicitv to Aquatic Animals
The data that are available according to the Guidelines concerning the
chronic toxicity of aniline are presented in Table 2, Four chronic toxicity
tests exposing freshwater organisms to aniline have been reported. The
cladoceran, ceriodaphnia dubia, was exposed to initial concentrations ranging
from 1.07 to 26.5 pg/L for seven days with daily renewed exposures (Spehar
1987). Survival was not significantly affected at any exposure concentration;
however, effects on young production were observed at 12.7 yg/L, but not at
8.1 pg/L. The chronic value, based upon reproductive impairment, is 10.1
pg/L. This number may be under-protective since it is based upon initial
measured concentrations of aniline and did not take into consideration that
the study showed nearly 100% loss of aniline from solution in 24 hr. A
companion acute test was conducted with the chronic study and resulted in a
48-hr EC50 of 44 pg/L. Division of this value by the chronic value generates
an acute-chronic ratio of 4.356 for Ceriodaphnia dubia.
Daphnia maona were exposed to aniline for 21 days in a renewal test
(Gersich and Milazzo 1988}. Mean concentrations for the exposures ranged from
12.7 to 168.6 pg/L for the five concentrations tested. Mean total
young/surviving adult and mean brood size/surviving adult were not
significantly different from the control organisms at 24.6 A/g/L but were
significantly different at 46.7 pg/L. Based"upon these two reproduction
endpoints, the chronic value is 33.9 »g/l. The companion acute value (48-hr
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EC50) used to compute an acute-chronic ratio was 170 yg/L (Gersich and Mayes,
1986). Division of this value by the chronic value of 33.9 pg/L results in an
acute-chronic ratio of 5.015.
A 90-da'y early life-stage test was conducted with rainbow trout (Spehar
1987). The test was started with newly fertilized embryos. After 56 days
(swim-up stage), wet weight was significantly reduced at concentrations of
4,000 pig/L and above. After 90 days of exposure, an effect was not seen at
4,000 pg/L but weight was reduced at 7,800 pg/L. Survival was reduced at only
the highest exposure concentration (15,900 pg/L). The chronic value for
rainbow trout is 5,600 pg/L, based upon growth. Spehar (1987) also conducted
a 96-hr acute test which resulted in an acute value of 30,000 pg/L. Division
of the acute value by the chronic value generates an acute-chronic ratio of
5.357.
The fathead minnow was exposed to aniline concentrations that ranged
from 316 to 2,110 pg/L in 32-day exposures (Russom 1993). Percentage normal
fry at hatch and survival at the end of the test did not differ significantly
from the control fish at any aniline concentrations. Growth (weight and
length) was significantly (p<0.05) reduced at aniline concentrations of 735
pg/L and greater, but not at 422 pg/L. Wet weight was reduced by 13.3% and
total length by 6.4% compared to control fish wet weight and total length at
735 jig/I.. The chronic value for this test, based upon growth, is 557 /ig/L.
The companion acute test resulted in a 96-hr LC50 of 112,000 pg/L (Geiger et
al. 1990). Division of this value by the chronic value results in an acute-
chronic ratio of 201.1.
The only chronic toxicity test with aniline and saltwater species was
conducted with the mysid, Hvsidopsis bahia (Thursby and Berry 1987b).
Ninety-five percent of the mysids exposed during a life-cycle test to 2,400
pg/L died and no'young were produced by the survivors. Reproduction of mysids
in 1,100 pg/L was reduced 94 percent relative to controls. No significant
effects were detected on survival, growth, or reproduction in mysids exposed
to <540 pg/L for 28 days. The chronic value for this species is 770.7 pg/L,
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based upon reproductive impairment. A comparison acute test was conducted
with the chronic test which resulted in an "acute value of 1,930 ^g/L.
Division of this value by the chronic value results in an acute-chronic ratio
of 2.504.
The Final Acute-Chronic Ratio of 4.137 is the geometric mean of the
acute-chronic ratios of 4.356 for the freshwater cladoceran, Ceriodaphnia
dubia. 5.015 for the freshwater cladoceran, Daphnia maqna. 5.357 for the
rainbow trout, Oncorhvnchus mvkiss, and 2.504 for the saltwater mysid,
Mvsidopsis bahia (Table 2)., The acute-chronic ratio of 201.1 for the fathead
minnow was not used in this calculation because, as described in the
Guidelines, this species is not acutely sensitive to aniline and its Species
Mean Acute Value is not close to the Final Acute Value (Table 3). Division of
the freshwater Final Acute Value of 56.97 ^g/L by 4.137 results in a
freshwater Final Chronic Value of 13.77 Aig/L. Division of the saltwater Final
Acute Value of 153.4 pg/L by 4.137 results in a saltwater Final Chronic Value
of 37.08 A/g/L. The freshwater Final Chronic Value is approximately 1.4 times
greater than the lowest freshwater chronic value of 10.1 pg/L for Ceriodaphnia
dubia. The saltwater Final Chronic Value is a factor of 21 times less than
the only saltwater chronic value of 770.7
Toxicitv to Aquatic Plants
Results of tests with two species of freshwater green alga exposed to
aniline are shown in Table 4. Sensitivity to aniline differed between the two
species. Four-day exposures with aniline and Selenastrum capricornutum showed
that the ECSOs ranged from 1,000 $*g/L (Adams et al. 1986) to 19,000 pg/L
(Calamari et al. 1980, 1982) with reduced growth as the effect. Slooff (1982)
determined an EC50 of 20,000 pg/L for an unidentified species of Selenastrum
with reduced biomass as the effect. The studies by Adams et al. (1986) were
conducted both with and without a carrier solvent (acetone). The lowest 96-hr
ECSOs were obtained from exposures using acetone. However, this relationship
was reversed when the exposure duration was increased to five and six days
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(Table 4). The green alga, Chlorella vulaaris, is considerably more tolerant
to aniline than selenastrum. In 14-day exposures, growth of £. vulaaris was
reduced 58% by 306,000 pg/l and 16% by 184,000 fjg/l, (Ammann and Terry 1985).
The study also demonstrated that aniline had significant effects upon
respiration and photosynthesis of the species. There are no acceptable plant
data for saltwater species for aniline. A Final Plant Value, as defined in
the Guidelines, cannot be obtained for aniline.
Bioaecumulation
Studies to determine the bioconcentration of aniline with three species
of organisms have been reported (Table 5). In all these studies, steady-state
bioconcentrations were not demonstrated. Daphnia macma bioconcentrated
aniline five times in a 24-hr exposure (Dauble et al. 1984, 1986), a green
alga 91 times in a 24- to 25-hr exposure (Hardy et al. 1985) and rainbow trout
507 times in a 72-hr exposure (Dauble et al. 1984). Because tests were not-of
sufficient duration according to the Guidelines, and no U.S. FDA action level
or other maximum acceptable concentration in tissue is available for aniline,
no Final Residue Value can be calculated.
Other Data
Other data available concerning aniline toxicity are presented in Table
5. Effects on two species of bacteria were seen at aniline concentrations
ranging from 30,000 to 130,000 j/g/L.
Three genera of algae were exposed to aniline. One species of bluegreen
algae, Microcvstis aeruainosa. (Bringmann and Kuhn 1976, 1978a,b), showed more
sensitivity to aniline than other species. Inhibition of cell replication of
this species was observed after an 8-day exposure to 160 /ig/L. Fitzgerald et
al. (1952) reported a 24-hr LC50 of 20,000 ngfL with the name species. A 66%
reduction of photosynthesis by the green algae, Selenastrum capricornutum, was
reported by Giddings (1979) after a 4-hr exposure to 100,000 pg/L of aniline.
Several species of protozoans were exposed to aniline. A 28-hr aniline
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exposure with Microreoma heterostoma showed that food ingestion was reduced at
20,000 /jg/L (Bringmann and Kuhn 1959a). Other species of protozoa were tested
and showed less sensitivity to aniline (Table 5).
The hydrazoan, Hvdra olicractis. showed sensitivity to aniline in a 48-hr
test. The LC50 for this species of 406 A«g/L was determined by Slooff (1983)
in a static, unmeasured test using river water. Other organisms such as
planarians fDuqesia luqubris), tubificid worms fTubificidael, and snails
(Lvmnea staanalis) were also tested and had much higher 48-hr LCSOs of
155,000, 450,000 and 800,000 nq/l>, respectively.
Cladocera appeared to be the group most sensitive to aniline. Spehar
(1987) reported a 48-hr LC50 of 132 pg/L for Ceriodaohnia dubia in an exposure
in which the organisms were fed their culturing ration. In the same study, a
LC50 of 44 ^g/L was determined for unfed Ceriodanhnia dubia. The difference
in results could have been due to the complexation of aniline by the food
and/or increased hardiness of the fed organisms. Daphnia maana was affected
(acoustic reaction and mortality) at aniline concentrations ranging from 400
to 2,000 A«g/L (Bringmann and Kuhn 1959a,b, 1960; Lakhnova 1975) for 48-hr
exposures. Calamari et al. (1980, 1982) found this species to be more
resistant to aniline with a reported 24-hr EC50 of 23,000 A/g/L.
Insects showed varying sensitivities to aniline. Puzikova and Markin
(1975) exposed the midge, Chironomus dorsalis. to aniline through its complete
life cycle and reported 100% survival at 3,000 jig/L and 5% survival at 7,800
Aig/L. Slooff (1983) exposed mayfly and mosquito larvae to aniline for 48 hr
and reported LCSOs of 220,000 and 155,000 vq/1*, respectively.
The toxicity values for rainbow trout in Table 5 are in general
agreement with those used in Table 1. Rainbow trout were exposed to aniline
by several workers using different exposure durations. Shumway and Palensky
(1973) found 100% mortality of rainbow trout at 100,000 /^g/L in a 48-hr
exposure and 100% survival at 10,000 pg/L. Lysak and Marcinek (1972) also
reported 100% mortality for a 24-hr exposure at 21,000 jig/L and observed no
mortality at 20,000 ^g/L. Abram and Sims (1982) determined the 7-day LCSO to
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be 8,200 pg/L in two separate tests using rainbow trout.
Several tests were run with aniline in dilution waters of different
water quality. Water hardness appeared to have little, if any, impact on
aniline toxicity -(Birge et al. 1979a,b). Young channel catfish, Ictalurus
punctatus, were exposed to aniline in waters with a four-fold difference in
hardness (53.3 and 197.5 mg/L as CaGO,). The resulting LCSOs indicated only a
slight decrease in toxicity with increasing hardness. In a similar test they
also exposed goldfish and largemouth bass, Micropterus salmoides, and reported
the opposite effect on toxicity. pH does not appear to affect toxicity of
aniline with aquatic organisms (Table 5).
The African clawed frog demonstrated varied effects over a broad range
of concentrations of aniline. Davis et al. (1981) and Dumpert (1987) observed
that aniline concentrations of 50 and 70 pg/L resulted in reduced epidermal
pigmentation or failure of larvae to develop normal pigmentation. In a
12-week exposure, Dumpert (1987) showed that 1,000 pg/L of aniline slowed
metamorphosis and reduced growth. At an exposure concentration of 10,000 pg/L
for 96-hr, 6% of the frog larvae developed abnormalities (Dumont et al. 1979;
Davis et al. 1981). Frog embryos had 50% teratogeny in 120- and 96-hr
exposures at 91,000 and 370,000 pg/L, respectively (Table 5). One hundred
percent mortality of immature frogs occurred during a 12-day exposure to
90,000 pg/L (Dumpert 1987) and 50% mortality during a 48-hr exposure to
560,000 pg/L (Slooff 1982; Slooff and Baerselman 1980).
Concentrations of the free amino acids aspartate, glutamate and alanine
in the sea anemone, Bunodosoma cavernata. increased after seven days of
exposure to aniline at 500,000 pg/L (Kasschau et al. 1980; Table 5). The
lethal threshold (geometric mean of the highest concentration with no
mortality and the next higher concentration) was 29,400 pg/L for sand shrimp,
Cranqon septemspinosa, and >55,000 for soft-shelled clams, My_a arenaria
(McLeese et al. 1979).
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Unused Data
Some data on the effects of aniline on aquatic organisms were not used
because the studies were conducted with species that are not resident in North
America or Hawaii (Freitag et al. 1984; Hattori et al. 1984; Inel and Atalay
1981; Juhnke and Ludemann 1978; Lallier 1971; Slooff and Baerselman 1980;
Tonogai et al. 1982; Yoshioka et al. 1986a). Chiou (1985b); Hermens et al.
(1985); Hodson (1985); Koch (1986); Newsome et al. (1984); Persson (1984);
Schultz and Moulton (1984); Slooff et al. (1983); Vighi and Calamari (1987)
compiled data from other sources. Results were not used where the test
procedures or test material, were not adequately described (Buzzell et al.
1968; Canton and Adema 1978; Carlson and Caple 1977; Clayberg 1917; Demay and
Menzies 1982; Kuhn and Canton 1979; Kwasniewska and Kaiser 1984; Pawlaczyk-
Szpilowa et al. 1972; Sayk and Schmidt 1986; Shelford 1917; Wellens 1982).
Data were not used when aniline was part of a mixture (Giddings and Franco
1985; Lee et al. 1985; Winters et al. 1977) or when the organisms were exposed
to aniline in food (Lee et al. 1985; Loeb and Kelly 1963).
Babich and Borenfreund (1988), Batterton et al. (1978), Bols et al.
(1985); Buhler and Rasmusson (1968), Carter et al. (1984), Elmamlouk et al.
(1974), Elmamlouk and Gessner (1976), Fabacher (1982), Lindstrom-Seppa et al.
(1983), Maemura and Omura (1983), Pedersen et al. (1976), Sakai et al. (1983),
and Schwen and Mannering (1982) exposed only enzymes, excised or homogenized
tissue, or cell cultures. Anderson (1944), and Bringmann and Kuhn (1982)
cultured organisms in one water and conducted tests in another. Batterton et
al. (1978) conducted a study in which organisms were not tested in water but
were tested on agar in the "algal lawn" test.
Results of one laboratory test were not used because the test was
conducted in distilled or deionized water without addition of appropriate
salts (Mukai 1977). Results of laboratory bioconcentration tests were not
used when the test was not flow-through or renewal (Freitag et al. 1985; Geyer
et al. 1981; Geyer et al. 1984) and BCFs obtained from microcosm or model
ecosystem studies were not used where the concentration of aniline in water
10
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decreased with time (Lu and Metcalf 1975; Yount and Shannon 1987). Douglas et
al. (1986) had insufficient mortalities to Calculate an LC50 and Sollmann
(1949) conducted studies without control exposures.
Summary
Data on the acute toxicity of aniline are available for nineteen species
of freshwater animals. Cladocera were the most acutely sensitive group
tested. Mean 48-hr ECSOs ranged from 125.8 ngfL for Ceriodaphnia dubia to 250
fjg/L for Daphnia macma. The planarian, Duqesia tiqrina. was the fourth most
sensitive species to aniline with a 96-hr LC50 of 31,600 pg/L.
Freshwater fish 96-hr LCSOs ranged from 10,600 to 187,000 ^g/L. Rainbow
trout, Oneorhvnchus mvkiss, were the most sensitive fish tested, with species
mean acute values of 26,130 pg/L. The bluegill, Lepomis macrochirus, was
nearly as sensitive to aniline as rainbow trout, with a 96-hr LC50 of 49,000
pg/L reported for this species. The fathead minnow, Pimephales promelas. and,
goldfish, Caraasius auratus. were the most tolerant fish species exposed to
aniline, with species mean acute values of 106,000 /jg/L and 187,000 pg/L,
respectively.
The most tolerant freshwater species tested with aniline was a midge,
Clinotanvpus pinquis, with a 48-hr LC50 of 477,000 pg/L. Developmental stages
of an amphibian, Xenopus laevis, had differing sensitivities to aniline. The
embryos were the most tolerant with a 96-hr LC50 of 550,000 pg/L and the
larvae had a 96-hr LC50 of 150,000 pg/L.
Data on the acute toxicity of aniline are available for eight species of
saltwater animals. Species Mean Acute Values ranged from >333,000 /jg/L for
larval winter flounder, Pseudopleuronectes americanus. to 610 /jg/L for larval
grass shrimp, Palaemonetes ouaio. Arthropods appear particularly sensitive to
aniline. There are no data to support the derivation of a salinity- or
temperature-dependent Final Acute Equation.
Chronic tests have been conducted with four species of freshwater
organisms. A chronic value of 10.1 pqfL for the cladoceran, Ceriodaphnia
11
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dubia, was based upon reproductive impairment:. A chronic value of 33.9 pg/L
for another cladoceran, Daohnia maona, was also based on reproductive
impairment. Rainbow trout were exposed for 90 days to aniline and the results
showed that survival was reduced at 15,900 pg/L and growth (wet weight) at
7,800 pg/L. The chronic value for trout of 5,600 /jg/L was based upon growth.
The fathead minnow was exposed for 32 days in an early life-stage test. The
chronic value of 557 pg/L was also based upon growth.
One saltwater chronic value was found. A chronic value of 770.7 pg/L
for the mysid, Mvsidopsis bahia. was based upon reproductive impairment.
Effects due to aniline have been demonstrated with two freshwater plant
species. The green alga, Selenastrum capricornutum. had ECSOs ranging from
1,000 to 19,000 .pig/L in 4-day exposures. Another green alga, Chlorella
vulgaris. was considerably more resistant to aniline, showing a growth
reduction of 58% by 306,000 pg/L in a 14-day exposure. No acceptable
saltwater plant data have been found. Final Plant Values, as defined in the
Guidelines, could not be obtained for aniline.
No suitable data have been found for determining the bioconcentration of
aniline in freshwater or saltwater organisms.
Acute-chronic ratio data that are acceptable for deriving numerical
water quality criteria are available for three species of freshwater animals
and one species of saltwater animal. The acute-chronic ratios range from
2.504 to 5.357 with a geometric mean of 4.137.
The freshwater Final Acute Value for aniline is 56.97 jug/L and the Final
Chronic Value is 13.77 /jg/L. The Freshwater Final Chronic Value is 1.4 times
greater than the lowest chronic value observed for one species of Cladocera
indicating that sensitive species of this group may not be adequately
protected if ambient water concentrations exceed this value. The saltwater
Final Acute Value for aniline is 153.4 pg/L and the Final Chronic Value is
37.08 Aig/L. Chronic adverse effects to the only saltwater species exposed to
aniline occurred at concentrations that are higher than the saltwater Final
Chronic Value which should be protective of saltwater organisms.
12
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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 for certain sensitive species of Cladocera,
freshwater organisms and their uses should not be affected unacceptably if the
four-day average concentration of aniline does not exceed 14 pg/L more than
once every three years on the average and if the one-hour average
concentration does not exceed 28 pg/L more than once every three years on the
average.
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, saltwater organisms and their uses should not be affected
unacceptably if the four-day average concentration of aniline does not exceed
37 pg/L more than once every three years on the average and if the one-hour
average concentration does not exceed 77 pg/L more than once every three years
on the average.
Implementation
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). Water quality criteria adopted in state water quality standards
could have the same numerical values as criteria developed under Section 304,
of the Clean Water Act. However, in many situations states might want to
adjust water quality criteria developed under Section 304 to reflect local
environmental conditions and human exposure patterns. Alternatively, states
13
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may use different data and assumptions than EPA in deriving numeric criteria
that are scientifically defensible and protective of designated uses. State
water quality standards include both numeric and narrative criteria. A state
may adopt a numeric criterion within its water quality standards and apply it
either state-wide to all waters designated for the use the criterion is
designed to protect or to a specific site. A state may use an indicator
parameter or the national criterion, supplemented with other relevant
information, to interpret its narrative criteria within its water quality
standards when developing NPDES effluent limitations under 40 CFR
122.44(d)(1)(vi).2
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 1991). The averaging periods of "one hour" and "four
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
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 1991). 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 1991), 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
available (U.S. EPA 1987, 1.991).
14
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