Technical Support Document for the Hazardous
Waste Identification Rule: Risk Assessment
for Human and Ecological Receptors
Volume I
Appendix B
Part 1 of 2
AtoH
Prepared tor
U.S. Environmental Protection Agency
Office of Solid Waste
Contract No. 68-02-0005,68-W3-0026
August 1995
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Appendix B
Toxicological Profiles for Ecological Receptors
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APPENDIX B
APPENDIX B
ECOTOXICOLOGICAL PROFILES
This appendix presents the ecotoxicological profiles for the 47 constituents of
ecological concern. Each profile is intended to present a complete chemical-stressor profile
for the ecological receptors and endpoints evaluated in this analysis. The profiles are
organized by chemical and, because they contain all references and data relevant to that
chemical, may be reviewed as "stand-alone" sections. In addition to the profiles, an expanded
discussion of the screening process used to identify the list of 47 priority constituents (see
Section 4.3.1.2) is presented below.
Identification of Constituents of Ecological Concern
B.I Conceptual Approach
Although any constituent may cause adverse effects to ecological receptors, some
constituents are likely to present significant risks to wildlife at environmental concentrations
that are considered acceptable for human exposure. For example, constituents that are highly
persistent and bioaccumulate in the food chain may pose higher risks to piscivorous wildlife
because the piscivores ingest a higher proportion of contaminated fish in the diet than do
humans. Similarly, modes of toxic action that are unique to wildlife (e.g., eggshell thinning;
stomatal closure) or exposure pathways that are unique to wildlife (e.g.. exposure via. gill
exchange) present risks to ecological receptors that have no human analog. Therefore, a
subset was selected from the 192 constituents evaluated for human health risk that represented
those chemicals most likely to be of ecological concern. It is crucial to recognize that this
priority list of constituents is not all-inclusive; chemicals lacking the requisite data to be
included in the priority list may adversely impact wildlife through a variety of exposure
pathways and scenarios. The decision to prioritize constituents for ecological risk assessment
was, in a real sense, a resource management decision. Given the time frame for the analysis.
it was not possible to research all (or most) of the 192 chemicals for the suite of ecological
receptors representing the generic terrestrial and aquatic ecosystems. Unlike human health
risk assessment, an Agency-approved data base (i.e., IRIS) is not yet available for
ecotoxicological benchmarks. Therefore, data collection activities were concentrated on a
smaller group of constituents judged to present more significant threats to ecological receptors
than to humans. . .
August 1995. B-l
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APPENDIX B
This "short list" of priority constituents shown in Table B-l was based largely on the
five stressor characteristics relevant to ecological receptors and the endpoints chosen for this
analysis. These characteristics were selected after an extensive review of the literature (e!g..
U.S. EPA, 1994; Colborn et al.,.1993) and are similar to the characteristics described in the
problem formulation stage of the Framework for Ecological Risk Assessment (U.S. EPA,
1992a). Although other stressor characteristics were considered, the characteristics described
in Table B-1 were chosen based on their usefulness in identifying chemicals that may be of
ecological concern at concentrations considered protective of human health. For each stressor
characteristic, available information on toxicity and.physicochemical properties was examined
to determine how the data could be used to identify chemicals of potential ecological concern.
Each characteristic was assigned an operational definition, and constituents failing under three
or more definitions were given the highest priority for developing ecological exit criteria. For
example, the frequency characteristic was defined in terms, of the adverse effects levels to
aquatic organisms with respect to human exposure to contaminated drinking water. Since
aquatic organisms live in constant contact with contaminated water, constituents with a
National Ambient Water Quality Criterion (AWQC) below the human health-based level
(HBL) for drinking water ingestion were flagged under frequency. Similarly, constituents
demonstrated to adversely affect reproductive success or disrupt the endocrine system
Table B-l. Stressor Characteristics Used to Identify Constituents of Ecological Concern
Stressor
characteristic
Intensity
Frequency
Timing
Scale
Mode of action
Description
Chemicals that bioaccumulate (and possibly biomagnify) in the
food chain present elevated exposures to certain predators
Chemicals may pose considerably higher risks to ecological
receptors that are exposed continuously
Reproductive and developmental chemicals elicit adverse
effects at sensitive life stages (e.g., gestation)
The spatial and, especially, the temporal scale for exposure is
likely to be increased for persistent chemicals
Chemicals may cause adverse effects to ecological receptors
with no analogous mechanism for humans (e.g., hatchability) j
August 1995
B-2
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APPENDIX B
(the so-called estrogen mimics) were flagged under timing (Colborn et al., 1993). The
operational definitions for each stressor characteristic and the selection rationale for the
priority "short list" are described in detail below.
B.2 Operational Definitions for Stressor Characteristics
B.2.1 Intensity
The intensity of exposure to upper trophic level predators may be significantly
increased for chemicals .that accumulate in the food chain. Chemicals that bioconcentrate
(i.e., uptake from contaminated medium) and/or bioaccumulate (i.e., uptake from
contaminated medium and food chain) may be present in contaminated prey items at
concentrations that are orders of magnitude above the concentration in surface water,
sediment, or soil. In addition, contaminants such as mercury and DDT that bioaccumulate
have been shown to biomagnify up the food chain (i.e., increasing concencentration with
trophic level). Although biological uptake of chemicals is a function of physiology (e.g.,
chemical assimilation efficiency, lipid fraction) and environmental chemistry (e.g., pH, FeOx,
foe), the exposure routes are largely determined by the physicochemical properties of
constituents such as log K,w and solubility. Because bioaccumulation results from all routes
of exposure that occur in nature (i.e., direct contact, direct ingestion, ingestion of
contaminated prey), and biomagnification has been demonstrated for so few constituents, the
potential exposure intensity was operationally defined as the bioaccumulation potential of
each contaminant. However, it should be noted that dietary exposure to upper trophic level
consumers may be significant even though a contaminant bioconcentrates wealdy in prey
items; exposure from consumption of prey items often exceeds exposure from the ingestion of
a contaminated medium.
Two chemical-specific attributes were used to represent bioaccumulation potential:
(1) data on bioaccumulation (or biomagnification) for terrestrial or aquatic organisms and (2}
log KgW values as a surrogate parameter for bioaccumulation in freshwater ecosystems. For
chemicals with a log K^ > 4.0, the potential to bioaccumulate was considered significant for
freshwater ecosystems (Thomann, 1989; Connell, 1988; 58 FR 20861; U.S. EPA, 1991g).
Although the "cutoff of log K^ > 4.0 is not a bright line under which bioaccumulation
cannot occur, scientific consensus on the relationship between log K^ and bioaccumulation
indicates that, for chemicals below a log K^w value of 4.0, uptake across the gills is the major
route of exposure in fish.
August 1995 B-3
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APPENDIX B
B.2.2 Frequency
Organisms that live in direct contact with contaminated media may be more highly
exposed by virtue of constant contact with biological membranes. Aquatic organisms such as
fish and daphnids may receive continuous exposure through gill exchange and, possibly, via
the food chain. Similarly, organisms that live in the soil or sediment may receive continuous
exposure through direct contact (e.g., earthworms, tubifex worms, plants) or through ingestion
of contaminated soil and other soil fauna. Since many species that make up "typical" soil
communities have somewhat limited ranges, exposure may be significant despite biological
barriers (e.g., exoskeletons of insects). As a result'of increased exposure frequency,
organisms living in close contact with a contaminated medium may be at substantially greater
risk to chemical stressors than organisms that spend a small fraction of the life-cycle in direct
contact with water, sediment, or soil.
As suggested above, the characteristic of frequency was operationally defined (for
aquatic species only) by comparing effects levels for aquatic organisms to effects levels in
drinking water for humans (i.e., health based levels). Two types of aquatic effects levels
were used: the National Ambient Water Quality Criteria (AWQC) for the protection of
aquatic life and chronic values for fish or daphnids identified in the open liteiature. The
AWQC are standards developed by the U.S. EPA to protect 95 percent of the species in an
aquatic community with approximately 50% confidence and are widely used in a variety of
regulatory programs and ecological screening analyses. For human health effects, the human
health-based levels, or HBLs, were used to represent acceptable concentrations in surface
water as a source of drinking water. The HBLs for drinking water (in mg/L) are screening
concentrations developed by RTI and used by the U.S. EPA Office of Solid Waste in a
number of applications. For drinking water, they are based maximum contaminant levels
(MCLs) when available, or on benchmarks for cancer (slope factors) and noncancer (reference
doses) effects, assuming that a 70 kg adult consumes 2 L of water per day. For carcinogenic
chemicals, it is also assumed that the averaging time, and exposure duration and frequency
are 70 years, 70 years, and 365 days/yr, respectively. Constituents were flagged under the
characteristic of frequency if the AWQC or other chronic aquatic effects level was below the
corresponding HBL for drinking water. For chemical flagged under this characteristic,
adverse ecological effects are clearly possible at concentrations considered protective of
humans.
B.2.3 Timing of Exposure
Recent findings in the scientific community strongly suggest that long-term exposures
to chemicals that disrupt the endocrine system of animals may have dire consequences on the
August 1995 B-4
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APPENDIX B
reproductive success of wildlife populations '(Thomas and Colborn, 1992; Colborn et aJ..
1993). Impacts include thyroid dysfunction in birds and fish; decreased fertility in birds, fish.
shellfish, and mammals; decreased hatching success in birds, fish, and turtles; behavioral
abnormalities in birds; demasculinizatidn and feminization of male fish; and compromised
immune systems in birds and mammals. The importance of endocrine disrupters has also
been acknowledged by the U.S. EPA. An EPA-sponsored workshop on endocrine disrupters
has been scheduled for April, 1995, to provide a forum for information exchange among a
diverse assembly of scientific specialties and organizations and to develop a national strategy
for research needed to understand the magnitude and nature of effects (60 FR 13271).
While acknowledging that the patterns of effects vary among species and constituents.
Thomas and Colborn (1992) present four important observations:
constituents may have entirely different effects on the embryo, fetus, or
perinatal organisms than on the adult;
the effects are most often manifested in offspring, not in the exposed parent;
the timing of exposure in the developing organism is crucial in determining its
character and future potential; and
although critical exposure occurs during embryonic development, obvious
manifestations may not occur until maturity.
Colbom et al., (1993) presented a list of 45 chemicals, including pesticides, metals,
and industrial chemicals, that have widespread distribution in the environment and are
reported to have reproductive and endocrine-disrupting effects. In conjunction with a
database on developmental and reproductive toxicants compiled by RTI, the list of 45 was
used to operationally define the timing characteristic. Constituents were flagged under this
stressor characteristic if: (1) they were shown to elicit reproductive and developmental effects
in more than one species, or (2) they were included on the list of endocrine disrupters.
Constituents that were common to the HWIR and the list of reproductive and endocrine-
disrupting chemicals were flagged under "timing of exposure." Although the absence of data
on reproductive/developmental effects does not indicate that these endpoints are unimportant
for a given chemical, the presence of such data confirms toxicological significance on
reproducing populations..
B.2.4 Scale
The spatial and temporal scale of exposure for ecological receptors is greatly increased
for constiuents that are persistent in the environment. For example, persistence maintains the
exposure concentration and increases the exposure duration (i.e., temporal scale); only dilution
August 1995 B-5
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APPENDIX B
via environmental transport acts to decrease the concentration of persistent chemicals. Even
if bioavailability is diminished through binding to sediment and soil components (e.g., f^),
these contaminants can exert their toxic effects on local populations (e.g., benthic dwellers,
earthworms) and, at times, be reintroduced into the ecosystem due to physical disturbances
(e.g., storms, dredging). In contrast, dilution and degradation act to decrease the exposure
concentrations of nonpersistent constituents.
Persistence may also extend the potential for exposure at sensitive life stages. For
example, exposure in birds and mammals during gestation is extended by the persistence of
PCDDs so that, even at localized, low-level exposures, the critical dose may be exceeded due
to short-term contact in a contaminated area. As a result, wildlife that utilize contaminated
areas for nesting, spawning, egg laying, etc. may be exposed in the short term to persistent,
relatively immobile chemicals. Thus, migratory species may suffer adverse reproductive
effects if contaminated areas are important part of their life-cycle.
In addition to increasing the temporal window for exposure, persistence may also
increase the likelihood of exposure to a larger number of populations by increasing the spatial
extent of contamination. Because persistent chemicals resist degradation, the potential for
transport is greatly increased. Detectable concentrations of mercury, PCDDs and other highly
persistent constituents have been documented in pristine waters far from any anthropogenic
source of contamination. Moreover, there is evidence that DDT, PCBs, and other chemicals
are being transported long distances over the globe via the atmosphere (Rapaport et al., 1985,
as cited in Colborn et al., 1993). Wildlife populations that are not in the immediate proximity
of a contaminant source may be exposed to low levels of contaminants. As Colborn et al.,
(1993) point out, low levels of exposure to persistent endocrine disrupters may cause severe
reproductive effects.
r
Based on the relationship between persistence and exposure described above, the
characteristic of scale was operationally defined using two chemical-specific attributes: (1)
half-life in excess of 1 year or (2) log K^, value above 4.5. Constituents possessing either of
these attributes were considered to be highly persistent. The half-life "cutoff of 1 year was a
convention based on the Handbook of Enviromental Degradation Rates (Howard et al., 1991)
and was applied to surface water and soil. The log K^w "cutoff of 4.5 was taken from the
the U.S. EPA Hazardous Ranking System (55 FR 51532), the principal mechanism for placing
sites on the National Priorities List (NPL) under the broad authority of the Superfund
Amendments and Reauthorization Act (SARA) of 1986. The ranking system uses a default
scale for log K^ and assigns the highest persistence value to chemicals with log K^ above
4.5.
August 1995 B-6
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APPENDIX B
B.2.5 Mode of Action
One of the best known examples of a "nonhuman" mode of action is the thinning of
eggshells associated with exposure to DDT. Long after the organochlorine pesticide was
banned in 1972, DDT-thinned eggshells continued to put many embryonic birds - including
bald eagles - at risk of being crushed to death (Raloff, 1994).
As with the characteristic of timing, the mode of action was operationally defined
based on the reproductive and developmental toxicity database compiled by RTI. Since
animals that are oviparous (e.g., egg-laying birds, amphibians) are more likely to experience a
unique mode of toxic action with respect to humans, particular attention was focused on the
avian toxicity data. Recognizing that other animal species and plants may experience unique
toxic effects (e.g., stomatal closure in plants by sulfur dioxide; imposition of male sex in
snails by tributyl tin), specific studies presenting unique modes of toxic action were also used
to flag constituents under this characteristic. However, there were relatively few chemicals
for which this information was available and most chemicals were flagged under mode of
action based on reproductive effects to birds (and- sometimes aquatic organisms).
B.3 Priority Constituents of Ecological .Concern
Physicochemical and lexicological screening data were reviewed for each of the 192
consituents of potential concern with respect to stressor characteristics. Thirty constituents
were flagged under three or more stressor characteristics' and identified as having the highest
priority for ecological concern. However, in evaluating the entire data set oo the remaining
162 constituents (see Table B-3), it became clear that the thirty priority chemicals were, in
part, an artifact of the available data. For example, a number of persistent constituents had
AWQC well below the HBLs for drinking water. In addition, several constituents flagged as
endocrine disrupters were also flagged under other characteristics such as frequency or mode
of action. Recognizing that: (1) including constituents flagged under three stressor
characteristics on the priority list was, in part, a function of data bias, and (2) excluding
constituents flagged under two characteristics overlooked some constituents of high ecological
significance (e.g., endocrine disrupters), the priority list was expanded.
1 Diethylstilbestrol (DES) was also flagged under three stressor characteristics. However, DES was not included
in the priority list of constituents of ecological concern because: (1) screening analyses indicated that humans are
significantly more sensitive to DES than test species and (2) DES has not been manufactured in the U.S. for over
20 years. Nevertheless. DES will be included in the next group of constituents for which ecological exit criteria will
be developed.
August 1995 B-7
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APPENDIX B
In general, the expansion included chemicals flagged under two characteristics,
although the frequency, timing, and scale were the three characteristics considered most
important in the second cut at priority constituents. The relationship between these
characteristics is particularly important since persistent chemicals are more likely to impact
ecological receptors through constant exposure or exposure during sensitive life stages. Thus,
the seventeen constituents added to the priority list included chemicals that were flagged
under frequency or timing and one other characteristic. Below, Table B-2.presents the 47
constituents that were considered the highest priority for ecological risk assessment and
includes the rationale for their inclusion in the ecological priority list.
August 1995 B-8
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APPENDIX B
Table B-2. Priority List for Constituents of Ecological Concern
constituent name
Acenaphthene
Aldrin
Antimony
Arsenic V .
3arium
3enz(a)anthracene
3enzo(a)pyrene
Beryllium
Bis(2-ethylhexyl)phthalate
Butylbenzylphthalate
Cadmium
Chlordane
Chromium VI
Chrysene
Copper ,
DDT (and metabolites)
Di-n-octyl .phthalate
Dieldrin
Diethyl phthalate
Dimethyl phthalate
Endosulfan
Endrin
Fluoranthene
rationale
AWQC well below HBL; sediment criterion proposed
flagged under three stressor characteristics
rep/dev effect in multiple species; highly persistent
flagged under three stressor characteristics
AWQC below HBL; highly persistent
AWQC well below HBL; persistent
flagged under four, stressor characteristics
AWQC below HBL; highly persistent
Flagged under four stressor characteristics
flagged under three stressor characteristics
flagged under five stressor characteristics
Flagged under four stressor characteristics
Flagged under four stressor characteristics
flagged under three stressor characteristics
AWQC well below HBL; highly persistent
flagged under five stressor characteristics
flagged under four stressor characteristics
flagged under five stressor characteristics
AWQC well below HBL; endocrine disrupter
chronic value for fish well below HBL; endocrine disrupter
AWQC well below HBL; endocrine disrupter
flagged under five stressor characteristics
AWQC well below HBL; persistent; sediment quality criteria
proposed
August 1995
B-9
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APPENDIX B
Table B-2. Priority List of Constituents of Ecological Concern (conf)
constituent name
ieptachlor
Heptachlor epoxide
iexachlbrobenzene
Hexachlorcyclopentadiene
Jndane
iexachlorophene
Cepone
^ead
Mercury
rtethoxychior
/lethyl parathion
Molybdenum
Nickel
'arathion
'entachlorobenzene
'entachlorophenol
Fotychlorinated biphenyls
Selenium
Silver
FCDD. 2.3,7,8-
foxaphene
rrichlorophenoxyacetic acid,
2.4.5- (2,4.5-T)
Vanacfium
Zinc
rationale
flagged under five stressor characteristics
flagged under four stressor characteristics
flagged under four stressor characteristics
flagged under three stressor characteristics
flagged under three stressor characteristics
flagged under three stressor characteristics
Flagged under three stressor characteristics
flagged under four stressor characteristics
flagged under five stressor characteristics
lagged under three stressor characteristics
rep/dev toxicant in multiple species; avian reproductive effects
AWQC below HBL; highly persistent
chronic value for daphnids below HBL: highly persistent .
endocrine disrupter; avian reproductive effects
lagged under three stessor characteristics
lagged under four stressor characteristics
flagged under four stressor characteristics
flagged under five stressor characteristics
chronic value for fish well below HBL; highly persistent
flagged under four stressor characteristics
flagged under five stressor'characteristics
endocrine disrupter; avian reproductive effects
AWQC well below HBL; highly persistent
flagged under three stressor characteristics
August 1995
B-10
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Table B-3. Stressor Characteristics Evaluate!* n> Prioritize Constituents of Ecological Concern
atreasor characteristics
CAS Number
83329
67641
75058
98862
107028
79061
107131
309002
107051
62533
7440360
7440382
7440393
56553
71432
92875
50328
205992
100516
100447
7440417
39638329
111444
117817
75274
75252.
71363
Htiti'.,/
Constituent name
Acenaphthene
Acetone
Acetonltrile '
Acetophenone
Acroleln
Acrylamlde
Actylonilnle
Aldrtn
Allyl chloride
Aniline
Antimony
Arsenic V
Barium
Benz(a)anthracene
Benzene
Benzldlne*
Benzo(a)pyrene
Benzo(b)fluoranthene
Benzyl alcohol
Benzyl chloride
Beryllium
Bis (2-chloroisopropyl) ether
Bls(2-chlorethyl)ethec
Bls(2-ethylhexyl)phthalate
Bromodlchloromethane
Bromolorm (Trlbromomethane)
Bulanol
Butyl-4,6-dintlrophenol. 2-sec (DinoseD)'
Intensity frequency timing scale mode ot action
operational definitions tor stres'sor characteristics
bioaccumulation potential,
reported or log Kow > 5
,
AWQC. chronic toxicity
values below HBLs
.
«
endocrine disruptor;
rep/dev effects in multiple
species
'
persistent as a function ol
1/2 lite or log Kow
avian reproductive effects
(e.g., eggshell thinning)
B I I
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Table B-3. Stressor Characteristics Evaluated to Prioritize Constituents of Ecological Concern
stressor characteristics
CAS Number
85687
7440439
75150
56235
57749
126998
106478
108907
510156
124481
67663
95578
7440473
218019
7440508
108394
95487
106445
98828
72548
72559
50293
84742
117840
2303164
53703
96128
95501
Constituent name
Butylbenzylphlhalale
Cadmium
Carbon dlsultide
Carbon telrachlohde
Chlordane
Chlofo-1,3-butadiene, 2- (Chloroprene)
ChlOfoanlllne. p-
Chlorobenzene
Chlocobenzllale
Chlorodlbromomethane
Chlorolorm
Chlorophenol. 2-
Chromium VI
Chrysene
Copper
C re sol, m-
Cresol. o-
Cresol, p-
Cumene
ODD
DDE
DDT
Dl-n-butyl phthalale
Dl-n-oclyl phthalale
Dlallale
Dibenz(a,h)anlhracene
Dibromo-3-chloiopropane, 1 .2-
Dlchlorobenzerie, 1.2-
Intenslty frequency timing scale mode ot action
operational definitions lor stressor characteristics
bioaccumulation potential.
reported or log Kow > 5
.
A WQC. chronic toxicity
values below HBLs
' '
endocrine disrupter;
rep/dev effects in multiple
species
.
.'
persistent as a function ot
1/2 lite or log Kow
.
avian reproductive ertects
(e.g.. eggshell thinning)
,
v
li-IJ
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Table B-3. Stressor Characteristics Evaluates .o Prioritize Constituents of Ecological Concern
atreasor characteristics
CAS Number
106467
91941
75718
75343
107062
75354
156592
156605
120632
94757
78875
542756
10061015
10061026
60571
84662
56531
60515
131113
57976
119937
105679
119904
99650
51285
121142
606202
12:1911
Constituent name
Dtchlorobenzene. 1,4-
Olchlorobenzldine. 3,3'-
Dichtorodifluoromethane
Dichloroethane. 1,1-
Dlchloroettiane, 1,2-
Dichloroflthylene, 1,1-
Dichloroelhylene. cis-1,2-
Dlchtoroethyiene. trans- 1.2-
Dlchlorophenol, 2,4-
Dlchlorophenoxyacetlc acid, 2,4- (2,4-0)'
Dichloropropane, 1,2-
Dlchloropropene, 1.3-
Dtchloropropene, els- 1.3-
Dlchloropropene, trans- 1,3-
DlekJrtn
Dlelhyl ptithalate
Diethylstilbestrol
CNmethoale
Dlmalhyl phlhalale
Dlmelhylbeaz(a)anthracene, 7,12-
DtmeUiylbenzldine, 3,3'- '
Dimelhylphenol, 2.4- '
Dlmethyoxybenzldine, 3.3'- '
Dinllrobenzene, 1.3-
Dlnllrophenol. 2,4-
Oinitiotoluene. 2,4-
Diriilrotoluene. 2,6
Dioxana. 1,4- "
Intensity frequency timing scale mode of action
operational definitions for stressor characteristics
bioaocumulation potential.
reported or log Kow > 5
'
AWOC. chronic toxicity
values below HBLs
t
-
endocrine disrupter,
rep/dev effects in multiple
species
/
persistent as a function ol
1/2 Hie or log Kow
-
a wan reproductive effects
(e.g.. eggshell thinning)
U
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Table B-3. Stressor Characteristics Evaluated to Prioritize Constituents of Ecological Concern
stressor characteristics
CAS Number
122394
298044
115297
72208
106698
110805
141786
60297
97632
62500
100414
106934
96457
206440
86737
50000
64186
110009
76448
1024573
87683
118741
319846
319857
58899
77474
67721
70304
Constituent name
Dlphenylamine* .
Disullolon
Endosullan
Endrin
EpicMorohydrin .
Ethoxyethanol, 2- "
Ethyl acetate
Ethyl ether
Ethyl methacrylate
Ethyl methanesultonale
Ethylbenzene
Ethylene Dlbromide
Ethylene Ihiourea
Fluoranthene
Fluorene
Formaldehyde
Formic Acid*
Furan
Heptachlor
Heptachlor epoxlde
Hexachloro- 1 ,3-butadlene
Hexachlorobenzene
Hexachlorocyclohexane, alpha- (alpha BHC)
Hexachlorocyclohexane, beta- (beta BHC)
Hexachlorocyclohexane. gamma- (Llndane)
Hexachlorocyclopentadlene
Hexachloroethane
Hexachlorophene"
Intensity frequency timing .. scale mode of action
operational definitions tor stressor characteristics
bioaocumulation potential,
reported or log Kow > 5
AWQC, chronic toxicity
values below HBLs
'
endocrine disrupter,
rep/dev effects in multiple
species
persistent as a function ol
1/2 life or log Kow
.
avian reproductive effects
(eg, eggshell thinning)
H-I4
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Table B-3. Stressor Characteristics Evaluates .j Prioritize Constituents of Ecological Concern
stressor characteristics
-
CAS Number
193395
78831
78591
143500
7439921
7439976
126987
67561
72435
74839
74873
78933
108101
80626
298000
56495
74953
75092
7439987
621647
86306
100754
930552
91203
91598
7440020
9B953
79469-
Constituent name
Indenofl ,2,3-cd) pyrene
Isobutyt alcohol
Isophorone
Kepone
Lead
Mercury
Methacrytonitrile
Methanol
Methoxychlor
Methyl bromide (Bromomethane)
Methyl chloride (Chloromethane)
Methyl ethyl ketone
Methyl Isobutyt kelone
Methyl methacrylale
Methyl parathlon
Melhylcholanthrene, 3-
Methylene bromide
Methyleoe chloride
Molybdenum
N-Nltfosodl-n propylamlne
N-Nitrosodiphenylamine
N-Nilrosopipertdine
N-Nltrosopyrrolldlne
Naphthalene
Naphthylamlne*
Nickel
Nitrobenzene
Nitiopiopane. 2-
Intensity frequency timing scale mode of action
operational definitions for stressor characteristics
bioaccumulation potential,
reported or log Kow > 5
s
AWQC, chronic toxteity
values below HBLs
endocrine disruptor-
rep/dev effects in multiple
species
persistent as a function ol
1/2 lite or log Kow
.
avian reproductive effects
(e.g., eggshell thinning)
*
'
U-15
-------
Table B-3. Stressor Characteristics Evaluated to Prioritize Constil uents of Ecological Concern
stressor characteristics
CAS Number
924163
55165
62759
10595956
152169
56382
608935
82688
87865
108952
62384
108452
298022
1336363
23950585
129000
110861
94597
7782492
7440224
57249
100425
1746016
. 95943
630206
79345
I2/1H4
1)8902
Constituent name
Nllrosodi-n-butylamlne
Nltrosodlethylamine
Nitrosodimethylamine
Nitrosomethylethylamine
Octamethylpyrophosphoramide
Parathion
Pentachlorobenzene
PenlachtoronJIrobenzene (PCNB)
Pentachlorophenor
Phenol
Phenyl mercuric acetale
Plienylenedlamlne. m- '
Phorale
Polychlorinated blphenyls
Pionamlde
Pyrene
Pyrtdlne"
Satrole
Selenium
Silver
Strychnine'
Slryene
TCDD, 2,3,7,8-
Tetrachlorobenzene. 1.2,4,5-
Tetrachloroethane, 1,1,1.2-
Telrachloroelhane. 1.1.2,2-
TeUacliloioethylene
1 elrachlorophenol. 2.3,4.6-
Intenslty frequency timing scale . mode of action
operational definitions for stressor characteristics
bioaccumulation potential,
reported or tog Kow > 5
AWQC. chronic toxlcity
values below HBLs
-
-
endocrine disruptor,
rap/dev effects i,i multiple
species
persistent as a lunction ot
1/2 lite or log Kow
avian reproductive ettects
(e.g., eggshell thinning)
-
,
IM6
-------
Table B-3. Strcssor Characteristics Evaluated *o Prioritize Constituents of Ecological Concern
atressor characteristics
CAS Number .
3689245
7440280
108883
95807
95534
106490
8001352
76131
120821
71556
79005
79016
75694
95954
88062
93765
93721
96184
99354
126727
7440622
75014
1330207
7440666
Constituent name
Telraethyldllhlopyrophosphate
Thallium (I)
Toluene
Toluenediamine, 2.4-
Toluidine, o- '
Toluldlne, p- '
Toxaphene
Trtchloro-1 ,2,2-lrifluoroethane, 1 ,1 ,2-
Trtchloroberuene, 1,2,4-
Trichloroelhane, 1,1,1-
Trichloroethane. 1.1,2-
Trichloroethylene
Trichlorofluoromethane
Trichlorophenol, 2.4;5-
Trtchtorophenol, 2,4.6-
Trtchlorophenoxyacetic acid. 2.4,5- (245-T)'
Trtchloropherioxypropkmlc acid, 2.4,5- (Sltvex
Tflchloropropane, 1,2,3-
Trinitrobenzene. sym-
Tris (2.3-dlbromopropyl) phosphate
Vanadium
Vinyl chloride
Xylenes (total)
Zinc
Intensity frequency timing scale ' mode of action
operational definitions lor stressor characteristics
bioaccumulation potential,
reported or log Kow > 5
-
AWQC. chronic toxicity
values below HBLs
.»'
endocrine disruptor.
rep/dev effects in multiple
species
persistent as a function ol
1/2 lite or log Kow
avian reproductive effects
(e.g.. eggshell thinning)
^
1J-I7
-------
APPENDIX B
Ac«naphthen« 7
a
Table 4. Biological Uptake Properties
ecological
pecsptor
fi&fi
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF. or
B8AF
BCF
.
BAF
BCF
BcF
BCF
ftpM baeed or
MJXila !><»<
-------
APPENDIX B Acenaphthene - 8
References
Academy of Natural Sciences. 1981. Early Life Stage Studies Using the Fathead Minnow
(Pimephalas promelas) to Assess the Effects of Isophorone and Acenaphthene. Final
report to U.S. EPA, Cinn., OH. Academy of Natural Sciences, Philadelphia, PA., 26 pp.
As cited in U.S. Environmental Protection Agency, 1993i. Sediment Quality Criteria for
the Protection of Benthic Organisms: Acenaphthene. Office of Water, Office of Research
and Development, Office of Science and Technology, Health and Ecological Criteria
Division, Washington, D.C., EPA-822-R-93-013.
AQUIRE (AOJJatic Toxicity Information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MM, 'June 1995.
Barrows, M.E., S.R. Petrocelli, KJ. Macek and J.J. Carroll. 1980. Bioconcentration and
Elimination of Selected Water Pollutants by the Bluegill Sunfish (Lepomis macrochirus)
In: Dyn., Exposure Hazard Assess. Toxic Chem., (Pap. Symp. 1978). As cited in
AQUIRE (AQUatic Toxicity Information REtrieval Database). 1994. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency,. Duluth, MN.
Cairns, Michael A. and Alan V. Nebeker. 1982. Toxicity of Acenaphthene and Isophorone to
Early Life Stages of Fathead Minnows. Arch. Environm. Contam. Toxicol. 11:703-707.
EG&G Bionomics. 1982. Acute toxicity of selected chemicals to fathead minnow, water flea
and mysid shrimp under static and flow-through test conditions. Final report to U.S. EPA.
EG&G, Bionomics, 790 Main SL, Wareham, MA. 13pp. As cited in U.S. Environmental
Protection Agency, 1993i. Sediment Quality Criteria for the Protection of Benthic
Organisms: Acenaphthene. Office of Water, Office of Research and Development, Office
of Science and Technology, Health and Ecological Criteria Division, Washington, DC,
EPA-822-R-93-013. , ' '
August 1995
-------
APPENDIX B Acenaphthene - 9
ERCO. 1981. Toxicity Testing Inter-Laboratory Comparison Early-Life Stage Test with
Fathead Minnow. Final Report to U.S. EPA, Cinn., OH and U.S. EPA, Duluth MN.
ERCO/Energy Resources Co., Inc., 185 Alewife Biook Parkway, Cambridge, MA. 47 pp.
As cited in U.S. Environmental Protection Agency, 1993L Sediment Quality Criteria for
the Protection of Benthic Organisms: Acenaphthene. Office of Water, Office of Research
and Development, Office of Science and Technology, Health and Ecological Criteria
Division, Washington, DC, EPA-822-R-93-013.
Geiger, D.L., C.E. Northcott, D.J. Call, and L.T. Brooke. 1985. Acute Toxicities of Organic.
Chemicals to Fathead Minnows (Pimephalas promelas), Vol. 2, Center for Lake Superior
Environmental Studies, University of Wisconsin, Superior, WI, 326 p. As cited in
AQUIRE (AOUatic Toxicity Information REtrieval Database), 1994. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Holcombe, Gary W., Gary L. Phillips, and James T. Fiandt. 1983. Toxicity of Selected
Priority Pollutant to Various Aquatic Organisms. Ecotoxicology and Environmental
Safety, 7:400-409. .
LeBlanc, G. A. 1980. Acute Toxicity to Priority Pollutants to Water Flea (Daphnia magna).
Bull. Environm. Contain. Toxicoi, 24:684-691.
Mackay, D., S. Paterson, and W. Y. Shiu. 1992. The effect of metals on the growth and
reproduction of Eisenia foetida (Oligochaeta^ Lumbricidae). Pedobiologia 24:129- 137.
Marine Bioassay Laboratories. 1981. Flow-through early-life stage toxicity tests with fathead
minnows (Pimephales promelas). Final report to U.S. EPA, Duluth, MN. Marine
Bioassay Laboratories, 1234 Highway One, Watsonville, CA. 71 pp.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. October 1994.
Randalll, T.L. and P.V. Knopp. 1980. Detoxification of specific organic substances by wet
oxidation. J. Water Pollut. Control Fed. 52:2117 - 2130. As cited in U.S. Environmental
Protection Agency, 1993L Sedimefc Quality Criteria for the Protection of Benthic
Organisms: Acenaphthene. Office of Water, Office of Research and Development, Office
of Science and Technology, Health and Ecological Criteria Division, Washington, DC,
EPA-822-R-93-013.
August 1995
-------
APPENDIX B Acenaphthene - 10
Stephan, C.E. 1993, Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. PB93-154672.
Suter, G.W. El and J.B. Mabrey, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
85OR21400 Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D.C
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
U.S. Environmental Protection Agency, 1978. In-depth Studies on Health and Environmental
Impacts of Selected Water Pollutants. Contract No. 68-01-4646.. As cited in U.S.
Environmental Protection Agency, 1980. Ambient Water Quality Criteria for
Acenaphthene. Criteria and Standards Div., EPA-440/5-80-015, 49 p.
U.S. Environmental Protection Agency. 1989. Mouse Oral Subchronic Study with
Acenaphthene. Study conducted by Hazelton Laboratories, Inc., for the Office of Solid
Waste, Washington, DC. As cited in IRIS (Integrated Risk Information System). 1994.
U.S. EPA, Office of Research and Development, Office of Health and Environmental
Assessment
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment Washington, D.C. January.
U.S. Environments Protection Agency. 1993i. Sediment Quality Criteria for the Protection
of Benthic Organisms: Acenaphthene. Office of Water, Office of Research and
Development, Office of Science and Technology, Health and Ecological Criteria Division,
Washington, DC, EPA-822-R-93-013.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effe&s on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Freshwater Toxic... - Acenaphthene
Cos No.: 83-32-9
Chemical Name
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
aacenaphthene
Species
fathead
minnow
fathead
minnow
fathead
minnow
fathead
minnow £
fathead
minnow '
fathead
minnow
blueglll
rainbow
trout
brown trout
channel
catfish
aquatic
oraanlsms
Type of
Effect
acute
acute
acute
early
lifestoge
early
lifestooe
chronic
acute-
acute
acute
acute
chron.
Description
LC50
LC50
NOEC
NOEC
NOEC
CV
LC50
LC50
LC50
LC60
FCV
Value
1.600
1.730
413
50
64
169
1.700
670
580
1.720
23
Units
ug/l
UQ/I
UQ/I
ug/l
ua/l
UQ/I
ug/l
ug/l
ug/l
ua/l
ug/l
Test Type
(static/ flow
through)
flow-through
NS
flow-through
NS
NS
NS
static
flow-through
flow-through
flow-through
NS
Exposure
Duration/
Tbnlna
96-hour
96-hour
96-hour
NS
NS
NS
96-hour
96-hour
96-hour
96-hour
NS
Reference
Holcombe et al..
1983
Gelger et al..
1985 as cited In
AQUIRE. 1994
Cairns &
Nebeker. 1982
Academy of
Natural Sciences.
198 las cited in
U.S.EPA. 19931
ERCO, 198 las
cited In U.SiEPA.
19931
U.S.EPA. 19931
U.S.EPA. 1978 as
cited U.S.EPA.
1980
Holcombe et al..
1983
Holcombe et al..
1983
Holcombe et al..
1983
U.S.EPA, 19931
Comments
temperature of water = 22.9 C
; ,
early-life-stage test
.
S
Reported value is the
geometric mean of six values
temperature of water = 1 2 C
temperature of water = 1 2 C
temperature of water = 22 9 C
freshwater FCV of 22.96 ug/l is
based on FAV=80.01 ug/l. final
ACR=3 484
-------
Freshwater Biological Uptake Measures - Acenaphthene
Cos No.: 8J-32-9
Chemical Name
acenaphthene
acenaohthene
Species
btuegitt
aauatic oraanisms
B-factor(BCF,
BAF. BUR
BCF
BCF
Value
387
37.9
Measured or
Predicted
(m,p)
m
p
Units
NS
NS
. ' Reference
Barrows et al., I960 as
cited In AQUIRE, 1994
U.S. EPA. 1993b
Comments
Normalized to 4.8% lipid. 28-day. How-
through test
Normalized to 1% lipid
NS = not specified
-------
Freshwater Toxlclfy - Acenaphthene
Cos No.: 83-32-9
Chemical Name
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenaphthene
acenapi ithene
acenaphthene
acenaphthene
flc^pachihune
Species
Daphnla
mogna
Daphnio
magna
Daphnia
magna
Daphnla
manna
Oaphnla
magna
Daphnla
magna
fathead
minnow
fathead
minnow
fathead
minnow
Type of
Effect
acute
acute
acute
r
acute
acute
acute
acute
acute
qcute
Description
LC50/EC50
LC50/EC60
LC60/EC50
.LC60/ICSO
LC50
EC60
LC60/EC60
LC60/EC60
LC5D/EC5Q
Value
320
1.300
120
3,450
41,000
41.200
1.500
3,100
1?QQ
Units
UQ/I
ua/i
ua/l
Jfl/1
up/I
UQ/I
up/I
uayl
uy/I
Test Type
(static/ flow
through)
static
static
flow-thrpupr
static
static
static
static
static
renewal
Exposure
Duration/
Tlmlna
NS
NS
NS
NS
48-hour
49-hour
NS
NS
NS
Reference
EG&G.
Btonomlcs. 1982
as cited In
U.S.EPA. 19931
EG&G.
Bionomics. 1982
as cited In
U.S.EPA. 19931
EG&G.
Bionomics. 1982
as cited In
U.S.EPA; 19931
Randall and
Knopp. 1980 as
cttedln U.S.EPA.
19931
LeBlonc, 1980
U.S.EPA. 1978 as
cited U.S.EPA.
I960
EG&G.
Bionomics. 1982
01 cttedln
U.S.EPA. 19931
Marine BkXKtay
Laboratories.
198 las cited In
U.S.EPA. 19931
Academy of
Natural Science.
1981 as cited In
U.S-EPA. 19931
Comments
:
-------
Terrestrial Toxlcli, Acenaphthene
Cos No.: 83-32-9
Chemical Nam*
acenaphthene
acenaptiihene
Species
rat
mousg
Endpolnt
acute
subchronlQ
Daacrlpllon
L050
1
NOAEL
Value
600
175
Units
mg/kg
mo-ko/dav
Exposure
Route (oral,
S.C., I.V., (.p.,
Injection)
I.P.
oral
Exposure
Duration /
Timing
NS
90-4
Reference
RTECS. 1994
U.S.EPA. 1989 as
cited In IRIS. 1994
Comments
Mica were gavaged dally with 0.175.350. or 700 mg/kg-d.
The reauila Indicated no treatment-related effects on survival.
cHnlcal slgna. and body weight changes. The LOAcL Is 390
mo/ka-d based on hepatotoxlcltv.
NS Nbt 8pe«(!«d
-------
Terrestrial Biological Uptak. .Measures - Acenaphthene
Cos No.: 83-32-9
Cht>f t flcoi Worn§
qcenophthene
Soecles
plant
B-factor
(BCF. BAF.
DMn
BCF
Value
0,2
Measured
or
Predicted
(n\.o)
p
Units
(ug/g DW
plant)/(ug/g
soil)
Reference
U.S. EPA. 1990e
Comments
Plant uptake from soil pertains to
foraaed plants
-------
APPENDIX B Aldrin 1
lexicological Profile for Selected Ecological Receptors
Aldrin
Cas No.: 309-00-2
Summary: This profile on aldrin summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem, lexicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e.. Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) -are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log K^ between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^,) for ihe generic freshwater ec -(system. Table 1 contains benchmarks
fu/ mammals and birds associated wiih the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. However, several chronic and subchronic toxicity
studies involving aldrin have been conducted using laboratory rats and mice. F'r.zhugh et al.
(1964-as cited in ATSDR, 1992) observed increased mortality in a chronic study in which rats
were orally administered aldrin at doses of 0.5 mg/kg-day and 2.5 mg/kg-day- Another chronic
study was identified in which rats were fed a diet c-ntaining aldrin for a period of 25 months
(Deichmann et al., 1967 as cited in ATSDR, 1992); Deichmann et al. reported a NOAEL of 0.25
mg/kg-day for survival. A chronic reproductive study was identified in which groups of male
and female Swiss white mice (120 days) were fed a diet that contained 3, 5, 10, or 25 ppm for
a period covering six generations (Keplinger et al., 1970). Keplinger et al. (1970) observed a low
August 1995
-------
Freshwater .city - Aldrin
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
Daphnia magna
Daphnia magna
Simocephalus
serrulatus
bluegill
striped bass
rainbow trout
fathead minnow
Endpolnt
immob.
mort
immob.
mort
mort
mort
mort
Description
EC50
LC50
EC50
LC50
LC50
LC50
LC50
Value
28
30
23 - 32
(27.2)
4.6 - 13.0
(7.37)
7.2-10
(8.96)
2.2 - 17.7
(4.51)
32
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
NA
NA
Exposure
Duration/
Timing
48 hours
1 .08 days
48 hour
96 hour
96 hour
96 hour
96 hour
Reference
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
Comments
NA = Not applicable
NS = Not specified
-------
Freshwater Biological Uptake Measures - Aldr.n
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
Species
fish
fish
fish
mosquitofish
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BCF
BCF
Value
10712
3162
2879
3140
Measured or
predicted
(m,p)
m
m
P
NS
Units
L/kg whole-
body
L/kg whole-
body
NS
NS
Reference
Garten and Trabalka, 1983
Garten and Trabalka, 1 983
Slephan, 1993
Metcalf et al., 1973
Comments
Flowing water; All estimates were calculated
based on published data, th« type of studies
from which the data were taken were not.
specified.
Microcosm.
Normalized to 1.0 % lipid.
Whole body value.
NS = Not specified
-------
APPENDIX B Aldrin. 2
pre-weaning pup survival and reported a LOAEL of 10 ppm for reproductive effects. Based on
the reference body weight (kg) and the recommended value for food consumption (kg/day) for
mice reported in Recommendations for and Documentation of Biological Values for Use in Risk
Assessment (U.S. EPA, 1988), a LOAEL of 2 mg/kg-day was calculated!
The LOAEL in the Keplinger et al. (1970) study was chosen to derive the lexicological
benchmark because (1) chronic exposures were administered via oral ingestion, (2) it focused on
reproductive toxicity as a critical endpoint, and (3) the study contained dose-response
information. The study by Fitzhugh et al. (1964 as cited ATSDR, 1992) was not selected for the
derivation of a benchmark because of the lack of dose-response data and because it does not
evaluate reproductive or developmental endpoints. Similarly, the study by Deichmann et al.
(1967 as cited in ATSDR, 1992) was not selected because the study lacked dose-response data
and because it did not evaluate reproductive or developmental endpoints.
The study value from Keplinger et al. (1970) was divided by 10 to provide a LOAEL-to-
NOAEL safety factor. This value was then scaled for species representative of a freshwater
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994):
w<
Benchmark^ = NOAEL, x - L
\bww)
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Keplinger et
al. (1970) study documented reproductive effects from aldrin exposure to male and female mice,
the mean body weight of both genders of representative species was used in the scaling algorithm
to obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of aldrin, as well as growth
or chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations, and during sensitive life stages. There were several study
values in the data set corresponding to renal and hepatic endpoints, which were equal to or lower
than the benchmark value. Most of the studies identified were conducted using laboratory rats
or mice and, as such, inter-species differences among wildlife species were not identifiable and,
therefore an inter-species uncertainty factor was not applied. Based on the data set for toxaphene
and because the benchmark is based on a LOAEL/10, the benchmarks developed from the
Keplinger et al. (1970) study were categorized as provisional with a "*" to indicate that some
adverse effects may occur at the benchmark level.
Birds: No subchronic or chronic studies were found for representative avian species in which
dose-response data were reported. However, chronic toxicity studies conducted using chickens,
quail, and pheasants were identified. In the study on quail, DeWitt (1956) also reported a
August 1995
-------
APPENDIX B Aldrin-3
NOAEL of 1 mg/kg-day for reproductive effects. DeWitt (1956) reported a LOAEL of 0.06
mg/kg-day for reproductive effects on pheasants. A subchronic study was identified in which
chicken eggs were injected with 0 to 2 rng/egg of aldrin on day 7 of incubation (Smith et al.,
. 1970). Smith et al., (1970) reported a NOAEL of 36.36 mg/kg (2 mg/egg) for egg hatchability
and morphological changes. In a chronic reproductive study (Brown et al., 1965), chickens were
fed a diet containing aldrin for a period of two years. No effects on fertililty and hatchability
were observed at 1 ppm (NOAEL). Based on the reference body weight (kg) and the
recommended value for food consumption (kg/day) for chickens reported in Recommendations
for and Documentation of Biological Values for Use in Risk Assessment (U.S. EPA, 1988), a
NOAEL of 0.07 mg/kg-day was calculated.
The LOAEL reported by DeWitt (1956) in his experiments on pheasants was used to calculate
the lexicological benchmark for birds because it focused on reproductive toxicity as a critical
endpoint and aldrin was administered via oral ingestion. The study by Smith et al. (1970) on
chicken eggs was not selected for benchmark derivation because data were not identified on
either: (1) direct absorption of dieldrin from direct contact with the eggs or (2) maternal transfer
from mother to egg. Without these absorption data, it is not possible to estimate the internal dose
from the injected dose. The study by Brown et al. (1965) was not selected because it was not
the lowest value in the data set for appropriate endpoints.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for aviari species were not identified. Thus, for the avian species
representative of a freshwater ecosystem, the LOAEL of 0.06 mg/kg-day from the DeWitt (1956)
study was divided by 10 to provide for a LOAEL to NOAEL safety factor, and scaled using the
cross-species scaling method of Opresko et al. (1994). Since the Dewitt
U956) study documented reproductive effects from aldrin on female pheasants, female body
weights for each representative species were used in the scaling algorithm to obtain the
lexicological benchmarks.
Data were available on reproductive and developmental effects as well as on growth or survival.
In addition, the data set contained studies that were conducted over chronic durations. No studies
conducted during sensitive life stages were located. There were no other values in the data set
which were lower than the benchmark value. Laboratory experiments of similar types were not
conducted on a range of avian species and, as such, inter-species differences among wildlife
species were not identifiable. Based on the avian data set for aldrin, the benchmarks developed
from the DeWitt (1956) study were categorized as provisional.
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
aldrin. Therefore, a Secondary Chronic Value (SCV) of 1.8 E-5 mg/1 was calculated using the
Tier H methods described in Section 4.2.5. Because the benchmark was derived using the Tier
II method, it was categorized as interim.
August 1995
-------
APPENDIX B Aldrin - 4
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECW) for a species of
freshwater algae, frequendy a species of green algae (e.g., Selenastrwn capricornutwn). Aquatic
plant data was not identified for aldrin and, therefore, no benchmark was developed.
Bent hie community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). The EQp number is the chemical concentration that may be present in sediment while still
protecting the benthic community from harmful effects from chemical exposure. Because no
FCV was available, a Secondary Chronic Value (SCV) was calculated as described in Section
4.3.5. The SCV reported for aldrin was used to calculate an EQp number of 43.9 mg aldrin per
kg organic carbon. Assuming a mass fraction of organic carbon for the sediment (f^ of 0.05,
the benchmark for the benthic community is 2.19 mg aldrin per kg of sediment. Because the
E(X number was set using a SCV derived using the Tier II method, it was categorized as
interim.
August 1995
-------
APPENDIX B
Aldrin . 5
Table I. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R»pr*Mnl*tiv+
SfHttfe*
mink
river otter
bald eagle
osprey
great blue heron .
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
Vajg*' togfat
*«y
0.052 (p')
0.031 (p')
0.004 (p)
0.005 (p)
0.005 (p)
0.006 (p)
0.006 (p)
0.013 (p)
0.06 (p)
0.09 (p)
Study
Sp*»J6*
mouse
mouse
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
£lf*cl
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study Vaiu*
ma/fca-day
1.2
1.2
0.06
0.06
0.06
0.06
0.06
0.06
0.06
0.06
Description
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
SF
' 10
10
10
10
10
10
10
10
10
10
Qrig(tt*t8«Ufl*
Keplingar et a)., 1970
Keplinger et al., 1970
DeWitt, 1956
DeWitt. 1956
Do Witt, 1956
DeWitt, 1956
DeWitt. 1956
OeWiS, 1956
OeWitt, 1956
OeWitt, 1956
Benchmark Category, a - adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftepfwrnttativ*
Sp*d»*
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark
Value*
»B/U
1 .8 E-05 (i)
10
2.19 (i) mg/kg
sediment
Study
Specie*
AWQC
organisms
AWQC
organisms
Description
scv
SCV x K^.
Original Sourc*
GLI, 1992 .
GLI, 1992
'Benchmark Category, a = adequate, p = provisional, i - interim; a '*' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
10 = Insufficient Data.
August 1995
-------
APPENDIX B Aldrin . 6
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Gpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to aldrin. Because
of the lack of additional mammalian toxicity studies, the same surrogate-species study
(Keplinger et al., 1970) was used to derive the aldrin lexicological benchmark for mammalian
species representing the terrestrial ecosystem. The study value from the Keplinger et al.
(1970) study was divided by 10 to provide for a LOAEL-to-NOAEL safety factor. This value
was then scaled for species in the terrestrial ecosystem using a cross-species scaling algorithm
adapted from Opresko et al. (1994). Since the Keplinger et al. (1970) study documented
reproductive effects from aldrin exposure to male and female mice, the mean of the body
weights for the gender of each representative species was used in the scaling algorithm to
obtain the lexicological benchmarks.
Based on the data set for toxaphene and because the benchmark is based on a LOAEL/10, the
benchmarks developed from the Keplinger et al. (1970) study were categorized as
provisional*, as in the aquatic ecosystem.
Birds: Two suitable subchronic or chronic studies were found for representative avian species
in which dose-response data were reported. As in the freshwaier ecosystem the pheasant
study by DeWitt (1956) was used to calculate ihe benchmarks for birds in the generic
terrestrial ecosystem. The study value of 0.06 mg/kg from the DeWitt (1956) study was
divided by 10 to provide a LOAEL-to-NOAEL safety factor. The LOAEL/10 was then
scaled for the representative species using the cross-species scaling algorithm adapted from
Opresko et al. (1994). Since the Dewitt (1956) study documented reproductive effects from
aldrin on female pheasants, female body weights for each represeniative species were used in
the scaling algorithm to obtain the lexicological benchmarks. Because the behnchmarks was
derived from a LOAEL/10, they were categorized as provisional.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
August 1995
-------
APPENDIX B Aldrin - 7
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for aldrin and, as a result, a benchmark
could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available. .
August 1995
-------
APPENDIX B
Aldrin - 8
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
R«prM*ntaitv*
Sp*C>**
dear mouse
short-tailed
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
white- tailed deer
red- tailed hawk
American kestrel
Northern
' bobowhite
American robin
American
woodcock
. plants
soil community
Benchmark
Vato«*
jngftg-d*?
0.140 (p*)
0.144 (p')
0.122 (p')
0.049 (p*)
0.036 (p')
0.034 (p')
0.01 7 (p*)
0.006 (p)
0.010 (p)
0.009 (p)
0.011 (p)
0.009 (p)
10
ID
Study
Spdcto
mouse
mouse
mouse
mouse
mouse
mouse
mouse
pheasant
pheasant
pheasant
pheasant
pheasant
-
-
Effect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
}
rep
rep
Study
Value
mo/kg*
d«y5
1.2 '
1.2
1.2
1.2
1 1.2
1.2
1.2
0.06
0.06
0.06
0.06
0.06
Description
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
'. LOAEL
LOAEL
-
9f
10
10
10
10
10
10
10
10
10
10
10
10
-
-
orT0in*i Sww*
Keplinger el al.,
1970
Keplinger et al.,
1970
Keplinger et al..
1970
Kepingeretal.,
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
DeWJtt, 1956
DeWitt, 1956
DeWitt 1956
DeWitt, 1956
DeWitt, 1956
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID= Insufficient Data
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
August 1995
-------
APPENDIX B Aldrin. 9
protective surface water and soil concentrations for constituents considered to bioconcentrate
anoVor bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The brief discussion below describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values.
As stated in section 5.3.2, the BAF/s for consituents of concern were generally estimated
using Thorhann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for aldrin. The
bioconcentration factor for fish was also estimated from the Thomann models (i.e., log Kow -
dissolved BCF/) and multiplied by the dissolved fraction (/j) as defined in Equation 6-21 to
determine the total bioconcentration factor (BCF/). The dissolved bioconcentration factor
(BCF/1) was converted to the BCF/ in order to estimate the acceptable lipid tissue
concentration (TC/) in fish consumed by piscivorous fish (see Equation 5-115). The BCF/
was required in Equation 5-115 because the surface water benchmark (i.e., FCV or SCV)
represents a total water concentration (C1). Mathematically, conversion from BCF/1 to BCF/
was accomplished using the relationship delineated in the Interim Report on Data and
Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S.
EPA, 1993i):
BCF/1 x fd = BCF/
Converting the predicted BCF;d of 3,133,284 L/kg LP to the BCF/ of 313,328 17kg LP was
not in close agreement with the single measured BCF/ of 41,300. It is difficult to determine
why the predicted value is larger than the measured value by almost a factor of 8. Although
the difference may be partly attributed to the limited data set on measured bioconcentration
factors, it is more likely the result of the rapid metabolism of aldrin to dieldrin in fish tissue.
Since dieldrin is both toxic and persistent, Stephan (1993) points out that a predicted BAF for
"aldrin plus toxic persistent metabolites" will likely be higher than a measured BAF for just
aldrin if the measured value does not account for toxic metabolites. Because measured
biological uptake factors probably do not account for toxic metabolites, a predicted value for
aldrin was preferred over the single measured value because the predicted value. Similarly,
the lower measured values cited in Garten and Trabalka (1983) for bioaccumulation in fish
August 1995
-------
APPENDIX B Aldrin - 10
were considered to be low estimates of the bioaccumulation potential of aldrin and its toxic
metabolites.
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of several
values, with sources in Table 4 (see master table). For earthworms and terrestrial
invertebrates, the bioconcentration factors were estimated as described in Section 5.3.5.2.3.
Briefly, the extrapolation method is applied to hydrophobic organic chemicals assuming that
the partitioning to tissue is dominated by lip ids. Further, the method assumes that the BAFs
and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data (in this case aldrin) is compared to the BBF for TCDD and
that ratio (i.e., aldrin BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial
vertebrates, invertebrates, and earthworms, respectively. For hydrophobic organic
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf
translocation, direct deposition on leaves and grasses, and uptake into the plant through air -
diffusion. For metals, empirical data were used to derive the BCF for aboveground forage
grasses and leafy vegetables.
August 1995
-------
APPENDIX B
Aldrin - 11
Table 4. Biological Uptake Properties
co logical
r»c*ptor
limnetic trophic
leva! 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
Ifpid-baMd or
whole-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole- body
whole- body
whole-plant
value
28,894,111 (d)
14,769,414 (d)
43,866 (t)
4 1.938. 806 (d)
48,382.161 (d)
52,309,538 (d)
3.2
0.037
0.3
0.0068
ource
predicted value based on
Thomann, 1 989. food chain
model
predicted value based on
Thomann, 1989, food chain
model
predicted value hasfftj on
Thomann, 1989 and adjusted to
estimate total BCF
predicted value based on
Thomann et a!., 1992, food web
model
predicted value based on
Thomann et at.. 1992. food web
model
predicted value based on
Thomann et al., 1992. food web
model
geometric mean of values in
Garten and Trabalka. 1983;
Clabom et al., 1956, 1960 as
cited in Kenaga. 1980
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDO
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDO
U.S. EPA, 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Aldrin. 12
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U!S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE (AOt/atic Toxicity_/nformation /?Etrieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Al-Hachim, G.M. 1971. .Effect of aldrin on the condition avoidance response and
electroshock seizure threshold of offspring from aldrin-treated mother.
Psychopharmacologia 21:370-373.
Brown, V.K.H., A. Richardson, J. Robinson, and D.E. Stevenson. 1965. The effects of aldrin
and dieldrin on birds. Food Cosmet. Toxicol. 3:675-679.
Claborn, H.V. 1956. Insecticide Residues in Meat and Milk. ARS-33-25. U.S. Department
of Agriculture. As cited in Kenaga, E.E, 1980, Correlation of bioconcentration factors of
chemicals in aquatic and terrestrial organisms with their physical and chemical properties,
Environmental Sci. Technol. 14(5):553-556.
Clabom, H.V., R.D. Radeleff, and R. C. Bushland. 1960. Pesticide Residues in Meat and
Milk. ARS-33-63. U.S. Department of Agriculture. As cited in Kenaga, E.E, 1980,
Correlation of- bioconcentration factors of chemicals in aquatic and terrestrial organisms
with their physical and chemical properties, Environmental Sci. Technol. 14(5):553-556.
Deichmann, W.B., M. Keplinger, and F. Sala et al. 1967. Synergism among oral
carcinogens: IV. The simultaneous feeding of four tumorigens to rats. Toxicol. Appl.
Pharmacol. 11:88-103.
Deichmann, W.B., W.E. MacDonald, and E. Blum et al. 1970. Tumorigenicity of aldrin,
dieldrin, and endrin in the albino rat Ind. Med. Surg. 39:426-434.
DeWitt, J.B. 1956. Chronic Toxicity to Quail and Pheasants of Some Chlorinated
Insecticides. Pesticide Toxicity, Vol. 4, No. 10, pp. 863-866.
Epstein, S.S., E. Arnold, and J. Andrea et al. 1972. Detection of chemical mutagens by the
dominant lethal assay in the mouse. Toxicol. Appl. Pharmacol. 23:288-325.
Fitzhugh, O.G., A.A. Nelson, and M.L. Quaife. 1964. Chronic oral toxicity of aldrin and
dieldrin in rats and dogs. Food Cosmet. Toxicol. 2:551-562.
August 1995
-------
APPENDIX B Aldrin . 13
57 FR 24152. June5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg%/day.
Garten, C. T., Jr., and J. R. Trabalka. 1983. Evaluation of models for predicting terrestrial
food chain behavior of xenobiotics. Environ. Sci. Technol. 26(10):590-595.
Great Lakes Water Quality Initiative (GLI), 1992. Tier II Water Quality Values for
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Keplinger, M.L., W.B. Deichmann, and F. Sala. 1970. Effects of combinations of pesticides
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Mehrotra, B.D., K.S. Moorthy, and S.R. Reddy et al. 1989. Effects of cyclodiene
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NCI (National Cancer Institute). 1978. Bioassay of alar in and dieldrin for possible
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Occupational Safety and Health, Washington, DC.
August 1995
-------
APPENDIX B Aldrin-14
Smith, S.I., C.W. Weber, and B.L. Reid. 1970. The effect of injection of chlorinated
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Suter n, G.W., J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
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August 1995
-------
APPENDIX B Aldrin - IS
U.S. EPA (U.S. Environmental Protection Agency). 1993b. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006. Office of
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Assessment of2j,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Life and Associated Wildlife.
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Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential Contaminants of
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Department of Energy.
World Health Organization (WHO), 1989. Environmental Health Criteria 91 - Aldrin and Dieldrin.
August 1995
-------
rerrestria! Toxicity - Aldrin
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
rat
rat
mouse
mice
-
mice
mouse
mouse
mouse
hamsters
quail
pheasants
chickens
Endpolnt
mort.
mort.
syst
terr. '
rep
rep
mort.
mort.
dev
rep
rep
rep
Description
NOAEL
LOAEL
LOAEL
AEL
LOAEL
LOAEL
NOAEL
LOAEL
AEL
NOAEL
LOAEL .
NOAEL
Value
4
8
2
25
-
1.2
0.5
1.3
26
50
1
0.06
1
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mo/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-diet
J
mg/kg-day
mg/kg-diet
Exposure
Route (oral,
s.c., l.v., l.p.,
Injection)
oral
oral
oil (gavage)
oil (gavage)
oral
oil (gavage)
oral
oral
oil (gavage)
oral
oral
oral
Exposure
Duration/Timing
6 weeks ad lib
6 weeks ad lib
5-7 days
once on gestation day 9
6 generations with 2
litters/generation
5 days, once/day
6 weeks ad lib
6 weeks ad lib
once on gestation day 7,
8, or 9
127 days
162 days +
2 years
Reference
NCI, 1978
NCI, 1978
AI-Hachim, 1971
Ottolenghi et al , 1974
Keplinger et at.. 1970
Epstein etal., 1972
NCI, 1978
NCI, 1978
Ottolenghi etal.. 1974
DeWitl, 1956
DeWitt, 1956
Brown etal., 1965
Comments
Effect = 2 of 10 died.
Effect = 'increased seizure
threshold in offspring'.
Effect = 'webbed feet.'
High litter mortality occurred at 25
mg/kg. At 10 mg/kg. low pre-
weaning pup survival.
Effect = 'decreased male fertility"
2 of 10 died. -
Increased fetal mortality.
No effects on egg production,
percentage fertility, or percentage
hatchability.
Egg production had virtually
ceased by the end of the tenth
week; however, it remained at
normal levels during the first six
weeks.
Fertility and hatchability were
normal at this dose level, (single
dose)
-------
Terrestrit /xicity - Aldrin
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
at
at
rat
ral
rat
rat
rat
ral
rat
rats
rat
rat
Endpolnt
mort.
mo rt
mort.
kidney
mbrt.
mort.
hepatic
sysl.
syst.
rep
neuro
neuro '
Description
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
Value
1.5
2.5
0.25
0.5
2.5
5
7.5
0.5
2.5
12.5
5
10
Units '
mg/kg-day
mg/Kg-day
mg/kg-day
mg/kg-day
mo/kg-day
mg/kg-day,
mg/kg-day
mg/kg-day
mg/Kg-day
mg/kg-diet
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p.,
njectlon)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oil (aavage)
oil (gavage)
Exposure
Duration/Timing
31 months, 7 days/week
31 months, 7 days/week
25 months ad lib
2 year
2 year ad lib
2 year ad lib
2 year ad lib
2 year ad lib
2 year ad lib
Three consecutive
generations
3 days, once/day
3 days, once/day '
Reference
Deichmann et al.,1970
Deichmann et al., 1970
Deichmann et al., 1967
Reuber, 1980
Fitzhughetal., 1964
Fitzhughetal.. 1964
Fitzhughetal., 1964
Fitzhughetal., 1964
Fitzhughetal., 1964
Treon and Cleveland.
1955
Mehrotra et al., 1989
Mahroira et al., 1989
Comments
-
An effect of 'decreased lifespan
in females' was observed.
-
An effect of Nephritis was noted.
An effect of 'increased mortality'
was observed.
Hepatic effects.
Renal.
Renal; effect = 'bladder
hemorrhages'.
At 12.5 mg/kg = 'marked
increase in mortality in pre-
weaning pups'; 'no effect on
reproductive capacity* at any
dose.
Effects = convulsions
-------
Terrestrial Toxicity - Aldrin
Cas No. 309-00-2
Chemical .
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
mammal
wild bird
tulvous whistling
duck
mallard
bobwhite
pheasant
mule deer
Endpolnt
mort.
morl.
mod.
mort.
mort.
mort.
mort.
Description
LD50
L050
LD50
LO50
LD50
LD50
LD50
Velue
39
7200
29.2
520
6.59
16.8
18.8-37.5
Units
mg/kg-
body wt.
ug/kg-body
wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wl.
mg/kg-
bodywl.
Exposure
Route (oral,
SC.. I.V.. I.D
Injection)
oral
oral
oral
oral
oral
oral
oral
Exposure
Ouratlon/Tlmlnq
NS
NS
NS
NS
NS
NS
NS
Rt ,rence
RTECS, 1994
RTECS, 1994
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
Peripheral nerve and sensation;
behavioral effects.
-
NS = Not specified
-------
Terrestrial xicity - Aldrin
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
egg (chicken)
rat
mouse
dog
rabbit
guinea pig
hamster .
pigeon
chicken
quail
duck
Endpolnt
dvp
mart.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort
Description
NOAEL
L050
LD50
LD50
LD50
LD50
LD50
L050
LD50
LD50
LD50
Value
2
39
44
65
50
33
100
56200
10
42100
520
Units
mg/egg
mg/kg-
body wt.
mg/kg-
bodywt.
mg/kg-
bodywt
mg/kg-
body wt.
mg/kg-
bodywt.
mg/kg-
bodywt.
ug/kg-body
wt.
mg/kg-
bodywl
ug/kg-body
wt.
mg/kg-
body w|.
Exposure
Route (oral,
S.C., I.V.. l.p.,
Inlectlon)
inject
oral
oral
oral
oral
oral
oral
oral
oral
oral
oial
Exposure
Duration/Timing
njected either prior to
incubation or after a 7-
day incubation period
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Smith etal., 1970
RTECS. 1994
RTECS, 1994
RTECS. 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RT£CS, 1994
Comments
Injection ot 5 mg of aldrin/egg by
(Dunachie and Fletcher. 1966)
resulted in only 50% hatchability.
- '
Behavioral effects.
.
-------
Terrestrial Biological. .ake Measures - Aldrin
Cas No. 309-00-2
Chemical
Name
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
aldrin
Species
cattle
cattle
swine
swine
swine
cattle (beef)
cattle (milk)
sheep
poultry
cow
swine
plants
B-factor
(BCF, BAF.
BMR
BCF
BCF
BCF
BCF
BCF
BTF
BTF
BAF
BAF
BAF
BAF
BCF
Value
2
3.5
2.4
3.8
1.4
0.085
0.02398
4.17
12.3
3.24
2.34
0.01
Measured or
predicted
(m,p)
m
m
m
m
m
m
m
P
P
P
P
P
Units
NS
NS
NS
NS
NS
NS
NS
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(ug/g DW
plant)/(ug/g
soil)
Reference
Claborn, et.al., 1960 as cited in
Kenaga, 1980
Claborn, et.al., 1960 as cited in
Kenaga, 1980
Claborn, et.al., 1956 as cited in
Kenaga, 1980
Claborn, et.al., 1956 as cited in
Kenaga, 1980
Clabom, et.al., 1956 as cited in
Kenaga, 1980
Travis and Arms, 1988
Travis and Arms, 1988
Garten and Trabalka, 1 983
Garten and Trabalka, 1983
Garten and Trabalka, 1 983
Garten and Trabalka, 1 983
U.S. EPA, 1990e
Comments
BTF = Biotransfer factors.
BTF = Biotransfer factors.
NA = Not applicable
-------
APPENDIX B Antimony-1
Toxicological Profile for Selected Ecological Receptors
Antimony
CasNo.: 7440-36-0
Summary: This profile on antimony summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population susiainabilily. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the data presented in the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ro) for the generic freshwater ecosystem. Table 1 coniains
benchmarks for mammals and birds associated wilh the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic planis, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several toxicily studies were identified which focused on Ihe effects of antimony
on laboratory mammals. Schroeder et al. (1969) exposed Long-Evans rats to 5 ppm of
poiassium antimony tartrate in their drinking water from weaning until natural death. A
decrease in ihe median lifespan was observed as well as abnormal serum glucose levels. The
ppm value was convened lo a daily dose ihrough the use of ihe geomean of ihe reported body
weights, 0.167 kg and daily water consumption given by ihe equation:
Water Consumption = 0.10W0'7377 (Nagy, 1987), where W is ihe body weighi in kg.
August 1995
-------
APPENDIX B Antimony - 2
The daily dose was calculated in this way as being equal to 0.8 mg/kg-day. In a separate
study by Schroeder et al. (1968), the effects of 5 ppm of antimony potassium tartrate in
drinking water was observed in randombred Charles River CD mice. A decrease in the
median lifespan of females and growth suppression in animals at 18 months of age was
observed at this dose. The ppm value was converted to a daily dose by using the geomean of
the reported body weights which was equal to 42.2 g and the daily water consumption
through the use of the equation presented above (Nagy, 1987). The daily dose was calculated
in this way to be 1.14 mg/kg-day. Rossi et al. (1987) observed reduced pup body weight
from the 10th to the 60th day of age in pups whose mothers had been exposed to 1 mg/dl
antimony trichloride. In this study, pregnant rats were exposed to 0.1 and 1 mg/dl antimony
trichloride in their drinking water from the first day of pregnancy until weaning (22nd day
after delivery) and a NOAEL of 0. 1 mg/dl was reported. Based on the geomean of the
reported body weights, 255 g, and daily water consumption estimated through the use of the
equation presented above (Nagy, 1987), a NOAEL of 0.162 mg/kg-day was calculated.
The studies by Schroeder et al. (1968) and (1969) were not selected for the derivation of a
benchmark because the studies did not evaluate reproductive or developmental endpoints. The
NOAEL in the Rossi et al. (1987) study was selected to derive the toxicological benchmark
because 1) it focused on developmental toxicity as a critical endpoint, 2) the study contained
adequate dose-response information and 3) chronic exposures were administered via oral
ingestion.
The study value from Rossi et al. (1987) was then scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
Benchmark^ = NOAEL. x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Rossi et al (1987) study documented developmental effects in pups of exposed female rats,
female body weights for each representative species were used in the scaling algorithm to
obtain the toxicological benchmarks. Based on the dataset for antimony, the benchmarks
developed from the Rossi et al. (1987) study were categorized as adequate.
Birds: No subchronic or chronic studies were identified which studied the toxicity effects of
orally ingested antimony in avian species.
Fish and aquatic invertebrates: The proposed Final Chronic Value (FCV) 3.0 E-02 mg/1
reported in the AWQC document for antimony (U.S EPA, 1980) was selected as the
August 1995
-------
APPENDIX B Antimony - 3
benchmark value protective of fish and aquatic invertebrates. Because the benchmark is
based on an FCV developed for a AWQC, it was categorized as adequate.
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). The aquatic
plant benchmark for antimony is 0.61 mg/1 based on a 4-day EC50 for chlorophyll A
inhibition in Selenastrum capricornutum (Suter and Mabrey, 1994). As described in Section
4.3.6, all benchmarks for aquatic plants were designated as interim.
Benthic community: The antimony benchmark protective of benthic organisms is pending a
U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
Terrestrial Biological Uptake Measun . - Antimony
Cas No. 7440-36-0
Chemical
Name
antimony
antimony
Species
plant
plants
B-factor
(BCF. BAF.
BMP)
BCF
BCF
Value
0.2
2.0 E-01
Measured
or
Predicted
(m.P)
P
m
units
(ug/g DW
plant)/(ug/g soil)
(ug / kg dw)/ (ug
/kg soil)
Reference
U.S. EPA, 1990e
Baesetal., 1984
Comments
-------
Freshwater Biological U^ .^e Measures - Antimony
Cas No. 7440-36-0
Chemical
Name
Antimony
Antimony
Antimony
-
Species
fish
bluegill
fish
B-factor
(BCF, BAF.
BMP)
BCF
BCF
BCF
Value
1
0
0
Measured
or
Predicted
(m,p)
m
m
m
Units
L/Kg
NS
L/Kg
Reference
U. S. EPA. 1992
Barrows et al., 1980 as cited in U.
S. EPA, 1993D
Stephan, 1993
Comments
Normalized to 3% lipid.
No increase above controls was detected
in whole body measurements.
-------
Freshwater Toxiclty - Anflrnony
C is No. 7440-36-0
Chemical
Name
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Species
aquatic
organisms
fish
daphnid
Fish
daphnid
minnow
Type of
Effect
chronic
chronic
chronic
chronic
chronic
acute
Description
SCV
CV
CV
EC20
EC20
NOEC
Value
104
1600
5400
2310
1900
6200
Units .
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS
NS
NS
NS
4 days
Reference
Suler and Mabrey, 1994
Suter and Mabrey, 1994
Suter and Mabrey. 1994
Suter and. Mabrey. 1994
Suter and Mabrey, 1 994
Heitmuller et at. 1981 as
cited in AQUIRE, 1995
Comments
-------
Terrestrial To. ..cy - Antimony
Cas No. 7440-36-0
rf>
1 r
I
Chemical
<*»_
48fpi«^
^l»£g£!
4
Antimony
Species
i
>,
1
rat
Endpoint
,
dev
Description
NOAEL
-^^
Value
0.162
Units
mg/kg-d
Exposure
Route (oral.
s.c., i.v.. i.p..
injection)
oral
. r~
»
Exposure Duration
/Timing
.
60 d
Reference
Rossi el al., 1987
Comments
doses were 0, 1 , 10 mg/l
0 162, 1 62 mg/kg-d) in
drinking water ad libitum.
For dose conversion used
body wl =255 g (in study).
and water consumption rate
(Nagy, 1987) Effect was
sig. reduced pup body
weight during suckling
period (0-22d)
-------
APPENDIX B
Antimony - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with a Freshwater Ecosystem
Rpj>r*»an|iHfv»
Specie* '
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring guU
kingfisher
BanchmwRVatua
rag/kg-d.
0.1 3 (a)
0.07 (a)
ID
ID
ID
.ID
ID
ID
ID
ID
W#H
Specie*
rat
rat
-
'
-
-
-
-
Cffeot
dev
dev
-
. . .
-
Study V«Jue
mg)fcg-d
0.162
0.162
-
-
-
-
-
-
Description
NOAEL
NOAEL
-
-
-
-
-
Sf
.
-
Ortalh»l«o*K»
Rossi etal., 1987
Rossi el al.. 1987
-
-
-
-
Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a () indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other advene effecti.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Representative :
Specie*
Fish and aquatic
invertebrates
aquatic
plants
benthic community
Benchmark -i
V*to«*
1 <"";«tft-"" -\
3.0 E-02 (a)
0.61 (i)
under review
Study ..
5p4cl«*
^ > f
aquatic
organisms
Setonastnim
capricomutum
Orfgtoa*
Vah*
- rtgrV * ;
3.0 E-02
0.61
Sttoitrttai
FCV
ECM
-
"* ss \ ^ % "" \ :
Ortfcfrii* Sou** ;
AWQC Table
Suter & Mabrey.
1994
-
'BenchmanX Category, a = adequate, p = provisional, i = interim: ID = insufficient data, a (') indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Antimony-5
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, several toxicity
studies were identified that focused on the effects of antimony on mammals.
Since no additional studies were identified which focused on reproductive or developmental,
toxicity in terrestrial wildlife, the same surrogate study (Rossi et al., 1987) was used to
calculate benchmark values for mammalian species representing the general terrestrial
ecosystem. The NOAEL from the Rossi et al. (1987) study was scaled for species in the
terrestrial ecosystem using the cross-species scaling algorithm adapted from Opresko et al.
(1994). Since the Rossi et al. (1987) study documented developmental effects from antimony
exposure to female rats, female body weights for each representative species were used in the
scaling algorithm to obtain the lexicological benchmarks. Based on the dataset for antimony,
the benchmarks developed from the Rossi et al. (1987) study were categorized as adequate.
Birds: No subchronic or chronic studies were identified which studied the toxicity effects of
orally ingested antimony in avian species.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the Lowest Observable Effects Concentration
(LOEC) values and then approximating the 10th percentile. If there were 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth.yield
reductions, or other effects reasonably assumed to impair the ability of a plant population to
sustain itself, such as a reduction in seed elongation. The benchmark for terrestrial plants was
5 mg/kg based on unspecified toxic effects on plants grown in a surface soil with the addition
of 5 ppm antimony (Kabata-Pendias and Pendias, 1984 as cited in Will and Suter, 1994).
This value was the lowest LOEC presented by Will and Suter (1994). The terrestrial plant
benchmark was categorized as interim, since less than 10 studies were presented by Will and
Suter (1994).
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Antimony - 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RtprMwtativ*
3p»ct«*
dear mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red (ox
raccoon
white-tailed
deer
red- tailed
hawk
American
kestrel
Northern
bobwhite
American
robin
American
woodcock
plant '
soil community
Benchmark
VaJu*ai#k«M
0.31 (a)
0.32 (a)
0.26 (a)
0.11 (a)
0.08 (a)
0.08 (a)
0.04 (a)
ID
ID
ID
ID
ID
5 mg/kg (i)
ID
SUMly
Sp«3««
rat
rat
rat
rat
rat
rat
rat
'
-
-
terrestrial
plants
Sitet
dev
dev
dev
dev
dev
dev
dev
-
unspecified
Study
V«a* ,
W0/KJHI
0.162
0.162
0.162
0.162
0.162
0.162
0.162
s.o
Dwcriptlott
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
-
'
'
-
LOEC
*?
-
-
-
-
-
Origin*! Souro* \
RoMietal., 1987
Rossi et al., 1987
ROM «t d., 1987
Rossi etal., 1987
Rossi etal., 1987
Rossi etal., 1987
Rossi et al., 1987
-
1
Kabata-Pwxtias and
Penotes, 1984 as
cited in Will and
Sutor. 1994
-
Benchmark Category, a - adequate, p = provisional, i = interim; ID= insufficient data; a (*) indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Antimony - 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystems, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following brief discussion describes the rationale for selecting the
biological uptake factors and provides the context for interpreting the biological uptake
values. .
The whole-body BCF for antimony was the measured value from Stephan (1993). BCF
values for muscle were not included because ecological receptors are likely to eat the whole
fish or, in the least, will not necessarily distinguish between the fillet and other parts of the
fish. The measured whole-body BCF for antimony in aquatic invertebrates was derived from
Stephan (1993). Insufficient data were identified to determine BCF values for terrestrial
vertebrates, terrestrial invertebrates and earthworms. A whole plant BCF value of 2.0 E-01
was derived from Baes et al. (1984). For metals, empirical data were used to derive the BCF
for above ground forage grasses and leafy vegetables. In particular, the uptake response slope
for forage grasses was used as the BCF for plants in the terrestrial ecosystem since most of
the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
cotoflleal
rvcwptor
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
ecF,sAi*+«r
BSAF
BCF
BCF
10
10
10
BCF
Kptd-battd of
whoto-body
whole
whole
-
-
whole-plant
vain*
0
0
2.0 E-01
court*
Stephan, 1993
Stephan, 1993
'
Baes eta), 1984
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Antimony-8
References
AQUIRE (AQUatic toxicity information REtrieval Database), 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Baes, C.F., R.D. Sharp, A.L. Sjoreen, and R.W. Shor. 1984. Review and Analysis of
Parameters and Assessing Transport of Environmentally Released Radionuclides
Through Agriculture. Oak Ridge National Laboratory, Oak Ridge, TN.
Barrows, M.E., S.R. Petrocelli, K.J. Macek, and J. Carroll. Bioconcentration and elimination
of selected water pollutants by bluegill sunfish (Lepomis macrochirus). Toxic Chemicals
379-392. As cited in U.S. EPA (Environmental Protection Agency). 1993b. Soil
Screening Fact Sheet (Second Draft August 12, 1993) Interim Guidance. Office of
Emergency and Remedial Response, Washington, DC. August.
Dieter, M.P., C.W. Jameson, M.R Elwell, J.W Lodge, M. Hejmancik, S.L Grumbein,
M. Ryan, A.C. Peters. 1991. Comparative toxicity and tissue distribution of antimony
potassium tartrate in rats and mice dosed by drinking water or intraperitoneal injection.
J. of Toxicology and Environmental Health, 34:51-82.
Fleming, A.J. 1982. The toxicity of antimony trioxide. Sponsored by E.I. Du Pont de
Nemours and Co., Wilmington DE. OTS215027. As cited in Syracuse Research
Corporation. 1990. Draft: Toxicological Profile for Antimony and Compounds. Prepared
for the Agency for Toxic Substances and Disease Registry (ATSDR). U.S. Public Health
Service.
57FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139). Draft
Report: A Cross-species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg 3/4/day.
Heitmuller, P.T., T.A. Hollister, and P.R. Parrish. 1981. Acute Toxicity of 54 Industrial
Chemicals to Sheepshead Minnows (Cyprinodon variegatus). Bull. Environ. Contam.
Toxicol. 27(5):596-604. As cited in AQUIRE (AQUatic toxicity Information REtrieval
Database), 1995. Environmental Research Laboratory, Office of Research and
Development, U.S. Environmental Protection Agency, Duluth, MN.
IRIS (Integrated Risk Information System), 1994. U.S. EPA, Office of Research and
Development, Office of Health and Environmental Assessment.
August 1995
-------
APPENDIX B , Antimony-9
Kabata-Pendias, A. and H. Pendias. 1984. Trace elements in soils and plants. CRC Press, Inc.
Boca Raton, Florida. As cited in Will, M.E and G.W. Suter H 1994. Toxicological
Benchmarks for Screening of Potential Contaminants of Concern for Effects on Terrestrial
Plants: 1994 Revision. DE-AC05-84OR21400. Office of Environmental Restoration and
Waste Management, U.S. Department of Energy, Washington, DC.
Luckey, T.D. and B. Venugopal. 1979. Metal toxicity in mammals (1): Physiologic and
chemical basis for metal toxicity. Plenum Press, N.Y.
M*rmo, E., M.H. Matera, R. Acampora, C. Vacca, D.De Santis, S. Maione, V. Susanna, S.
Chieppa, V. Guarino, R. Servodio, B. Cuparencu, and F. Rossi. 1987. Prenatal and
Postnatal metal exposure: Effect on vasomotor reactivity development of pups. Curr.
Ther. Res. 42:823-838.
Nagy, K.A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol. Mono. 57:111-128.
D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Ridgway, L.P. and D.A Kamofsky. 1952. The effects of metals on the chick embryo:
toxicity and production of abnormalities in development Ann. N.Y. Acad.
Rossi, F, R. Acampora, C. Vacca, S. Maione, M.G. Matera, R. Servodio, and E. Marmo.
1987. Prenatal and postnatal antimony exposure in rats: Effect on vasomotor reactivity
development of pups. Teratogenesis, Carcinogenesis, and Mutagenesis 7:491-496.
Schroeder, H.A., M. Mitchener, J.J. Balassa, M. Kanisawa and A.P. Nason. 1968. Zirconium,
niobium, antimony and fluorine in mice: Effects on growth, survival and tissue levels-/.
Nutrition, 95:95-101.
Schroeder, H.A., M. Mitchener, A.P. Nason. 1970. Zirconium, niobium, antimony,
vanadium, and lead in rats: Life term studies. J. Nutrition, 100:59-68.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. Office of Research andDevelopment, U.S.
Environmental Research Laboratory. PB93- 154672. Springfield, VA.
Sunagawa, S. 1981. Experimental studies on antimony poisoning. Igaku kenkyu 51:129-142.
As cited in Syracuse Research Corporation. 1990. Draft: Toxicological Profile for
Antimony and Compounds. Prepared for the Agency for Toxic Substances and Disease
Registry (ATSDR). U.S. Public Health Service,
August 1995
-------
APPENDIX B Antimony-10
Suter n, G.W. and J.B Mabrcy. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Syracuse Research Corporation. 1990. Draft: Toxicological Profile for Antimony and
Compounds. Prepared for the Agency for Toxic Substances and Disease Registry
(ATSDR). U.S. Public Health Service.
U.S. Environmental Protection Agency. 1980. Ambient Water Quality Criteria for Antimony.
Criteria and Standards Division.Washington, D.C.
U.S.Environmental Protection Agency (EPA, 1989). Ambient Water Quality Criteria
Document: Addendum for Antimony (Draft Report (Final)). Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1992. TSC1292 Criteria Chart. Region IV.
Water Management Division, 304 (a) Criteria and Related Information for Toxic
Pollutants. December.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993b. Soil Screening Fact Sheet (Second
Draft August 12, 1993) Interim Guidance. Office of Emergency and Remedial
Response, Washington, DC. August.
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
June 1992.
Venugopal, B. and T.D. Luckey. 1978. Metal toxicity in mammals (2): Chemical toxicity of
metals and metalloids. Plenum Press, N.Y.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision.
DE-AC05-84OR21400. Office of Environmental Restoration and Waste Management,
U.S. Department of Energy, Washington, DC.
August 1995
-------
Terrestrial Toxicity - Antimony
Cas No. 7440-36-0
Chemical
Name
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Antimony
Species
rat
rat
rat
rat
rat
dog
*
dog
dog
mouse
rat
rat
Endpoinl
longevity
cardio
cardio
lepatic
dev
gastro
neuro
gastro
chronic
chronic
neuro
Description
PEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
PEL
PEL .
LOAEL
Value
0.35
0.0748
0.748
418
0.0748
84
6,644
84
1.14
0.8
0.162
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-d
Exposure
Route (oral,
s.c., i.v., i.p.,
injection)
oral
oral
oral
oral
oral
gavage
gayage _
gavage
oral
oral
oral
Exposure Duration
/Timing
> 100 days
30 days
30 days
24 weeks
21 days + gestation
days 0-21
32 days
32 days
32 days
Weaning until natural
death^
Weaning until natural
death.
60 d
Reference
Schroeder et al., 1970 as
cited in IRIS, 1992
Marmo et al., 1987 as cited
n ATSDR, 1992
Marmo et al., 1987 as cited
n ATSDR, 1 992
Sunagawa, 1981 as cited in
ATSDR, 1992
Rossi et al., 1987 as cited
in ATSDR. 1992
Fleming, 1982 as cited in
ATSDR, 1992
Fleming, 1982 as cited in
ATSDR, J992
Fleming, 1 982 as cited in
ATSDR. 1992
Schroeder et at ,1968
Schroeder et al, 1969
Rossi etal, 1987
Comments
Decreased longevity and
}lood glucose; altered
cholesterol levels.
t
Decreased hypertensive
response in newborns. No
clear dose response
Decreased RBC count and
cloudy swelling in hepatic
cords.
Decreased maternal weight
gain.
Severe diarrhea.
Muscle weakness.
Severe diarrhea.
Decreased median
litespans of females and
growth suppression in
animals at 18 months of
age.
Single dose given;
decreased longevity and
lifespan; abnormal serum
cholesterol.
decreased .hypotensi ve
resonse. doses were 0,
0.162, and 1 .62 mg/kg-d
-------
APPENDIX B Arsenic - 1
lexicological Profile for Selected Ecological Receptors
Arsenic
] . ' Cas No.: 744-03-82
Summary: Arsenic exists as a trivalent species (arsenic III) and as a pentavalent species (arsenic
V). The speciation of arsenic is dependent on numerous environmental factors, such as pH,
Eh.and temperature (Eisler, 1988). Under reducing conditions, arsenic (V) is reduced to the
arsenic (HI) form and methylated. Although trivalent arsenic has been shown to be more toxic
to mammals, the pentavalent species is the dominant species of arsenic in aerobic soils and
aquatic environments (Eisler, 1988). As 80 % of arsenic released to the environment is released
to the soil (EPA 1982c as cited in ATSDR, 1993), this profile will focus primarily on the more
dominant arsenic (V) species with reference to arsenic (in) where necessary. This profile
summarizes the lexicological benchmarks and biological uptake measures (i.e., bioconcentration,
bioaccumulation, and biomagnification factors) for birds, mammals, daphnids and fish, aquatic
plants and benthic organisms representing the generic freshwater ecpsystem and birds, mammals,
plants, and soil invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for
birds and mammals were derived for developmental, reproductive or other effects reasonably
assumed to impact population sustainability. Benchmarks for daphnids, benthic organisms, and
fish were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFsj are also summarized for the ecological receptors, although some
BAFs for the freshwater ecosystem were calculated for organic constituents with log Kow between
4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire lexicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most current
information and may differ from the data presented in the support document for the Hazardous
Waste Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contain benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contain
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks for Arsenic
Mammals: Only two studies were identified which investigated the effects of chronic oral
exposure to arsenic (V) in. mammals. In a two-year study, rats were fed arsenic as sodium
arsenate at doses ranging from 31.25 to 400 ppm (Byron et al., 1967). Rats in the group
receiving 62.5 ppm did not differ from the controls, however rats fed 125 ppm exhibited
increased weight loss. Based on these results, a NOAEL of 62.5 ppm and a LOAEL of 125 ppm
August 1995
-------
APPENDIX B Arsenic-2
were inferred for growth effects. Since no information was provided on daily food consumption
or body weight, conversion from ppm (mg/kg-diet) to mg/kg-day required the use of an
allometric equation:
Food Consumption = 0.056(W°-6611) where W is body weight in kg (Nagy, 1987).
Using the geomean of the reported body weight of the male rats, 0.489 kg, the NOAEL of 62.5
ppm was converted to 4.73 mg/kg-day and the LOAEL of 125 ppm was calculated to be
equivalent to 9.46 mg/kg-day. In the same study (Byron et al., 1967), dogs were also fed arsenic
as sodium arsenate for two years at doses of 5, 25, 50 and 125 ppm. Dogs fed doses of 50 ppm
or less showed no signs of clinical or pathological toxicity, however, reduced survival and
increased weight loss were observed in those given 125 ppm. These results suggest a NOAEL
of 50 ppm and a LOAEL of 125 ppm for pathological effects. Using the same procedure as
above and an average body weight of 9 kg, the NOAEL of 50 ppm was converted to 1.3 mg/kg-
day and the LOAEL of 125 ppm was calculated. to be 3.3 mg/kg-day.
Although both studies provide evidence for the toxicity of chronic exposure to arsenic (V), the
rat study focused on growth during a critical life stage, an endpoint likely to more directly impact
the fecundity of a population than pathological effects. Therefore, the study NOAEL of 4.73
mg/kg-day was chosen for calculation of the mammalian benchmark value. This value was
scaled for species that were representative of a freshwater ecosystem using a cross-species scaling
algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL, x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BWt is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Byron et al.
(1967) study documented growth effects in both male and female rats, the mean body weight of
both genders of representative species were used in the scaling algorithm to obtain lexicological
benchmarks. Data were available on the reproductive and developmental effects of arsenic (V)
as well as growth or chronic survival. In addition, the data set contained studies which were
conducted over chronic and subchronic durations and during sensitive life stages. Based on the
data set for arsenic (V), benchmarks developed from the Byron et al. (1967) study were
categorized as adequate, with a "*" to indicate that some adverse effects have been observed at
the benchmark level. It should also be .noted that arsenic (HI) has been observed as being more
toxic to mammalian species (Eisler, 1988). Toxicological benchmarks were based on studies
focusing on arsenic (V) since it is likely to be the most prevalent species in aquatic
environments.
Birds: Two studies were identified which investigated arsenic (V) toxicity in avian wildlife. In
a two-part study, Stanley et al. (1994) examined arsenic's effect on the reproduction and
August 1995
-------
APPENDIX B Arsenic-3
development of mallard ducks by feeding adult mallards 25, 100 and 400 ug As/g for 4 weeks
prior to mating. While no signs of toxicity were observed in the two lower dose groups, ducks
treated with 400 ug/g exhibited delayed egg laying and lowered duckJing production. In addition,
the eggs of the 400 ug/g group weighed less than the eggs of the control group and showed signs
of eggshell thinning. Based on these results, a NOAEL of 100 ug/g and a LOAEL of 400 ug/g
can be inferred for reproductive effects. Since no information on body weight or food intake
was provided, converting the dietary doses from ug/g-diet to mg/kg-day required the use of the
allometric equation:
i
Food consumption = 0.301(W°'751) where W is weight in kg (Nagy, 1987)
Assuming an average weight of 1.162 kg (EPA, 1988), the NOAEL of 100 ug/g was calculated
to be 5.51 mg/kg-day and the LOAEL of 400 ug/g was calculated as 22 mg/kg-day. The
ducklings which hatched from the eggs of the treated parents were also fed 25, 100 and 400 ug
As/g food for 14 days after hatching. Although no effects were seen at dose levels of 25 and
100, those in the 400 ug/g dose group had decreased growth rates and body and liver weights
suggesting a NOAEL of 100 ug/g and a LOAEL of 400 ug/g for developmental effects. Neither
body weight nor food consumption data were provided for conversion from ug/g-diet to mg/kg-
day. Therefore, assuming an average body weight of 0.24 kg (Lokemoen et al., 1990) and using
the allometric equation from above, a NOAEL of 8.3 mg/kg-day and a LOAEL of 33.3 mg/kg-
day were calculated for developmental effects in the ducklings. In another study, mallard
ducklings were given arsenic in doses of 30, 100 or 300 ppm beginning the day after hatching
until 10 weeks of age (Camardese et al., 1990). Although reduced growth was seen in female
ducklings given 30 ppm, only male ducklings in the 300 ppm exhibited decreases in growth
compared to controls, suggesting a LOAEL of 30 ppm for pathological effects. Using the
allometric equation presented above and a body weight of 0.78 kg (Lokemoen et al., 1990), the
LOAEL of 30 ppm was calculated to be 9.6 mg/kg-day.
The Camardese et al. (1990) study was not considered suitable for the derivation of a benchmark
value since pathological effects do not clearly indicate that the fecundity of a wildlife population
could be impaired. The NOAEL value of 5.51 mg/kg-day inferred from the Stanley et al. (1994)
study on adult mallard ducks was selected over the NOAEL of 8.3 mg/kg-day derived from the
Stanley et al. .(1994) duckling study since it was more conservative. The NOAEL of 5.51 mg/kg-
day was then scaled using the cross-species scaling algorithm adapted from Opresko et al. (1994).
Although the procedure in the Stanley et al. (1994) study dictated the exposure of both male and
female adult mallards, the reproductive effects were primarily documented in female mallards.
Therefore, female body weights for each representative species were used in the scaling algorithm
to obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of arsenic (V)/as well as
chronic survival. In addition the data set contained studies conducted over chronic and
subchronic durations. Based on the avian data set for arsenic (V), the benchmarks developed from
Stanley et al. (1994) were categorized as adequate with a "*" to indicate that some adverse
effects have been observed at the benchmark level-
August 1995
-------
APPENDIX B Arsenic - 4
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 1.9E-01 reported in the
AWQC document for arsenic (III) was used since it is a nationally accepted standard and none
\vas available for arsenic (V). A Secondary Chronic Value (SCV) of 8.11E-3 mg/1 was reported
by Suter and Mabrey, (1994) for arsenic (V). Since the benchmark value selected is based on
an FCV for Arsenic (in) developed for a AWQC, it was categorized as adequate.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or 2) an effective concentration (ECXX) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum capricornutwri). Suter and Mabrey (1994)
reported a benchmark of 4.8 E-02 based on EC50 tests conducted on Scenedesmus obliquus. As
described in Section 4.3.6, all benchmarks for aquatic plants were designated as interim. Arsenic
(ID) has been observed as being less toxic to aquatic plants than the more prevalent, pentavalent
species.
Benthic community: The arsenic (V) benchmark protective of benthic organisms is pending a
U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Arsenic 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with a Freshwater Ecosystem - Arsenic (V)
R«pr*s«at*Uv*
SpMit*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
B*nchm«rK
V*lue* «o*ff-
d,y
3.94 (a*)
2.35 (a')
3.93 (a*)
4.96 (a*)
4.70 (a')
5.51 (a')
6. 15 (a')
12.29(a')
5,76 (a')
9.24 (a*)
&U*f
Sptofe*
rat
rat
duck
duck
duck
duck
duck
duck
duck
duck
Eltoot
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study Y*Iu»
rag/Jcg-day
4.73
4.73
5.51
5.51
5.51
5.51
5.51 .
5.51
5.51
5.51
Otwfptfon
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
8?
' -
-
-
-
WjNtfSoww*
Byron et al., 1967
Byron etal., 1967
Stanley etal., 1994
Stanley etal., 1994
Stanley etal., 1994
Stanley et al.,1994
Stanley etal., 1994
Stanley etal., 1994
Stanley etal., 1994
Stanley etal., 1994
'Benchmark Category, a » adequate, p » provisional, i = interim, ID = insufficient data; a (*) indicates (hat the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with a Freshwater Ecosystem - Arsenic (V)
RapojjwitatiVB (
Sped**
Fish and aquatic
invertebrates
aquatic plants
benthic community
BwiehrafcrK
Value*
mflA,
1.9E-01 (a)
4.8 E-02
under review
stooyspoctes
aquatic
organisms
Sconedesmus
obliquus
Otigimi
Value
mgft.
1.9E-01
4.8 E-02
-
DescripSon.
FCV
ECM
OriojndSowo*
AWQC Table
Suter & Mabrey,
1994
'
Benchmark Category, a - adequate, p = provisional, i = interim, ID = insufficient data; a (') indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Arsenic-6
II. lexicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective media
concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains benchmarks for mammals,
birds, plants and soil invertebrates representing the generic terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks for Arsenic (V)
Mammals: As discussed in the rationale for the freshwater ecosystem, there were two possible studies
from which to estimate a benchmark value. Since no additional studies were identified, the Byron et al.
(1967) study used to calculate a freshwater mammalian benchmark was also used for the terrestrial
ecosystem. The NOAEL from the Byron et al. (1967) study was scaled for species representative of the
terrestrial ecosystem using the cross-species scaling algorithm to obtain the lexicological benchmarks.
Based oh the data set for arsenic, the benchmarks developed from the Byron et al. (1967) study were
categorized as adequate with a "*" to indicate that some adverse effects have been observed at the
benchmark level. It should also be noted that arsenic (III) has been observed as being more toxic to
mammalian species (Eisler, 1988). Toxicological benchmarks were based on studies focusing on arsenic
(V) since it is likely to be the most prevalent species in the aerobic terrestrial ecosystem.
Birds: Additional avian toxicity data were not identified for birds representing the terrestrial ecosystem
therefore, the Stanley et al. (1994) study used in the freshwater ecosystem discussion above, was also used
to calculate terrestrial, avian benchmark values. The Stanley et al. (1994) focused on the reproductive
effects of arsenic (V) in adult mallard ducks. The NOAEL from the Stanley et al. (1994) study was scaled
for avian species representative of the terrestrial ecosystem using the cross-species scaling algorithm to
obtain the lexicological benchmarks. Based on the data set for arsenic, the benchmarks developed from
the Stanley et al. (1966) study were categorized as adequate with a "*" to indicate that some adverse
effects have been observed at the benchmark level.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root length. As presented in Will and Suter (1994), -phytotoxicity benchmarks were selected by
rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or fewer
values, the 10th percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield
reductions, or other effects reasonably assumed to impair the ability of a plant population to sustain itself,
such as a reduction in seed elongation. The terrestrial benchmark of 10 mg/kg was based on studies
focusing on the effects of arsenic (HI) and (V) (Will and Suter, 1994). The terrestrial plant benchmark
of 10 mg/kg is categorized as interim, since the value is based on less than 10 values.
Soil Community: Adequate data with which to derive a benchmark protective of the soil community were
not available.
August 1995
-------
APPENDIX B
Arsenic - 7
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem Arsenic (V)
R«pT»*«ntattv»
Specie*
dear mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
Benchmark ][ Study
V«iu«« ] 3p*ti«*
ntg/kgHJay ||
10.65 (a')
10.95 (a')
9.28 (a*)
3.76 (a*)
2.71 (a4)
2.57 (a')
1.30 (a*)
5.46 (a*)
, 9.58 (a*)
8.92 (a')
10.69 (a*)
8.50 (a*)
10.0 mg/kg
(').
ID
rat
rat
rat
rat
rat
rat
rat
mallard
mallard
mallard
mallard
mallard
terrestrial
plants
-
Eftot :
growth
growth
growth
growth
growth
growth
growth
rep
rep
rap
rep
rep
growth/
yield
-
Study
Value
' mgfcg*
d.y
4.73
4.73
4.73
4:73
4.73
4.73
4.73
5.64
5.64
5.64
5.64
5.64
10 mg/kg
Description
»
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOEC
-
SF
-
-
-
-
-
-
-
-
-
-
-.
, QdefcutSmire*
^ ;
" *
Byron et «!., 1967
Byron et al., 1967
Byron et al., 1967
Byron etal. 1967
Byron etal.. 1967
Byron et al., 1967
Byron et al.. 1967
Stanley et al., 1994
Stanley el al., 1994
Stanley et at., 1994
Stanley et al., 1994
Stanley et at., 1994
Will&Suter, 1994
'Benchmark Category, a adequate, p « provisional, i = interim; ID = insufficient.data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Arsenic - 8
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for ecological receptor categories: fish in the limnetic or littoral
ecosystem, aquatic invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and
plants. For metals, BCFs are whole-body bioconcentration factors and refer to total surface water
concentrations (versus freely dissolved concentrations). Consequently, all calculations of
acceptable tissue concentrations (TC) represent whole-body concentrations. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values.
The whole-body BCF for arsenic in fish is derived from the geometric mean of two measured
values, 3 and 4 (Stephan, 1993). BCF values for muscle were not included because ecological
receptors are likely to eat the whole fish or, in the least, will not necessarily distinguish between
the fillet and other parts of the fish. Insufficient data were identified to determine BCF values
for aquatic invertebrates, terrestrial vertebrates and earthworms. A whole plant BCF value of 6.0
E-02 was derived from U.S. EPA. (1992e). For metals, empirical data were used to derive the
BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake response
slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem since most
of the representative plant-eating species feed on wild grasses.
August 1995
-------
APPENDIX B
Arsenic 9
Table 4. Biological Uptake Properties
ootogiea!
receptor
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF,«r
BSAF
BCF
BCF
BCF
iipki-I»M«i of
whole-body
whole
whole
-
-
. .
whole-plant
vaiu*
3.5
3.5
10
10
10
6.0 E-02
pure*
Stephan. 1993
Stephan, 1993
.
U.S.EPA, 1992e
d = refers to dissolved surface water concentration'
t = refers to total surface water concentration
10 » refers to insufficient data
August 1995
-------
APPENDIX B Arsenic - 10
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APPENDIX B Arsenic - 11
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APPENDIX B Arsenic - 12
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APPENDIX B Arsenic-13
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Luckey, T.D. and B. Venugopal. 1979. Metal toxicity in mammals (1): Physiologic and
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August 1995
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APPENDIX B Arsenic - 14
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Opresko, D.M;, B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
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Passino, D. R. M., and A. J. Novak. 1984. Toxicity of arsenate and DDT to the cladoceran
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August 1995
-------
APPENDIX B Arsenic. 15
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/
Stephan, C.E. '1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
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August 1995
-------
APPENDIX B Arsenic - 16
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arsenic, estimated by six methods and response of com (Zea mays L.) Soil Sci. Soc. Am.
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World Health Organization, 1981. Environmental Health Criteria- Arsenic Environmental
Aspects, Geneva, 1989.
August 1995
-------
Terrestrial Toxicity - Arsenic
Cas No. 7440-38-2
Chemical
Name
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic
arsenic (III)
arsenic (III)
Species
mouse
mouse
mouse
dog
cat
mouse
-
mouse
mouse
rat
rat
rat
Endpoint
hislo
let
let
no effect
chronic
terat
!?P .
feyerat
lerat
growth
growth
Description
LOAEL
NOAEL
''
LOAEL
NOEL
AEL
AEL
AEL
LOAEL
NOAEL
NOAEL
LOAEL
Value
0.5
20
40
30
1.5 __
400
5
10
17.5
62.5
125
Units
mg/kg
mg/kg-day
mg/kg-day
n^g .
mg/kg-body
weight
mg/kg-body
weight
mg/kg-diet
mg/kg
mg/kg-diet
ppm
ppm
Exposure
Route (oral,
s.c., i.v.,
i.p..
injection)
'P-
oral
(gavage)
oral
(gavage)
oral
oral
oral
oral
'P-
oral
oral
oral
Exposure
Duration
/Timing
made 30 hr
after
treatment
gestation
days 8^1 5
Gestation
days 10 or
12
90 days
/
NS
Days 7 to 16
of gestation
3
generations
One of days
7 to 12 of
gestation.
7
generations
2 years
2 years -
Reference '
Deknudt et al., 1986
-
Baxley et al., 19B1
Baxleyetal., 1981
Hood. 1985 as cited in
Eisler, 1988
Pershagen and
Vahter. 1979 as cited
in Eisler, 1988
Hood, 1985 as cited in
Eisler. 1988
Pershagen and
Vahter, 1979 as cited
in Eisler, 1988
Hood. 1972
Frost etal., 1964 as
cited in WHO, 1981
Byron et al., 1967
Byron el al., 1967
Comments
Increase of micronucleated
erythrocytes.
Treatment was given on
gestation days 8- 15,
however, significant
increases in fetal mortality
were seen only in the
groups treated on gestation
days 10 or 12.
Arsenic III as cacodylic
acid and methanearsonic
acid; no ill effects.
Produced cleft palate,
delayed skeletal
ossification and fetal weight
reduction.
Reduced liner size.
Increased resorptions,
malformations and
decreased fetal weights.
Arsenic as arsinilic acid.
Doses were 250, 125, 62.5,
31.25, 15.63 ppm
Enlargement of the
common bile duct and
increased weight loss.
-------
Terrestrial Toxicity - Arsenic
Cas No. 7440-38-2
Chemical
Name
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V) '
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
Species
rat
rat
hamster
cat
hamster
hamster
hamster
mouse
monkey
chicken
Endpoint
mortality,
growth
mortality,
growth
terat
chronic
let
let
let
terat, rep
mortality
embryonic
Description
NOAEL
LOAEL
AEL
AEL
NOAEL
LOAEL
LOAEL
AEL
LOAEL
NOAEL
Value
4.73
-
9.46
20
1.5
2
8
15
~
25
2.8
6,3-3
Units
mg/kg-day
mg/kg-day
mg/kg
mg/kg-body
weight
mg/kg-body
weight
mg/kg-body
weight
mg/kg-body
weight
mg/kg-body
weight
mg/kg-day
ug/embryo
Exposure
Route (oral,
s.c., i.v.,
i.p,
injection)
oral
oral
i.v.
oral
i.v.
i.v.
i.v.
i.p.
oral
NS
Exposure
Duration
./Timing
2 years
2 years
Day 8 of
gestation
NS
Day 8 of
gestation
Day 8 of
gestation
Day 8 of
gestation.'
One of days
6 to 12 of
gestation
1 year
NS
Reference
Byron et_aj., 1967
Byron e\a\.. 1967
Perm and Carpenter,
1968
Pershagen and
Vahter, 1979 as cited
in Eisler. 1988
Pershagen and
Vahter, 1979 as cited
in Eisler, 1988
Pershagen and
Vahter, 1979 as cited
in Eisler. 1988
Fermetal., 1971 '
Hood and Bishop,
1972
Heywood and Sortwell,
1979 as cited in
ATSDR. 1993
'
NRCC, 1978 as cited
in Eisler, 1988
Comments
Doses were 400, 250. 125,
62.5, and 3 1.25 ppm
Reduced survival,
enlargement of the
common bile duct and
increased weight loss.
High incidence of
exencephaly.
Increased incidence of
malformation and
resorption.
Increased malformation
and resorption rates.
Increased fetal resorptions
Decreased fetal weights
and an increase in fetal
malformations seen in mice
treated on gestation days 6
Plli ._.
Threshold for embryo
malformations at specified
dose range. Dosing route
and duration was not
specified in Eisler.
-------
Terrestrial 7. ,ity - Arsenic
Cas No. 7440-38-2
Chemical
Name
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
-
arsenic (III)
arsenic (V)
Species
mouse
rat
mouse
mallard
mallard
Calitornia
quail
Ring-necked
pheasant
rat
Endpoint
rep
see
comments
mortality
acute
acute
acute
acute
kidney
Description
AEL
AEL
AEL
LC50
LC50
LC50
LC50
AEL
Value
5
0.38
0.4
323
500
47.6
3B6
1200
Units
ppm
mg/kg-day
mg/kg-day
mg/kg
mg/kg
mg/kg
mg/kg
ug/kg-day
Exposure
Route (oral,
s:c.. i.v.,
i.p.,
injection)
oral
oral
oral
oral
oral
oral
oral .
oral
Exposure
Duration
/Timing
3
generations
Weaning
until natural
death.
Weaning
until natural
death.
NS
32 d
NS
NS
6 weeks
Reference
Schroeder and
Mitchener, 1971
Schroeder et al., 1968
Schroeder and
Balassa, 1967
NAS, 1977 as cited in
Eisler, 1988;NRCC.
1978 as cited in Eisler,
1988
NAS, 1977 as cited in
Eisler. 1988
Hudson et al., 1984 as
cited in Eisler, 1988
Hudson etal., 1984 as
cited in Eisler, 1988
Jauge and Del-Razo,
1985
Comments
Increase in the ratio, of
males to females and a
reduction in litter size.
No effects on growth rates,
longevity or survival;
increased serum
cholesterol levels,
increased incidence of
abnormal liver cells and
significantly different
fasting serum glucose
levels.
Increased mortality, and
decreased life-span and
longevity.
-..
Reduction in renal
excretion of uric acid was
significant at 6 weeks of
treatment for As V,
however, for As III renal
excretion of uric acid was
significantly reduced in the
first 3 weeks.
-------
Freshwater Toxicity - Arsenic
Cos No. 7440-38-2
Chemical
Name
arsenic (V)
arsenic (V)
arsenic
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (III)
arsenic (V)
Species
cladoceran
daphnid
rainbow trout
rainbow trout
rainbow trout
daphnid
rainbow trout
brook trout
goldfish
goldfish
daphnid
Type of
Effect
acute
acute
subchronic
chronic
subchronic
chronic
acute
acute
acute
acute
acute
Description
EC50
EC50
NOEL
NOEL
PEL
LOEC
LC50
LC50
LC50
LC50
LC50
Value
0.85+/-0.12
49.6 +/ 9.0
1600
'°
120
520
540 -
10,440
18,618
490
7,400
Units
mg/L
mg/L
PP.m
PPm
PPm
"9/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
static
static
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
96 hours
48 hours .
8 weeks
16 weeks
8 weeks
3 weeks _
28 days
262 hours
336 hours
7 days
48 hours
Reference
Passino and Novak,
1984
Passino and Novak,
1984
Cockell and Hilton,
1985
Cockell and Hilton,
1985^
Cockell and Hilton,
1985
Biesinger and
Christensen, 1972
Birge, 1979 as cited
in U. S. EPA. 1980
Cardwell et al.. 1976
as cited in U. S.
EPA. 1980
Cardwell et al.. 1976
as cited in U. S.
EPA. 1980
Birge, 1979 as cited
in U.S. EPA, 1980
Biesinger and
Christensen, 1972
Comments
Arsenic as
dimethylarsinic acid and
p-aminobenzenearspnic
acid.
Arsenic as disodium
arsenate.
Arsenic as disodium
arsenate and arsenic
trioxide; toxicity
responses included feed
refusal, growth
depression and impaired
feed efficiency..
16% reproductive
impairment.
Embryo-larval stage.
Juvenile stage.
Embryo-larval stage.
Without food.
-------
Terrestrial TV ,,ty - Arsenic
Gas No. 7440-38-2
Chemical
Name
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (III),
(V)
arsenic (III),
(V)
Species
mallard duck
mallard duck
mallard
duckling
mallard
duckling
mallard
duckling
d99
dog
Endpoint
rep
rep
dev
dev
dev
path
path
Description
LOAEL
NOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
Value
22
5.51
8.3
33.3
9.6
1.3
3.3
Units
mg/kg-day
mg/kg-day
mg/kg-day
m9/k9:.(?ay
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., i.V.,
i.p..
injection)
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
4 weeks
prior to
mating
4 weeks
prior to
mating
14 days after
hatching .
14 days after
hatching
1 0 weeks
2 years
2 years
Reference
Stanley el al., 1994
Stanley el al, 1994
Stanley el al., 1994
Stanley et al., 1994
Camardese et al.,
1990
Byron et^al., 1967
Byron et al., 1967
Comments
belayed egg laying,
reduced egg weight ,
caused eggshell thinning
and lowered duckling
production.
Decreased duckling growth
and body and liver weights.
Reduced growth in female
ducklings. Male ducklings
only showed reduced
growth after treatment with
300 ppm As.
-------
Freshwater Biological Uptake Measures - Arsenic
Cos No. 7440-38-2
Chemical
Name
arsenic (III)
arsenic (III)
arsenic
arsenic
arsenic
arsenic
arsenic
arsenic
arsenic (III)
arsenic (III)
arsenic (V)
arsenic
arsenic
arsenic
Species
daphnid
daphnid
daphnid
daphnid
daphnid
mosquito fish
mosquito fish
mosquito fish
rainbow trout
bluegill
rainbow trout
fathead
minnow
bluegill
fish
B-factor
(8CF, BAF,
BMP)
BCF
BCF
BAF
BAF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
BCF
BCF
BCF
Value
50
219
1658+/-463
2175+/-290
736 W: 104
21 W- 6
19 W- 7
49 +/- 24
0
4
0
3
4
44
Measured
or
Predicted
(m,p)
m
m
m
m
m
m
m
m
NS
NS
NS
NS:
NS
m
Units
L/g .
L/g
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
L/Kfl
Reference
Spehar et al., 1980
Spehar etal., 1980
Isensee et aL, J973
Isensee et al., 1973
jsenseeetal.,^1973
Isensee etal., 1973_
Isensee etal, 1973
Isensee et al., J973
Spehar et al., 1980
U.S. EPA, 1978 as cited in
U. S. EPA, 1980
Spehar et §1.^1980
Defoe et al.. 1982 as cited
inJJ. S. EPA. 1993 _
Barrows et al.. 1980
U. S7EPA. 1992
Comments
Exposure period was 21 days to 970 ug
As/L.
Exposure period was 21 days to 96 ug
As/L.
Arsenic as cacodylic acid; 2-day exposure
period.
Arsenic as dimethylarsine (oxygen); 2-day
exposure period.
Arsenic as dimethylarsine (nitrogen); 2-
day exposure period.
Arsenic as cacodylic acid; 2-day exposure
period.
Arsenic as dimethylarsine (oxygen); 2-day
exposure period.
Arsenic as dimethylarsine (nitrogen); 2-
day exposure period.
Exposure period was 28 days; whole body
BCF.
Exposure period was 28 days; whole body
BCF.
Exposure period was 28 days; whole body
BCF.
Whole body BCF.
Whole body BCF.
Normalized to 3% lipid.
-------
Freshwater K Jty - Arsenic
Cas No. 7440-38-2
Chemical
Name
arsenic (V)
arsenic
arsenic
arsenic (Ml)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (V)
arsenic (III)
arsenic (III)
Species
daphnid
daphnid
fathead
minnow
aquatic
organisms
aquatic
organisms
fish
daphnid
fish
daphnid
fish
daphnid
Type of
Effect
acute
acute
acute
chronic
chronic
chronic
chronic
chronic
chronic
chronic
chronic
Description
LC50
LC50
LC50 .
NAWQC
NAWQC
cv
CV
EC20
EC20
CV
CV
Value
3800
1900
9900
190
0.77
891.6
450
1500
>932
2962
914.1
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS.
NS
Exposure
Duration
/Timing
48 hours
48 hours
96 hours
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Mount and Norberg,
1 984 as cited in
AQUIRE
Mount and Norberg,
1 984 as cited in
AQUIRE
Dyeretal., 1993 as
cited in AQUIRE,
1994
U.S. EPA, 1980
Suteretal., 1992
Suteretal.. 1992
Suteretal., 1992
Suteretal., 1992
Suteretal.. 1992
Suteretal., 1992
Suteretal , 1992
Comments
-------
Terrestrial Biological U, xe Measures - Arsenic
Cos no. 7440-38-2
Chemical
Name
arsenic
I
Species
plant
B-faclor
(BCF. BAF.
BMP)
BCF
Value
0.65
Measured
or
Predicted
(m.p)
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Barium - 1
Toxicological Profile for Selected Ecological Receptors
Barium
Cas No.: 7440-39-3
Summary: This profile on barium summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals and fish representing the generic freshwater and terrestrial ecosystems.
Toxicological benchmarks were derived for developmental, reproductive or other effects
reasonably assumed to impair population growth and survival. Bioconcentration factors
(BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs)
are also summarized for the ecological receptors, although BAFs for the freshwater ecosystem
were calculated for organic constituents with log Kow between 5 and 6.5. For the terrestrial
ecosystem, these biological uptake measures also include terrestrial invertebrates (i.e. insects
and earthworms). In addition, the entire toxicological data base compiled during this effort is
presented at the end of this profile and includes additional studies and existing regulatory
benchmarks (e.g., National Ambient Water Quality Criteria or NAWQC). The entire
toxicological data base compiled during this effort is presented at the end of this profile. This
profile represents the most current information and may differ from the technical support
document for the Hazardous Waste Identification Rule (HWIR): Risk Assessment for Human
and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammais and birds associated with the fresh water ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No studies were identified which investigated the effects of barium toxicity on
mammalian species. Therefore, benchmarks for mammalian species could not be derived.
Birds: Few toxicity studies were identified that focused on the effects of vanadium toxicity
in avian species. Rigway and Karnofsky (1952) injected eight-day old White Leghorn chick
embryos with a single 20 mg dose of barium chloride and observed an inhibition of toe
growth in 50% of the treated embryos surviving to 18 days of age. In another study, Johnson
et al. (1960) fed female chicks with 0, 250, 500, 1000, 2000, 4000 and 8000 ppm barium
hydroxide or barium acetate from the first day of age to 4 weeks. At dosages above 1000
ppm barium, a depression in growth was observed with higher dosages resulting in increased
mortality. As the weight and species of the female chicks were not included in the .study, the
August 1995
-------
APPENDIX B Barium - 2
reported male body weight at 7 weeks was compared to reference body weight values at 7
weeks (U.S. EPA, 1988) so as to determine which species of chicken was likely to have
utilized in the study. Through this method of deduction, New Hampshire chickens were
assumed to be similar in weight to the species tested, if not the actual species. Based on the
geomean of the reference body weight of female New Hampshire chicks (U.S.EPA, 1988) and
daily food consumption given by the equation:
Food Consumption = 0.075W0'8449 , where W is the body weight in kg (Nagy, 1987).
The NOAEL of 1000 ppm was converted to 102.5 mg/kg-day in ihis way.
The Rigway and Karnofsky (1952) study was not selected for the derivation of a lexicological
benchmark for birds because ihe dose was administered via injection and extrapolation from
the injection lo ihe oral rouie of exposure would increase the uncertainty associated with the
value. The Johnson et al (1960) study was selected, as it 1) contained .clear dose-response
data, 2) focused on growih during a critical life stage, and 3) chronic exposures were
administered via oral ingestion.
The study value from ihe Johnson el al. (1960) study was scaled for species
representative of a freshwater ecosystem using a cross-species scaling algoriihm adapted from
Opresko el al. (1994):
(bw.
Benchmark = NOAEL. x 1
IK,
where NOAEL, is ihe NOAEL (or LOAEL/10) for ihe tesi species, BWW is ihe body weighi
of ihe wildlife species, and BW, is ihe body weighi of ihe tesi species. This is ihe same
defauli meihodology EPA provided for carcinogeniciiy assessments and repoitable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Johnson et al (1960) study documented effecis from barium exposure lo female chicks, mean
female body weighi of ihe represeniative species were used in ihe scaling algoriihm to obtain
lexicological benchmarks. Based on the daia sei for barium, the benchmarks developed from
Johnson et al (1960) were categorized as adequate.
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
barium. Therefore, a Secondary Chronic Value (SCV) of 1.0 E+00 mg /I (AQUIRE) was
utilized. Because an SCV was utilized, ihe benchmark was categorized as interim.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effecis conceniration (LOEC) for vascular aquatic
planis (e.g., duckweed) or 2) an effective conceniration (ECXX) for a species of freshwater
algae, frequently a species of green algae, (e.g., Selenastrum capricornutwn). Data were not
identified in Suter and Mabrey (1994) or AQUIRE. As described in Section 4.3.6, all
benchmarks for aquatic plants were designated as interim.
August 1995
-------
APPENDIX B Barium - 3
Benthic community: The barium benchmark protective of benthic organisms is pending a
U.S.EPA review of acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Barium - 4
Table 1. Toxicoiogical Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative
Spocfot
mink
river otter
bald eagle
osprey
great blue heron
. mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
Value* »g/kg*
day
ID
ID
40.89 (a)
51.60 (a)
48.88 (a)
58.08 (a)
64.00 (a)
127.92 (a)
59.94 (a)
96. 19 (a)
Study
Specie*
Chicken
Chicken
Chicken
Chicken
Chicken
Chjcken
Chicken
Chicken
Effect
-
growth
growth
growth
growth
growth
growth
growth
growth
Study Value
mfl/kfl-day
1.03E+02
1 .03E+02
1 .03E+02
1.03E+02
1.03E+02
1.03E+02
1.03E+02
1.03E+02
Description
-
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
-
Ofigfn*l$ouro«
Johnson et al., 1960
Johnson et al., 1960
Johnson et al.. 1960
Johnson et al., 1960
Johnson et al., 1960
Johnson et al., 1960
Johnson et al., 1960
Johnson et al., 1960
'Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates (hat the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicoiogical Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fleprB«mt««v«
Spade*
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark
VWu*»
mg/L
1.0 E+00(i)
ID
under review
Study
Specie*
aquatic
organisms
-
Origin**
V»lw
mg/L
1 .0 E+00 .
-
-'
Description
SCV
-
Original Sourc*
AOUIRE
' -
"Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Barium - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable sub-
chronic or chronic studies were identified which focused on the reproductive or
developmental toxicity of barium on mammalian species.
Birds: As in the freshwater ecosystem, the Johnson et al. (I960) study was used to calculate
the benchmarks for birds in the generic terrestrial ecosystem. The NOAEL of 1.03E+02
mg/kg-day from the Johnson et al. (1960) study was scaled for the representative species
using the cross-species scaling algorithm adapted from Opresko et al. (1994). Since the
Johnson et al. (1960) study documented effects on female chicks, mean female body weights
for each of the representative species were used in the scaling algorithm to obtain the
lexicological benchmarks. Based on the data set for barium, the benchmarks developed from
Johnson et al. (1960) were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 values, the 10th percentile LOEC was used. Such LOECs
applied to reductions in plant growth, yield reductions, or other effects reasonably assumed to
impair the ability of a plant population to sustain itself, such as a reduction in seed
elongation. The benchmark for terrestrial plants was 500 mg/kg, based on a Lowest
Observable Effects Concentration (LOEC) of 500 mg/kg which resulted in a reduction in the
shoot weight of barley and the growth of bush beans (Phaseolus vulgaris L.) after 14 days
(Chaudhry et al, 1977 as cited in Will and Suter, 1994). Since less than 10 values were
presented by Will and Suter (1994), with the benchmark being the lowest LOEC value
identified, the terrestrial plant benchmark of 500 mg/kg-day was categorized as interim.
Soil community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Barium - 6
Table 3. Toxicologicai Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Representative
Specie*
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bob white
American robin
American
woodcock
plants
soil community
Benchmark
Value*
jn0/k0-
-------
APPENDIX B
Barium 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystems, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentratiori factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following brief discussion describes the rationale for selecting the
biological uptake factors and provides the context for interpreting the biological uptake
values.
Insufficient data were identified to determine BCF values for fish, littoral invertebrates,
terrestrial vertebrates, terrestrial invertebrates and earthworms. A whole plant BCF value of
1.5E-01 was derived from U.S. EPA (1992e). For metals, empirical data were used to derive
the BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake
response slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem
since most of the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
«cQtogic«i
nceptor
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF,BAF,«f
BSAF
-
.. -
-
BCF
irphttwMd of
whofe-body
'
whole-plant
valu*
ID
ID
ID
ID
ID
1.SE-01
ourc*
'
-
-
U.S. EPA, 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Barium-8
References
AQUIRE (AQUatic Toxicity Information REtrieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
, Duluth, MN. '
ASTER Ecotoxicity Profile. 1992. U.S. EPA (Environmental Protection Agency),
Environmental Research Laboratory-Duluth, MN.
Biesinger,K.E. and Glenn M. Christensen. 1972. Effects of various metals on survival,
growth, reprbduction, and metabolism of Daphnia magna. J. Fisheries Res. Bd. of
Canada, V29(12).
Chaudhry, P.M., A. Wallace and R.T. Mueller. 1977. Barium toxicity in plants. Commun Soil
Sci. Plant Anal. .8(9):795-97. As cited inWill, M.E and G.W. Suter II. 1994.
Toxicological Benchmarks for Screening of Potential Contaminants of Concern for Effects
on Terrestrial Plants: 1994 Revision. DE-AC05-84OR21400. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg 3/4/day.
ICAIR, Life systems, Inc. 1987. Final Draft for the Drinking Water Criteria Document on
Barium. Prepared for U.S. Environmental Protection Agency (EPA), Office of Drinking
Water, Criteria and Standards Division, Washington, DC.
Leblanc, G.A. 1980. Acute toxicity of priority pollutants to water flea (Daphnia magna). Bull
Environ. Contam. Toxicol. 10(5):291-294.
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
Nagy, K.A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:111-128.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Ridgway.L.P and D.A. Karnofsky. 1952. The effects of metals on the chick embryo: Toxicity
and production of abnormalities in development. Ann. N.Y. Acad. Sci. 55:203.
August 1995
-------
APPENDIX B Barium - 9
Suter n, G.W., M.A.'Futrell, and G.A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
. Ridge Reservation, Oak Ridge, Tennessee. DE93-OQ0719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Tarasenko, N. Y., O.A Pronin, A.A. Silayev. 1977. Barium Compounds as industrial Poisons
(an Experimental study). J. Of Hygiene, Epidemiology, Microbiology and Immunology
21(4)361-373.
Tardiff, R.G, M Robinson, N.S. Ulmer. 1980. Subchronic oral toxicity of BaCl2 in rats. /. of.
Environmental Path and Tox. 4:267-275.
U.S. EPA (Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993c. Integrated Risk Information System.
June, 1995.
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2): Chemical toxicity of metals.
and metalloids. Plenum Press, N.Y., 1978.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
World Health Organization (WHO). 1990. Environmental Health Criteria 107: Barium.
Published under the joint sponsorship of the United Nations Environment Programme, the
International Labour Organisation, and the World Health Organisation.
August 1995
-------
Terrestrial "i Jty - Barium
Cas No. 7440-39-3
Chemical
Name
Barium
hydroxide or
barium
acetate
Barium
hydroxide
and barium
acetate
Barium
Barium
carbonate
Barium
carbonate
Species
chick
chick
rat
rat
male rat
Type of
Effect
growth
growth
path
growth
rep
Description
LOAEL
NOAEL
NOAEL "
LOAEL
LOAEL
Value
210
102.5
250
5.2
5.2
Units
mg/kg-day
mg/kg-day
PPm
mg/m3
mg/m3
Exposure
Route (oral,
s.c.. i.v., i.p.,
injection)
oral
oral
oral
inhalation
inhalation
Exposure Duration
/Timing
4 weeks
4 weeks
4, 8 and 13 weeks
4 month, 6x a week for
4 hrs/day
4 months-
. ' Reference
Johnson etal., 1960
Johnson etal., 1960
Tardjff el :al., 1980
Tarasenko et al ,1977
Tarasenko et al., 1977
Comments
Depression in growth. '
No adverse effects
observed in food
consumption, clinical signs,
bodyweight, or
hematological parameters.
Considerable drop in weight
increase, higher arterial
pressure, drop in
hemoglobin, leukocystosis
and thrombopenia,
decreased blood sugar
level, increased phosphorus
in blood and increased
concentration of calcium in
urine.
A decrease in total number
of spermatozoids. in the %
of motile forms and the
duration of their motility;
reduced osmotic resistance
of the spermatozoids; an
increased number of ducts
with desquamated
epithelium.
Barium - Page 7
-------
Terrestrial Toxicity - Barium
Cas No. 7440-39-3
Chemical
Name
Barium
carbonate
Barium
chloride
Species
female rat
chick embryo
Type of
Effect
rep
dev
Description
LOAEL
PEL
Value
13.4
20
Units
mg/m3
mg
Exposure
Route (oral,
s.c.. i.v., i.p..
injection)
inhalation
single
injection
Exposure Duration
/Timing
4 months
Reference
Tarasenko et al., 1977
Ridgway and Kamofsky,
1952
Comments
Shortening of mean
duration of estrous cycle,
disturbances in
morphological structure of,
ovaries, gave birth to
underdeveloped offspring
with considerable mortality
and slow increase in weight
within first 2 months.
inhibition in toe growth
Bariur "'age 8
-------
igica! Uptake Measures - Barium
Cos No. 7440-39-3
Chemical
Name
Species
B-factor
(BCF. BAF.
BMP)
Value
Measured
or
Predicted
(m,p)
Units
Reference
Comments
-------
Freshwater i .city - Barium
Cos No. 7440-39-3
Chemical
Name
barium
barium
barium
barium
barium
barium
Species
aquatic
organisms
daphnid
daphnid .
daphnid
daphnid
daphnid
s
Type of
Effect
chronic
chronic
acute
acute
acute
chronic
Description
NAWQC
CV
LC50
LC50
Lp50
LOEC
Value
109
20336
>530000
410000
14,500
5,800
Units
ug/L
ug/L
ug/L
"9A
"g/L-
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS.
NS
NS
Exposure
Duration
/Timing
NS
NS
24 hours
48 hours
48 hours
3 weeks
Reference
Suterejjd.. 1992
Suteretal.. 1992
LeBlanc, 1980 as cited in
AQUIRE, 1994 :
LeBlanc, 1980 as cited in
AQUIRE, 1994
Biesinger & Christensen,
1972
Biesinger & Christensen,
1972
. Comments
"
Without food.
16% reproductive impairment.
-------
APPENDIX B Benz(a)anthracene -1
lexicological Profile for Selected Ecological Receptors
Benz(a)anthracene
CasNo.: 56-55-3
Summary: This profile on benz(a)anthracene summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and.
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e.. Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e:g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in die
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind lexicological benchmarks used lo derive protective
media concentrations (C^) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic toxicity studies regarding wildlife mammalian
exposure to benz(a)anthracene were identified. Since no laboratory studies focusing on
reproductive or other critical endpoints were available, a mammalian benchmark for
freshwater ecosytems was not derived
August 1995
-------
Terrestrial Biological bh ...Ke Measures - Barium
Cos No. 7440-39-3
Chemical
Name
barium
Species
plant
B-lactor
(BCF. BAF.
BMP)
BCF .
Value
0.15
Measured
or
Predicted
. (m_.P) ._.
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPAL1990e
Comments
-------
APPENDIX B Benz(a)anthracene - 2
Birds: No toxicity studies documenting terrestrial avain exposure to benz(a)anthracene were
identified.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is
available for benz(a)anthracene. Therefore the Tier II method described in Section 4.3.5 was
used to calculate an SCV of 0.025 mg/L. Tier II values or Secondary Chronic Values (SCV)
were developed so that aquatic benchmarks could be established for chemicals with data sets
that did not fulfill all the requirements of the National AWQC. Because the benchmark was
based on an SCV, this benchmark was categorized as interim.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECU) for species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricomutum).
Adequate data sufficient for the development of benchmark values were not identified in
Suter and Mabrey (1994) or in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
octanol-carbon partition coefficient (K^.) to determine protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The SCV, for benz(a)anthracene was used to calculate an EQp value of 9.73 mg
benz(a)anthracene/kg organic carbon. Assuming a mass fraction of organic carbon for the
sediment (f,,,.) of 0.05, the benchmark for the benthic community is 0.49 mg/kg sediment.
Since the EQp number was based on an SCV, the sediment benchmark was categorized as
interim.
August 1995
-------
APPENDIX B
Benz(a)anthracene - 3
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Spvctea
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
myncg-wy
ID
ID
ID
ID
ID
ID
ID
ID
ID
' ID
Study
Spcda*
-
-
-
-
-
.
-
-
-
-
Effect
-
-
-
-
-
-
-
-
-
-
Study Valw
ing/kg-day
-
-
-
-
-
-
-
-
-
-
DMcHptton
-
-
-
-.
-
-
'
-
-
SF
-
-
-
-
-
-
-
-
-
Original Soum
-
-
-
-
-
9
-
'
Benchmark category, a 3 adequate, p = provisional, i = interim; a indicates trial the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
B^M1^^^MM«M^M
nvptwvnBwv
fish and
aquatic
invertebrates
aquatic plants
benthic
community
Bwicnmrti
V«hM
ngfl
0.025 (i)
ID
0.049 0)
Sh*y«p*dM
aquatic
organisms
aquatic
organisms
f^BMwfa«lfcw*
t^WBr^BDn
scv
.
SCVxK.
OrtgM
Some*
AQUIRE,
1995
'
AOUIRE,
1995
'Benchmark Category, a = adequate, p = provisional, i = mtenm; a "' indicates .that the benchmark value
was an order of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Benz(a)anthracene 4
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (Cpn,) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants, and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As discussed previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies documenting mammalian exposure to benz(a)anthracene were
identified. Since no additional laboratory mammal studies focusing on reproductive or other
critical endpoints were identified, a mammalian benchmark for terrestrial ecosystems was not
calculated. ...
Birds: No toxicity studies documenting terrestrial avain exposure to benz(a)anthracene were
identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints- ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10* percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for benz(a)anthracene and, as a result, a
benchmark could not be developed.
Soil Community. Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Benz(a)anthracene - 5
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
R«prM«nt>dv« 3p»elM
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed dear
red-tailed hawk
kestrel
American robin
A-nerican woodcock
plants
soil community
mgAm^tay
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
>P
Study Spa*.
-
-
- .
-
-
-
-
-
-
-
-
fflKt
-
-
-
-
-
-
-
.
-
-
-
-
-
Study Vita*
m0*04qr
-
-
-
.
-
-
-
-
.
-
-
-
-
DMulpaon
-
-
-
-
-
- '
-
-
-
-
-
*P
-
-
-
-
-.
-
-
-
-
- '
-
-
-
Ortghwiaeam
'Benchmark Category, a = adequate, p = provisional, i = interim; a ~ indicates that the bechmark value was an order of magnitude
or mots above the NEL or LEI for other adverse effects.
ID = Insufficient Data .
171. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial invertebrates, and plants. Each
value is identified as whole-boy or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
August 1995
-------
APPENDIX B Benz(a)anthracene - 6
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K,,w values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). The following discussion describes the rationale for selecting the
biological uptake factors and provides the context for interpreting the biological uptake values
presented in Table 4.
As stated in section 5.3.2, the BAP/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. However, these models were considered inappropriate to estimate BAF/s for
benz(a)anthracene because they fail to consider metabolism in fish. A number of studies have
demonstrated that polycyclic aromatic hydrocarbons (PAHs) are readily metabolized in the
tissue of fish (see Polycyclic Aromatic Hydrocarbon Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review. U. S. Fish and Wildlife Service Biol. Rep. 85(1.11).
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemical assuming that the partitioning to tissue is
dominated by lipids. For hydrophobic organic constituents, the bioconcentration factor for
plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion. As with the aquatic ecosystem,
these biological uptake values should be interpreted with caution since they do not address
metabolism of benz(a)anthracene in animal tissue.
August 1995
-------
APPENDIX 8
Benz(a)anthracene - 7
Table 4. Biological Uptake Properties
ftotopteal
fvc0pli)r
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAF, or
BSAF
BAP
BAF
BCF
BAF
BAF
-
BAF
BAF
BAF
BAF
ffabkiM^Ad ar
whow body
liptd
fipid
lipki
liptd
Rpid
-
whole-body
whole-body
whole-body
whole-plant
WhM
800 (t)
800 (t)
800 (t)
BOO(t)
800 ( t)
ID
5.9E - 03
5.6 E -03
4.5 E -02
2.0 E -02
oure*
measured; Stephen 1993
measured; Stephan 1993
measured; Stephan 1993
measured; Stephan 1993
measured; Stephan 1993
-
cate
cate
calc
. U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = insufficient data
August 1995
-------
APPENDIX B
Benz(a)anthracene - 8
References
Brunstrom, B., D. Broman, and C. Naf. 1991. Toxicity and EROD-inducing potency of 24
polycyclic aromatic hydrocarbons (PAHs) in chick embryos. Arch Toxicoi, 65:485-489.
Newsted, J. L. and J. P. Giesy. 1987. Predictive models for photoinduced acute toxicity of
polycyclic aromatic hydrocarbons to Daphnia Magna, Strauss (Cladocera, Crustacea).
Environmental Toxicology and Chemistry, Vol. 6, pp. 445-461.
Sbuthworth, G. R., J. J. Beauchamp and P. K. Schmieder. 1978. Water Res., 12:973-7. As
cited in Hazardous Substance Database (HSDB), National Library of Medicine, 1994.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Biodccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MM. PB93-154762. .'
Suter II, G. W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C. "
Thomann, R. V. 1989. Bioaccumulation model of organic chmeical distribution in aquatic
food chains. Environ. ScL Technol. 23(6): 699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulationin aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615 - 629.
Trucco, R.G., F.R. Engelhardt, and B. Stacy. 1983. Toxicity, accumulation and clearance of
aromatic hydrocarbons in Daphnia Pulex. Environ. Pollut. Ser. A Ecol. BioL 31(3):191-
202. As cited in AQUIRE (AOUatic Toxicity Information REtrieval Database).
Environmental Research Laboratory, Office of Research ,and Development, U.S.
Environmental Protection Agency, Duluth, MN.
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office,of
Health and Environmental Assessment. Washington, D.C. January.
August 1995
-------
..; -fri
APPENDIX B Benz(a)anthracene - 9
Will, M. E. and G. W. Suter n. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Biological Uptake, xmires - Benr(a)anthracene
Cas No.: 56-55-3
Chumlcal Name
Benz(a)anthracene
Soocles
plant
B-ractor
(BCF, BAF,
BMF)
BCF
VahM
002
Measured
' or
Predicted
(m.o)
P
Unlti
(ug/g DW
plant)/(ug/g
soil)
Referanc«
1 U.S. EPA, I990e
Commant*
Plant uptake from soil pertains to
forocjed plants
-------
Freshwater Toxicity - Benz(a)anthracene
CasNo.: 56-55-3
Chemical Name
benz(a)anthracehe
benz(a)anthracene
Speclei
aquatic
organisms
daphnia
maana
Type of
Effect
chron
acute
Description
scv
LC50
Value
0.027
JO
Unlit
ug/i
ua/i
TwtType
(italic/ flow
through)
NS
NS
Exposure
Duration/
Timing
NS
4-davs
Reference
Suter and
Mabrey. 1994
Truccoetal..
1983 as cited In
AQUIRE, 1995
Comments
1
NS » Not Specified
-------
APPENDIX B Benzo(a)pyrene - 2
late (16-18d) gestation with 100 or 150 ug BaP per gram of body weight. Urso and
Gengozian observed reduction in the immune capacity of F, generation mice, which
corresponded to increased tumor incidence in later life. In another study involving laboratory
mice exposure to BaP, Mattison (1980) reported primordial oocyte destruction after a single
ip injection of 80 mg/kg. MacKenzie and Angevine (1981) investigated the effect of daily
oral doses of 0, 10, 40, and 160 mg BaP/kg on days 7-16 of gestation on maternal body
weight, pregnancy maintenance, fetal development, and survival of CD-I mice. For .the F,
generation mice exposed in utero to 10 mg BaP/kg, there was a marked reduction of gonadal
weight and reduced reproductive capacity.
The studies by Urso and Gengozian (1980) and Mattison (1980) were considered unacceptable
for the derivation of a wildlife benchmark value because the intraperitoneal exposure route is
not consistent with probable wildlife exposure routes and the studies lacked sufficient dose-
response data. These two studies were presented to provide a relative perspective for doses at
which lexicological impacts occur. The study by MacKenzie and Angevine (1981) was used
to extrapolate a benchmark value for mammals associated with the aquatic ecosystem. In the
data set on mammalian toxicity, the MacKenzie and Angevine (1981) study was the only
identified study to examine oral exposure to BaP, contain sufficient information to establish a
dose-response curve, and evaluate reproductive and developmental endpoints.
The study value from the MacKenzie and Angevine (1981) study was divided by 10 to
provide a LOAEL-to-NOAEL safety factor. This value was then scaled for representative
species in the freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994)
bw
Benchmark,, = NOAEL x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose. Since the MacKenzie and
Angevine (1981) study documented reproductive effects from benz(a)pyrene exposure to
female and male mating rats, the mean body weight of both genders of representative species
was used in the scaling algorithm to obtain the lexicological benchmarks.
August 1995
-------
APPENDIX B Benzo(a)pyrene - 3
Data were available on the reproductive and developmental effects of benzo(a)pyrene, as well
as chronic survival. There were several acute study values in the data set which were lower
than or approximately equal to, the benchmark value. All of the studies identified were
conducted using laboratory mammals, and since, inter-species differences among wildlife
species were not identifiable, an inter-species uncertainty factor was not applied. Based on
the data set for benzo(a)pyrene and because the benchmark is based on a LOAEL/10, the
benchmarks developed from the MacKenzie and Angevine (1981) study were categorized as
provisional, with a "*" to indicate that adverse effects may occur at the benchmark level.
Birds: Since the minimum data set of at least three avian species was not fulfilled,
toxicological benchmarks for benzo(a)pyrene exposure to representative avian species could
not be calculated.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for benzo(a)pyrene. Therefore, the Tier II method described in Section 4.3.5 was
used to calculate an SCV of 1.3E-5 mg/L. Tier n values or Secondary Chronic Values (SCV)
were developed so that aquatic benchmarks, could be established for chemicals with data sets
that did not fulfill all the requirements of the National AWQC. Because the benchmark is
based on an SCV, this benchmark was categorized.as interim, with an "*" to indicate that
sensitive species of fish may exhibit adverse effects at the benchmark level.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwri).
Aquatic plant data was not identified for aldrin and, therefore, no benchmark was developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^) to determine a protective sediment concentration
(Stephan, 1993). The EQP number is the chemical concentration that may be present in
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. The SCV for benzo(a)pyrene was'used to calculate an EQP number of 13.5 mg
benzo(a)pyrene /kg organic carbon. Assuming a mass fraction of organic carbon for the
sedimetot (f^.) of 0.05, the benchmark for the benthic community is 0.67 mg/kg. Since the
EQp number was based on an SCV, the sediment benchmark is categorized as interim.
August 1995
-------
APPENDIX B
Benzo(a)pyrene 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative
Specie*
mink
river otter
bald, eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herrring gull
kingfisher
Benchmark Value'
mg/kg-d
0.44 (p-)
0.26 (p«)
10
ID
ID
ID
ID
ID
ID
ID
Study
Spedee
mouse
mouse
-
-
-
-
-
-
-
Effect
rep
rep
-
-'
-
-
-
-
-
-
Study Value
mg/kg-d
10
10
-
-
-
-
-
-
-
LOAEL
LOAEL
-
-
-
-
-
-
-
-
SF
10
10
-
-
-
. -
-
-
.
-
Original Source
MacKenzie & Angevine, 1981
MacKenzie & Angevine, 1981
-
t
-
-
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a '*' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ROprtftMllBIIW
Species
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark Value
mo/l
1.3E-05(i*)
ID
0.67 (i) mg/kg
sediment
Study Spedee
Daphnia pulex
-
Daphnia pulex
Description
scv
SCVxK,,.
Original
Source
AQUIRE, 1995
-
AQUIRE, 1995
Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Benzo(a)pyrene - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains .
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to benzo(a)pyrene.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (MacKenzie & Angevine, 1981) was used to derive the benzo(a)pyrene toxicological
benchmark for mammalian species representing the .terrestrial ecosystem. The study value
from the MacKenzie & Angevine (1981) study was divided by 10 to provide a LOAEL-to-
NOAEL safety factor. This value was then scaled for species in the terrestrial ecosystem
using a cross-species scaling algorithm adapted from Opresko et al. (1994). Since the
MacKenzie and Angevine (1981) study documented reproductive effects from benzo(a)pyrene
to female and male mating rats, the mean body weight of both genders of representative
species was used in the scaling algorithm to obtain the toxicological benchmarks. Based on
the data set for benzo(a)pyrene and because the benchmark is based on a LOAEL/10, the
benchmarks developed from the MacKenzie and Angevine (1981) study were categorized as
provisional, with a "*" to indicate that adverse effects may occur at the benchmark level.
Birds: Since the minimum data set of at least three avian species was not fulfilled, a
toxicological benchmark for benzo(a)pyrene exposure to representative avian species could
not be calculated.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for benzo(a)pyrene and, as a result, a
benchmark could not be developed.
August 1995
-------
APPENDIX B
Benzo(a)pyrene - 6
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plant
soil community
BmtaiMi* VWu»«
IDQFK^v
1.2(p')
1.2(p')
' 1.0 (p')
0.42 (p-)
0.30 (p-)
0.29 (p--)
0.15(p')
ID
ID
ID
ID .
ID
ID
ID
Study 8p*dM
mouse
mouse
mouse
mouse
mouse
mouse
mouse
-
-
-
-
-
-
-
Effect
rep
rep
rep
rep
rep
rep
rep
-
-
-
-
-
-
'
Study VMut
mgftfrd
10
10
10
10
10
10
10
-
.
-
.
-
» 1 Jin
wcnpoon
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
.
-
-
-
-
Sf
10
10
10
10
10
10
10
-
-
.
-'
-
Origin*! Soura
MacKenzie &
Angevine, 1981.
MacKenzie &
Angevine, 1981.
MacKenzie & .
Angevine, 1981.
MacKenzie &
Angevine, 1981.
MacKenzie &
Angevine, 1981.
Mackenzie &
Angevine, 1981.
MacKenzie &
Angevine, 1981.
-
-
-
-
Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order of magnitude or
more above the NEL or LEL for other adverse effects.
ID = Insufficient Data '
August 1995
-------
APPENDIX B Benzo(a)pyrene - 7
III. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors .are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemical* with log K^ values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for consituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. However, these models were considered inappropriate to estimate BAF/s for
benzo(a)pyrene (BaP) because they fail to consider metabolism in fish. A number of studies
have demonstrated that polycyclic aromatic hydrocarbons (PAHs) such as BaP are readily
metabolized in the tissue of fish (see Polycyclic Aromatic -Hydrocarbon Hazards to Fish,
Wildlife, and Invertebrates: A Synoptic Review. U. S. Fish Wildlife Service Biol. Rep.
85[1.11]. Stephan (1993) noted that unpublished field data by Burkard resulted in predicted
BAPs of 17 to 228 for four PAHs with three and four rings for fish with 5% lipids, and
suggested that it seems unlikely that PAHs with five rings would have BAPs greater than
1,000. Converting the BAPs to BAF/s (i.e., dividing by lipid fraction of 0.05) results in a
BAF,d range of 340 to 4,560. The geometric mean of these values (1,245) was rounded to a
BAF/ of 1,000 to represent a default value for BaP and other five or four ring PAHs.
Considering that PAH levels in fish are usually low and that the higher molecular weight
PAHs do not seem to accumulate in fish, a BAF/ of 1,000 appears to be a reasonable,
although not overly conservative, value for bioaccumulation. The bioconcentration factor for
BaP in fish was assumed to be equivalent to the BAF/ of 1,000, however, because it is not
known whether fish metabolize BaP more rapidly via the gut or gills, it is difficult to
August 1995
-------
APPENDIX B Benzo(a)pyrene 8
determine .whether it over- or underestimates the actual bioconcentration of BaP. In short,
selecting one biological uptake of 1,000 for BaP and similar PAHs represents a best estimate
(i.e., central tendency) of the bioaccumulation and bioconcentration of this class of
compounds. However, steady-state measured data on biological uptake of BaP (and most
PAHs) are very limited at this time and these default values for BAF/ and BCF/ should be
interpreted with caution.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e., BaP BBF/TCDD BBF)
is multiplied by the TCDD standard for terrestrial, vertebrates, invertebrates, and earthworms,
respectively. For hydrophobic organic constituents, the bioconcentration factor for plants was
estimated as described in Section 6.6.1 for above ground leafy vegetables and forage grasses.
The BCF is based on roufe-to-leaf translocation, direct deposition on leaves and grasses, and
uptake into the plant through air diffusion. As with the aquatic ecosystem, these biological
uptake values should be interpreted with caution since they do not address metabolism of BaP
in animal tissue.
August 1995
-------
APPENDIX B
Benzo(a)pyrene - 9
Table 4. Biological Uptake Properties
ecological receptor
limnetic trophic level 4 fish
limnetic trophic level 3 fish
fish
littoral trophic level 4 fish
littoral trophic level 3 fish
trophic level 2 invertebrates
terrestrial vertebrates
terrestrial invertebrates
earthworms
plants
BCF, BAF,
orBSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
HpkMwsedor
wholt oooy
lipid
lipid
lipid
lipid
lipid
lipid
whole- body
whole-body
whole- body
whole-plant
value
1,000.(t)
1,000(1)
1,000(t)
1,000 (t)
1,000(t)
-
0,017
0.016
0.13
0.011
source
conservative default value for PAHs based on field
BAFs in Stephan, .1993
conservative default value for PAHs based on field
BAFs in Stephan, 1993
conservative default value for PAHs based on field
BAFs in Stephan, 1993
same as value as in the limnetic ecosystem
same value as in the limnetic ecosystem
possible values under review
estimated based on beef biotransfer ratio with
2,3,7,8-TCDD
estimated based on beef biotransfer ratio with
2,3,7,8-TCDD
estimated based on beef biotransfer ratio with
2,3.7,8-TCDD
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Benzo(a)pyrene -10
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE (AQItalic Toxicity_/nformation /?£trieval Database). 1995. Environmental Research
Laboratory, Office of.Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Agency for Toxic Substances and Disease Registry (ATSDR), 1988. Toxicological Profile
for Benzp(a)pyrene. U.S. Department of Health and Human Services.
Barbieri, O., E. Ognio, O. Rossi, S. Astigiano, and L. Rossi. 1986. Embryotoxicity of
Benzo(a)pyrene and Some of its Synthetic Derivatives in Swiss Mice. Cancer Research,
46:94-98.
Bulay, O.M. and L.W. Wattenberg. 1970. Carcinogenic Effects of Subcutaneous
Administration of Benzo(a)pyrene During Pregnancy on the Progeny (34993). Proc. of
the Soc. ofExp. Biol. and Med., Vol. 135, pp. 84-86..
Connell, D.W. and G. Schuurmanri. 1988. Evaluation of Various Molecular Parameters as
Predictors of Bioconcentration in Fish. Ecotoxicology and Environmental Safety. 15;
324-335.
Connolly, J.P. 1991. Application of a Food Chain Model to Polychlorinated Biphenyl
Contamination of the Lobster and Winter Flounder Food Chains in New Bedford Harbor.
Environ. Sci. Technol., 25: 760-770.
Frank, Allan P., Peter F. Landrum, and Brian J. Eadie. 1986. Polycyclic Aromatic
Hydrocarbon Rates of Uptake, Depuration, and Biotransformation by Lake Michigan
Stylodrilus Heringianus. Chemosphere 15(3):317-330.
Freitag, D., L. Ballhom, H. Geyer, and F. Korte. 1985. Environmental Hazard Profile of
Organic Chemicals. Chemosphere 14(10): 1589-1616.
August 1995
-------
APPENDIX B Benzo(a)pyrene -11
Hoffman, David J., and Martha L. Gay. 1981. Embryotoxic Effects of Benzo(a)pyrene,
Chrysene, and 7,12-Dimethylbenz(a)anthracene in Petroleum Hydrocarbon Mixtures in
Mallard Ducks. Journal of Toxicology and Environmental Health 7:775-787.
Hose, Jo Ellen, James B. Hannah, Harold W. Puffer, and Marsha L. Landolt. 1984.
Histologic and Skeletal Abnormalities in Benzo(a)pyrene-treated Rainbow Trout Alevins.
Arch. Environ. Contain. Toxicol. 13:675-684.
Hose, J. E., J. B. Hannah, D. DiJulio, M. L. Landolt, B. S. Miller, W. T. Iwaoka, and S. P.
Felton. 1982. Effects of benzo(a)pyrene on early development of flatfish. Arch. Environ.
Contam. Toxicol. 11:167-171.
Hazardous Substance Database (HSDB). 1992.
International Agency for Research on Cancer (IARC). 1983. I ARC Monographs on the
Evaluation of the Carcinogenic Risk of Chemicals to Humans - Polynuclear Aromatic
Compounds, Part 1, Chemical, Environmental, and Experimental Data. Volume 32.
Jimenez, B. D., C. P. Cirmo, and J. F. McCarthy. 1987. Effects of feeding and temperature
on uptake, elimination and metabolism of benzo(a)pyrene in the bluegill sunfish (Lepomis
macrochirus). Aquat. Toxicol. 10(l):41-57. As cited in AQUIRE (AQUatic Toxicity
Information REtrieval Database). 1995. Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
Johnsen, S., J. Kukkonen, and M. Grande. 1988. Influence of natural aquatic humic
substances on the bioavailability of benzo(a)pyrene to Atlantic salmon. Sci. Total
Environ. 81/82:691-702. As cited in AQUIRE (AOUatic Toxicity Information REtrieval
Database). 1995. Environmental Research Laboratory, Office of Research and
Development, U.S. Environmental Protection Agency, Duluth, MN.
Johnsen, S., J. Kukkonen, and M. Grande. 1989. Influence of natural aquatic humic
substances on the bioavailability of benzo(a)pyrene to Atlantic salmon. Sci. Total
Environ. 81/82:691-702. As cited in AQUIRE (AOUatic Toxicity Information REtrieval
Database). 1995. Environmental Research Laboratory, Office of Research and
Development, U.S. Environmental Protection Agency, Duluth, MN.
August 1995
-------
APPENDIX B Benzo(a)pyrene -12
Kenaga, E.E. 1982. Predictability of Chronic Tbxicity from Acute Tox'icity of Chemicals in
Fish and Aquatic Invertebrates. Environmental Toxicology and Chemistry, Vol. , pp. 347-
358. .
Kuhnhold, W.W., and F. Busch. 1978. On the uptake of three different types of hydrocarbons
by salmon eggs (Salmo salar L.). Meeresforsch. 26:50-59. As cited in Eisler, R. 1987.
Polycyclic Aromatic Hydrocarbon Hazards to Fish, Wildlife, and Invertebrates: A
Synoptic Review. U. S. Fish Wildl. Serv. Biol. Rep. 85(1.11). 81 pp.
Landrum, Peter F., Brian J. Eadie, and Warren R. Faust. 1991. Toxicokinetics and toxicity
of a mixture of sediment-associated polycyclic aromatic hydrocarbons to the Amphipod
Diporeia Sp. Environmental Toxicology and Chemistry 1.0:35-46.
Leversee, G.J., J.P. Geisy, P.P. Landrum, S. Bartell, S. Gerould, M. Briino, A. Spacie, J.
Bowling, J. Haddock, and T. Fannin. 198.1. Disposition of Benzo(a)pyrene in Aquatic
Systems Components: Periphyton, chironomids, daphnia, fish. Pages 357-366 ]n M.
Cooke and A.J. Dennis (eds.). Chemical Analysis and Biological Fate: Pplynuclear
Aromatic Hydrocarbons. Fifth International Symposium. Battelle Press, Columbus, Ohio.
As cited in Eisler, R. 1987. Polycyclic Aromatic Hydrocarbon Hazards to Fish, Wildlife,
and Invertebrates: A Synoptic Review. U. S. Fish Wildl. Serv. Biol. Rep. 85(1.11). 81
PP-
Lu, Po-Yung, Robert L. Metcalf, Nancy Plummer, and Douglas Mandel. 1977. The
environmental fate of three carcinogens: Benzo(a)pyrene, benzidine, and vinyl chloride
evaluated in laboratory model ecosystems. Arch. Environ. Contain. Toxicol. 6:129-142.
Mackenzie, Karen M., and D. Murray Angevine: 1981. Infertility in mice exposed in utero
to benzo(a)pyrene. Biology of Reproduction 24:183-191.
Mattison, Donald R. 1980. Morphology of oocyte and follicle destruction by polycyclic
aromatic hydrocarbons in mice. Toxicology and Applied Pharmacology 53:249-259.
McCarthy, John F. 1983. Role of paniculate organic matter in decreasing accumulation of
polynuclear aromatic hydrocarbons by Daphnia magna. Arch. Environ. Contam. Toxicol.
12:559-568.
August 1995
-------
APPENDIX B Benzo(a)pyrene -13
McCarthy, John F., and Braulio D. Jimenez. 1985. Reduction in bioavailability to bluegills
of polycyclic aromatic hydrocarbons bound to dissolved humic material. Environmental
Toxicology and Chemistry 4:511-521.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substance) Database. March 1994.
Neff, J.M. 1979. Polycyclic aromatic hydrocarbons in the aquatic environment. Applied
Science Publ. Ltd., London. 262pp. As cited in Eisler, R. 1987. Polycyclic Aromatic
Hydrocarbon Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review. U. S.
Fish Wildl. Serv. Biol. Rep. 85(1.11). 81 pp.
Newsted, John L., and John P. Giesy. 1987. Predictive models for photoinduced acute
toxicity of polycyclic aromatic hydrocarbons to Daphnia Magna, Strauss (Cladocera,
Crustacea). Environmental Toxicology and Chemistry 6:445-461.
Sabourin, T.D. and R.E. Tullis. 1981. Effect of Three Aromatic Hydrocarbons oh
Respiration and Heart Rates of the Mussel, Mytilus californianus. Bull. Environm.
Contam. Toxicol., 26:729-736.
Spacie, Anne, Peter F. Landrum, and Gordon J. Leversee. 1983. Uptake, depuration, and
biotransformation of anthracene and benzo(a)pyrene in bluegill sunfish, 1982.
Ecotoxicology and Environmental Safety 7:330-341.
Stephan, C.E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. PB93-154672.
Suter II, G.W., M.A. Futrell, and G.A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste. Management, U.S. Department of Energy, Washington, DC.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
August 1995
-------
Terrestrial Toxlclty - Benzo(a)pyrene
Cos No.: 50-32-8
Chemical Nam*
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
benzo(a)pyrene
Species
rat
mouse
dog
monkey
rabbit
guinea pig
hamster
mammal
Endpolnt
acute
acute
acute
acute
acute
acute
acute
acute
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
20
114
1
2
115
500
1157
200
Unite
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ng/kg-body wt.
ug/kg-body wt.
ng/kg-body wt.
Exposure
Rout* (oral,
B.C., I.V., l.p.,
Inlectlon)
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration /
Timing
NS
NS
NS
NS
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
RTECS. 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
Comments
N
0
-------
APPENDIX B , 8enzo(a)pyrene -14
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
Trucco, R.G., F.R. Engerhardt, and B. Stacey. 1983. Toxicity, Accumulation and Clearance
of Aromatic Hydrocarbons in Daphnia pulex. Environ. Pollut. Ser. A Ecol. Biol. 31(3):
191-202. As cited in AQUIRE (AOUatic Toxicity Information REtrieval Database).
1995. Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN.
U.S. Environmental Protection Agency. 1980. Ambient Water Quality Criteria for
Polynuclear Aromatic Hydrocarbons. Criteria and Standards Division, Washington, DC.
202pp. .
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment. Washington, D.C. January.
U.S. Environmental Protection Agency. 1992. 304(a) Criteria and Related Information for
Toxic Pollutants. Water Management Division - Region IV.
Urso, Paul, and Nazareth Gengozian. 1980. Depressed humoral immunity and increased
tumor incidence in mice following in utero exposure to benzo[a] pyrene. Journal of
Toxicology and Environmental Health 6:569-576.
Will, ME', and G.W. Suter, 1994. lexicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Freshwater Toxlciry - Benzo(a)pyrene
Cos No.: 50-32-8
Chemical Name
benzo[a]pyrene
benzo[a]pyrene
benzo(a)pyrene
benzolalpyrene
Species
rainbow trout
(alevins)
rainbow trout
(alevins)
Daphnla
pulex
Sand sole
eqqs
Type of
Effect
dvp
dvp
mortality
emb
.
Description
NOEC
LOEC
LC50
AEL
Value
0.00000008
0.00000021
5
0.1
Uunits
ng/ml
ng/ml
ua/1
PPb
Test Type
(static/ now
through)
static
static
NS
static
Exposure
Duration /
Timina
36-day
36-day
4-day
48-hour
Reference
Hose et al., 1984
Hose et al.. 1984
Trucco et al., 1983
as cited in
AQUIRE, 1995
Hoseetal.. 1982
Comments
Effects - pycnosis and abnormal
erylhrocytes were observed.
Could not determine if these
effects would have an adverse
effect on population.
Effects - pycnosis, necrosis of
skeletal muscle and spinal cord.
Exposure to 0.10 ppb BAP
resulted in decreased hatching
success in sand sole eqqs.
NS = not specified
-------
Terrestrial Toxlclty . nzo(a)pyrene
Cos No.: 50-32-8
Chemical Nam*
benzo(a]pyrene
benzo(a]pyrene
benzol a]pyrene
benzofalpyrene
Species
CD-1 mice
C3H/AnF
mice
mallard duck
mice
EndDoInt
rep, fertility
immun.
embryotoxic
rep
Description
LOAEL
LOAEL
LOAEL
AEL
Value
10
100
0.036
80
Units
mg/kg-day
ug/q-bodywt.
mg/kg-egg wt.
mg/ka-bodv wt.
Exposure
Routs (oral,
Infection)
oral
l.p.
applied to
eggshell
surface
i.p.
Exposure
Duration /
Timing
days 7- 16 of
gestation
1M3dor 16-
18dof
gestation
1 - 18 days of
incubation
single i.p.
injection
Rsfsrsncs
MacKenzie and
Angevine, 1980
Urso and
Gengozian, 1980
Hoffman and Gay.
1981
1
Mattison. 1980
Comment*
Doses were 0.10.40, 160 mg/kg. At 1 0 mg/kg reduction of
gonadal weight, reduced fertility and reproductive capacity
were observed among offspring.
Reduction in immune capacity in F1 generation mice -
resulting In increased tumor incidence in later life.
Study doses were 0.002. 0.01 and 0.05 mg/egg (equivalent to
0.036. 0.18. 0.9 mg/kg fresh weight), significant reduction of
embryonic growth and an increased incidence of abnormal
survivors.
Chemical was dissolved in com oil. Mice were sacraficed 6
days after Injection. Effect - primordial oocvte destruction.
NS = not specified
-------
Freshwater B!o!og!ca! Uptake Measures - B«nzo(a)pyrene
Cos No.: 50-32-8
Chemical Name
benzo|a]pyrene
benzo[a]pyrene
benzo[a]pyrene
benzo(a]pyrene
benzo(a]pyrene
benzofajpyrene
benzo(a]pyrene
b6nzo|a)pyrene
Soecles
mosquito fish
attantic salmon; egg
bluegill
bluegill
bluegill
bluegill
bluegill
blueqill
B-factor (BCF,
BAF. BMR
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Valuo
930
71
12
2,657
4,900
490
3.208
608
Measured or
Predicted
(m.o)
m
m
m
m
P
P
m
m'
Unit*
NS
NS
NS
ml/g
NS
NS
NS
NS
Reference
Luetal, 1977
Kuhnhold and Busch,
1978 as cited in Eisler.
1987
Leversee et al., 1981 as
cited in Eisler, 1987
McCarthy and Jimenez,
1985
Spacie et al., 1983
Spacieetal , 1983
Jimenez et al., 1987 .as
cited in AQUIRE, 1994
Jimenez et at.. 1987 as
cited in AQUIRE. 1994
Comments
Exposure period = 3 days
Exposure period = 168 hours
Exposure period = 4 hours
Test water absent of dissolved humic
material. Flow-through water system.
Estimated BCF value from measured
uptake and depuration rates lor 4-hour
exposure (Ku/Kd). Vaue includes parent
compound plus metabolites.
Estimated BCF value from measured
uptake and depuration rates (or 4-hour
exposure (Ku/Kd). Vaue includes parent
compound only.
2-day exposure study. Flow-through
test. Life stage of fish = 10-15 G.
2-day exposure study. Flow-through
test. Life staqe of fish = 10-15 G.
-------
Freshwater Biological Uptakt asures - Benzo(a)pyrene
Cos No.: 50-32-8
Chemical Name
benzo[a]pyrene
benzo(a]pyrene
benzo[a]pyrene
benzo[a]pyrene
benzo[a]pyrene
benzo[a]pyrene
benzo[a]pyrene
benzol alpvene
Species
bluegill
bluegill
atlantic salmon
atlantic salmon
golden ide
ns
lish
clam (Rangia
cuneata)
B-factor (BCF.
BAF.BMF)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
377
367
2,310
2,310
4BO
2,985
30
9 to 236
Measured or
Predicted
(m,p)
m
m
m
m
m
P
m
m
Units
NS
NS
NS
NS
ug/g/
ua/g
NS
NS
NS
Reference
Jimenez et al.. 1987 as
cited in AQUIRE, 1994
Jimenez et al.. 1987 as
cited in AQUIRE, 1994
Johnsen et al., 1988 as
cited in AQUIRE, 1994
Johnsen et al., 1988 as
cited in AQUIRE, 1994
Freitagetal., 1985
U.S.EPA, 1993a
U.S.EPA, 1992
Nett, 1979 as cited in
Eisler. 1987; U.S. EPA.
1980
Comments
2-day exposure study. Flow-through
test. Life stage of fish = 10-15 G
2-day exposure study. Flow-through
test. Life stage of fish = 10-15 G.
2-day exposure study. Static test. Life
stage of fish = 2 G
2-day exposure study. Static test. Life
stage of fish = 2 G.
This fish species represents an
intermediate position between a trout
and a carp.
BCF normalized to 1% lipid
Normalized to 3% lipids:
Exposure period = 24 hours
-------
Terrestrial Biological Uptake Measures - Benzo(a)pyrene
Cos No.: 50-32-8
Chemical Name
benzo[a)pyrene
Species
plant
B-factor
(BCF. BAF.
BMR
BCF
Value
0.01 1
Measured
or
Predicted
(m.o)
P
Units
(ug/g DW
plant)/(ug/g
soil)
Reference
U.S. EPA, IWOe
Comments
Plant uptake from soil pertains to
leafy veatabtes
-------
Freshwater Biological Uptak >asures - Benzo(a)pyrene
Cos No.: 50-32-8
Chemical Name
benzo|a]pyrene
benzofajpyrene
benzo[a]pyrene
benzofajpyrene
benzolalpyrene
Species
Oaphnia magna
Daphnia magna
Daphnia magna
Pontoporeia hoyi
(amphipod)
Stylodrilus
herinqianus
B-factor (BCF,
BAF. BMP)
\
BCF
BCF
BCF
BAF (soil)
BAF
Value
2,837
12,761
8,000
2.3 - 7.2
676
Measured or
Calculated
(m,c)
m
m
m
m
m
Units
NS
NS
mUg
nmol/g/
nmol/g
NS
Reference
Leversee el al., 1981 as
cited in Eisler, 1987
Newsted and Giesy,
1987
McCarthy, 1983
Landrum el al, 1991
Frank et al, 1986
Comments
Exposure period = 6 hours
24-hour exposure
24-hour exposure, no paniculate
organics (yeast) present in water.
Sediments dosed with a mixture ot
PAH's at four concentrations. Calculation
ot BAF equals the concentration in the
organism divided by the concentration in
the sediment .
BAF was calculated from an equation of
measured values, equation considered
both water and sediment uptake.
NS = not specified
-------
APPENDIX B Beryllium-1
Toxicological Profile for Selected Ecological Receptors
Beryllium
Cas No.: 7440-41-7
Summary: This profile on beryllium summarizes the toxicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagmfication factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagmfication factors (BMFsj are also'
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire toxicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic 'or chronic studies were identified which studied the effects
of beryllium toxicity on reproductive or developmental endpoints in mammalian species.
Birds, No suitable subchronic or chronic studies were identified which studied the effects of
beryllium toxicity in avian species.
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
beryllium. Therefore, a Secondary Chronic Value (SCV) of 5.1 E-03 mg/1 as reported by Suter
and Mabrey (1994) was utilized. Because the benchmark selected is based on a SCV, rather
than an FCV, it was categorized as interim.
August 1995
-------
APPENDIX B Beryllium - 2
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or 2) an effective concentration (ECXX) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum cdpricornutwn). The aquatic plant
benchmark for beryllium is 100 mg/1 based on a reduction in autotrophic growth rates of
Chlorella vannieli (Suter and Mabrey, 1994). As described in Section 4.3.6, all benchmarks for
aquatic plants were designated as interim.
Benthic community: The beryllium benchmark protective of benthic organisms is pending a U.S.
EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Beryllium 3
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative
Specfee
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
. herring gull
kingfisher
Benchmark
Value* »B*a>
««y
ID
ID .
ID
ID
ID
ID
ID
ID
ID
ID
Study
Sp«d«*
-
-
-
-
eiuct
-
-
-
-
Study Vatu*
me/kfl-day
-
-
-
-
DMerlpflon
-
-
-
'
-
se -
.
'
Oriflto** Source
.-
-
-
-
-
-
-
. -
-
'Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEI for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Rapre tentative
Sp»«*»
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark
Va»«e*
mg/t
5.1 E-03(i)
100 (i)
under review
Study
Species
aquatic
organisms
aquatic
plants
Original
Value
mg/l
5.1 E-03
100
Description
scv
LOEC
-
Original Soon*
Suter & Mabrey,
1994
Suter & Mabrey,
1994
-
'Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (') indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL tor other adverse effects.
August 1995
-------
APPENDIX B Beryllium-4
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were identified which studied the effects of orally administered
beryllium on reproductive or developmental endpoints in mammalian species.
Birds: As noted in the freshwater ecosystem discussion, no suitable subchronic or chronic
studies were identified which studied the effects of beryllium.toxicity in avian species.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the Lowest Observable Effects Concentration
(LOEC) values and then approximating the 10th percentile. If there were 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth.yield
reductions, or other effects reasonably assumed to impair the ability of a plant population to
sustain itself, such as a reduction in seed elongation. The benchmark for terrestrial plants was
10 mg/kg, based on a Lowest Observable Effects Concentration (LOEC) of 10 ppm, which
resulted in unspecified toxic effects (Kabata-Peridias and Pendias, 1984 as cited in Will and
Suter, 1994). Since only one value was presented by Will and Suter (1994), with the
benchmark being a LOEC value, the terrestrial plant benchmark of 10 mg/kg-day was
categorized as interim.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Beryllium - S
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
ft*j*94*ftt*tfv*
.. Sped**
doer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
Bwtcfontrk
Vife»«
«l«fl«8Hbjf
10
. ID
ID
10
10
ID
ID
ID
ID
ID
ID
ID
10mg/kg (i)
ID
Study
3p»oi*»
-
-
-
-
-
.
-
-
-
-
terrestrial
plants
-
Effect
-
-
-
-
-
unspecified
Study
V«lu»
'mgfa>
*** .
.
10
BmcfipBoit
.
-
-
-
-
-
-
LOEC
-
«F
-
-
-
-
-
-'
.
-
Or^hwJ Sourw
.
-
'
.
-
'
.
-
-
-
Kabata-Pendias
and Pendias, 1984
as cited in Will &
Suter, 1994
Benchmark Category, a » adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Beryllium 6
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconceritration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values.
The whole-body BCF value for beryllium was the geomean of measured values (Stephan,
1993). Insufficient data were identified to determine the BCF value in aquatic invertebrates,
terrestrial vertebrates, terrestrial invertebrates and earthworms. A whole plant BCF value of .
1.0 E-02 was derived from U.S. EPA (1992e). For metals, empirical data were used to derive
the BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake
response slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem
since,most of the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
ecological
nceptor
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAIvor
BSAF
BCF
BCF
-
BCF
ItpfaMMMd of
whoto-boxly
whole
-
-
whole-plant
value
19
ID
ID
ID
ID
1.0 E-02
tourc*
Stephan. 1993
-
-
U.S. EPA. 19929
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Beryllium-7
References
AQUERE (AQUatic Toxicity_Information REtrieval Database), 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
ASTER Ecotoxicity Profile. 1992. U.S. EPA (Environmental Protection Agency),
Environmental Research Laboratory-Duluth, MN.
Barrows, M.E., S.R. Petrocelli, K.J. Macek, and J. Carroll. Bioconcentration and elimination
of selected water pollutants by bluegill sunfish (Lepomis macrochirus). Toxic Chemicals
379-392. As cited in U.S. EPA (Environmental Protection Agency). 1993b. Soil
Screening Level Fact Sheet (Second draft August 12, 1993) Interim Guidance. Office of
Emergency and Remedial Response, Washington, DC.August.
HSDB (Hazardous Substances database). 1992.
Kabata-Pendias,A., and H. Pendias. 1984. Trace elements in soils and plants. CRC Press, Inc.
Boca Raton, Florida. As cited in Will, M.E and G.W. Suter II. 1994. Toxicological
Benchmarks for Screening of Potential Contaminants of Concern for Effects on Terrestrial
Plants: 1994 Revision. DE-AC05-840R21400. Office of Environmental Restoration and
Waste Management, U.S. Department of Energy, Washington, DC.
Leonard, A. and R. Lauwerys. 1987. Mutagenicity, carcinogenicity and teratogenicity of
beryllium. Mutation Res. 186:35-42.
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
Mathur, R., S.Sharma, S. Mathur and A.O. Prakash. 1987. Effect of beryllium nitrate on early
and late pregnancy in rats. Bull. Environ. Contam. Toxicol. 38:73-77.
Ridgway, L.P and D.A. Kamofsky. 1952. The effects of metals on the chick embryo: Toxicity
and production of abnormalities in development. Ann. N.Y. Acad. Sci. 55:203.
Stephan, C.E. 1993. Derivation of Proposed Human Health and Wildlife
Bioaccumulation Factors for the Great Lakes Initiative. Office of Research and
Development, U.S. Environmental Research Laboratory. PB93-154672. Springfield, VA.
August 1995
-------
APPENDIX B Beryllium - 8
Suter 13, G.W., M.A. Futrell, and G.A. Kerchner. 1992. lexicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Suter, G.W., and J.B. Mabrey. 1994. Toxicological benchmarks for screening potential
contaminants of concern for effects on aquatic biota: 1994 revision. ES/ER/TM-96/R1
Office of Environmental Restoration and Waste Management, U.S Department of Energy,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1984. Review Draft: Health Assessment
Document for Beryllium. EPA 600/8-84-026A. Office of Health and Environmental
Assessment, Washington, DC. December.
U.S. EPA (Environmental Protection Agency). 1992. TSC1292. Criteria Chart. Region IV.
Water Management Division, 304(a) Criteria and Related Information for Toxic Pollutants.
December.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S.EPA (Environmental Protection Agency). 1993b. Soil Screening Level Fact Sheet
(Second draft August 12, 1993) Interim Guidance. Office of Emergency and Remedial
Response, Washington, DC. August.
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
April. .
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2): Chemical toxicity of metals
and metalloids. Plenum Press, N.Y., 1978.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
World Health Organization, 1990. Environmental Health Criteria 106: Beryllium. Published
under the joint sponsorship of the United Nations Environment Programme, the
International Labour Organisation, and the World Health Organization.
Wren, C.D., H.R. Maccrimmon and B.R. Loescher. 1983. Examination of bioaccumulation
and biomagnification of metals in a precambrian shield lake. Water, Air, Soil Pollution
19:277-291.
August 1995
-------
Terrestrial Toxicity - Beryllium
Cas No. 7440-41-7
Chemical
Name
beryllium
beryllium .
Species
rat
rat
Type of
Effect
lei
fet
Description
PEL
PEL
Value
0.316
0.316
Units
Exposure
Route (oral,
s.c., i.v., i.p..
injection)
mg/kg-day |i.v.
mg/kg-day
i.v.
Exposure Duration
/Timing
Day 1 jrt j}estation
Day 1 1 following
mating.
Reference
Mathur et al., 1987 as cited
in WHO, 1990
Mathur et al.. 1987 as cited
in WHO, 1990
Comments
Offspring died 2-3 days
after delivery.
All fetuses were resorbed.
-------
Freshwater T>_ ity - Beryllium
Cos No. 7440-41-7
Chemical
Name
Beryllium
Beryllium
Beryllium
Beryllium
Beryllium
Species
aquatic
organisms
fish
daphnid
fish
daphnid
Type of
Effect
chronic
chronic
chronic
chronic
chronic
Description
NAWQC
CV
CV
EC20
EC20
Value
0.61
57
5.3
148
3.8
Units
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
JStatic/Flow
Through)
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS
NS
NS
NS
Reference
Suteretal., 1992
Suteretal., 1992
Suterelal., 1992
Suterelal.. 1992
Suteretal.. 1992
Comments
-------
Freshwater Biological Uptake Measures - Beryllium
Cos No. 7440-41-7
Chemical
Name
Beryllium
Beryllium
Species
fish
bluegill
B-factor
(BCF, BAF,
BMP)
BCF
BCF
Value
19
19
Measured
or
Predicted
(m,p) '
m
m
Units
IJKg
NS
Reference
U. S. EPA, 1992
Barrows et al., 1980 as
cited in U.S. EPA, 1993b
Comments
Normalized to 3% lipid.
Whole body BCF.
-------
Terrestrial Biological Up ..e Measures - Beryllium
Cas No. 7440-41-7
Chemical
Name
beryllium
Species
plant
B-factor
(BCF. BAF.
BMP)
BCF
Value
0.01
Measured
or
Predicted
. (m-P)
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Butylbenzyl phthalate - 1
lexicological Profile for Selected Ecological Receptors
Butylbenzyl phthalate
Cas No.: 85-68-7
Summary: This profile on butylbenzyl phthalate summarizes the lexicological benchmarks
and biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire toxicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several studies were identified which investigated the effects of butylbenzyl
phthalate exposure to mammals. Rats were exposed to dietary butylbenzyl phthalate at 17,
51, 159, 470 and 1417 mg/kg-day (NTP, 1985). After 26 weeks, increased liver-to-body
weight and liver-to-brain weight ratios were seen in the 470 mg/kg-day treatment group. A
NOAEL of 159 mg/kg-day and a LOAEL of 470 mg/kg-day were reported for these
pathological effects. Another study exposed male rats to doses of butylbenzyl phthalate
ranging from 0-25,000 mg/kg-diet for 90 days (NTP, 1981). Rats in the 25,000 mg/kg-diet
treatment group exhibited depressed weight gain and testicular degeneration from which, a
LOAEL of 25,000 mg/kg-diet can be inferred for pathological and reproductive effects. Since
August 1995
-------
APPENDIX B Butylbenzyl phthaJate - 2
no information was provided on daily food consumption or body weight, conversion from
mg/kg-diet to mg/kg-day required the use of an allometric equation:
Food consumption = 0.056(W°-6611) where W is body weight in kg (Nagy, 1987).
Assuming a body weight of 0.4 kg, the LOAEL of 25,000 mg/kg-diet was converted to
1909.81 mg/kg-day. Agarwal et al. (1985) fed male rats 0.625, 1.25, 2.5, and 5% butylbenzyl
phthalate for 2 weeks. Rats exposed to levels of 2.5% butylbenzyl phthalate and higher
exhibited reductions in total body, thymus, testis, epididymis, prostate and seminal vesicle
weight Based on these results, a LOAEL of 2.5% and a NOAEL of 1.25% butylbenzyl
phthalate could be inferred. Daily food consumption and body weight information were not
provided. Therefore, using the allometric equation presented above and assuming a body
weight of 0.3 kg, a LOAEL of 2000 mg/kg-day and a NOAEL of 1000 mg/kg-day were
calculated. .
The NTP (1981) study measures chronic reproductive effects that may impair the fecundity
of a wildlife population. Therefore, the study LOAEL of 1909.81 mg/kg-day was chosen for
derivation of a benchmark value. The NTP (1985) study was not considered suitable for
derivation of a benchmark value because of the uncertainty surrounding the critical endpoim.
While increases in liver-to-body weight and liver-to-brain weight ratios may cause inimical
effects, the results of the study do not clearly indicate such effects could impair the
sustainability of a population. The Agarwal et al. (1985) study also was not selected because
the exposure duration was too short to be appropriate for a chronic toxicity study.
The LOAEL value from the NTP (1981) study was divided by 10 to provide a LOAEL-to-
NOAEL safety factor. This value was then scaled for species representative of a freshwater
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL, x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the, body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
NTP (1981) documented reproductive effects from butylbenzyl phthalate exposure to male
rats, male body weights of the representative species were used in the scaling algorithm to
obtain lexicological benchmarks.
Data were available on reproductive, developmental, growth and survival endpoints for
butylbenzyl phthalate exposure. In addition, the data set contained acute and chronic toxicity
studies that were conducted during sensitive life stages. The data set contained a study values
for pathological and developmental endpoints (NTP, 1985 and Lake et al., 1978) that were
approximately an order of magnitude lower than the benchmark value. Based on the data set
August 1995
-------
APPENDIX B Butylbenzyl phthalate 3
for butylbenzyl phthalate, the benchmarks developed from the NTP (1981) study were
categorized as provisional, with a "*" to indicate that some adverse effects have been
observed at the benchmark level.
Birds: No suitable subchronic or chronic studies were found for butylbenzyl phthalate
toxicity in avion species. Thus, benchmarks for avian species could not be derived.
One acute study was identified which found that 0.05 ml of butylbenzyl phthalate injected
into fertilized eggs produced no embryonic malformations (Bower, 1970 as cited in I ARC,
1982). However, this study was not considered suitable for calculation of a benchmark value
because data were not identified on (1) the direct absorption of butylbenzyl phthalate from
direct contact with the eggs or (2) on the maternal transfer of butylbenzyl from mother to
egg. Without sufficient absorption data, it is not possible to distinguish maternal transfer of
butylbenzyl from the applied dose.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for butylbenzyl phthalate. Therefore, the Tier II methodology described in Section
4.3.5 was used to calculate a Secondary Chronic Value (SCV) of 16 mg/L for butylbenzyl
phthalate. Teir n values or SCV were developed so that aquatic benchmarks could be derived
for chemicals with data sets that did not fulfill all the requirements of the National AWQC.
Because it is based on an SCV, the benchmark was categorized as interim.
Aquatic plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwn).
Adequate data sufficient for the development of benchmark values were not identified in
Suter and Mabrey (1994) or in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQJ method. The EQ method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
octanol-carbon partition coefficient (K^ to determine the protective sediment concentration
(Stephan, 1993). The EQ number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The SCV for butylbenzyl phthalate was used to calculate an EQp value of 349 mg
butylbenzyl phthalate/kg organic carbon. Assuming a mass fraction of organic carbon for the
sediment (f^.) of 0.05, the benchmark for the benthic community is 17.4 mg/kg sediment.
Since the EQp number was based on an SCV, the sediment benchmark was categorized as
interim.
August 1995
-------
APPENDIX B
Butylbenzyl phthalate - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
n*»ieeenuttv«
' QpesteC s
mink
river otter
bald eagle
osprey
great blue heron
mallard
lessor scaup
kingfisher
spotted sandpiper
herring gul
V**»*«0fte*
*y
1 65.87 (p')
92.35 (p')
10
ID
10
10
10
10
10
10
toutf
.. 30eol0e
rat
rat
-
8*«
rep. dvp
rep. dvp
-
-
-
Study ₯e*»
wo*9>4*
1909.81
1909.81
' '
i
QM^faltttl
, ^"^^^ifflJMWW*
LOAEL
LOAEL
-
.
-
-
* \
r-
10
10
-
-
-
-
s % >
ffitrtmrf ^ffwoif :
IARC, 1982
as cited in
HSDB, 1994
IARC, 1982
as cited in
HSDB, 1994
'
-
-
-
'Benchmark Category, a - adequate, p = provisional, i - interim; a "" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B
Butylbenzyl phthaiate - 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
xsr
fish and aquatic
invertebrates
aquatic plants
benthic community
8*ncfim«fc
Vt&Mf
16 (i)
No data
17 "'
«
aquatic
organisms
aquatic
organisms
*****
SCV
scv
-
'
AQUIRE. 1995
-
AQUIRE, 1995
IL
Benchmark Category, a = adequate, p = provisional, i - interim; a "' indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^ for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem. .
Mammals: Becasue of the lack of additional mammalian toxicity studies, the same surrogate-species
study (NTP, 1981) was used to derive the butylbenzyl lexicological benchmark for mammalian species
representing the general terrestrial ecosystem. The study value was scaled for species in the terrestrial
ecosystem using the cross-species scaling algorithm adapted from Opresko et al. (1994). Since the
NTP (1981) documented reproductive effects from butylbenzyl phthaiate exposure to male rats, male
body weights of the representative species were used in the scaling algorithm to obtain lexicological
benchmarks. Based on the data set for butylbenzyl phthaiate, the benchmarks developed for the
terrestrial ecosystem were categorized as provisional.
Birds: Adequate data with which to derive a benchmark protective of the avian community were not
identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root lengths. As presented in Will and Suter (1994), phytotoxicity benchmarks were selected
by rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or
fewer values for a chemical, the lowest LOEC was used. If there were more than 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or
other Effects reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for butylbenzyl
phthaiate and, as a result, a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Butylbenzyl phthalate 6
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
SpMfMI
do«r mouM
short-tailed
shrew
. meadow vote
Eastern
cottontail
red fox
raccoon
white-tailed daor
red- tailad hawk
American kestrel
Northern
bobowhita
American robin
American
woodcock
plants
soil fauna
8*80}WMKfC :
tthM*
: paa*"**
409.09 (p*)
420.62 (p*)
341. 77 (p')
1 44.40 (p*)
. 107.16 (p«)
103.13 (p*)
51.44 (p«)
ID
ID
ID
ID s
ID
No data
No data
aujov
fiCMSfi^MI
rat
rat
rat
rat
rat
rat
rat
-
BfMi
rap
rep
rap
rap
rep
rep
rep
-
-
-
-
' -
Nrtik
*»
1909.81
1909.81
1909.81
1909.81
1909.81
1909.81
1909.81
-
-
!- v * v ..
Pt-inifrittiii
-'-X "st\\* -"-'
s v--- , »
LOAEL .
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
-
-
"'*»
'\
10
10
10
10
10
10
10
-
-
-
-
- CMBfeftiSaiffM
'v ** ,,"* *"
IARC, 1982
(ARC, 1982
IARC. 1982
(ARC. 1982
IARC, 1982
IARC, 1982
IARC. 1982
-
-
-
-
-
'Benchmark Category, a » adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse affects.
ID - Insufficient Data
m. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAJFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and sources
are presented in Table 4 for ecological receptor categories: trophic level 3 and 4 fish in the limnetic
and littoral ecosystems, general fish (BCF only), aquatic invertebrates, earthworms, other soil
invertebrates, terrestrial invertebrates, and plants. Each value is identified as whole-boy or lipid-based
and, for the generic aquatic ecosystems, the biological uptake factors are designated with a "d" if the
August 1995
-------
APPENDIX B Butylbenzyl phthalate - 7
value reflects dissolved'water concentrations, and a "t" if the value reflects total surface water
concentrations. For organic chemicals with log Kow values below 4, bioconcentration factors (BCFs)
in fish were always assumed to refer to dissolved water concentrations (i.e., dissolved water
concentration equals total water concentration). The following discussion describes the rationale for
selecting the biological uptake factors and provides the context for interpreting the biological uptake
values presented in Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated using
Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for butyl benzyl
phthalate. The bioconcentration factor for fish was also estimated from the Thomann models
(i.e., log Kow - dissolved BCF/) and multiplied by the dissolved fraction (/"d) as defined in
Equation 6-21 to determine the total bioconcentration factor (BCF/). The dissolved
bioconcentration factor (BCF/1 ) was converted to the BCF/ in order to estimate the
acceptable lipid.tissue concentration (TC/) in fish consumed by piscivorous fish (see Equation
5-115). The BCF/ was required in Equation 5-115 because the surface water benchmark (i.e.,
FCV or SCV) represents a total water concentration (C1). Mathematically, conversion from
BCF/1 to BCF/ is accomplished using the relationship delineated in the Interim Report on
August 1995
-------
.APPENDIX B Butyibenzyl phthaJate - g
Data and Methods for Assessment of2J,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic
Wildlife (U.S. EPA, 19931):
BCF,d x fd = BCF/
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of several
values with sources in Table 4 (see master table). For earthworms and terrestrial
invertebrates, the bioconcentration factors were estimated as described in Section 5.3.5.2.3.
Briefly, the extrapolation method is applied to hydrophobia organic chemicals assuming that
the partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs
and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data (in this case chlordane) is compared to the BBF for TCDD
and that ratio (i.e., chlordane BBF/TCDD BBF) is multiplied by the TCDD standard for
terrestrial vertebrates, invertebrates, and earthworms, respectively. For hydrophobia organic
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf
translocation, direct deposition on leaves and grasses, and uptake into the plant through air
diffusion.
August 1995
-------
APPENDIX B
Butylbenzyl phthalate - 9
Table 4. Biological Uptake Properties
ootooiofi
i receptor
limnetic trophic
leveUfish
limnetic fropnic
level 3 fish
fish
littoral trophic
leveUfish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms .
plants
BCF.BAF.or
BSAF
BAF
\
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
lipkMMwad or
wnoMMMoy
lipid
lipid
lipid
lipid
lipid
lipid
whola-body
whole- body
whoia-body
whote- plant
witw .
28.432 ( d)
28.346 (d)
23.774 (t)
26.782 (d)
28,191 (d)
55,622(d)
3.2 E-04
3.0E-O4
2.4 E -03
1.1 E-01
pure*
pradiclad valua basad on
Tnomann, 1989, food chain
model
predicted value based on
Thomann, 1989, tood chain
model
predicted valua based on
Thomann, 1989 and adjusted to
estimate total BCF
predicted value based on
Thomann at al., 1992. food web
model
predicted value based on
Tnomann at al.. 1992. food web
model
predicted value based on
Thomann at al., 1992, food web
model
cafe
caic
caic
U.S. EPA. 1990*
d = refers to dissolved surface water concentration
t - refers to total surface water concentration
August 1995
-------
APPENDIX B Butylbenzyl phthalate - 10
References
Agarwal, D.K., R.R. Maronpot, J.C. Lamb IV, and W.M. Kluwe. 1985. Adverse effects of
butyl benzyl phthalate on the reproductive and hematopoietic systems of male rats.
Toxicology. 35: 189 -206.
Bower, R.K., Haberman, S. and Minton, P.D. (1970) Teratogenic effects in the chick
embryo caused by ester of phthalic acid. Pharmacol. exp. Ther., 171, 314-324. As cited
in I ARC Monographs on the Evaluation of the Carcinogenic Risk of Chemical to Humans
Some Industrial Chemicals and Dyestuffs. 1972-Prcsent V29 280 (1982).
Buccafusco, R. .!., S.J. Ells, and G. A. LeBlanc. 1981. Acute toxicity of priority pollutants
to bluegill (Lepomis macrochirus ). Bull Environ. Contam. Toxicol. 26(4):446-452. As
cited in AQUIRE (AQUatic Toxicity Information REtrieval Database). 1995.
Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN.
Gledhill, W. E., R. G. Kaley, W. J. Adams, O. Hicks, P. R. Michael, V.W. Saeger, and G.
A.LeBlanc. 1980. An environmental safety assessment of butyl benzyl phthalate.
Environ. Sci. Technol. 14 (3):301-305.
International Agency for Research on Cancer. I ARC Monographs on the Evaluation of the
Carcinogenic Risk of Chemical to Humans - Some Industrial Chemicals and Dyestuffs.
1972-Present V29 280 (1982).
Lake, B.C., R.A. Harris, P. Grasso and S.D. Gangolia. 1978. Studies on the metabolism
of biological effects of n-butyl benzyl phthalate in the rat. Prepared by British Industrial
Biological Research Association for Monsanto, Report No. 232/78, June. As cited in
U.S. EPA (Environmental Protection Agency). IRIS (Integrated Risk Information System).
March 1994. ^
Nagy, K. A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:11-128.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
NTP (National Toxicology Program). 1981. Carcinogenesis Bioassay of Di(2-
ethylhexyl)Adipate (CAS No. 103-23-1) (Technical Report Series No. 212) (DHHS
Publication No. (NIH) 81-1768), U.S. Department of Health and Human Services, Public
Health Service, National Institute of Health, Research Triangle Park, NC. As cited in
I ARC Monographs on the Evaluation of the Carcinogenic Risk of Chemical to Humans
Some Industrial Chemicals and Dyestuffs. 1972-Present V29 280 (1982).
August 1995
-------
APPENDIX B Butyibenzyl phthaJate - 11
NTP (National Toxicology Program). 1985. Twenty-six week subchronic study and modified
mating trial in F344 rats. Butyl benzyl phthalate. Final Report Project No. 12307-02, -
03. Hazelton Laboratories America, Inc. Unpublished study. As cited in U.S. EPA
(Environmental Protection Agency). IRIS (Integrated Risk Information System). March
1994.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratoy, Oak Ridge, Tennessee.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. .
Suter n, G. W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615-629.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D.C. January.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxiciiy - Benzyibui,. ^hihaiate Cas No.: 85-88-7
Chemical
Name
butylbenzyl
phthalate
butylbenzyl
phthalale '
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
Species
rat
rat
rats
rat
rat
rat
rat
rat
guinea pig
chicken
Endpolnt
liver
liver
re£
rep
rep
dev
dev
acute
behv. dvp
emb
Description
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
LD50
LD50
NOEL
Value
159
470
25000
0.001
0.002
160
480
2330
13,750
0.05
Units
mg/kg-day
mg/kg-day
ppjn
mg/kg-day
mg/kg-day
mg/kg-day
mg/k£-day
mg/kg-body
wl.
mg/kg-body
wl.
ml
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
gastric
intubation
gastric
intubation
oral
oral
injection
Exposure
Duration
/Timing
26 weeks
26 weeks
90 days
14 days
14 days
14 days
14 days
NS
NS
single dose
Reference
NTP, 1985 as cited in
IRIS. 1994
NTP, 1985 as cited in
IRIS, 1994
NTP, 1981
Agarwal et al., 1985
Agarwal etal., 1985
Lake etal., 1978 as
cited in IRIS, 1994
Lake etal., 1978 as
cited in IRIS, 1994
RTECS, 1994
RTECS. 1994
Bower, 1 970 as cited in
IARC, 1 982
Comments
(300gBWand 17g/day
consumption)
Increased liver/body wl. and
liver/brain wt.
Depressed body wt. gain and
testicular degeneration.
Reduction in total body, thymus,
testis, epididymis, prostate and
seminal vesicle weight.
No liver or testicular effects were
observed at this dose level.
Testicular atrophy was observad
No malformations when injected
into 32 fertilized hens' eggs.
-------
Freshwater Toxicity Benzylbutyl phthalates Cas No.: 85-68-7
Chemical
Name
butylbenzyl
phthalate
butylbenzyl
phthalate
butylbenzyl
phthalate
Species'
Daphnia
magna
Bluegill
fathead
minnow
NS = Not Specified
Type of
Effect
immob.
mort.
mort.
Description
EC50
LC50
LC50
Value
3700
43,000
2320
Units
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
Exposure
Duration
/Timing
48-hour
96-hour
96-hour
Reference
Gledhill et al.. 1980
Buccafusco et al., 1981 as cited
in AQUIRE, 1995
Gledhill et al., 1980
Comments
-------
Freshwater Bioiogicai uptake Measures, -tenzyibutyi phthaiates Cas No.: 85-68-7
Chemical Name
butylbenzyl
phthalate
butylbenzyl
phjhalate
butylbenzyl
phthalate
NS = Not specified
Species
fish
fish
fish
B-factor
(BCF. BAF,
BMP)
BCF
BCF
BCF
Value
270.50
138.1*
414'
* = BCF values may have come from a single source.
Measured
or
Predicted
(m.p)
P
m
m
Units
NS
NS
L/kg
Reference
Stephan, 1993
Stephan, 1993
U.S. EPA, 1992
Comments
Normalized to 1.0% lipids.
Normalized to 1 .0% lipids.
Normalized to 3% lipid.
-------
Terrestrial Biological Uptake Measures - Benzylbutyl phthalate Cas No.: 85-68-7
Chemical
Name
butylbenzyl
phthalate
Species
plant
B -factor
(BCF. BAF.
BMP)
BCF
Value
0.11
Measured
or
Predicted
(m.p)
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B DEHP - 1
lexicological Profile for Selected Ecological Receptors
Bis(2-ethylhexyl) phthalate (DEHP)
Cas No.: 117-81-7
Summary: This profile on bis(2-ethylhexyl) phthalate, or DEHP summarizes the
toxicological benchmarks and biological uptake measures (i.e., bioconcentration,
bioaccumulation, and biomagnification factors) for birds, mammals, daphnids and fish, aquatic
plants and benthic organisms representing the generic freshwater ecosystem and birds,
mammals, plants, and soil invertebrates in the generic terrestrial ecosystem. Toxicological
benchmarks for birds and mammals were derived for developmental, reproductive or other
effects reasonably assumed to impact population sustainability. Benchmarks for daphnids,
benthic organisms, and fish were generally adopted from existing regulatory benchmarks (i.e.,
Ambient Water Quality Criteria). Bioconcentration factors (BCFs), bioaccumulation factors
(BAFs) and, if available, biomagnification factors (BMFs) are also summarized for the
ecological receptors, although some BAFs for the freshwater ecosystem were calculated for
organic constituents with log K,,w between 4 and 6.5. For the terrestrial ecosystem, these
biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire toxicological data base compiled during this effort is presented at
the end of this profile. This profile represents the most current information and may differ
from the information presented in the technical support document for the "Hazardous Waste
Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (CL,,) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several studies investigating DEHP toxicity in" mammals were identified. Shiota
& Nishimura (1982) fed pregnant mice DEHP at concentrations of 0.05, 0.1, 0.2, 0.4 and
1.0% throughout gestation. Fetal mortality increased in a dose-related manner, and was
significantly higher in mice with diets of 1.0, 0.4 and 0.2% DEHP. The average daily dose of
DEHP administered was calculated from food intake and body weight At all doses except
0.05%, the percentage of resorptions and dead fetuses differed significantly from the control
group. Consequently, a NOAEL of 70 mg/kg-day and a LOAEL of 190 mg/kg-day (0.05%
and 0.1% DEHP respectively) was inferred for fetotoxic effects.
In a similar investigation, adult rats were injected with DEHP at concentrations of 0.1% and
0.2 % the acute LD50 value on the fifth, tenth and fifteenth day of gestation (Singh, 1972).
August 1995
-------
Terrestrial Biologica fake Measures - DDT
Gas No. 50-29-3
Chemical Name
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
Species
earthworm
earthworm
cattle
swine
cattle (beef)
cattle (milk)
sheep
poultry
small birds
rodents
cow
swine
plants
B-factor
(BCF, BAF,
BMF)
BCF
BCF
BCF
BCF
BTF
BTF
BAF
BAF
BAF
BAF
BAF
BAF
BCF
Value
0.7
0.1
0.9
0.4
0.0281
0.00239
0.59
12.3
0.04
2.45
1.12
0.74
0.01
Measured or
predicted
(m,D)
P
m
m
m
m
m
P
P
P
P
P
P
P
Units
NS
NS
NS
NS
NS
NS
kg fat/ kg
diet
kg fat/ kg .
diet
kg fat/ kg
diet
kg fat/ kg
diet
kg fat/ kg
diet
kg fat/ kg
diet
(ug/g DW
plant)/(ug/g
soil)
Reference
Beyer and Gish, 1980
Beyer and Gish, 1980
Clabom, el.al , 1960 as cited
in Kenaga, 1 980
Clabom, et.al., 1956 as cited
in Kenaga, 1980
Travis and Anns, 1988
Travis and Anns, 198B
Garten and Trabalka, 1983
Garten and Trabalka, 1983
Garten and Trabalka, 1983
Garten and Trabalka, 1 983
Garten and Trabalka, 1983
Garten and Trabalka, 1983
U.S. EPA. 1990e
Comments
BTF = Biotransfer factors. .
BTF = Biotransfer factors.
Percent lipid not specified.
Percent lipid not specified.
Percent lipid not specified.
Percent lipid not specified.
Percent lipid not specified.
Percent lipid not specified.
NS = Not specified
-------
APPENDIX B . ' DEHP - 2
With increasing concentrations of DEHP, the number of resorptions increased and average
fetal weight decreased as compared to the control group. Another mammalian study conducted
by Carpenter et al. (1953) found guinea pigs maintained on a diet of 19 mg DEHP/kg-day for
1 year exhibited increases in liver weight. Carpenter et al. (1953) also investigated the effects
of chronic DEHP exposure to rats. In a 2 year study, rats fed a diet with DEHP at a dose of
200 mg/kg-day had increases in liver and kidney weights. No effects were seen in rats
maintained on a 60 mg/kg-day diet.
Both studies by Carpenter et al. (1953) were not considered suitable for calculation of a
benchmark value because increases in liver and kidney weights may not impair the fecundity
of an entire population. The Singh (1972) and Shiota & Nishimura (1982) studies both
present fetotoxic effects exhibited by mammals when exposed to DEHP at a critical lifestage.
In each case, DEHP exposure resulted in effects which could impair the fecundity of a
wildlife population. However, the NOAEL value in Shiota & Nishimura (1982) was chosen to
derive the toxicological benchmark because (1) chronic exposures were administered via oral
ingestion, (2) the study contained a NOAEL with sufficient dose-response information, and
(3) studies providing a NOAEL are generally preferred to studies to providing a LOAEL.
While Singh's study (1972) reports reproductive toxicity as a critcal benchmark, it was not
selected because the extrapolation from the injection route of exposure to typical wildlife
exposure is unfounded. Furthermore, based on the number of doses, the Shiota & Nishimura
study (1982) provides clearer dose response data than the Singh (1972) study.
The study value from the Shiota & Nishimura (1982) study was then scaled for species
representative of a freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994):
( bw >/4
Benchmark = NOAEL, x L
VKJ ,
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and importable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Shiota & Nishimura study (1982) documented reproductive effects from DEHP exposure to
female mice, the representative body weights of the females species were used in the scaling
algorithm to obtain toxicological benchmarks.
Data were available on reproductive, developmental, growth and survival endpoints for DEHP
exposure. In addition, the data set contained studies which were conducted over acute and
chronic durations and during sensitive life stages. Therefore, based oh the data set for DEHP,
the benchmarks developed from the Shiota & Nishimura (1982) were categorized as
adequate.
August 1995
-------
APPENDIX B DEHP-3
Birds: Toxicity data were not identified involving DEHP toxic ity in avian species. Thus,
benchmarks for avian species could not be derived. .
Fish and aquatic invertebrates: A review of the literature revealed that an AWQG is not
available for DEHP. Therefore, the Tier II method described in Section 4.3.5 was used to
calculate a Secondary Chronic Value (SCV) of 5.5 mg/L. Tier n values or SCV were
developed so that aquatic benchmarks could be established for chemicals with data sets that
did not fulfill all the requirements of the National AWQC. Because the benchmark is based
on an SCV, this benchmark was categorized as interim.
Aquatic plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for species of
freshwater algae, frequently a species of green algae (e.g., Selenastrwn capricornutum).
Adequate data Sufficient for the development of benchmark values were not identified in
Suter and Mabrey (1994) or in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (ECO method. The EQ method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. Because no FCV was available, a Secondary Chronic Value (SCV) was calculated
as described in Section 4.3.5. The SCV reported for DEHP was used to calculate an EQp
number of 116,000 mg DEHP/kg organic carbon. Assuming a mass fration of organic carbon
for the sediment (f^) of 0.05, the benchmark for the benthic community is 5,810 mg DEHP/
kg of sediment Because the EQp number was set using a SCV derived using the Tier II
method, it was categorized as interim.
August 1995
-------
APPENDIX B
DEHP- 4
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R«y«nrUrtv»
mink
riv»r otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring guM
kingfisher
w^WvwWtWFH
v«iM»*ffl9*a-
-------
APPENDIX B
DEHP- 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and aquatic
invertebrates
aquatic plants
benlhic community
Benchmark
5.5 (i)
No data
5,800 (i)
aquatic
organisms
aquatic
organisms
scv
SCVXKoc
AQUIRE. 1995
AQUIRE, 1995
'Benchmark Category, a = adequate, p - provisional, i = interim; a "' indicate* that the benchmark value
was an order of magnitude or more above the NEL or LEL for other adverse effects.
n. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^ for the general terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem.
Mammals: Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Shiota & Nishimura, 1982) was used to derived the DEHP lexicological benchmark for
mammalian species representing the terrestrial ecosystem. The study value was scaled for species in
the terrestrial ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994).
Since the Shiota & Nishimura study (1982) documented reproductive effects from DEHP exposure to
female mice, the representative body weights of the female species were used in the scaling algorithm
to obtain lexicological benchmarks. Based on the data set for DEHP from Shiota & Nishimura
(1982), the benchmarks developed for the terrestrial ecosystem were categorized as adequate.
Birds: Adequate data with which to derive a benchmark protective of the avian community were not
identified. ,
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root lengths. As presented in Will and Suter (1994), phytotoxicity benchmarks were selected
by rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or
fewer values for a chemical, the lowest LOEC was used. If there were more than 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or
other effects reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for DEHP and, as a
result, a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
DEHP- 6
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Bsy»«ntrtv»
Specie*
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red- tailed hawk
American Kestrel
Northern
bobowhite
American robin
American
woodcock
plants
soil community
8*eehraaift
VahM*
«9ft8-*y
79.74 (a)
8 1.99 (a)
66.62(a)
28.15 (a)
20.89 (a)
20.10 (a)
10.03 (a)
ID
ID
ID
ID
ID i
ID
ID
Study
mice
mice
mice
mice
mice
mice
mice
-
-
Etftt*
rep
rep
rep
rep
rep
rep
rep
-
-
'-
*«*
«*tt
«8*r
4y
70
70
70
70
70
70
70
-
-
ttaMdptfeft
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
'
-
-
-
.
.
-
ar
-
-
-
-
-
-
Shiota &
Nishimura, 1982
Shiota&
Nishimura, 1982
Shiotai
Nishimura. 1982
Shiotai
Nishimura. 1982
Shiota &
Nishimura. 1982
Shiota and
Nishimura. 1982
Shiota &
Nishimura, 1982
-
'
'Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID - Insufficient Data
m. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
August 1995
-------
APPENDIX B DEHP- 7
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic chemicals
with log K,,w values below 4, bioconcnetration facctors (BCFs) in fish were always assumed to refer to
dissolved water concentrations (i.e., dissolved water concentration equals total water concentration).
The following discussion describes the rationale for selecting the biological uptake factors and
provides the context for interpreting the biological uptake values presented in Table 4.
Because the log Kow for DEHP is above 6.5 (i.e., 7.5), the Thomann (1989) and Thomann etal.,
(1992) models were not used to estimate bioaccumulation factors. For extremely hydrophobic
constituents, the Agency has stated that reliable measurements of ambient water concentrations
(especially dissolved concentrations) are not available and that accumulation of these constituents in
fish or other aquatic organisms cannot be referenced to a water concentration as required for a BCF or
BAF (U.S. EPA, 1993i). Since no measured BAF was available, a measured BCF identified in
Stephan (1993) was used as a BAF since DEHP, like other phthalates, is capable of being metabolized
by aquatic organisms
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates and earthworms
were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation method is applied to
hydrophobic organic chemicals assuming that the partitioning to tissue is dominated by lipids. For .
hydrophobic organic constituents, the bioconcentration factor for plants was estimated as described in
Section 6.6.1 for above ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf
translocation, direct deposition on leaves and grasses, and uptake into the plant though air diffusion.
August 1995
-------
APPENDIX B
DEHP- 8
Table 4. Biological Uptake Properties
r*»pt>r
limnetic trophic
level 4 fiih
limnetic trophic
level 3 fish
fish
littoral, tropiuc
level 4 fiih
lioonl trophk
level 3 fiih
lioonl trophic
level 2
invertebrate*
terrestrial
vertebrates
terrestrial
inveitehnles
earthworms
plants
8CFrRAr,«r
BSAF
BAF
BAF
BCF
BAF
BAF
-
BAF
BCF
BCF
BCF
tfptebMtfw
wfcgtebady
lipid
lipid
lipid
lipid.
lipid
whole-body
whole -body
whole -body
whole -plant
, **»
2.400 ( 1)
2.400 (t)
2.400 (1)
2.400 (t)
2.400 (t)
10
3JE-01
3.3 E-01
2.7
1.9 E -03
""" '
no meuured BAF; bued on
measured BCF (Stephan.1993)
no meuured BAF: bued on
meuured BCF (Slephin.1993)
no measured BAP, based on
measured BCF (Stephan.1993)
no measured BAF; based on
measured BCF (Stephan.1993)
no measured BAP, based on
measured BCF (S(ephan.l993)
.
calc
calc
calc
U.S. EPA. 1990e
d = refers to dissolved surface water concentration
t =refen to total surface water concentration
ID = Insufficient Data
August 1995
-------
APPENDIX B DEHP - 9
References
Adams, W. J. and B. B. Heidolph. 1985. "Short-cut chronic toxicity estimates using
Daphnia magna." IN R.D. Cardwell, R. Purdy, and R. C Bahner (eds.), Aquatic
Toxicity and Hazard Assessment, Seventh Symposium. ASTM, Philadelphia, PA. As
cited in Suter, G.W. n and J.B. Mabrey, 1994. Toxicological benchmarks for screening
potential contaminants of concern for effects on aquatic biota: 1$94 revision. DE/AC05-
84OR21400 Office of Environmental Restroation and Waste Management, U.S.
Department of Energy, Washington, D.C.
Agarwal, D. K., W. H. Lawrence, and J. Autian. 1985. Antifertilityand mutagenic effects in
mice from parenteral administration of di-2-ethylhexyl phthalate (DEPH). Journal of
Toxicology and Environmental Health. 16:71-84.
AQUIRE (AQUatic Toxicity Information REtrieyal Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN. June 1995.
Birge, W. J., J. A. Black, and A. G. Westerman. 1978. Effects of polychlorinated
biphenyl compounds and proposed PCB - replacement products on embryo-larval stages of
fish and amphibians. Res. Rep. No 118. University of Kentucky, Water Resour. Res.
Inst., Lexington, KY. 33pp. (U.S. NTIS PB - 290711).
Carpenter, G. P., C. S. Well, and H. F. Smyth, Jr. 1953. Chronic oral toxicity of di (2-
ethylhexyl) phthalate for rats, guinea pigs and dogs. /. Ind. Hyg. Occup. Med. 219-
226. pp
IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Humans -
Some Industrial Chemicals and Dyestuffs. 1972-Present V29 280 (1982). As cited in
National Library of Medicine. HSDB (Hazardous Substance Database). 1994.
Nakamura, Y., Y. Yagi, I. Tomita, and K. Tsuchikawa. 1979. Teratogenicity of di-(2-
ethylhexyl) phthalate in mice. Toxicology Letters. 4:113 -117.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. 1994.
Opresko, D. M., B.E. Sample, G.W. Suter EL 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. DE-AC05-84OR21400. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
August 1995
-------
APPENDIX B DEHP - 10
Passino, D. R. M. and S. B. Smith. 1987. Acute bioassays and hazard evaluation of
representative contaminants detected in Great Lakes fish. Environ. Toxicol. Chem.
6(11):901-907. As cited in AQUIRE (AOUatic Toxicity Information REtrieval
Database), Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency, Duluth, MN. June 1995.
Peters, J.W. and R.M. Cook. 1973. Effect of phthalate esters on reproduction in rats.
Environmental Health Perspectives 3(91):91-94.
Seth, P.K., and S.P. Srivastava, D.K. Agarwal, and S.V. Chandra. 1975. Effect of di-2-
ethylhexyl on rat gonads. Environmental Research. 12:131-138.
Shiota, K., MJ. Chou and H. Nishimura. 1980. Embryotoxic effects of di-2-ethylhexyl
phthalate (DEPH) and di-n-butyl phthalate (DBP) in mice. Environmental Perspectives.
22:245-253.
Shiota, K., and H. Nishimura. 1982. Teratogenicity of di(2-ethylhexyl)phthalate (DEHP) and
di-n-butyl phthalate (DBP) in mice. Environ. Health Perspectives 45:65- 72.
Singh, A. R., W. H. Lawrence, and J. Autian. 1972. Teratogenicity of phthalate esters in
rats. Journal of Pharmaceutical Sciences 61(l):51-55.
Suter, G.W. II and J.B. Mabrey, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restroation and Waste Management, U.S.
Department of Energy, Washington, D.C.
Stephan, C.E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. PB93-154672.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment. Washington, D.C. Janauary.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division - Region IV.
U.S. EPA (Environmental Protection Agency). 1994. Ambient Aquatic Life Water Quality
Criteria for Di-2-Ethylhexyl Phthalate. Health and Ecological Criteria Division, Office
of Science and Technology, Office of Water, Washington, D.C.
August 1995
-------
APPENDIX B DEHP - 11
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxlcity - DEHP Cos No.: 117-81-7
bis(2-ethylhexy1)
phttialate
bis(2-ethylhexyl)
phttialate
di (2-ethylhexyl)
phthalat0
bis(2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalate
di (2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalate
i
mouse
i
i
mouse
guinea pigs
rat
rat
rat
mouse
fet
fet
iver
liver
liver
rep
(eto
NOAEL
LOAEL
LOAEL
LOAEL
NOAEL
PEL
NOAEL .
!
t
70
190
19
195
60
5
17.86
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
ml/kg
mg/kg-day
oral
oral
oral
oral
oral
L?
oral
hroughout
gestation
hroughout
gestation
1 year
2 years
2 years
3 injections on
days 1,5. and 10
Day 7 of gestation
Shiota & Nlshimura,
1982
Shiota & Nishimura,
1982
Carpenter et al.. 1953
Carpenter et al., 1953
Carpenter et al., 1953
Sethetal., 1976
Nakamura et al., 1979
No embryotoxic effects of
DEHP were observed at this
dose level.
Fetal resorptions and fetal
malformations including
intrauterine growth retardation
and delayed ossification.
Increased relative liver weights.
Increased liver apd kidney
weights and retarded growth
were observed at this dose
level.
/
No effects were observed at
this dose level.
Activity of succinic
dehydrogenase and adenosine
triphosphatase were
significantly reduced, which
could account for degeneration
of sperm-producing structures.
No significant differences were
observed between this dose
level and untreated controls.
-------
Terrestrial Toxlcity - L..HP Cas No.:l 17-81-7
Chemical Name
bis(2-ethylhexyl)
phthalale '
bis(2-ethylhexy))
phthalale
bis(2-ethylhexyl)
phthalale
bis(2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalale
bis(2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalate
Species
rat
mouse
rabbit
rats
mouse
rat
rat
rat
Endpolnt
acute
acute
acute
rep
rep
rep. let
systemic
systemic
Description
LD50
LD50
LD50
LOAEL
LOAEL
PEL
NOAEL
LOAEL
Value
30,600
30 .
34
2
1
5
7500
15,000
Units
mg/kg-body wt.
g/kg-body wt.
g/kg-body wt.
ml/kg BW
ml/kg
g/kg-body wt.
PEP! ..
ppm
Exposure
Rout* (oral,
S.C., I.V.. l.p.,
Injection)
oral
oral
oral
i.p.
s.c. injection
i.p. injection
oral
oral
Exposure
Duration /Timing
NS
NS
NS
3,6, and 9 days of
gestation
days 1, 5, 10 prior
to mating
gestation days 5,
10&15
90 days
90 days
Reference
RTECS, 1994
RTECS. 1994
RTECS, 1994
Peters and Cook, 1 973
Agarwal et al., 1985
I ARC , 1982 as cited in
HSDB. 1994.
Shaffer et al., 1945 as
cited in IARC, 1982
Shaffer etal, 1945 as
cited in IARC, 1982
Comments
.
Implantation and parturition are
affected at this dose level.
Preimplantation losses and
early fetal deaths were
significantly increased at all
dose levels (1, 2, 5, and 10
ml/kg)
Resorptions, gross
abnormalities, fetal death or
decrease in fetal size.
No effects were observed at
this dose level.
Body weight gain was
observed at this dose level.
-------
Freshwater Toxicity - DEHP
CasNo.: 117-81-7
Chemical
Name
bis(2-
ethylhexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
bis(2-
ethythexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
bis(2-
ethylhexyl)
phthalate
NS = Not Sf
Species .
aquatic
organisms
fish
daphnid
lish
daphnid
Daphnia
magna
large mouth
bass
rainbow trout
jecified
Type of
Effect
chronic
chronic
chronic
chronic
chronic
immob.
acuia
acute
Description
scv
cv
cv
EC20"
EC20
EC50
LC50
LC50
Value
32.2
8.4
<3
>54
<3
133
32900
139.500-
149,200
(142,889)
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
.NA-
NS
NS
NS
NS
48-hour
NS
NS
Reference
Adams and Heidolph,
1 985 as cited in Suter
and Mabrey , 1 994
Suter and Mabrey,
1994
Suter and Mabrey,
1994
Suter and Mabrey,
1994
Suter and Mabrey,
1994
Passino and Smith,
1987 as cited in
AQUIRE, 1995
Birgeetal.. 1978
Birgeetal., 1978
Comments
.
-------
Terrestrial Toxicity -. .HP Cas No.: 117-81-7
bis(2-ethylhexyl)
phthalate
bis(2-ethylhexyl)
phthalate
mouse
rats
NS = Not Specified
feto
ter
LOAEL
LOAEL
35.71
10 j
-
n^a^ay
ml/kg
oral
i.p.
Day 7 of gestation
Day 5. 10, 15 of
gestation
Nakamura el al.. 1979
Singh etal., 1972
1 1 .2% fetal mortality was
observed at this dose level.
However, gross and skeletal
abnormalities including
elongated and fused ribs, etc.,
occurred only at 1 .0 ml/kg^
Resorptions and malformed
fetuses were observed at this
dose level, the highest of two
dose levels.
-------
Terrestrial Bioiogicai Uptake Measures - DEKP
Cos No.: 117-8!-7
Chemical
Name
bis(2-
ethylhexyl)
phthalate
Species
plant
B-factor
(BCF. BAF,
BMP)
BCF
Value
0.0051
Measured
or
Predicted
(m,p)
P
Units
(ug/g DW
plant)/(ug/g
soil)
Reference
U.S. EPA,
1990e
Comments
-------
Freshwater Biological Uptake K._ jsures - DEHP Cas No.:l 17-81 -7
Chemical
Name
bis(2-
ethylhexyl)
phthalate
Species
fish
B -(actor
(BCF. BAF.
BMP)
BCF
Value
130.00
Measured
or
Predicted
(m,p)
m
Units
Ukg
Reference
U.S. EPA,
1992
Comments
Normalized
to 3% lipid.
-------
APPENDIX B Cadmium - 1
Toxicological Profile for Selected Ecological Receptors
Cadmium
Cas No.: 7440-43-9
Summary: This profile on cadmium summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the
freshwater ecosystem were calculated for organic constituents with log Kow between 4 and
6.5. For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire toxicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from the data presented in the technical support document
for the Hazardous Waste Identification Rule (HW1R): Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
i
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
i
Mammals: Numerous studies were identified on the effects of cadmium toxicity to
mammalian species. In a study by Loeser and Lorke (1977), dogs given food containing
cadmium at doses of 0.02, 0.06, 0.2, and 0.6 mg/kg-day for a 3 month period exhibited no
behavioral or developmental effects. From this study, a NOEL of 0.6 mg/kg-day was inferred
for dog exposure to cadmium. Sorell and Graziano (1990) exposed female rats to cadmium
via drinking water at doses of 5, 50, and 100 ppm on gestation days 6 through 20. Growth
retardation, as expressed in decreased fetal and maternal weights, was noted at the two higher
doses. Based on the recommended body weight of 0.35 kg and water consumption of 0.046
I/day for Sprague-Dawley rats (U.S EPA, 1988), a NOAEL of 0.66 mg/kg-day and a LOAEL
of 6.6 mg/kg-day were calculated for developmental effects. Sutou et al. (1980) assessed
cadmium toxicity in rats exposed to 0.1, 1.0 and 10 mg/kg-day over a period of six weeks,
August 1995
-------
APPENDIX B Cadmium - 2
including a three-week mating period and up to day 20 of gestation. No effects were seen in
the groups exposed to 0.1 and 1.0 mg/kg-day, however, at 10 mg/kg-day, the number of
embryonic implantations and live fetuses decreased significantly. In addition, surviving
fetuses from the 10 mg/kg-day treatment group exhibited decreases in body weight, body
length and tail length as well as delayed ossification of the vertebrae. These results suggest a
NOAEL of 1.0 mg/kg-day and a LOAEL of 10 mg/kg-day for developmental effects.
Although Sutou et al (1980) and Sorell and Graziano (1990) reported similar NOAELs, the
NOAEL of 1.0 mg/kg-d from the Sutou et al. (1980) study was chosen to derive the
mammalian toxicological benchmarks because it contained sufficient dose-response
information and focused on developmental endpoints at a critical lifestage. In terms of
population sustainability, the decreased fetal body weight observed by Sorcll and Graziano et
al. (1990) was not as significant as the decreased embryonic implantations and live fetuses
reported by Sutou et al. (1980). Although dogs are members of the same taxonomic Order
(Camivora) as the representative species, the Loeser and Lorke (1977) study does not provide
clear dose-response information. While the studies by Sorell and Graziano et al. (1990) and
Loeser and Lorke (1977) were not chosen for the development of a toxicological benchmark,
they do illustrate the dose range at which cadmium toxicity occurs.
The study value from the Sutou et al.. (1980) was scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Oprcsko et al.
(1994)
Benchmark = NOAEL. x
bw >/4
w,
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Sotou et al. (1980) study documented developmental effects from cadmium exposure to
mating male and female rats, the mean body weight for both genders for each representative
species was used in the scaling algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive, developmental, and growth effects of cadmium. In
addition, the data set contained studies which were conducted over chronic and subchronic
durations and during sensitive life stages. The data set does not support an uncertainty factor
to account for inter-species differences in toxicological sensitivity. The study value selected
from the Sotou et al. (1980) was a NOAEL based on a developmental endpoint that was
within an order of magnitude of the lowest identified NEL or LEL. Based on the data set for
cadmium, the benchmarks developed from the Sotou et al. (1980) study were categorized as
adequate.
August 1995
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APPENDIX B Cadmium - 3
Birds: Three studies were identified that investigated cadmium toxicity in avian species. The
effects on avoidance response to fright stimuli were assessed in one-week-old black ducks fed
4 or 40 ppm cadmium (Heinz et al., 1983). "No information on daily food consumption rates
were provided therefore, the use of an allometric equation was required to convert the doses
from dietary ppm to mg/kg-day:
Food consumption = 0.0582(W°'651) where W is body weight in kg (Nagy, 1987).
Assuming a body weight of 0.053 kg, doses for this study were calculated as 0.1 and 1
mg/kg-day. Ducklings fed 0.1 mg/kg-day ran longer distances away from a fright stimulus
than the control group or the 1 mg/kg-day ppm group. The authors could not explain why
effects were seen at the lower dose level and not at 1 mg/kg-day.
Richardson et al. (1974) investigated the effects of cadmium on Japanese quail given an oral
dose of approximately 75 mg/kg-diet from hatching until 4 or 6 weeks of age. Since daily
food consumption was not provided the allometric equation presented above was used to
convert the cadmium dose to mg/kg-day. Using a body weight of 0.08 kg, the dietary dose
was estimated at 10.5 mg/kg-day. After 4 weeks of exposure, quail exhibited signs of
testicular hypoplasia, growth retardation and severe anemia and after 6 weeks of exposure,
both heart ventricles were hypertrophied. In another study, dietary cadmium was given to
mallard duck hens at 0.19, 1.9, and 19 mg/kg-day for up to 90 days (White & Finley, 1978).
No effects in egg laying were seen at the lower dose levels, however, egg production was
suppressed in the group given 19 mg/kg-day. Based on these results a LOAEL of 19 mg/kg-
day and a NOAEL of 1.9 mg/kg-day can be inferred for reproductive effects.
All of these investigations indicate effects that could impair the survival of a wildlife
population. However, the study by Richardson et al. (1974) was not considered suitable for
derivation of a benchmark value because of insufficient dose response information. Since
behavioral effects were observed at the lower dose and not at the higher dose, the Heinz et al.
.(1983) study also did not establish a clear dose response relationship. Therefore, the White
and Finley (1978) NOAEL of 1.9 mg/kg-day was chosen for estimation of an avian
benchmark value. '
The principles for allometric scaling were, assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian
species representative of a freshwater ecosystem, the NOAEL of 2.0 mg/kg-day from the
White and Finley (1978) study was scaled using the cross-species scaling method of Opresko
et al. (1994).
Data were available on the reproductive and developmental effects of cadmium, as well as on
behavioral effects potentially effecting survival. Laboratory experiments of similar types were
not conducted on a range of avian species and as such, inter-species differences among
wildlife species were not identifiable. There were no other values in the data set which were
lower than the benchmark value. Based on the avian data set for cadmium, the benchmarks
developed from the White and Finley (1978) study were categorized as adequate.
August 1995
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APPENDIX B Cadmium - 4
Fish and Aquatic Invertebrates: The Final Chronic Value (FCV) for cadmium of 1.1 E-3
mg/1 was selected as the benchmark protective of fish and aquatic invertebrates (U.S. EPA,
1986). The FCV for cadmium is a function of water hardness and is calculated using the
equation e(i-28[Nhardnessj].3.828) (U s EpA> 1986)i j^^ng a water hardness of 100 mg/i.
Since the benchmark is based on the FCV developed for the AWQC and was within an order
of magnitude of the lowest adverse effect levels for daphnids, this benchmark was categorized
as adequate.
\
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEQ or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (EC,^) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
The aquatic plant benchmark for copper is 2E-03 mg/1 based on reduced population growth
rate of Asterionella formosa (Conway, 1977 as cited in Suter & Mabrey, 1994). As
described in Section 4.3.6, all benchmarks for aquatic plants were designated as interim.
Benthic community: The cadmium benchmark protective of benthic organisms is pending a
U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Cadmium -5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with a Freshwater Ecosystem
Representative
vK:f8pecte« >:'::
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
"Benchmark
Value mg/*g-d
0.82 (a)
0.49 (a)
1.4 (a)
1.7 (a)
1.6 (a)
1.9 (a)
2.1 (a)
4.3 (a)
1.9 (a)
3.2 (a)
'';$ Study :!ipi;
I ."i-SpecietK-.
rat
rat
mallard duck
matardduck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
Effect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study Value
mg/kg>d
1.0
1.0
1.9
1.9 '
1.9
1.9
1.9
, 1.9
1.9
1.9
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
-
-
'
Origin*! Soon*
Sutou et a)., 1980
Sutouetal., 1980
While et a!.. 1978
White etal., 1978
Whitoatal., 1978
White et a!.. 1978
While et a!., 1978
Whin et al.. 1978
White et al., 1978
Whttaetal.. 1978
Benchmark Category, a > adequate, p « provisional, i « interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL (or other adverse effects.
August 1995
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APPENDIX B
Cadmium -
Table 2. Toxicological Benchmarks for Representative Fish
Associated with a Freshwater Ecosystem
RBpr**anlattv«
Specie* .
fish and aquatic
invertebrates
aquatic plants
benihic
community
Benchmark
Vafu«
tngfL
1.lE-03(a)
2.0E-03 (i)
under review
Study
Specie*
aquatic
organisms
aquatic
plants
.
Description
FCV
CV
Origin*
Sourc*
U.S. EPA. 1986
Conway, 1977 as
cited in Suter &
. Mabrey, 1994
.
IL
'Benchmark Category, a * adequate, p > provisional, i - interim: a "" indicates that the benchmark.value we* an order
of magnitude or more above the NEL or LEL for other adverse effects. -
Toxicological Benchmarks for Representative Species in the Generic TerrestriaJ
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C,) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to cadmium.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Sutou et al., 1980) was used to derive the cadmium toxicological benchmark for
mammalian species representing the terrestrial ecosystem. The study value from the Sutou et
al. (1980) study was scaled for species representative of a terrestrial ecosystem using a cross-
species scaling algorithm adapted from Opresko et al. (1994). Since the Sutou et al. (1980)
study documented reproductive effects from cadmium exposure to mating male and female
rats, the mean body weight for both genders for each representative species was used in the
scaling algorithm to obtain the toxicological benchmarks. Based on the data set for cadmium,
the benchmarks developed from the Sutou et al. (1980) study were categorized as adequate.
Birds: Additional avian toxicity data were not identified for birds representing the terrestrial
ecosystem therefore, the White and Finley (1978) study on reproductive effects in mallards
used in the freshwater ecosystem was also used to calculate a benchmark value. The NOAEL
of 1.9 mg/kg-day from White and Finley (1978) was scaled for species representative of a
terrestrial ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994). Based on the avian data set for toxaphene, the benchmarks developed from the White
and Finley (1978) study were categorized as adequate.
August 1995
-------
APPENDIX B Cadmium.?
Plants; Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation. The
selected benchmark for phytotoxic effects of cadmium in soils is 3 mg/kg (Will & Suter,
1994). Since the study value selected is the 10th percentile of more than 10 LOEC values,
the terrestrial plant benchmark for cadmium is categorized as provisional.
Soil Community: For the soil community, the toxicological benchmarks were established
based on methods developed by the Dutch National Institute of Public Health and
Environmental Protection (RIVM). In brief, the RIVM approach estimates a concentration at
which the no observed effect concentration (NOEC) for 5 percent of the species within the
community is not exceeded. A minimum data set was established in which key structural and
functional components of the soil community (e.g., decomposer guilds, grazing guilds)
encompassing different sizes of organisms (e.g., microfauna, mesofauna, and macrofauna)
were represented. Measurement endpoints included reproductive effects as well as measures
of mortality, growth, and survival. The derived cadmium benchmark deemed protective of the
soil community is 0.685 mg/kg. Since the cadmium data set contains NOECs and/or LOECs
for at least five of the representative species outlined in the minimum soil data set, the soil
community benchmark is categorized as provisional.
August 1995
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APPENDIX B
Cadmium 8
Table 3. ToxicologicaJ Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Representative
Specie*
doer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed
deer
red- tailed hawk
American
kestrel
Northern
bob white
American robin
American
woodcock
plants
soil community
'Benchmark
Value maftfl-d
2.2 (a)
2.3 (a)
1.9 (a)
0.78 (a)
0.56 (a)
0.53 (a)
. 0.27 (a)
1.9 (a)
3.4 (a)
3.1 (a)
3.7 (a)
3.1. (a)
3 (p) mg/kg
0.685 (p) mg/kg
Study
Sp«oiM
rat
rat
rat
rat
rat
rat
rat
mallard duck
mallard duck
malard duck
mallard duck
maflard duck
terrestrial
plant
soil
invertebrates
Effect
dev
dev
dev
dev
dev
dev
dev
rep
rep
rep
rep
rep
growth/yeild
chronic
Study
Value
mfl/kfl-d
1.0
1.0
1.0
1.0
. 1.0
1.0
1.0
1.9
1.9
1.9
1.9
1.9
3 mg/kg
0.685
mg/kg
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
10th percentile
LOEC
NOEC
SF
-
-
-
-
-
-
-
,-
-
OrfeJn* Souro.
Sutou et al.. 1980
Sutou at al.. 1980
Sutou etal.. 1980
Sutou el al.. 1980
Sutou etal.. 1980
Sutou el al.. 1980
Sutou et al., 1980
While etal.. 1978
Whteetal.. 1978
Write etal.. 1978
White et al.. 1978
White etal.. 1978
WMIandSuter,
1994
Aldenberg and
Slob. 1993
Benchmark Category, a >
magnitude or more above
i adequate, p « provisional, i » interim; a "' indicates that the benchmark value was an order of
the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Cadmium. 9
in. Biological Uptake Measures
This section presents biological uptake measures (i.e. BCFs, BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconccntrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for. selected ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values presented
in Table 4.
The whole-body BCF for cadmium was the geometric mean of 15 measured values, with most
values supplied from two studies by Kumada et al. (1973, 1980). The values ranged from 20
to 12,000 and the mosquito fish appeared to be somewhat of an outlier species relative to the
other measured values. BCF values for muscle were not included because ecological
receptors are likely to eat the whole fish or, in the least, will not necessarily distinguish
between the fillet and other parts of the fish. Data on bioconcentration in aquatic
invertebrates are under review. Studies on bioaccumulation/bioconcentration in terrestrial
vertebrates and invertebrates have been identified and are currently being reviewed.. For
metals, empirical data were used to derive the BCF for aboveground forage grasses and leafy
vegetables. In particular, the uptake-response slope for forage grasses was used as the BCF
for plants in the terrestrial ecosystem since most of the representative plant-eating species
feed on wild grasses.
August 1995
-------
APPENDIX B
Cadmium - 10
Table 4. Biological Uptake Properties
co logical
receptor
fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
8CF.BAF.or
BSAF :
BCF
BCF
BAF
BCF
BCF
BCF
lipid-tated or
whole-body
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
187(1)
-
3.5
0.14
. ' eourc* ; /. ..^^- >;
geometric mean of 15 measured
values lor whole-body BCFs as
cited in master table (e.g.;
Kumada et al.. 1973)
data under review
data under review
data under review
geometric mean of measured
values from Da vies, 1983;
Helmke. 1979
U.S. EPA, 1992e
d « refers to dissolved surface water concentration
t > refers to total surface water concentration
August 1995
-------
APPENDIX B Cadmium - 11
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Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Suter n, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Sutou, S., K. Yamamoto, H. Sendota, K. Tomomatsu, Y. Shimizu, and M. Sugiyama. 1980.
Toxicity, fertility, teratogenicity, and dominant lethal tests in rats administered cadmium
subchronically: I. Toxicity Studies. Toxicology and Environmental Safety 4:39-50.
Sutou, S., K. Yamamoto, H. Sendota, and M. Sugiyama. 1980. Toxicity, fertility,
teratogenicity, and dominant lethal tests in rats administered cadmium subchronically: n.
Fertility, teratogenicity, and dominant lethal tests. Toxicology and Environmental Safety
4:51-56.
Taylor, D. 1983. The significance of the accumulation of cadmium by aquatic organisms.
Ecotoxicology and Environmental Safety 7:33-42.
August 1995
-------
APPENDIX B Cadmium - 15
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
U.S. EPA (Environmental Protection Agency). 1986. Ambient Water Quality Criteria for
Cadmium. U.S. Environmental Protection Agency, Washington, DC. Publication No.
EPA-440/5-86-001.
U.S.EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
.Documentation of Biological Values for Use in Risk Assessment. EP A/600/6- 87/008.
Environmental Criteria and Assessment Office, Office of Health and Environmental
Assessment, Office of Research and Development, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1990. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, D.C.
Webb (ed.). 1979. The Chemistry, Biochemistry, and Biology of Cadmium. Elsevier/North-
Holland Biomedical Press, pp. 340-413.
White, D. H., and M. T. Finley. 1978. Uptake and retention of dietary cadmium in mallard
ducks. Environmental Research 17:53-59.
Will, M.E. and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Williams, D. R., and J. P. Gisey, Jr, 1978. Relative importance of food and water resources
to cadmium uptake by Gambusia afflnis (Poeciliidae). Environmental Research 16:326-
332.
August 1995
-------
Freshwater R ry - Cadmium
Cas No. 7440-43-9
Chemical
Name
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
Species
Daphnia
magna
rainbow
trout
rainbow
trout
rainbow
trout
striped bass
aquatic
organisms
fish
daphnid
fish
daphnid
atlantic -
salmon
atlantic
salmon
NS = Not specified
Endpolnt
immob.
mort.
mort.
mort.
mort.
chronic
chronic
chronic
chronic
chronic
dvp
dvp
,
Description
EC50
LC50
LOEC
NOEC
LC50
AWQC
CV
CV
EC20
EC20
LOEC
NOEC .
Value
24 4 - .
355.3
(129.6)
2.10-7.71
(4.56)
1.74-5.16
(3.56)
1.25-2.57
1100
1.1
i.7
0.15
1.8
0.75
2
0.2
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
48-hour
96 hour
100 days
100 days
96-hour
NS
NS
NS
NS
NS
NS
NS
Reference
Stuhlbacher etal., 1993
as cited in AQUIRE, 1995
AQUIRE. 1995
Daviesetal., 1993 as
cited in AQUIRE, 1995
Daviesetal.. 1993 as
cited in AQUIRE, 1995
Rehwoldt et al., 1972 as
cited in AQUIRE, 1995
U.S. EPA, 1986
Suter and Mabrey, 1994
Suter and Mabrey, 1 994
Suter and Mabrey, 1994
Suter and Mabrey, 1 994
Peterson etal., 1983
Peterson etal., 1983
Comments
-------
Freshwater Biological Uptake Measures - Cadmium
Cas No. 7440-43-9
Chemical
Name
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
Species
daphnids
fish
fish
rainbow trout
rainbow trout
mosquito fish
mosquito fish
NS = Not specified
B-factor
(BCF. BAF,
BMF)
BAF
BCF
BCF
BCF
BCF
BCF
BCF
Value
0.45
<20
64
49,16, 19,
19
12,000, 500,
380, 245,
200
4,100.00
7,440.00
Measured
or
Predicted
(m,p)
m
m
m
m
m
m
m
Units
NS
LAg
L/kg
L/kg whole- .
body (wet
weight)
L/kg whole-
body (wet
weight)
NS
NS
.
Reference
Ferard et al., 1983
Taylor, 1 983
U:S. EPA, 1992
Kumada et al., 1980
Kumada et al., 1973
Williams & Giesy, 1978
Giesy et al., 1977 as cited
in Eisler, 1985
Comments
BAF is for uptake from food only.
Not clear how BCF was derived.
Whole body BCF; 20 weeks of
exposure to 1 ppb or 4 ppb Cd stearate,
or 4 ppb of Cd acetate or Cd chloride.
Whole body BCF; 30 weeks of
exposure to 0.00001 , 0.0001 , 0.001 ,
0.0022, 0.0048 ppm "cadmium solution'
Whole body BCF; 8 weeks of exposure
to 0.02 ppb Cd.
Whole body BCF; 26 weeks of
exposure to 5.0 ppb Cd.
-------
APPENDIX B Cadmium - 16
Zhuang, Y. and H.E.-Allen. 1994. Cadmium Mobilization Resulting from Sediment
Aeration. "Preprint Extended Abstract", Presented Before the Division of Environmental
Chemistry, American Chemical Society, pp. 110-112.
August 1995
-------
Terrestrial Toxicity - Cadmium
Cas No. 7440-43-9
Chemical
Name
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
cadmium
NS = Not spe
Species
mouse
rat
rat
rat
rat
rat
mouse
dog
mallard hen
mallard hen
Japanese
quail
American
black ducks
cilied
Endpolnt
NS
rep
rep, dev '
rep, dev
let
fet
rep
dvp, behv
rep
rep
dvp
dvp
Description
LD50
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
AEL
NOEL
LOAEL
NOAEL
AEL
AEL
Value
890
0.014
1
10
0.4
4
QJ
0.6
19
1.9
75
4
....
Units
mg/kg-body
wt.
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
PPm
mg/kg-day
mg/kg-day
ppm
ppm
Exposure
Route (oral,
S.C., I.V.,
Injection)
oral
oral (water[
oral
oral
oraljwater)_
oral (water)
oral (water)
oral (diet)
oral
oral
oral
oral
Exposure Duration
/Timing
NS
90 days
6-9 weeks
6-9 weeks
gestational days 6-20
gestational days 6-20
3 generations
3 months
90 days
90 days
Birth to 6 weeks
parents - 4 mo. < egg
laying; ducklings - up
to 6 weeks
Reference
RTECS, 1994
Dixonetal., 1976
Sutouetal., 1980
Sutouetal., 1980
Sorell and Graziano, 1990
Sorell and Graziano, 1 990
Schroeder and Mitchener,
197J
Loeser and Lorke, 1977
White and Finley, 1978
White and Finley, 1978
Richardson et al, 1974
Heinz etal., 1983
Comments
No adverse effects on male
reproduction
decreased fetal weight
reproductive failure
egg production suppressed
testicular hypoplasia,
anemia, hypertrophy of
heart ventricles
altered avoidance behavior-
hyperresponsiveness
-------
Terrestrial Biological Up i Measures - Cadmium
Cas No. 7440-43-9
Chemical
Name
cadmium
cadmium
cadmium
Species
plant
earthworms
earthworms
NS = Not specified
B-lactor
(BCF. BAF.
BMP)
BCF
BCF
BCF
Value
0.18
1-7.5
4-5
Measured
or
Predicted
im-Pl_ .
£_
m
NS
units
(ug/g DW
plant)/(ug/g
soil)
NS
NS
Reference
U.S. EPA 1990e
Davies, 1983
Helmke, 1979
Comments
Data obtained from varying
distances to points of soil
contamination.
-------
APPENDIX B Chlordane-1
Toxicological Profile for Selected Ecological Receptors
Chlordane
Cas No.: 57-74-9
Summary: This profile on chlordane summarizes the toxicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (B.AFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
KOW between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire toxicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the data presented in the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. However, several chronic and subchronic toxicity
studies involving chlordane have been conducted using rats, mice and other laboratory
August 1995
-------
APPENDIX B Chlordane - 2
animals. Keplinger et al. (1970) conducted a six generational study to assess the reproductive
impacts of dietary chlordane exposure to mice at levels of 25, 50, and 100 ppm. At 50 ppm,
viability was significantly lower in the 4th and 5th generations and the fertility index was
significantly lower in the second and fifth generations. The 25 ppm level appeared to have
little or no significant effect on the generations of mice. The NOEL of 25 ppm was
converted to a daily dose of 4.37 mg/kg-d using the mean body weight and daily food
consumption equations for mice reported in Recommendations for and Documentation of
Biological Values for Use in Risk Assessment (U.S. EPA, 1988a). An additional chronic
study was.identified in which male and female rats were fed a 9 month diet containing 2.5 or
25 ppm of chlordane (Ortega et al., 1957). Liver cell abnormality was noted in males at the
lower dose. Using a recommended body weight for mature, male rates of 0.505 kg
(U.S.EPA, 1988) and the daily food consumption equation of FrrO.OSeW066" (Nagy, 1987),
the 2.5 ppm level was converted to a daily dose of 0.176 mg/kg-d. In a 2-year study by Ingle
(1952), male and female Osborne-Mendel rats were fed oral dietary doses of 5, 10, 30, 150,
300 ppm of chlordane. These rats were mated and all the resulting litters were normal as to
the individuals and the litter size. However, the pups from rats dosed at 150 ppm and 300
ppm that remained with their lactating mothers showed definite symptoms of chlordane
toxicity, resulting in some instances of death. Therefore, a NOAEL of 30 ppm for
reproductive effects from maternal transfer was derived from the Ingle study (1952). Using
the proper body weight and food intake assumptions for Osborne-Mendel strain rats (U.S.
EPA, 1988), the 30 ppm was converted to 2.29 mg/kg-d. Also, the International Research
and Development Corporation (1967 as cited in WHO, 1984) conducted a two-year feeding
study in which beagle dogs were fed chlordane at levels of 0, 0.3, 3.0, 15 or 30 mg/kg diet.
The IRDC reported a NOAEL of 3 mg/kg in the diet (equivalent to 0.075 mg/kg body
weight) for hepatic endpoints in dogs.
The studies by Keplinger et al. (1970) and Ingle (1952) documented reproductive toxicity to
laboratory mammals at similiar levels of chlordane exposure (4.37 mg/kg-d and 2.29 mg/kg-d,
respectively). For both of these studies, there were (1) chronic exposures administered via
oral ingestion, (2) sufficient dose-response information, and (3) critical reproductive endpoints
examined. However, the Kiplinger et al. (1970) study was chosen as the benchmark study on
the basis of its multi-generational dosing regime, which provides a useful indicator of a
chemical's effects on succeeding populations. Also, regarding the study by Ingle (1952),
there is additional uncertainly concerning the rate of maternal transfer of chlordane via
lactating female rats from a laboratory diet versus maternal transfer in mammalian wildlife.
The IRDC (1967 as cited in WHO, 1964) and Ortega et al. (1957) studies did not examine
reproductive or developmental endpoints. While the studies by IRDC (1967 as cited in
WHO, 1964) and Ortega et al. (1957) were not chosen for the development of a toxicological
August 1995
-------
APPENDIX B Chlordane-3
benchmark, they do illustrate the dose range at which chlordane toxicity occurs.
The study value from Keplinger et al. (1910) was scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994)
Benchmarkw = NOAEL, x L
'
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57FR 24152). Since the
Keplinger et al. (1970) study documented reproductive effects from toxaphene exposure to
male and female mice, the mean male and female body weight of each representative species
was used in the scaling algorithm to obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of chlordane, as well as on
growth and chronic survival. In addition, the data set contained studies which were mostly
chronic in nature. The majority of the studies identified were conducted using laboratory rats
or mice and as such, inter-species differences among wildlife species were not identifiable.
Therefore, an inter-species uncertainty factor was not applied. There were several study
values in the data set which were at least an order of magnitude lower than the benchmark
value. These values corresponded to effects on hepatic and growth endpoints. Based on the
data set for chlordane, the benchmarks developed from Keplinger et al. (1970) were
categorized as adequate, with a "*" to indicate that adverse effects may occur at the
. benchmark level.
Birds: No subchronic or chronic studies focusing on reproductive or developmental effects
from chlordane exposure to avain species were identified. Sources reviewed for avian toxicity
information included: Chlordane Hazards to Fish, Wildlife, and Invertebrates: A Synoptic
Review (FWS, 1990); Chlordane (WHO, 1984); an on-line search of the TOXLIT and DART
databases; and an extensive library search at National Institute for Environmental Health
Sciences (NEEHS) library.
Fish and aquatic invertebrates: The chronic AWQC of 4.3 E-6 mg/L is based on Final
Residue Value (U.S. EPA, 1989). The FRY was not considered to be appropriate for the
August 1995
-------
APPENDIX B Chlordane - 4
r
development of a benchmark for daphnids because it is intended to protect fish and other .
wildlife, which consume aquatic organisms, from the adverse effects of chemicals that may
bioconcentrate. Also, the FRY was not an appropriate benchmark value because residues and
bioaccumulation are already taken into account by the Thomann efal. (1992) model. The
Final Chronic Value (FCV) of 1.7E-4, as presented in the AWQC document (U.S. EPA,
1980) was selected as the benchmark protective of fish and aquatic invertebrate in the generic
freshwater ecosystem. Because the benchmark was based on a FCV derived for the AWQC,
this benchmark is categorized as adequate.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Aquatic plant data was not identified for chlordane and, therefore, no benchmark was
developed. .
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^) to determine a protective sediment concentration
(Stephan, 1993). The EQP number is the chemical concentration that may be present in
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. The FCV reported in the AWQC document (U.S. EPA, 1980) was used to
calculate a EQP number of 10.34 mg chlordane /kg organic carbon. Assuming a mass fraction
of organic carbon for the sediment (f^) of 0.05, the benchmark for the benthic community is
0.517 mg/kg. Since the EQP number was based on a FCV established for the AWQC, the
sediment benchmark is categorized as adequate.
August 1995
-------
APPENDIX B
Chlordane - 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem . /
D«MVAMAPlt«lllM
HopfUMIWIW
SpteiM
mink
river otter
bald eagle
osprey
great blue heron
mallard .
lesser scaup
herring gull
spotted sandpiper
kingfisher
Bvkchmvk
Value*
mgfte-fey
1.9 (a')
1.1 (a')
ID
ID
ID
ID
ID
ID
ID
ID
Study
Spacha
" mouse
mouse
-
'
-
-
-
'
-
-
Effect
rep
rep
-
-.
-
-
-
-
-
Study VsiiM
mgflcg-day
4.4
4.4
-
-
-
-
-
-
-
Ooscr Iption
NOEL
NOEL
-
-
-
-
-
-
-
-
. SF
-
-
-
-
-
-
-
-
-
Original Source
Keplinger et al., 1970
Keplinger et al.. 1970
-
-
-
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data '
August 1995
-------
APPENDIX B
Chlordane - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
R « nfra^hia
oproonuRivv
SpocJM
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
V«tu»'
fAQrl*
1 .7E-04 (a)
ID
0.517(a) mg/kg
sediment
StudySpactM
AWQC
organisms
-
AWQC
organisms
uMCflption
FCV
-
FCV x Kj
Original Source
U.S. EPA, 1980
U.S. EPA, 1980
II.
Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was
an order of magnitude or more above the NEL or LEI for other adverse effects.
ID = Insufficient Data .
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic .
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No additional chronic or subchronic studies were found for mammalian wildlife in
which dose-response data were reported for reproductive or developmental endpoints.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Keplinger et al., 1970) was used to derive the toxicological benchmark for mammalian
species representing the terrestrial ecosystem. The study value from Keplinger et al. (1970)
was scaled for species representative of a terrestrial ecosystem using a cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the Keplinger et al. (1970) study
documented reproductive effects from toxaphene exposure to female and male mice, the mean
body weight of both genders was used in the scaling algorithm to obtain the toxicological
benchmarks. Based on the data set for chlordane, the terrestrial benchmarks developed from
Keplinger et al. (1970) were categorized as adequate, with a "*" to indicate that adverse
effects may occur at the benchmark level.
August 1995
-------
APPENDIX B Chlordane-7
Birds: Although numerous sources were reviewed for toxicity information, no subchronic or
chronic studies were identified for representative or surrogate avian exposure to chlordane..
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, 'such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for chlordane and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Chlordane - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
napieaentaHve
SpaofeM
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American
woodcock
plants
soil community
. Benchmark
Value* mg/kg-
day
5.1 (a')
5.2 (a-)
4.4 (a')
1.8 (a')
1.3 (a')
1.2 (a')
0.62 (a')
ID
ID
ID
ID
ID
ID
ID
Steely
Spades
mouse
> mouse
mouse
mouse
mouse
mouse
mouse
-
-
-
-
-
-
Effect
rep
rep
rep
rep
rep
rep
rep
-
-
.
-
-
-
Original
Value mgftg-
day
4.4
4.4
4.4
4.4
4.4
4.4
4.4
-
-
-
-
-
-
-
Deeorlptton
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
-
-
-
SF
-
-
-
-
-
-
-
-
-
-
Original Source
Keplinger et at.,
1970
Keplinger et al..
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
Keplinger et al.,
1970
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a '*' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Chlordane - 9
III. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K^w values belowr4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log K,,w values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for chlordane. The
predicted BAF,d for trophic level 4 fish in both the limnetic and littoral ecosystems is
approximately twice the geometric mean (4,387,500) of the three measured values presented
in Derivation of Proposed Human Health and Wildlife Bioaccumulation Factors for the Great
Lakes Initiative (Stephan, 1993). The geometric mean of the measured values includes both
the alpha and gamma isomers and was based on data from Oliver and Niimi (1985 and 1988)
for rainbow trout and salmon. The bioconcentration factor for fish was also estimated from
the Thomann models (i.e., log K^ - dissolved BCF/) and multiplied by the dissolved fraction
(/d) as defined in Equation 6-21 to determine the total bioconcentration factor (BCF/). The
dissolved bioconcentration factor (BCF/1 ) was converted to the BCF/ in order to estimate the
acceptable lipid tissue concentration (TCI) in fish consumed by piscivorous fish (see Equation
5-115). The BCF/ was required in Equation 5-115 because the surface water benchmark (i.e.,
FCV or SCV) represents a total water concentration (C). Mathematically, conversion from
BCF,d to BCF/ is accomplished using the relationship delineated in the Interim Report on
August 1995
-------
APPENDIX B Chlordane - 10
Data and Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic
Wildlife (U.S. EPA, 1993i):
BCF," x rd = BCF;
Converting the predicted BCF,d of 870,963 LAg LP to the BCF/ of 242,777 L/kg LP was in
reasonable agreement (i.e., within a factor of 4) of the geometric mean of five measured BCF/
values presented in the master table on chlordane (geometric mean = 311,800).
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of several
values with sources in Table 4 (see master table). For earthworms and terrestrial
invertebrates, the bioconcentration factors were estimated as described in Section 5.3.5.2.3.
Briefly, the extrapolation method is applied to hydrophobic organic chemicals assuming that
the partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs
and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to .Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data (in this case chlordane) is compared to the BBF for TCDD
and that ratio (i.e., chlordane BBF/TCDD BBF) is multiplied by the TCpD standard for
terrestrial vertebrates, invertebrates, and earthworms, respectively. For hydrophobic organic
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses: The BCF is based on route-to-leaf
translocation, direct deposition on leaves and grasses, and uptake into the plant through air
diffusion.
August 1995
-------
APPENDIX B
Chlordane - 11
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
Itpld-bttad or
whole Dooy
lipid
lipid
lipid'
lipid
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
9,255,382 (d)
4,893,454 (d)
242,777 (t)
9,580,446 (d)
8,889,256 (d)
10,200,120 (d)
0.49
0.01
0.85
0.014
ource
predicted value based on Thomann, 1989,
food chain model
predicted value based on Thomann, 1989,
food chain model
. predicted value based on Thomann, 1989
and adjusted to estimate total BCF
predicted value based on Thomann et al.,
1992, food web model
predicted value based on Thomann et al.,
1 992, food web model
predicted value based on Thomann et al.,
1992, food web model
geometric mean of values in Garten and
Trabalka, 1983; Clabom et al., 1956, 1960
as cited in Kenaga, 1980
estimated based on beef biotransfer ratio
with 2,3,7,8-TCOD
estimated based on beef biotransfer ratio
with 2,3,7,8-TCDD
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Chlordane 12
References
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TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
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Arruda, J.A., M.S. Cringan, D. Gilliland, S.G. Haslouer, J.E. Fry, R. Broxterman, and K.L.
Brunson. 1987. Correspondence between urban areas and the concentrations of chlordane
in fish from the Kansas River. Bull. Environ. Contam. Toxicol. 39:563-570.
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of soil applied DDT, dieldrin and heptachlor. J. AppL Ecol. 17:295-307.
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Cardwell, R.D. et al. 1977. Acute and Chronic Toxicity of Chlordane to Fish and
Invertebrates. EPA/600/3-77/019. Chemico Process Plants Co., El Monte, California
Claborn, H.V., R.D. Radeleff, and R.C. Bushland. 1960. Pesticide Residues in Meat and
Milk. ARS-33-63. U.S. Department of Agriculture. As cited in Kenaga, E.E, 1980,
Correlation of bibconcentration factors of chemicals in aquatic and terrestrial organisms
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Claborn, H.V. 1956. Insecticide Residues in Meat and Milk. ARS-33-25. U.S. Department
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chemicals in aquatic and terrestrial organisms with their physical and chemical properties,
Environmental Sci. Technol. 14(5):553-556.
August 1995
-------
APPENDIX B Chlordane-13
Colombo, J.C., M.F. Khalil, M.Arnac, and A.C. Horth. 1990. Distribution of Chlorinated
Pesticides and Individual Polychlorinated Biphenyls in Biotic and Abiotic Compartments
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Deichmann, W.B. and M.L. Keplinger. 1966. Effect of Pesticides on Reproduction of Mice.
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Eisler, R. (ed.), 1990. Chlordane Hazards to Fish, Wildlife, and Invertebrates: A Synoptic
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Equivalence of mg/kg3/4/day.
FAOAVHO (Food Agriculture Organization/ World Health Organization). 1968. Evaluations
of Some Pesticides Residues in Food. Food and Agriculture Organization in the United
Nations, Rome.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol. 14:525-534. As
cited in WHO (World Health Organization). 1984. Chlordane, Environmental Health
Criteria 34, Geneva, Switzerland.
Garten, C.T., and J.R. Trabalka. 1983. Evaluation of models for predicting terrestrial food
chain behavior of xenobiotics. Environ. Sci. Technol. 26(10):590-595.
Goodman, L.R., D.J. Hansen, J.A. Couch, and J. Forester. 1978. Effects of heptachlor and
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Proceedings of the Thirtieth Annual Conference. 1976. Southeastern Association of
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Derivations of proposed human health and wildlife bioaccumulation factors for the
Great Lakes Initiative. PB93-154672. Environmental Research Laboratory, Office of
Research and Development, Duluth, MN.
Ingle, L. 1952. Chronic Toxicity of Chlordan to Rats. Arch. Ind. Hyg. Occ. Med. 6: 357-367
August 1995
-------
APPENDIX B Chlordane -14
Ingle, L. 1965a. Effects of 1-hydroxychlordene when incorporated into the diets of rats for
224 days. Prepared for Velsicol Chemical Corporation. Department of Zoology,
University of Illinois, Urbana, Illinois. As cited in WHO (World Health Organization).
1984. Chlordane, Environmental Health Criteria 34, Geneva, Switzerland.
Ingle, L. .1965b. Monograph on Chlordane-Tqxicological and Pharmacological Properties.
Library of Congress, Card Number 65-28686A. Food and Drug Library, University of
Illinois, Urbana, Illinois.
IRDC (International Research and Development Corporation). 1967. Chlordane, Two-Year
Chronic Feeding Study in the Beagle Dog. Prepared for Velsicol Chemical Corporation,
Report 163-001. International Research and Development Corporation, Mattawan,
Michigan. As cited in WHO (World Health Organization). 1984. Chlordane,
Environmental Health Criteria 34, Geneva, Switzerland.
Kawano, M., T. Inoue, T. Wada, H. Hidaka, and R. Tatsukawa. 1988. Bioconcentration and
Residue Patterns of Chlordane Compounds in Marine Animals: Invertebrates, Fish,
Mammals, and Seabirds. Environmental Science and Technology, 22:792-797.
Keplinger, M.L., W.B. Deichmann, and F. Sala. 1970. Effects of combinations of pesticides
on reproduction in mice. In: Pesticides Symposia, 6th and 7th Inter-American Conf.
Toxicol. Occup. Med., Halos and Associates, Inc., Coral Gables, Florida, pp. 125-138.
Lundholm, C.E. 1988. The Effects of DDE, PCB, and Chlordane on the Binding of
Progesterone to its Cytoplasmic Receptor in the Eggshell Gland Mucosa of Birds and the
Endometrium of Mammalian Uterus. Comparative Biochemistry and Physiology, Vol.
89C, No. 2, pp. 361-368.
NCI (National Cancer Institute). 1977. Bioassay of Chlordane for Possible Carcinogenicity.
NCI Carcinogenesis Technical Report Series, Number 8.
Oliver, B.G., and A.J. Niimi. 1985. Bioconcentration factors of some halogenated organics
for rainbow trout: Limitations in their use for predictions of environmental residues.
Environ. Sci. Technol. 22:388-397. As cited in Stephan, C.E. 1993. Derivations of
proposed human health and wildlife bioaccumulation factors for the Great Lakes
Initiative. PB93-154672. Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
August 1995
-------
APPENDIX B Chlordane - 15
Oliver, E.G., and A.J. Niimi. 1988. Trophodynamic analysis of polychlorinated biphenyl
congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ.
Sci. Technol. 22:388-397.
Ortega, P., W.J. Hayes, and W.F. Durham. 1957. Pathologic changes in the liver of rats
after feeding low levels of various insecticides. Am. Med. Assoc. Arch. Pathol. 64:614-
622.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Smith, V.E., J.M. Spurr, J.C. Filkins, and J.J. Jones. 1985. Organochlorine contaminants of
wintering ducks foraging on Detroit River sediments. J. Great Lakes Res. 11:231-246.
Stephan, C.E. 1993. Derivations of proposed human health and wildlife bioaccumulation
factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Stohlman, E.F., W.T.S, Thorp, and M.I. Smith. 1950. Toxic action of chlordane. J. Ind.
Hyg. Occup. Med. 1:13-19. As cited in WHO (World Health Organization). 1984.
Chlordane, Environmental Health Criteria 34, Geneva, Switzerland.
Suter II, G.W., J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Oak Ridge, TN
Talamantes, F. and H. Jang. 1977. Effects of Chlordane Isomers Administered to Female
Mice During the Neonatal Period. Journal of Toxicology and Environmental Health,
3:713-720,
Thomann, R.V. 1989.. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
August 1995
-------
APPENDIX B Chlordane -16
Travis, C.C., and A.D. Arms. 1988. Bioconcentration of organics in beef, milk, and
vegetation. Environ. Sci. Technol. 22(3):271-274.
Truhaut, R., J.C. Gak, and C. Graillot. 1974. Study of the modalities and action mechanisms
of organochlorine insecticides. I. Comparative study of the acute toxicity in hamster and
rat. J. Eur. Toxicol. 7:159-166. As cited in WHO (World Health Organization). 1984.
Chlordane, Environmental Health Criteria 34, Geneva, Switzerland.
U.S. EPA (U.S. Environmental Protection Agency). 1980. Ambient Water Quality Criteria
for Chlordane. ECAO-C-625. Environmental Criteria and Assessment Office, Office of
Water Regulations and Standards, Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1986a. Health Effects Assessment for
Chlordane. EPA540/1-86-023. Office of Emergency and Remedial Response,
Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1986b. Carcinogenicity Assessment of
Chlordane and Heptachlor/Heptachlor Epoxide. EPA/600/6-87/004. Office of Health and
Environmental Assessment, Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1987. Drinking Water Criteria
Document for Heptachlor, Heptachlor Epoxide, and Chlordane (Final). PB89-192157.
Environmental Criteria and Assessment Office, Cincinnati, Ohio, 303 p.
U.S. EPA (U.S. Environmental Protection Agency). 1988a. Recommendations for and
Documentation of Biological Values for use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1988b. United States Environmental
Protection Agency Office of Drinking Water Health Advisories. Chlordane. Rev.
Environ. Contam. Toxicol. 104:47-62. As cited in U.S. Fish and Wildlife Service, 1990,
Chlordane Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review, Contaminant
Hazard Reviews Report 21, Biological Report 85 (1.21), U.S. Department of the Interior.
Washington, D.C.
August 1995
-------
APPENDIX B , Chlordane-17
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
. Health Risks Associated with Indirect. Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (U.S. Environmental Protection Agency). 19935. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Science and Technology, Office of Water, Washington, DC.
U.S. EPA (Environmental Protection Age.ncy). 1993c. Technical Basis for Deriving Sediment
Quality Criteria for Nonionic Organic Contaminants for the Protection of Benthic
Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of Water,
Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1993i. Interim Report on Data and
Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Life and
Associated Wildlife. EPA/600/R-93/055. Office of Research and Development,
Washington, DC.
Veith, G.D., D.L. DeFoe, and B.V. Bergstedt. 1979. Measuring and estimating the
bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36:1040-1048.
Veith, G.D., D.L. DeFoe, and B.V. Bergstedt. 1979. Measuring and estimating the
bioconcentration factor of chemicals in fish. /. Fish. Fes. Board Can. 36:1040-1048.
Ware, G.W. (ed.). 1988. Reviews of Environmental Contamination and Toxicology.
Continuation of Residue Reviews, United States Environmental Protection Agency, Office
of Drinking Water Health Advisories, Springer-Verlag, New York, Vol. 104.
Welch, R.M. 1948. Tests of the toxicity to sheep and cattle of certain of the newer
insecticides. J. Econ. Entomol. 41:36-39.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial To y - Chlordane
Cas No. 57-74-9
Chemical
Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
mice
rat
rat (m)
rat (f)
mouse (m)
mouse (f)
mice
dog
rat
rat
Endoolnt
rep
hepatic
body wt.
body wl.
mort.
mort.
fet, ter
hepatic
hepatic.
renal, repro.,
cardiac
growth
Description
NOEL
LOAEL
LOAEL
LOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Value
4.37
0.176
20.3
12.1
29.9
63.8
2.3
0.075
0.25
0.73
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-dav
mg/kg-day
mg/kg-d
Exposure
Route (oral,
S.C., I.V., l.p..
Inlectlon)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
6 generations
9 months
80 weeks
80 weeks
80 weeks
80 weeks
2 years
2-year study
2 years
407 days
Reference
Keplinger et al., 1970
Ortega etal., 1957
NCI, 1977
NCJ. 1977
NCI, 1977
NCI, 1977
Ingle, 1952
IRDC, 1967 as cited in
WHO, 1984
Ingle, 1952
Ambrose et al., 1953
Comments
No significant effects were seen
on fetus viability and fertility.
Doses were 25 and 2.5 ppm. >
Liver cell abnormality noted in
males at the lower dose.
Decreased body weight gain was
noted in high-dose males.
Decreased body weight gain was
noted at both dose levels of
females.
Increased mortality incidence was
noted for male mice.
No mortality differences were
noted for female mice.
Doses were 5, 10, 30. 150, and 300
ppm. Reproductive effects from
maternal transfer (lactation) to
offspring were noted' at 150 and
300 ppm. 30 ppm equivalent to
2.3 mg/kg-d
No liver effects were observed at
this dose level. Sufficient dose-
response.
No symptoms of toxicity, gross or
histopathologic changes in the
liver, kidneys, lungs, pancreas,
testes, ovaries, heart, or spleen
were noted at 5 ppm. (this dose)
Sufficient dose-response.
Suggestion of retardation in
growth of males fed equivalent
dose of 1 .45 mq/kq-d.
Chlordane - Page 9
-------
Terrestrial Toxicity - Chlordane
Cas No. 57-74-9
Chemical
Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
rat
mouse
rabbit
hamster
chicken
duck
mammal
mallard
California
quail
pheasant
female rat
male rat
rabbit
rabbit
male rat
female rat
Endpolnt
mort.
moil.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
200
145
100
1720
220
1200
180
1200
14.1
24.0 - 72.0
530
205
<780
1100-1200
335
430 .
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral.
B.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
dermal
dermal
dermal
dermal
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Gaines, 1969 as cited in
WHO, 1984
Gaines, 1969 as cited in
WHO, 1984
Ingle, 1965a as cited in
WHO, 1984
Ingle, 1965a as cited in
WHO, 1984
Gaines, 1969 as cited in
WHQJ984
Gaines, 1969 as cited in
WHO, 1984
Comments
-
'
p - Pane 10
-------
Terrestrial Tc .y - Chlordane
Cas No. 57-74-9
Chemical
Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
rat
rat
rat
rabbit
hamster
goat
sheep
chicken.
mallard
cow
Endpolnt
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
200-590
263
350
100-300
1720
180
500-1000
220-230
1200.
25-90
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
B.C., I.V., l.p..
Inlectlon)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Ambrose et al., 1953
and Ingle, 1965aas
cited in WHO 1984
Bucketal., 1973
Truhaut et al., 1974 as
cited in WHO. 1984
Stohlman et al., 1950 as
cited in WHO, 1984
Truhaut etal.,. 1974 as
cited in WHO, 1984
Welch, 1948
Welch, 1948
FAO/WHO, 1968
Bucketal., 1973
Bucketal., 1973
Comments
'
NS=Nol specified
Chlordane - Page 11
-------
Freshwater Toxicity - Chlordane
Cas No. 57-74-9
Chemical
Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
fish
aquatic
organisms
aquatic
organisms
fish
daphnids
fish
daphnids
fathead
minnow
Endpolnt
chron
chron
chron
chron
chron
chron
chron
mort
Description
NOAEL
AWQC
FCV
CV
CV
EC20
EC20
LC50
Value
<0.1
0.0043
0.17
1.6
16
<0.25
12.1
69 - 180
(112.2)
Units
mg/ kg fresh
weight tissue
ug/L
ug/L
ug/L
ug/L
ug/L
uo/L
uo/L
Test type
(static/ flow
through)
NS
NS
NS
NS
NS
NS
NS
NA
Exposure
Duration/
Tlmlnq
NS
NS
NS
NS
NS
NS
NS
4 days
Reference
Arrudaetal , 1987
U.S. EPA, 1980
U.S. EPA, 1980
Suter and Mabrey, 1994
Suter and Mabrey, 1 994
Suter and Mabrey, 1994
Suter and Mabrey, 1 994
AQUIRE. 1995
Comments
NS = Not specified
-------
Freshwater Biological I . ie Measures - Chlordane
Cas No. 57-74-9
Chemical
Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
lathead
minnow
fish
fish
fish
fish
fish
fish
fish
rainbow trout
rainbow trout
salmon
carp
salmon ids
B-factor
(BCF, BAF,
BMP)
BCF
BAF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
BAF
BSAF
BSAF
Value
37,800
8318
1894
2218
4357
4974
2577
2379
175000
9500
50802
46
2
Measured or
predicted
(m,p)
m
m
P
m
m
m
m
m
m
m
m
P
P
Units
NS
L/kg
NS
NS
NS
NS
NS
NS
NS
NS
NS
ug/g
UQ/Q
Reference
Veithetal., 1979
Garten and Trabalka, 1983
Stephan, 1993
Goodman et al., 1978 as
cited in Stephan, 1993
U.S. EPA, 1978 as cited in
Stephan, 1993
Veithetal., 1979
Oliver and Niimi, 1985 as
cited in Stephan, 1993
Oliver and Niimi, 1985 as
cited in Stephan, 1 993
Oliver and Niimi, 1985 as
cited in Stephan, 1993
Oliver and Niimi, 1985 as
cited in Stephan, 1993
Oliver and Niimi, 1985 as
cited in Stephan, 1993
Smith et. al., 1985
Oliver and Niimi, 1988
Comments
Microcosm; All estimates were calculated
based on published data; the type of studies
from which the data were taken were not
specified.
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1 % lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
NS = Not specified
-------
Terrestrial Biological Uptake Measures - Chlordane
Cas No. 57-74-9
Chemical Name
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
chlordane
Species
earthworm
earthworm
cattle
cattle
swine
swine
cattle (beef)
cattle (milk)
sheep
poultry
rodents
cow
swine
plants
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BTF
BTF
BAF
BAF
BAF
BAF
BAF
BCF
Value
3
0.24
0.5
0.1
0.32
0.9
0.0074
0.00037
0.89
3.3
0.35
0.32
0.54
260
Measured or
predicted
(m,p)
P
m
m
m
m
m
m
m
P
P
P
P
P
P
Units
NS
NS
NS
NS
NS
NS
NS
NS
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
tat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(ug/gWW
piant)/(ug/mL
soil water)
Reference
Beyer and Gish, 1980
Beyer and Gish, 1980
Claborn et.al., 1960 as cited in
Kenaga, 1980
Claborn et.al., 1960 as cited in
Kenaga, 1980
Claborn et.al., 1956 as cited in
Kenaga, 1980
Claborn et.al., 1956 as cited in
Kenaga, 1980
Travis and Arms, 1988
Travis and Arms, 1988
Garten and Trabalka, 1983
Garten and Trabalka, 1983
Garten and Trabalka, 1983
Garten and Trabalka, 1983
Garten and Trabalka, 1983
U.S. EPA, 1990e
Comments
BTF = Biotransfer factors.
BTF = Biotransfer factors.
NS = Not specified
-------
APPENDIX B Chromium (VI) .
Toxicological Profile for Selected Ecological Receptors
Chromium (VI)
Cas No.: 7440-47-3
Summary: This profile on chromium (VI) summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized.for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5: For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from ihe technical support document
for the Hazardous Waste Identification Rule (HWIR): Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated wilh the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several studies were identified which investigated the effects of chromium (VI)
exposure in mammals. Mice given oral doses of 250, 500 and 1000 ppm Cr (VI) as
potassium dichromate in drinking water during gestation days 1 through 19 had increased pre-
implantation and post-implantation losses as well as decreased liiter size (Trivedi et al., 1989).
These ppm values were converted to daily doses by using the reported body of weighi .03 kg
'and daily water consumption given by the equation:
C = 0.10W°'7377 (Nagy, 1987)
August 1995
-------
APPENDIX B Chromium (VI) - 2
where C is the water consumption rate and W is the body weight of the test species, which
was reported as being 0.03 kg.. The equivalent daily doses were calculated as being equal to
58, 117, and 233 mg/kg-day. A LOAEL of 58 mg/kg-day was reported for reproductive
effects. Decreases in motor activity and balance were seen in rats given oral doses of
chromium (VI) at 700 mg/1 as sodium chromate for 28 days. However, no adverse effects on
motor activity were exhibited by the treatment group receiving 70 mg/1 (Diaz-Mayans et'al.,
1986). These mg/1 values were converted to daily doses through the use of the allometric
equation (Nagy, 1987) presented above. Based on the reported body weight of 0.26 kg, the
water consumption rate was estimated as being 0.035 I/day. A NOAEL of 70 mg/1, estimated
as being equivalent to 10.2 mg/kg/day, and a LOAEL of 700 mg/1, equivalent to 102 mg/kg-
day, were reported for these neurological deficits. Zahid et al. (1990) fed mice potassium
dichromate at doses of 100, 200 and 400 ppm. After 7 weeks of treatment, the treatment
group receiving 100 ppm sodium dichromate exhibited reduced sperm counts and
degeneration of the outer cellular layer of seminiferous tubules. Morphologically altered
sperm were seen in the rats receiving 200 ppm sodium dichromate. The 100 ppm
concentration was converted to a daily dosage value by multiplying the food consumption per
animal (0.0075 kg/day as was measured in the study) by the ppm value (100) and dividing by
the geomean of the reported body weights of the mice. Based on these conversions, a
LOAEL of 32.6 mg/kg-day was reported for alterations in reproduction.
The Trivedi et al. (1989) study was not chosen for the development of a mammalian
benchmark because the derived LOAEL was not the lowest value in the data set and- would
therefore not be appropriate as the benchmark value. The Diaz-Mayans et al. (1986) study
was not selected because it focused on neurological impairment as the primary endpoint rather
than reproductive or developmental endpoints. As the Zahid et al (1990) study considered
reproductive effects, illustrated a clear dose-response relationship and represents the lowest
LOAEL identified for reproductive effects, it was chosen for the derivation of a benchmark
value. The selected study LOAEL, 32.6 mg/kg-day was divided by 10 to provide a LOAEL-
to-NOAEL safety factor. The LOAEL-to-NOAEL safety factor of 10 was applied to provide
a conservative estimate of the NOAEL. The LOAEL/10 from Zahid et al. (1990) was then
scaled for species that were representative of the generic freshwater ecosystem using a cross-
species scaling algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL. x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Zahid et al. (1990) study documented reproductive effects on male mice, male body weights
for each representative species were used in the scaling algorithm to obtain lexicological
August 1995
-------
APPENDIX B Chromium (VI). 3
benchmarks. Based on the data set for chromium, and since study value was a LOAEL rather
than a NOAEL, the benchmarks developed from the Zahid et al. (1990) study were
categorized as provisional.
Birds: No sub-chronic or chronic studies demonstrating adequate dose-response relationships
were identified which studied the effects of chromium (VI) toxicity in avian species.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) 1.1 E-02 mg/1 reported in
the AWQC document for chromium (U.S. EPA, 1980) was selected as the benchmark value
protective of fish and aquatic invertebrates. Because the benchmark is based on an FCV
developed for a AWQC, it was categorized as adequate with a "*" to indicate that adverse
effects may occur at the benchmark level.
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest pbserved effects concentration (LOEC) for vascular aquatic
plants (e,g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrwn capricornutum). The aquatic
plant benchmark for chromium (VI) is 0.002 mg/1 based on the incipient inhibition of
Microcystis aeruginosa. As described in Section 4.3.6, all benchmarks for aquatic plants
were designated as interim.
Benthic community: The chromium (VI) benchmark protective of benthic organisms is
pending a U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for
metals.
August 1995
-------
APPENDIX B
Chromium (VI) - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Raprm«nUttv«
8p*of«»
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
Value* ro8/fco>
day
1.18(p)
0.75{p)
ID
ID
10
ID
ID
ID
ID
ID
Study
Sp«ci«*
<
mouse
mouse
-
-
-
Elteot
rep
rep
-
-
-
-
-
Study Value
nig/fcg-day
32.6
32.6
- .
-
-
-
'
-
Dttcriptton
LOAEL
LOAEL
9F
10
10
-
'- .
Oria^wtSoWC* :
Zahid et al. 1990
Zahid et al. 1990
-
-
-
'Benchmark Category, a « adequate, p « provisional, i = interim; ID = insufficient data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
RaptMvnittiv*
SP«C!M
fish and aquatic
invertebrates
aquatic plants
benthic community
Beoctonarfc
Vote**
,witfl.
1.1 E-02(a)'
0.002 (i)
under review
- -fcudy
Sp«da*
aquatic
organisms
Microcystis
aaniginosa
Original
Valu»
mg/t
1.1E-02
0.002
Oaacfiptton
FCV
. CV
OrtgfnaJ Soorca
AWQC Table
Suter & Mabrey,
1994
-
'Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates that the
benchmark value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Chromium (VI) - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As discussed in the rationale for the freshwater ecosystem, there were three
possible studies from which to estimate a benchmark value. Since no additional studies of
terrestrial mammals were identified, the same surrogate study (Zahid et al., 1990) was used to
calculate benchmark values for mammalian species representing the general terrestrial
ecosystem. A LOAEL-to NOAEL safety factor of 10 was needed to estimate a NOAEL. The
calculated NOAEL was then scaled for species in the terrestrial ecosystem using the cross-
species scaling algorithm adapted from Opresko et al. (1994). Since the Zahid et al. (1990)
study documented reproductive effects in male mice, male body weights for each
representative species were used in the scaling algorithm to obtain the lexicological
benchmarks. Based on the data set for chromium (VI), the benchmarks developed from the
Zahid et al (1990) study were categorized as provisional.
Birds: No sub-chronic or chronic studies demonstrating adequate dose-response relationships
were identified which studied ihe effects of chromium (VI) toxicity in avian species.
Plants: Adverse effects levels for terreslrial planis were identified for endpoinls ranging
from perceni yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the Lowest Observable Effecis Conceniration
(LOEC) values and then approximating the 10th percentile. If there were 10 values, Ihe lOih
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield
reductions, or other effects reasonably assumed to impair the ability of a plant, population to
sustain itself, such as a reduction in seed elongation. The benchmark for terreslrial planis
was 1.8 mg/kg based on EC50 values of lettuce in loam soil. This was the lowest LOEC
presented by Will and Suter (Adema and Henzen, 1989 as cited in Will and Suter, 1994).
The phytotoxicity benchmark was categorized as interim as there were less than 10 values
presented by Will and Suter (1994).
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Chromium (VI) - 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RepniMnurtv*
. SpftCi** -
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
ftancbnvwt
Vtiur
rngfttfr*
3.38(p)
3.52(p)
2.79(p)
1.23(p)
0.85(p)
O.B1(p)
0.39
-------
APPENDIX B
Chromium (VI). 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values.
The whole-body, fish BCF value for chromium VI was the geometric mean of the measured
values, 0.13 and 2.8 (Stephan, 1993). BCF values for muscle were not included because
ecological receptors are likely to eat the whole fish or, in the least, will not necessarily
distinguish between the fillet and other parts of the fish. Insufficient data were identified to
determine the BCF value in aquatic invertebrates, terrestrial vertebrates, terrestrial
invertebrates and earthworms. A whole plant BCF value of 7.5 E-03 was derived from Baes
et al. (1984). For metals, empirical data were used to derive the BCF for aboveground forage
grasses and leafy vegetables. In particular, the uptake response slope for forage grasses was
used as the BCF for plants in the terrestrial ecosystem since most of the representative plant-
eating species feed on wild grasses.
Table 4. Biological Uptake Properties
«ca logical
r*o*pter
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
6CF,SA?,ar
BSAF
BCF
-
-
-
BCF
Itpid-baMd «w
wfw>4*-bo
-------
APPENDIX B Chromium (VI) - 8
References
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APPENDIX B Chromium (VI) - 9
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APPENDIX B Chromium (VI) - 10
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-------
APPENDIX B Chromium (VI) . 11
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«>»$^
August 1995
-------
APPENDIX B Chromium (VI) - 12
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
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Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1990. Noncarcinogenic effects of
Chromium:Update to the health assessment document. EPA/600/8-87/048F. Office of
Research and Development, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993. Derivations of Proposed Human
Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative. PB93-
154672. Environmental Research Laboratory, Office of Research Development, Duluth,
MN.-
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
September, 1993.
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2): Chemical toxicity of metals
and metalloids. Plenum Press, N.Y., 1978.
Wittbrodt P.R. and C.D. Palmer. 1994. Reduction of Cr (VI) by soil fulvic acid. Presented
before the Division of Environmental Chemistry, American. Chemical Society, San Diego,
CA.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
840R21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
August 1995
-------
APPENDIX B Chromium (VI) - 13
Zahid, Z. R., Z. S. Al-Hakkak, and A. H. H. Kadhim. 1990. Comparative effects of trivalent
and hexavalent chromium on spermatogenesis of the mouse. Toxicology and
Environmental Chemistry 25:B 1-136. As cited in Agency for Toxic Substances and
Disease Registry (ATSDR). 1993. Toxicological Profile for Chromium. Public Health
Service, U.S. Department of Health and Human Services, Atlanta, GA.
August 1995
-------
Terrestrial Tox. / - Chromium
Cas No. 7440-47-3
Chemical
Name
chromium
chromium III
chromium III
chromium VI
chromium VI
chromium VI
chromium VI
Species
common tern
black duck
rat
rat
chicken
mouse
rat
Type of
Effect
dev, rep
behv
dev
path
dev
"
rep
neuro
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
NOAEL
Value
<8
200
50000
2.4
8
58.3
10.21
Units
ug/g-DW
ppm
"9/9
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C.. I.V.. i.p.,
injection)
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
NS
5 months
NS
1 year
21 days
gestation
days 1-19
28 days
Reference
Cysteretal.. 1986
Heinz and Haseltine,
1981
Ivankovic and
Preussmann, 1975
MacKenzie et al., 1958
as cited in IRIS. 1992
Rornoser etjil., 1961
Trivedi etal., 1989
Diaz-Mayans et al.,
1986
Comments
Common Tern clutch size,
reproductive success and
growth of young were equal to
or greater than the control, a
less contaminated area. .
No significant difference in the
avoidance response of
ducklings to a fright stimulus
was detected.
50,000 FW
No significant effects were seen
on appearance, weight gain, or
food consumption, and there
were no pathological changes
in the blood or other tissues.
No adverse effects in survival,
growth or food utilization
efficiency.
Increase in fetal resorption and
post-implantation loss.
No effect on motor activity.
-------
Terrestrial Toxicity - Chromium
Cas No. 7440-47-3
Chemical
Name
chromium VI
chromium VI
NS = Not
specified
Species
rat
mouse
Type of
Effect
neuro
rep
Description
LOAEL
LOAEL
Value
102.1
32.6
Units
mg/kg-day
mg/kg-day
Exposure
Route (oral.
s.c., l.v., i.p.,
injection)
oral
oral
Exposure
Duration
/Timing
28 days
35 days
Reference
Diaz-Mayans et at.,
1986
Zahidetal., 1990
Comments
Decreased motor activity.
Reduced sperm count;
degeneration of outer cellular
layer of seminiferous tubules.
-------
Freshwater Toxicily - Chromium
Cas No. 7440-47-3
Chemical
Name
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
Species
brook trout
fathead
minnow
lake trout
channel
catfish
bluegill
white sucker
northern pike
walleye
Type of
Effect
dev.rep
dev.rep
dev.rep
dev.rep
dev.rep
dev.rep
dev.rep
Description
CV
CV
CV
CV
CV
CV
CV
dev.rep |CV
Value
200-350
1000-3950
105-194
150-305
522-1122
290-538
538-963
>2161
Units
ug/L
ug/L
ug/L
UQ/L
ug/L
ug/L
ug/L
-
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS .
NS
NS
Exposure
Duration
/Timing
-
NS
NS
NS
NS
NS .
NS
NS
NS
Reference
EPA. 1980 as cited in
Eisler. 1986.
Pickering, 1980 as cited in
Eisler. 1986.
Sauter et al., 1976 as cited
in EislerL1986.
Sauter et al., 1976 as cited
in Eisler, 1986.
Sauter et al., 1976 as cited
injEisler, 1986.
Sauter et al., 1976 as cited
in Eisler, 1986.
Sauter et al., 1976 as cited
in Eisler, 1986.
_
Sauter et al.. 1976 as cited
in Eisler, 1986.
Comments
-------
Freshwater To iy-Chromium
Cas No. 7440-47-3
Chemical
Name
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
chromium VI
Species
aquatic
organisms
fish
daphnid
fish
daphnid
fathead
minnow
striped bass
bluegill
daphnid
rainbow trout
rainbow trout
Type of
Effect
chronic
chronic
chronic
chronic
chronic
acute
acute
acute
acute
dev, rep
dev.rep
Description
AWQC
CV
CV
EC20
EC20
LC50
LC50
LC50
LC50
CV
CV
Value
11
73.18
6.132
51
0.5
37,000 -
52,000
(44,157)
17,700
113,000
435
51-105
200-350
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
"9/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS
NS
NS_
NS
96-hour
96-hour
96 hours
24 hours
,
NS
NS
Reference
U.S. EPA, 1986
Suteretal., 1992
Suter et al.. 1992
Suteretal . 1992
Suteretal., 1992
Reusink et al., 1975 as
cited in AQUIRE, 1995
Rehwoldtetal.. 1973 as
cited in AQUIRE, 1995
U.S. EPA, 1980 as cited in
Eisler.J986
Jouany et al., 1982 as cited
inEisler, 1986
Sauter et al., 1976 as cited
inEisler, 1986
EPA, 1980 as cited in
Eisler, 1986.
Comments
Water hardness=44 mg
CaCO3/L.
For 34 mg CaCO3/L
For 45 mg CaCO3/L
-------
l-reshwater To. ,y - Chromium
Cos No. 7440-47-3
Chemical
Name
chromium III
chromium III
chromium III
chromium III
chromium III
chromium III
chromium III
- Species
Cladoceran
Fathead
minnow
Rainbow
trout
aquatic
organisms
fish
daphnid
fish
NS = Not specified
Type of
Effect
dev.rep
dev.rep
dev.rep
chronic
chronic
chronic
chronic
Description
cy
CV
CV
NAWQC
CV
CV
EC20
Value
47-93
750-1400
30-157
210
68.B3
<44
89
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS .
NS
NS
NS
NS
NS
Reference
EPA. 1980 as cited in
Eisler, 1986.
EPA, 1980 as cited in
Eisler, 1986.
Stevens & Chapman, 1985
as cited in Eisler, 1986.
U.S. EPA, 1986
Suleretal.,1992
SuteretaJ., 1992
SuteretaJ.. 1992
Comments
-------
Freshwater Biological Uptake Measures - Chromium
Cas No. 7440-47-3
Chemical
Name
chromium
chromium
chromium VI
Species
rainbow trout
rainbow trout
fish
NS = Not specified
B-faclor
(BCF, BAF,
BMP)
BCF
BCF
BCF
Value
0.13
2.80
16
Measured
or
Predicted
(m.p)
NS
NS
NS
Units
NS
NS
LAg
Reference
Buhleretal., 1977
Calamari, 1982
U.S. EPA. 1992
Comments
Muscle BCF.
Muscle BCF.
Normalized to 3% lipid.
-------
Terrestrial Biological Up. .a Measures - Chromium
Cas No. 7440-47-3
.
Chemical Name
chromium VI
Species
plant
B-Iaclor
(BCF. BAF.
BMP)
BCF
Value
1.1
Measured
or
Predicted
(ro.P)
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPA, 1990e
Comments
..»'-.'
I
-------
Terrestrial Biological Up. .e Measures - Chromium .
Cos No. 7440-47-3
.
Chemical Name
chromium VI
Species
plant
B-factor
(BCF, BAF,
BMP)
BCF
Value
1.1
Measured
or
Predicted
(m,p)
P
units
(ug/g DW
plant)/(ug/g soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Chrysene-1
Toxicological Profile for Selected Ecological Receptors
Chrysene
CasNo.: 218-01-9
Summary: This profile on chrysene summarizes the toxicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), .bioaccumulation factors (BAFs) and, if available, biomagnification
factors (BMFs) are also summarized for the ecological receptors, although some BAFs for the
freshwaterecosystem were calculated for organic constituents with log K,,w between 4 and 6.5.
For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire toxicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from the information presented in the technical support
document for the "Hazardous Waste Identification Rule (HWIR): Risk Assessment for Human
and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Adequate toxicity data focusing on critical endpoints pertinent to population
sustainability were not identified. Therefore, benchmarks protective of the mammalian
community in a freshwater ecosystem were not derived.
August 1995
-------
APPENDIX B Chrysene-2
Birds: The only identified avian study documented a NOAEL for chrysene exposure to
mallard ducks. Hoffman and Gay (1981) observed embryotoxic effects from a one-time
application of chrysene to the shell surface of mallard eggs at 72 hours of development
(observation period through day 18 of incubation). Eggshell surfaces were coated with a 10
ul of a synthetic petroleum hydrocarbon mixture containing 0.05%, 0.15% and 0.5%
chrysene. Although the application of a "mixed" solution may seem to confound any results,
a control group receiving 10 ul per egg of the petroleum hydrocarbon mixture minus any
chrysene concentration was maintained. Hoffman and Gay (1981) documented a significant
reduction of embryonic growth and an increased incidence of abnormal survivors for
eggshells exposed to 0.15% chrysene. The 0.05% dose was inferred as the NOAEL, since
this dose did not result in a further reduction of mallard embryo survival.
There are particular concerns associated with using the Hoffman and Gay (1981) study to
extrapolate an avian benchmark, particularly with respect to. the application method and one-
time dosing regime. The laboratory application of pipetting the chrysene mixture, directly onto
the eggshell surface corresponds to a worst-case wildlife exposure scenario involving the
same absorption rate, exposure during the same critical life stage, and direct contact between
the egg and chrysene-contaminated soil. Thus, the study provides a very conservative
estimate of the effects of chrysene exposure on mallard eggs in the environment. Derivation
of a benchmark value is geared to the modeling of a chronic or multiple-day critical life stage
dose, not a single, one-time dose to an egg. Consequently, avian benchmark values for the
freshwater ecosystem were not derived.
Fish and aquatic invertebrates'. A review of the litereature revealed that AWQC and
adequate toxicity data focusing on critical endpoints pertinent to fish and aquatic invertebrate
population sustainability were not identified for chrysene. There is insufficient data and, thus
no benchmark was derived.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Adequate data for the development of benchmarks for chrysene were not identified in Suter
and Mabrey (1994) or in AQUIRE,
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
August 1995
-------
APPENDIX B
Chrysene - 3
octanol-carbon partition coefficient (K^.) to determine protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from harmful effects from chemical
exposure. No FCV or SGV was reported for chrysene and, therefore, no benchmark was
developed.
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Repraeentatiw
Specie*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark Value*
mg/kg-day
. ID
ID
ID
ID
ID
ID _
ID
ID
ID
ID
Study
Spedee
-
-
-
-
-
.
-
-
-
. -
Effect
-
-
-
-
-
-
-
-
-
Study Value
rng/kg-day
-
-
-
-
-
-
-
-
-
Description
-
-
-
_
.
-
. -
-
'
SF
-
-
-
-
-
-
-
-
-
-
Original Source
.
-
-
-
.
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = intermim; a '*' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B
Chrysene - 4
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Rtprwtntdiw
SpiOlM
fish and
aquatic
invertebrates
aquatic plants
bentfiic
community
BwicfvMrii
VtflM
mgfl
ID
No data
10
StudySpcdM
.
-
.
Description
.
-
.
Origin!
Soum
.
-
II.
'Benchmark .Category, a = adequate, p = provisional, i = intermim; a "*" indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other advene effects.
ID = Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem. . - .
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As discussed previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic toxicity studies were found for mammalian wildlife exposure to
chrysene. Since no additional laboratory mammal studies focusing on reproductive or other
critical endpoints were identified, a mammalian benchmark for terrestrial ecosystems was not
derived.
August 1995
-------
APPENDIX B Chrysene-5
Birds: For reasons previously mentioned in the freshwater ecosystem discussion, the
Hoffman and Gay (1981) studywas not used to calculate a benchmark value for avian species.
No additional toxicity studies documenting avian exposure to chrysene were identified and,
therefore avian benchmarks for the terrestrial ecosystem were not developed.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for chrysene and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Chrysene - 6
Table 3. Toxicological Benchmarks for Representative Mammals
and Birds Associated with Terrestrial Ecosystem
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plants
soil community
BwieimMfk Valur
m0kg4ay
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
No data
No data
Study SpodM
-
-
-
-
-
-
-
-
-
-
-
-
-
- '
cflMt
-
-
-
-
-
-
-
.
-
' -
-
'
-
-
Study V«ta»
utgflig-dtr
-
-
-
-
-
-
-
-
-
-
-'
-
-
. -
Description
-
-
-
-
-
.
-
-
-
-
-
-
-
8F
-
.
-
-
-
-
-
-
-
-
-
-
-
Origin*) Sourca
.
-
-
-
-
. -
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = intermim; a '" indicates that the benchmark value was an
order of magnitude or more above Hie NEL or LEL for other adverse effects.
ID = Insufficient Data
III. Biological Uptake Measures
This section presents the biological uptake measures (i.e., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcnetrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: tropic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertbrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are deignated with a "d" if the value reflects dissolved water
August 1995
-------
APPENDIX B Chrysene - 7
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K^ values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log K^, values above 4, the BCFs
were assumed to refer to total water concentrations and concentrations in fish. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values presented in Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. However, these models were considered inappropriate to estimate BAF/s for
chrysene because they fail to consider metabolism in fish. A number of studies have
demonstrated that polycyclic aromatic hydrocarbons (PAHs) such as chrysene are readily
metabolized in the tissue of fish (see Polycyclic Aromatic Hydrocarbon Hazards to Fish,
Wildlife, and Invertebrates: A Synoptic Review. U. S. Fish and Wildlife Service Biol. Rep.
85[1.11]. The BAF/s selected for fish in the limnetic and littoral ecosystems for chrysene are
from Stephan (1993). This document contains unpublished field data by Burkard with
predicted BAPs of 17 to 228 for four PAHs for fish with 5% lipids. Steady-state measured
data on biological uptake of chrysene (and most PAHs) are very limited at this time and
should be interpreted with caution. Since no measured fish BCf values were identified, the
fish BAF reported by Stephan (1993) was used for bioconcentration factor for fish.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, earthworms and
terrestrial invertebrates were estimated as described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobic organic chemicals assuming that, the
partitioning to tissue is dominated by lipids. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
1
APPENDIX B
Chrysene - 8
Table 4. Biological Uptake
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
. invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BCF
BCF
BCF
UpkMMMdor
wboksbody
lipid
lipid
lipid
lipid
lipid
-
whole-body
whole-body
whole-body
whole-plant
value
800 (t)
800 (t)
800 (t)
800 (t)
800(0
ID
6.7 E-03
6.5E-03
5.2E-02
1.9E-02
source
measured; Stephan. 1993
measured; Stephan, 1993
Based on a geometric mean of
field BAF for 2 ring PAHs
(Stephan. 1993)
measured; Stephan. 1993
measured; Stephan. -1993
-
calc
calc
calc
U.S. EPA. I990e
d = refers to dissolved surface water concentration
t = refers to total surface water concnetration
ID = Insufficient Data
August 1995
-------
APPENDIX B Chrysene-9
References
AQUIRE (AQUatic Toxicity Information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN. '
Eastmond, D. A., G. M. Booth, and M. L.Lee, 1984. Toxicity, accumulation, and elimination
of polycyclic aromatic sulfur heterocycles in Daphnia magna. Arch. Environ. Contam.
Toxicoi, 13(1): 105-111.
Hoffman, D. J. and M. L. Gay, 1981. Embryotoxic effects of benzo(a)pyrene, chrysene, and
7,12-dimethylbenz(a)anthracene in petroleum hydrocarbon mixtures in Mallard ducks.
Journal of Toxicology and Environmental Health, 7:775-787.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. October 1994.
/ '
National Library of Medicine. HSDB (Hazardous Substance Database). 1994.
Newsted, J. L. and J. P. Giesy, 1987. Predictive models for photoinduced acute toxicity of
polycyclic aromatic hydrocarbons to Daphnia Magna, Strauss (Cladocera, Crustacea).
Environmental Toxicology and Chemistry, Vol. 6, pp. 445-461.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research Laboratory,
Office of Research and Development, Duluth, MN. PB93-154672.
Suter, G. W. Et, and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of
Potential Contaminants of Concenr for Effects of Aquatic Biota: 1994 Revision. DE-
AC05-84OR21400; Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
Thomann, R. V. 1989. Bioaccumulation model of organic chmeical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6): 699-707.
Thomann, R. V., J. P. Connolly, .and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulationin aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:61.5 - 629,
August 1995
-------
APPENDIX B Chrysene-10
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment. Washington, D. C. January.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ERfTM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxicity - Chrysene
Cos No.: 218-01-9
Chemical Name
chrysene
chrysene
Species
mouse
mallard duck
Endpolnt
acute
embryotoxic
Description
LD50
NOAEL
Value
>320
0.09
Units
mg/kg
mg/kg-egg wt.
Exposure
Route (oral,
S.C., I.V., l.p.,
Injection)
i.p.
applied to
eggshell
surface
Exposure
Duration /
Tlmlna
NS
1 -18 days of
incubation
Reference
RTECS. 1994
Hoffman and Gay,
1981
Comments
-
Study doses were 0.005. 0.015 and 0.05 ug/egg (equivalent
to 0.09, 0.27, 0.9 ug/kg fresh weight), significant reduction of
embryonic growth and an increased incidence of abnormal
survivors.
NS = Not Specified
-------
Freshwater Biological Up.-ste Measures - Chrysene
Cos No.: 218-01-9
Chemical Name
chrysene
chrysehe
chrysene
Species
aquatic organisms
Daphnia magna
Daphnia maqna
B-factor (BCF,
BAF. BMP)
BCF
BCF
BCF
Value
1,762
6,088.40
5,500
Measured 01
Predicted
(m,p)
P
P
p
Units
NS
NS
NS
Reference
U.S.EPA, 1993b
Newsled & Geisy, 1987
Eastmond et al., 1984
Comments
BCF normalized to 1 % lipid
.
Maximum predicted BCF is reported.
NS = Not Specified
-------
Terrestrial Biological Uptake Measures - Chrysene
Cos No.: 218-01-9
Chemical Name
-
chrysene
Species
plant
B-f actor
(BCF. BAF.
BMR
BCF
Value
0.019
Measured
or
Predicted
(m.o)
p
Untts
(ug/g DW
plant)/(ug/g
soil)
Reference
U.S. EPA, 1990e
Comments
Plant uptake from soil pertains to
foraaed plants
-------
APPENDIX B Copper - 1
Toxicological Profile for Selected Ecological Receptors
Copper
Cas No.: 7440-50-8
Summary: This profile on copper summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For .the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule.
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for Ihe generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated wilh ihe freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: One suitable chronic study, documenting mammalian wildlife exposure to copper, was
identified. Developmenlal endpoinis were investigated in mink mating pairs fed a diel of 25, 50,
100 or 200 ppm copper for 357 days (Aulerich el al., 1982). Allhough no adverse effecis were
see.i at the lowesl dose, increased mortality of offspring from birth to 4 weeks occurred in the
group given 50 ppm. Therefore, a NOAEL of 25 ppm and a LOAEL of 50 ppm were inferred
based on these developmental effects in young mink. Since the mink's food consumption was
not provided, an allometric equation was required lo estimate daily food intake:
Food consumption = 0.235(W°'822) where W is body weight in grams (Nagy, 1987).
August 1995
-------
APPENDIX B DDT - 1
Toxicological Profile for Selected Ecological Receptors
DDT
Cas No.: 50-29-3
Summary: This profile on DDT summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. However, several chronic and subchronic toxicity
studies involving DDT and its metabolites (i.e., DDE and/or DDD) have been conducted using
laboratory rats. Oral acute toxicity values presented in the Great Lakes Initiative (EPA, 1993b)
demonstrate that the rat is among the most sensitive of the mammalian species tested for the
acute effects of DDT. Liver toxicity was observed in a subchronic study (Mitjavila et al., 1981)
in which male rats were administered DDT at 14.5 mg/kg-day by gavage for a period of 52 days.
Liver toxicity was also observed in rats fed DDT at concentrations of 0, 1, 5, 10, or 50 ppm for
1-27 weeks, and a NOAEL of 0.05 mg/kg-day (1 ppm) was estimated (Laug et al., 1950). A
chronic reproductive study was identified in which rats were fed a diet that contained 0, 10, 50,
100, and 160 ppm DDT for a period of two years (Fitzhugh, 1948). Fitzhugh (1948) observed
sw^^
August 1995
-------
Terrestrial Biological bt ..*e Measures - Copper
Cas No. 7440-5O8
Chemical
Name
copper
Species
plant
B-factor
(BCF, BAF.
BMP)
BCF
Value
0.9
Measured
or
Predicted
.-JGIPL.
P
units
(ug/gDW
plant)/(ug/g soil)
Reference
U.S. EPA, I990e
Comments
-------
APPENDIX B DDT-2
the number of litters,-number of live young at birth, average weight at birth, and the number of
young surviving the weaning period and reported a NOAEL of 10 ppm for reproductive effects.
The 10 ppm level was convened a daily dose of 0.815 mg/kg, assuming an average female body
weight of 0.33 kg (U.S. EPA, 1988) and a food consumption rate based on the following
equation:
Food intake = 0.056W0'6611, where W is body weight in kg (U.S. EPA, 1988).
The NOAEL in Fitzhugh's study was chosen to derive the toxicological benchmark because (1)
it focused on reproductive toxicity as a critical endpoint and (2) chronic exposures were
administered via oral ingestion. The study by Mitjavila et al., (1981) was not selected because
of insufficient dose-response data and it did not evaluate reproductive or developmental
endpoints. Similarly, the study by Laug et al., (1950) was not selected for the derivation of a
benchmark because the study was subchronic (27 weeks) and did not evaluate reproductive or
developmental endpoints. Nevertheless, these studies illustrate the dose ranges at which DDT
toxicity occurs.
Based on the NOAEL of 0.815 mg/kg-day in the Fitzhugh (1948) study, the benchmarks for
representative freshwater mammals were estimated using a cross-species scaling algorithm
adapted from Opresko et al. (1994).
( b\v V4
Benchmark = NOAEL. x L
VKJ
This is the same methodology the EPA uses in carcinogenicity assessments and reportable
quantity documents for adjusting animal data to an equivalent human dose. Since the Fitzhugh
(1948) study documented reproductive effects to weanling rats from DDT exposure via maternal
transfer, female body weights for each representative species were used in the scaling algorithm
to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental, effects of DDT, as well as chronic
survival. In addition, the data set contained studies which were conducted over chronic and
subchronic durations and during sensitive life stages. There were several study values in the data
set which were lower than or approximately equal to the benchmark value. These values
corresponded to effects on behavioral and hepatic endpoints. All of the studies identified were
conducted using laboratory rats or mice and as such, inter-species differences among wildlife
species were not identifiable. Therefore, an inter-species uncertainty factor was not applied.
Based on the data set for toxaphene and because the NOAEL for reproductive and
developmental effects was at least an order of magnitude above the lowest NOAEL for
nonreproductive or nondevelopmental effects, the benchmarks for DDT were categorized as
adequate*.
Birds: Several studies were identified on mallards, kestrels, and pelicans which focused on the
reproductive effects of DDT and/or DDE and DDD. Studies on both DDT and its metabolites .
were considered for deriving toxicological benchmarks for birds in freshwater ecosystems. Heath
August 1995
-------
APPENDIX B DDT - 3
et al. (1969) demonstrated significant reduction in eggshell thickness for mallards ingesting DDT
and/or DDE (10 to 40 ppm in feed) for a period ranging from 5 weeks prior to egg laying
through two years. Heath et al., (1969) observed the following reproductive endpoints in a
dietary study in female mallards: percent cracked eggs, embryo mortality, hatchling survivability,
and number of ducklings per hen. A LOAEL of 0.58 mg/kg-day (10 ppm) was recorded for
DDE; a LOAEL of 1.45 mg/kg-day (25 ppm) and a NOAEL of 0.58 mg/kg-day (10 ppm) were
recorded for DDT. The differences in potency observed by Heath et al., (1969) were DDE >
DDD > DDT. Similar eggshell thinning was observed at 20 ppm (1.16 mg/kg-day) and no effect
was observed at 2 ppm (0.116 mg/kg-day) after dietary administration of DDT to female mallards
(Davison and Sell, 1974). In American kestrels, Peakall et al. (1973) measured eggshell
thickness, breaking strength, and permeability after dietary administration of 3, 6, and 10 ppm
of DDE. Significant effects on each of these endpoints were observed at the lowest dietary
concentration. Using a default female kestral body weight of 127 g (U.S. EPA, 1993a) and a
daily food intake rate of 0.037 kg/d, the LOAEL determined for this study is 0.87 mg/kg-d (3
ppm). Anderson et al. (1975) studued the reproductive success of brown pelicans off the coast
of California in the early 1970's. During the period of observation, combined concentrations of
DDT, DDD, and DDE in the food source steadily declined from 4.27 ppm (wet weight) to 0.15
ppm. At 0.15 ppm, Anderson et al. (1975) determined that the fledgling rate was 30 percent
below the rate necessary to maintain a stable population. Based on these findings, a LOAEL of
0.15 ppm (0.028 mg/kg-day) for total DDT for reproductive success (U.S. EPA, 1993b).
The study by Peakall et al. (1973) was judged to be the most appropriate for the derivation of
a benchmark value for avian species representing the freshwater ecosystem. Although the pelican
study was conducted in the field, Anderson et al., (1975) noted that the pelicans were also
exposed to PCBs, lead and mercury at consistent levels throughout the life of the study. This
is a significant factor since recent findings suggest that low-levels of chemicals that impact the
endocrine system (e.g., PCBs, lead, and mercury) may impair reproductive success and, possibly,
act in an additive or synergistic manner (Colbom et al., 1993; Thomas and Colbom, 1992). The
Peakall et al. (1973) study on kestrels was selected to derive the DDT lexicological benchmark
for birds because: (1) confounding exposures to other endocrine disrupters (besides DDT) were
not addressed in the pelican study, (2) the kestrel is a representative species and is taxonomically
similar to the eagle, osprey and hawk, and (3) the pelican typically dwells in coastal areas and
is more representative of marine or .estuarine birds than birds associated with the generic
freshwater or terrestrial ecosystems.
The LOAEL of 0.87 mg/kg-day (3 ppm) from the Peakall et al. (1973) was selected, as detailed
above, for benchmark derivation. A LOAEL-to-NOAEL safety factor of 10 was applied to
account the uncertainty between the study LOAEL and the desired NOAEL. The principles for
allometric scaling were assumed to apply to birds, although specific studies supporting allometric
scaling for avian species were not identified. Thus, for the avian species representative of a
freshwater ecosystem, the LEL/10 of 0.087 mg/kg-day from the Peakall et al. (1973) study was
scaled using the cross-species scaling method of Opresko et al. (1994).
Data were available on the reproductive and developmental, effects of DDT, as well as growth
or chronic survival. In addition, the data set contained studies which were conducted over
August 1995
-------
APPENDIX B DDT - 4
chronic and subchronic durations and during sensitive life stages. There were no study values
in the data set which were more than an order of magnitude lower than the benchmark value.
Due to the limited range of species which were tested , inter-species differences among wildlife
species could not be identified. Therefore, an inter-species uncertainty factor was not applied.
Based on the data set for DDT and because the benchmarks are based on a LEL/10, the
benchmarks were categorized as provisional.
Fish and aquatic invertebrates: Based on the Final Residue Value (FRY), the AWQC for DDT
is 1.0 E-6 mg/1 (57 FR 60911). The FRY was not considered to be appropriate for the
development of a benchmark for daphnids because it is intended to protect fish and other wildlife,
which consume aquatic organisms, from the adverse effects of chemicals that may bioconcentrate.
Also, the FRY was not an appropriate benchmark value because residues and bioaccumulation
are already taken into account by the Thomann et al. (1992) model. Instead of the FRY, the
Secondary Chronic Value (SCV) was selected as the benchmark protective of daphnids. The
SCV, as calculated by GLI (1992), is 1.3 E-5 mg/1. Because the benchmark is based on an SCV
developed for the Great Lakes Initiative, the benchmark is categorized as interim.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no observed
-effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). The
benchmanrk for aquatic plants was reported as 3.0 E-4 mg/1 for growth and morphology effects
(Suter and Mabrey,.1994). As described in Section 4.3.6, all benchmarks for aquatic plants were
designated as interim.
Benthic community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbbn partition coefficient (K,,,.) to determine a chemical concentration that may be present in
the sediment while still protecting the benthic community (Stephan, 1993). The EQp number is
the best recommendation of a chemical concentration that may be present in the sediment while
still protecting the benthic community from harmful effects resulting from possible chemical
exposure. The Secondary Chronic Value (SCV) of 1.3 E-5 mg/1 for fish and aquatic invertebrates
FCV developed for the AWQC. This value was used to calculate an EQp number of 81.9 mg
DDT /kg organic carbon. Assuming a mass fraction of organic carbon for the sediment (f^ of
0.05, the benchmark for the benthic community is 4.1 mg/kg. Since the EQp number was based
on a SCV and not an FCV, the sediment benchmark is categorized as interim.
August 1995
-------
APPENDIX B
DDT-5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ffoprMentaiiv*
9p*oi+»
mink
river otter
bald eagle
osprey
groat blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
V«!U** mg/kB"
<>«y
0.67 (a')
0.38 (a*)
0.04 (p)
, 0.05 (p)
0.04 (p)
0.05 (p)
0.06 (p)
0.1 1(p)
0.05 (p)
0.08 (p)
Study
Sptcif*
rat
rat
kestrel
kestrel
kestrel
kestrel
kestrel
kestrel
kestrel
kestrel
Effect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study Value
me'ks-d*?
0.815
0.815
0.87
0.87
0.87
0.87
0.87
0.87
0.87
0.87
Dwcrtptiofl
NOAEL
NOAEL
. LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
*
\-.
-
-
10
10
10
10
10
10
10
10
OrtgNHSovw*
Pitzhugh, 1948
Fitzhugh, 1948
Peakall et al.,
1973
Peakall etal.,
1973
Peakall etal.,
1973
Peakall etal.,
1973
Peakall et al..
1973
Peakall et al.,
1973
Peakall etal.,
1973
Peakall etal..
1973
Benchmark Category, a = adequate, p * provisional, i = interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
DDT-6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
{topr«*entatlv«
SfMQlO*
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
V«Ju»*
IBflfl
1.3E-05(i)
3.0E-04 (i)
4.1 (i) mg/kg
sediment
Study Sp«cfe*
AWQC species
aquatic plants
AWQC species
D«»cflpUon
SCV
cv
SCV
OfiQtavf Seure*
GLI, 1992
Suter and Mabrey,
1994
GLI. 1992
IL
Benchmark Category, a - adequate, p « provisional, i = interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to DDT and its
metabolites. Because of the lack of additional mammalian toxicity studies, the same
surrogate-species study (Fitzhugh, 1948) was used to derive the DDT toxicological benchmark
for mammalian species representing the terrestrial ecosystem.
Based on the NOAEL of 0.815 mg/kg-day in the Fitzhugh (1948) study, the benchmarks for
representative terrestrial ecosystem mammals were estimated using a cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the Fitzhugh (1948) study documented
reproductive effects to weanling rats from DDT exposure via maternal transfer, female body
weights for each representative species were used in the scaling algorithm to obtain the
toxicological benchmarks.
Data were available on the reproductive and developmental, effects of DDT, as well as
chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. There were several study
values in the data set which were lower than or approximately equal to the benchmark value.
These values corresponded to effects on behavioral and hepatic endpoints. All of the studies
August 1995
-------
APPENDIX B DDT - 7
identified were conducted using laboratory rats or mice and as such, inter-species differences
among wildlife species were not identifiable. Therefore, an inter-species uncertainty factor
was not applied. Based on the data set for DDT and because the NOAEL for reproductive
and developmental effects was at least an order of magnitude above the lowest NOAEL for
nonreproductive or nondevelopmental effects, the benchmarks for DDT were categorized as
adequate*.
Birds: No additional avian toxicity studies were identified for species representing the
generic terrestrial ecosystem. Therefore, the LOAEL of 0.87 mg/kg-day (3 ppm) from the
Peakall et al. (1973) was selected for benchmark derivation. A LOAEL-to-NOAEL safety
factor of 10 was applied to account the uncertainty between the study LOAEL and the desired
NOAEL. As was the case for birds in the freshwater ecosystem, allometric scaling of the
LOAEL/10 was performed using the method of Opresko et al. (1994).
Data were available on the reproductive and developmental, effects of DDT, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. These values
corresponded to effects on behavioral and developmental endpoints. Due to the limited range
of species which were tested , inter-species differences among wildlife species could not be
identified. Therefore, an inter-species uncertainty factor was not applied. Based on the data
set for DDT and because the benchmarks are based on a LOAEL/10, the benchmarks were
categorized as provisional.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for DDT and, as a result, a benchmark
could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
DDT-8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
ft*pf«*«ntaiiv*
Sp«clM
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plants
soil community
Benchmark
VaKM* mgftfi*
-------
APPENDIX B DDT - 9
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, ,bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
Because the log Kow for TCDD is above 6.5 (i.e., 6.91), the Thomann (1989) and Thomann et
al., (1992) models were not used to estimate bioaccumulation factors. For extremely
hydrophobic constituents, the Agency has stated that reliable measurements of ambient water
concentrations (especially dissolved concentrations) are not available and that accumulation of
these constituents in fish or other aquatic organisms cannot be referenced to a water
concentration as required for a BCF or BAF (U.S. EPA, 1993i). However, extremely
hydrophobic constituents can be measured in sediments and aquatic life and, because these
chemicals tend to partition to lipids and organic carbon, a biological uptake factor that reflects
the relationship between sediment concentrations and organism concentrations may be more
appropropriate. Consequently, the biota-sediment accumulation factor (BSAF) is the preferred
metric for accumulation in the littoral aquatic ecosystem for extremely hydrophobic chemicals
(e.g., chemicals with > log Kow of ~ 6.5). Unfortunately, a BSAF for DDT was not identified
in the literature. Therefore, the lipid-based bioaccumulation factors for fish and invertebrates
in the limnetic ecosystem were taken from the Great Lakes Water Quality Initiative Technical
Support Document for the Procedure to Determine Bioaccumulation Factors - July 1994 (U.S.
EPA, 1994b). The document indicated that the basis for the DDT BAF;d s was unpublished
work by Burkhard (1994) in which log BAFs were calculated from measured values for
sculpin, alewives, and small smelt for trophic level 3. For trophic level 4, Burkhard
calculated a log BAF (a BAF/1) from measured values corresponding to a log Kow of about
6.59. Although the BAF,d calculated by Burkard did not correspond to the log Kow used for
DDT in this analysis, the "measured" BAF,ds were considered to be the most appropriate
values for bioaccumuation of DDT. In addition, subsequent analyses of log K,,w data on DDT
suggest that 6.91 may be too high and that a more reasonable estimate is probably 6.5, the
geometric mean of values estimated using the slow-stir technique (Karickhoff and Truesdale,
unpublished data). As with toxaphene, the same BAF^s that were used for trophic levels 3
August 1995
-------
APPENDIX B DDT - 10
and 4 in the limnetic-ecosystem were considered appropriate for the littoral ecosystem,
although some differences in food chain transfer are likely.
The biocohcentration factor (BCF/) for fish was estimated as the geometric mean of two
measured values presented in Stephan (1993). The geometric mean BCF/ of 433,900 is
approximately a factor of 3 higher than the BCF/ estimated from the BCF* = log Kow
relationship and adjusted for the dissolved fraction (/j) as defined in Equation 6-21 (assuming
log K^ ~ 6.6). Nevertheless, the difference between the two values was considered
relatively insignificant given the inherent uncertainties in BCF measurement and modeling
techniques.
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of a number of
measured values with sources shown in Table 4 (see master table). For terrestrial
invertebrates, the bioconcentration factor was estimated as described in Section 5.3.5.2.3.
Briefly, the extrapolation method is applied to hydrophobic organic chemicals assuming that
the partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs
and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard.. The beef biotransfer factor (BBFs) for a
chemical lacking measured data is compared to the BBF for TCDD and that ratio (i.e.,
pentachlorobenzene BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial
vertebrates, invertebrates, and earthworms, respectively. For earthworms, a measured BCF
value from Beyer and Gish (1980) was selected. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
DDT - 11
Table 4. Biological Uptake Properties
ecological!
- receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
lipid-bas«d or
whole-body
lipid .
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
100,000,000 (d)
53,700,000 (d)
433,900 (t)
100,000,000 (d)
53,700,000 (d)
-
0.82
0.097
0.26
0.0039
»ourc«
measured value from Cook,
1994 as cited in U.S. EPA.
1994b
measured value from Cook.
1994 as cited in U.S. EPA,
1994b
geometric mean of measured
values in Stephan, 1993
same value as in the limnetic
ecosystem
same value as in the limnetic
ecosystem
insufficient data
geometric mean of measured
values (e.g., Garten and
Trabalka. 1983; Clabom et.al.,
1956, 1960 u chad in
Kenaga, 1980)
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
measured value from Beyer
and Gish, 1980
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t 3 refers to total surface water concentration
August 1995
-------
APPENDIX B DDT-12
References
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TCDD and TCDF in Pulp and Paper Sludge. Prepared for Qssi Meyn, U.S.
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^
August 1995
-------
APPENDIX B DDT - 13
Davison, K.L., and J.L. Sell. 1972. Dieldrin and p'p'-DDT effects on egg production and
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20.
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Heinricks, W.J., R.J. Gellert, J.L. Bakke, and N.L. Lawrence. 1971. DDT administered to
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August 1995
-------
APPENDIX B DDT - 14
Jarvinen, A.W., M.J. Hoffman, and T.W. Thorslund. 1977. Long-term toxic effects of DDT
food and water exposure on fathead minnows (Pimephales promelas). J. Fish. Res. Bd.
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Kolaja, G.J. 1977. The effect on DDT, DDE and their sulfonated derivatives on eggshell
formation in the mallard duck. Bull. Environ. Contain. Toxicol. 17(6):697-701.
Laug, E.P., A. Nelson, G. Gitzhugh, and F. Kunze. 1950. Liver cell alteration and DDT
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Environmental Research Laboratory, Office of Research and Development, U.S.
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AQUIRE (AQUatic Toxicity /nformation /?£trieval Database). 1995 Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
McCain, B.B., S.L. Chan, M.M. Krahn, D.W. Brown, M.S. Myers, J.T. Landahl, S. Pierce,
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MitjavUa, S., G. Carrera, R.-A. Boigegrain, and R. Derache. 1981. Evaluation of the toxic
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469.
Mitral, P.K., H.C. Agarwal, and M.K. Pillai. 1980. Tolerance, Uptake, and Metabolism of
DDT by the Freshwater Flea Simocephalus sp. (Cladocera). Indian J. Exp. Biol. 18(11):
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1995. Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN.
August 1995
-------
APPENDIX B DDT - IS
N1OSH (National Institute for Occupational Safety and Health). 1992. General Toxicity File
for DDT in Registry of Toxic Effects of Chemical Substances (RTECS database).
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Water Quality Initiative Criteria Documents for the Protection of Wildlife (Proposed)
DDT; Mercury; 2.3,7,8-TCDD; PCBs, EPA-822-R-93-007, Office of Science and
Technology, Office of Water, Washington, DC.
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14:74-81.
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multiple generations of beagle dogs. Arch. Environ. Contam. Toxicol. 6:83-101.
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MN.
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egg production, mortality, fertility, hatchability and pesticide content of yolks in Japanese
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August 1995
-------
APPENDIX B DDT - 16
Smith, V.E., J.M. Spurr, J.C. Filkins, and J.J. Jones. 1985. Organochlorine contaminants of
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i
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Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
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August 1995
-------
APPENDIX B DDT - 17
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Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
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of o,p'-DDT on reproduction and lactation in the rat. Bull. Environ. Contam. Toxicol.
6:471-479.
August 1995
-------
Freshwatci /icity - DDT
Cas No. 50-29-3
Chemical
Name
DDT
Species
fathead
minnow
Endpolnt
mort.
'-
Description
MATC
Value
0.36-1.5
Units
ug/L
Test type
(static/ flow
through)
complete life
cycle test
Exposure
Duration/
Timing
NS
Reference
Jarvinen at al.. 1977
Comments
early juvenile; mortality
and hatchability
NS = Not specifed
-------
Freshwater Biological Uptake Measures - DDT
Cas No. 50-29-3
Chemical Name
DDT
DDT
DDT
DDT
DDT
Species
carp
salmon ids
white croaker
fish
fish
B-factor
(BCF, BAF,
BMP)
BSAF
BSAF
BSAF
BAF
BAF
Value
0.84
1.21
14.90
30,903
1,913.862
Measured or
predicted
(m,p)
P
P
P
P
p
Units
ug/g
ug/g
ug/g
Ukg
NS
Reference
Smith etal., 1985
Oliver and Niimi, 1988
McCain etal., 1992
Garten and Trabalka, 1 983
Stephan. 1993
Comments
Microcosm; All estimates were
calculated based on published data.
the type of studies from which the
data were taken were not specified.
Normalized to 5.0% lipid. Trophic level
4 fish
NS = Not specified
-------
Terrestrial .Jclty - DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
/
DDT
DDT
DDT
DDT
DDT
Species
bullfrog
mallard
California
quail
Japanese
quail
pheasant
Sandhill
crane
rock dove
Iroq
Endpolnt
mod
mort
mort
mort
mort
mort
mort
mort
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
72000
72240
595
841
1334
71200
74000
7600
Units
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
bodv wt.
Exposure
Route (oral,
s.c., l.v., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/
Timing
NS
NS
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1993c
U.S. EPA, 1993C
U.S. EPA, 1993c
U.S. EPA, 1993c
U.S. EPA, 1993C
U.S. EPA, 1993C
U.S. EPA, 1993c
RTECS. 1994
Comments
NS = Not specified
-------
Freshwater Toxicity DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
DDT
DDT ,
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
Species
aquatic
organisms
aquatic
organisms
ish
daphnids
fish
Daphnia
carinata
Daphnia
maqna
Daphnia
magna
Daphnia
maqna
Daphnia pulex
Simocephalus
serrulatus
Simocephalus
sp.
Brook trout
Northern pike
channel
catfish
bluegill
striped bass
rainbow trout
fathead
minnow
Endpolnt
chronic
chronic
chronic
chronic
chronic
immob.
immob.
immob.
rep
immob.
immob.
immob.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
AWQC
SCV
CV
CV
EC20
EC50
EC50
EC50
EC50
EC50
EC50
EC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
Value
0.001
0.04
0.73
0.016
0.35
12
0.68 - 4.0
(141)
0.67
0.50 - 0.75
(0.58)
0.36 - 2.67
(1.04)
2.5-2.8
(2.65)
5.8
1 .8 - 20.0
(8.55)
1.7
3.3-17.5
(11.85)
1.2-16.0
(4.95)
0.53
1.5-18.0
(7.16)
8.5 - 45
(19.42)
Units
ug/L
ug/l
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration/
Timing
NS
NS
NS
NS
NS
48 hour
48 hour
14 day
14 day
48 hour
48 hour
48 hour
96 hour
96 hour
96 hour
96 hour
96 hour
96 hour
4 days
Reference
57 FR 60848
Suter and Mabrey, 1 994
Suter and Mabrey, 1 994
Suter and Mabrey, 1 994
Suter and Mabrey, 1 994
Santharam et al., 1976 as
cited in AQUIRE, 1995
AQUIRE, 1995
Maki et al., 1975 as cited in
AQUIRE, 1995
Maki etal., 1975 as cited in
AQUIRE, 1995
AQUIRE, 1995
Sanders et al., 1966 as
cited in AQUIRE, 1995
Mittal et al., 1980 as cited in
AQUIRE, 1995
AQUIRE, 1995
Marking, 1966 as cited in
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE. 1995
Comments
-------
Terrestrial .icily - DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
Species
dog
monkey
cat
rabbit
guinea pig
hamster
rat
rabbit
guinea pig
rat
mouse
Endpolnt
mort
mort
mort
mort
mort
mort
mort
mort
mort
mort
mort
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
150
200
250
250
150
>5
1931
300
1000
0.91
32
Units
mg/kg-
bodywt.
mg/kg-
body wt.
mg/kg-
body wl.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
Exposure
Route (oral,
8.C., I.W., l.p.,
Infection)
oral
oral
oral
oral
oral
oral
dermal
dermal
dermal
i-P.
i.p.
Exposure
Duration/
Timing
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
19936
NIOSH. 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA.
1993b
NIOSH. 1992 as
cited in U.S. EPA,
1993b
Comments
-------
Terrestrial Toxicity - DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
Species
rat
rabbit
guinea pig
rat
mouse
dog
monkey
cat
rabbit
rat
mammal
Endpolnt
mort
mort
mort
mort
mort
mort
mort
mort
mort
mort
mort
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
1500
250
900
68
6.85
150
50
40
50
300
200
Units
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wl.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
Exposure
Route (oral,
S.C., I.V., l.p.,
Inlectlon)
s.c.
s.c.
s.c.
i.v.
i.v.
i.v.
i.v.
i.v.
i.y.
NS
NS
Exposure
Duration/
Timing
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
NIOSH, 1992 as
cited in U.S. EPA.
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
19935
NIOSH, 1992 as
cited in U.S. EPA.
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH. 1992 as
cited in U.S. EPA,
1993b
NIOSH. 1992 as
cited in U.S. EPA.
1993b
NIOSH, 1992 as
cited in U.S. EPA,
1993b
NIOSH. 1992 as
cited in U.S. EPA,
1993b
NIOSH. 1992 as
cited in U.S. EPA,
1993b
Comments'
-------
Terresiriat .icity - DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT
Species
rat
ral
rat
rats
rats
rat
rats
rats
mice
mice
dogs
rats
Endpolnt
hepatic
hepatic
hepatic
rep
rep
rep
rep
dev
rep
rep
rep
rep
Description
LOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
AEL
NOAEL
LOAEL
NOAEL
NOAEL
Value
14.5
0.25
0.05
4.07
0.615
14.2
200
1
25
100
10
9.04
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-diet
mg
ppm
ppm.
mg/kg-day
mq/kq-dav
Exposure
Route (oral,
8.C., I.V., l.p.,
Inlectlon)
oral
oral
oral
oral
oral
oral
oral
s.c.
oral
oral
oral '
oral
Exposure
Duration/
Timing
52 days
1 -27 weeks
1 -27 weeks
0.01 kg/day
0.01 kg/day
2 generations
NS
Days 2, 3, 4 of
post-natal life
6-generations
6-generations
3-generations
2-9 months
Reference
Mitjavila et al., 1981
Laugetal., 1950
Laugetal., 1950
Fitzhugh, 1948
Fitzhugh, 1948
Ottoboni, 1969
Hayes, 1976
Henrichs et al., 1971
as cited in WHO.
1979
Keplingeretal., 1970
Keplinger et at., 1970
Ottoboni et al.. 1977
Wrennetal.. 1971
Comments
Liver toxicity was observed at this
level.
Liver toxicity was observed at this
level.
Liver toxicity was not observed at
this level.
Reproductive effects were reported.
No reproductive effects were
reported.
Rats reproduced normally at this
level.
Rats reproduced normally at this
level.
Abnormal effects were observed at
this level.
No effect on fertility, gestation,
viability, lactation, and survival at this
level.
At this dose, there was a slight
reduction in lactation and survival in
some generations, but the effect was
not progressive.
No effect on gestation, fertility,
success of pregnancy, litter size,
lactation ability of the dams, viability
at birth, survival to weaning, sex
distribution, growth of pups,
morbidity, mortality, or organ/body
weight ratios.
No effect on reproduction.
-------
Terrestrial Toxicity - DDT
Cas No. 50-29-3
Chemical
Name
DDT
DDT
DDT
DDT
DDT
DDT
DDT
DDT.
DDT
DDT
DDT
DDT
DDT
Species
female
mallards
Female
mallards
mallards
mallards
mallards
American
kestrels
brown
pelicans
Japanese
quail
Japanese
quail
white
leghorn
hens
rat
rat
mouse
Endpolnt
rep
rep
rep
rep
rep
rep
rep
rep
rep, mort.
rep
mort
mort
mort
Description
LOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LD50
LD50
LD50
Value
1.16
0.116
2.91
1.45
0.58
0.87
0.028
200
400
200
87
152.3
135
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-diet
mg/kg-diet
mg/kg-diet
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
Exposure
Route (oral,
B.C., I.V., l.p.,
nlectlon)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/
Timing
0.0582 kg/day
0.0582 kg/day
0,0582 kg/day
2 years
2 years
NS
.66 kg/day; 5
yr. study
NS
NS
12 weeks
NS
NS
NS
Reference
Davison and Sell,
1974
Davison and Sell,
1974
Kolajaetal., 1977
Heath et. al., 1969
Heath et. al., 1969
Peakall et al., 1973
Anderson et al., 1975
Smith el al., 1969
Smith el al., 1969
Davison and Sell.
1972
NIOSH. 1992 as
cited in U.S. EPA.
1993b
Mijavila et al , 1981
NIOSH. 1992 as
cited in U.S. EPA.
1993b
Comments
There was a significant reduction in
eggshell thickness at this level.
There was no reduction in eggshell
thickness at this level.
Eggshell thickness and weight were
significantly reduced at this level.
Reproductive success was impaired
at this level.
No effect on reproduction at this
level.
Effects were observed on eggshell
thickness, breaking strength, and
permeability.
Reproductive success was impaired
at this level.
There was no effect on hatchability
or fertility of eggs.
50% mortality among dosed birds;
survivors showed a decline in
hatchability and fertility after 30 days.
No effect on average egg production
per bird,- egg weight, dry shell weight,
shell thckness, and shell calcium.
-------
APPENDIX B Di-n-octyl phthalate - 1
lexicological Profile for Selected Ecological Receptors
Di-n-octyl phthalate
Cas No.: 117-84-0
Summary: This profile on di-n-octyl phthalate summarizes the lexicological benchmarks
and biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms ,
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Di-n-octyl Phthalate
Mammals: Two studies were identified which investigated the effects of di-n-octyl phthalate
exposure to laboratory mammals. Mann et al. (1985) fed male rats di-n-octyl phthalate at a
daily dose of 20000 g/kg-dieL After 3 weeks, the rats exhibited increases in liver size, but no
other signs of toxicity were observed. Another study observed the effects on mice fed 1.25%,
2.50% and 5% dietary di-n-octyl phthalate 7 days prior to mating and throughout a 98-day
mating period (Heindel et al., 1989). No effects on reproductive function were seen at any
administered doses of di-n-octyl phthalate..
Neither of the studies above were considered suitable for derivation of a benchmark value
because of the uncertainty surrounding the critical endpoint. Liver enlargement may affect
the lifespan of an individual organism, but it is unclear as to whether these effects would
impair the fecundity of an entire population. Since adequate toxicity data focusing on critical
August 1995
-------
APPENDIX B Di-n-octyl phthalate - 2
endpoints pertinent to population sustainability were not identified, benchmarks protective of
the mammalian community in a freshwater ecosystem were not derived.
Birds: No subchronic or chronic toxicity studies were identified for di-n-octyl phthalate
exposure to avian species, and therefore, no benchmark was developed.
Fish and aquatic invertebrates: A review of the literature revealed that adequate data with
which to derive a benchmark protective of the fish and aquatic invertebrate community were
not identified.
Aquatic plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Adequate data for the development of benchmarks for di-n-octyl phthalate were not identified
in Suter and Mabrey (1994) or in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined using the
Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value (FCV) or other
chronic water quality measures, along with the fraction of organic carbon and the octanol-carbon
partition coefficient (K^.) to determine a protective sediment concentration that may be present in the
sediment while still protecting the benthic community from harmful effects from chemical exposure
(Stephan, 1993). No FCV or other chronic water quality measures were identified and therefore, no
benchmark was developed.
August 1995
-------
APPENDIX B
Di-n-octyl phthalate 3
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
$MK^MI
mink
river otter
bald eagle
osprey
great MM heron
malard
lesser scaup
potted sandpiper
herring gul
kingfisher
v«ta*'»«*o-
«**
10
ID
ID
ID
ID
ID
ID
ID
ID
ID
Sttrfy
At^^JL^*
^ni^^*^^^»
.
-
ItfMt
-
-
-
-
.
SMoyVfthM
m&*4*t
'
-
^
.-*-
-
.-
«F
+
.
-
0ri0to*ft*tt«* "
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL tor other adverse effects.
ID = Insufficient Data
. Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and aquatic
invertebrates
aquatic plants
benthic community
ID
No data
ID
Study
Specie
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" Indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = insufficient Data
August 1995
-------
APPENDIX B Di-n-octyl phthalate - 4
n. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^ for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem.
Mammals: As discussed in the freshwater ecosystem section, no suitable subchronic or chronic toxicity
studies focusing on reproductive or other critical endpoints were identified for di-n-octyl phthalate.
Thus, mammalian benchmarks protective of the terrestrial ecosystem were not derived.
Birds: No avian toxicity studies were identified and therefore, benchmark values were not derived.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root lengths. As presented in Will and Suter (1994), phytoioxicity benchmarks were selected
by rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or
fewer values for a chemical, the lowest LOEC was used. If there were more than 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or
other effects reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for di-n-ocyl
phthalate and,'as a result, a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Di-n-octyl phthaJate 5
Table 3. Toxicologjcal Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
4MpfWMMMHf9>
flpicjjf
deer mouse
short-tailed
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
.. _,
red- tailed hawk
American kestrel
Northern
bODOWrMte
American robin
American
plants
toil community
ft nil !*%
BeaCnRiaiK
₯**«*
£^*JftMi|_4t||M-
^^F^T^^r
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
ft»*
3pap»ai
-
'
-
-
-
'
-
-
-
-
-
-
-
-
-
-
-
-
.-
«tt*
J|M|flBMb>
^rwr^*
'. *» r
-
-
-
-
.
-
-
-
-
-
-
' ,,X5V-Y, "
. ^ IW--/-' \AS
l^^^^l^^^^*
.
-
-
-
"
V
-
-
'
-
-
^ 1^'s
j.5<-;s-'r ?-
Six /fe-'^-
.
-
-
-
.
-
-
-
-
,;v"-- % N r %"\ ?" "
' 0*feto»i8<«iro»' "
^ ^', "<- >
^;~f ' *
.
.
Benchmark Category. « adequate, p » provisional, i « intanm; "' indicalec tat the benchmark value was an order o(
magnifejdB or more above the NEL or LEL for other adverse effects.
10 . Insufficient Data
m. Biological Uptake Measures
This section presents the biological uptat&-measures (i.e., BCFs, and BAFs) used to derive protective
surface water and soil concnetrations for constituents considered to bioconcnetrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and sources
are presented in Table 4 for ecological receptor categories: tropic level 3 and 4 fish in the limnetic and
littoral ecosystems, general fish (BCF only), aquatic invertebrates, earthworms, other soil invertbrates,
terrestrial vertebrates, and plants. Each value is idenfieid as whole-body or lipid-based and, for the
generic aquatic ecosystems, the biological uptake factors are deignated with a "d" if the value reflects
dissolved water concentrations, and a "t" if the value reflects total surface water concentrations. For
organic chemicals with log K,,w values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer dissolved water concentrations (i.e., dissolved water concentration equals total water
August 1995
-------
APPENDIX B Di-n-odyi phthalate - 6
concentration). For organic chemicals with log Kow values above 4, the BCFs were assumed to refer
to total water concentrations and concentrations in fish. The following discussion describes the
rationale for selecting the biological uptake factors and provides the context for interpretting the
biological uptake values presented in Table 4.
Because the log Kow for di-n-octyl phthalate is above 6.5 (i.e., 7.5), the Thomann (1989) and
Thomann etal., (1992) models were not used to estimate bioaccumulation factors. For extremely
hydrophobic constituents, the Agency has stated that reliable measurements of ambient water
concentrations (especially dissolved concentrations) are not available and that accumulation of these
constituents in fish or other aquatic organisms cannot be referenced to a water concentration as
required for a BCF or BAF (U.S. EPA, 1993Q. Since no measured BAF was available, a measured
BCF identified in Stephan (1993) was used as a BAF since di-n-octyl phthalate, like other phthalates,
is capable of being metabolized by aquatic organisms
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, earthworms and terrestrial
invertebrates were estimated as described in Section 5.3J.2.3. Briefly, the extrapolation method is
applied to hydrophobic organic chemicals assuming that the partitioning to tissue is dominated by
lipids. For hydrophobic organic constituents, the bioconcentration factor for plants was estimated as
described in Section 6.6.1 for above ground leafy vegetables and forage grasses. The BCF is based on
route-to-leaf translocation, direct deposition on leaves and grasses, and uptake into the plant through
air diffusion.
August 1995
-------
APPENDIX B
Di-n-octyl phthalate - 7
1
Table 4. Biological Uptake Properties
ecotogicH
reoeptor
level 4 fish
level 3 fish
fish
.
littonl toprac
leveUfwh
littoral trophic
level 3 fish
Mural trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
ffiWt0bfetiaM
(MfthWOfTllt
pfents
BCF,BAP,<»
BSAf
BAF
BAF
BCF
BCF
BAF
BAF
BAF
BAF
.BAF
ItpfcMMMd or
wlifllftAAdhf
lipid
Upid .
lipid
lipid
lipid
whole-body
wtMto-booy
whoto-body '
wtMto-plant
IMAM
2.400 (t)
2,400 (t)
2. 400 (t)
2, 400 (t)
2, 400 (t)
10
3.9 E - 01
3.7 E 01
3.0
4.4
** -AAttA^to
^^^W^MP
no nwMtnd BAF; buad on
iWMured BCF (Staphan. 1993)
no nwitturad BAF; bated on
nwwurad BCF (Suphm. 1993)
no nw«ur*d BAF; buad on
measured BCF (Stephen, 1993)
no measured BAF; based on
measured BCF (Stephen, 1993)
no.maasured BAF; based on
measured BCF (Stephen, 1993)
ceJc
ceJc
catc
U.S. EPA, 1990e
d refers to dfesofeed surface water concentration
t «refers to total surface, water ooocerrtreeon
ID = buafiickat D«u
August 1995
-------
APPENDIX B Di-n-octyl phthalate - 8
References
Heindel, J.J., O.K. Gulati, R.C Mounce, S.R. Russell and J.C. Lamb IV. 1989.
Reproductive toxicity of three phthalic acid esters in a continuous breeding protocol.
Fundamental and Applied Toxicology. 12: 508 -518.
.4
Korhonen, A., K. Hemminki and H. Vainio. 1983. Embryotoxic effects of phthalic acid
derivatives, phosphates and aromatic oils used in the manufacturing of rubber on three day
chicken embryos. Drug and Chemical Toxicology. 6(2): 191 -207.
Mann, A.H., S. C. Price, F.E. Mitchell, P. Grasso, R. H. Hinton, and J.W. Bridges. 1985.
Toxicology and Applied Pharmacology. 77:116-132.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
National Library of Medicine. HSDB (Hazardous Substance Database). 1994.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93 - 154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter H, G. W. and J. B. Mabrey 1994. Toxicological Benchmarks for Screening of
Potential Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-
ACOS-84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
Thomann, R. V. 1989. Bioaccumulation model of organic chmeical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6): 699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615 - 629.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of'Health and Environmental Assessment, Washington, D. C. January.
U.S. EPA (Environmental Protection Agency). 1994. Integrated Risk Information System.
March.
August 1995
-------
APPENDIX B Di-n-octyl phthalate 9
Will, M. E. and G. W. Suter n. 1994. lexicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Freshwater Toxicity - Di-n-octyl phthalate Cas No.: 117-84-0
Chemical
Nam*
di-n-octyl
phthalale .
di-n-octyl
phthalate
di-n-octyl
phthalate
di-n-octyl
phthalate
NS = Not Spe
Species
fish
daphnid
fish .
daphnid
dfied
Type of
Effect
chronic
chronic
chronic
chronic
J
Description
cv
cv
EC20
EC20
Value
3822
708
<100
310
Units
ug/L
ug/L
ug/L
ug/L
,}
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
.
Exposure
Duration
/Timing
NS
NS
NS
NS
Reference
Suter and Mabrey, 1994
Suter and Mabrey, 1994
Suter and Mabrey, 1994
Suter and Mabrey, 1994
. '
Comments
...
-------
Freshwater Toxieiiy - Dl-n-oc ,. phthaiate Gas No.: 117-84-0
Chemical
Nam*
di-n-octyl -
phthaiate
di-n-octyl
phthaiate
di-n-octyl
phthaiate
di-n-octyl
phthaiate
Species
fish
daphnid
fish
daphnid
NS = Not Specified
Type of
Effect
chronic
chronic
chronic
chronic
.1
Description
CV
CV
EC20
EC20
Value
3822
708
<100
310
Units
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
-
Exposure
Duration
/Timing
NS
NS
NS
NS
Reference
Suter and Mabrey. 1994
Suter and Mabrey. 1994
Suter and Mabrey, 1994
Suter and Mabrey, 1994
Comments
,
-------
Terrestrial Toxiclty - Di-n-octyl phthalate Cas No.: 117-84-0
Chemical
Name
di-n-octyl
phthalate
di-n-octyl
phthalate
di-n-octyl
phthalate
di-n-octyl
phthalate
di-n-octyl
phthalate
Species
chicken
rat
mouse
J
rat
mouse
NS = Not Specified
Endpolnt
emb
liver
rep
acute
acute
Description
NOEL
FEL
NOAEL
LD50
LD50
Vslus
20
2.400
0.03
47
6.513
Units
ugmol/egg
mg/kg-day
mg/kg-day
g/kg-body wt.
g/kg-body wt.
Exposure
Routs (oral,
S.C.. I.V., l.p..
Injection)
NS
oral
oral
oral
oral
Exposure
Duration
/Timing
NS
3 weeks
7 days prior to
and throughout
mating period.
NS
NS
Reference
Korhonen et a!..
1983
Mannetal.. 1985
Heindel et al.. 1989
RTECS. 1994
RTECS. 1994
Comments
No embryotoxic
effects observed.
Liver enlargement
effects on
reproductive
function at any
dose.
-------
Terrestrial Biological Uptake Measure. Ji-n-octy! phthalate Gas No.: 117-84-0
Chemical
Name
di-n-octyl
phthalate
Species
plant
B-lactor
(BCF. BAF.
BMP)
BCF
Value
460,000
Measured
or
Predicted
(m,p)
P
units
(ug/g DW
plant)/(uo/g soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Dieldrin-1
Toxicological Profile for Selected Ecological Receptors
Dieldrin
CasNo.: 60-57-1
Summary: This profile on dieldrin summarizes the toxicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the
freshwater ecosystem were calculated for organic constituents with log K^ between 4 and
6.5. For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire toxicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from the data presented in the technical support document
for the Hazardous Waste Identification Rule (HWIR): Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. However, several chronic and subchronic toxicity
studies involving dieldrin have been conducted using laboratory rats and mice. Fetotoxicity
was observed in a subchronic study where pregnant mice were administered 1.5, 3.0, and 6.0
August 1995
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APPENDIX B . Dieldrin - 2
mg dieldrin/kg-day by gastric intubation on days 7-16 of gestation (Chernoff et al., 1974).
From this research, Chemoff reported a NOAEL of 1.5 mg/kg-day based on fetal skeletal
abmormalities resulting from the two higher doses. A chronic reproductive study was
identified in which female rats were fed a diet containing 2.5, 12.5, and 25.0 ppm dieldrin for
three generations (Treon and Cleveland, 1955). Treon and Cleveland (1955) reported reduced
number of pregnancies and a moderate increase in mortality among the offspring of the rats
exposed to 2.5 ppm (0.189 mg/kg-d) dieldrin. The mg/kg-day value for Treon and Cleveland
(1955) was calculated from the reported ppm-dose using the reference body weight (kg) and
the recommended value for food consumption (kg/day) for rats reported in Recommendations
for and Documentation of Biological Values for Use in Risk Assessment (U.S. EPA, 1988).
Reproductive toxicity was also observed in 220 female rats (Harr et al., 1970) fed dieldrin in
10 two-fold concentrations (ranging from 0.08 to 40 ppm). In this lifetime observational
study, Harr et al., (1970) reported a NOAEL of 0.014 mg dieldrin/kg-day (equivalent dietary
dose of 0.24 ppm), based on dam survival, conception rate, pup survival, and weaned litter
size.
The value of 0.014 mg/kg-day (Harr et al., 1970) was selected to derive the mammalian
lexicological benchmark because: (1) chronic exposures were administered via oral ingestion,
(2) it focused on reproductive toxicity as a critical endpoint, (3) the study contained sufficient
dose-response information and (4) the study represented the lowest reproductive endpoint in
the dataset. The study by Treon and Cleveland (1955) was not selected because the LOAEL
of 0.189 mg/kg-day was not as protective of the representative species as the Harr et al.
(1970) study. Similarly, the study by Chernoff et al., (1974) was not selected because the
NOAEL for fetotoxicity in mice (1.5 mg/kg-day) was two orders of magnitude greater than
the NOAEL for reproductive effects in rats (0.014 mg/kg-day) as reported by Harr et al.,
(1970).
The selected NOAEL value was then scaled for species representative of a freshwater
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994)
Benchmark^ = NOAEL. x L
(bww
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
August 1995
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APPENDIX B Dieldrin - 3
Harr et al. (1970) study documented reproductive effects from dieldrin exposure to male and
female rats, male and female body weights for each representative species were used in the
scaling algorithm to obtain the lexicological benchmarks.
Data were available on the reproductive and developmental, effects of dieldrin, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. All of the studies
identified were conducted using laboratory rats and mice and as such, inter-species differences
among wildlife species were not identifiable. Therefore, an inter-species uncertainty factor
was not applied. Based on the data set for dieldrin, the benchmarks developed were
categorized as adequate.
Birds: Subchronic and chronic toxicity studies involving dieldrin have been conducted using
chickens and mallard ducks. Reduced hatchability and morphological changes were observed
in chicken eggs (Smith et al., 1970) injected with 0, 1.25, 2.5, 5, and 10 mg/egg of dieldrin
either prior to incubation, or after a 7-day incubation period and a NOAEL of 45.45 mg/kg
(2.5 mg/egg) was estimated. Nebeker et al. (1992), conducted research to determine the
developmental effects of dieldrin administered to 1-day old mallard ducklings^ through dietary
exposure. Nebeker et al., (1992) recorded a NOAEL of 0.08 mg/kg-day for growth
impairment after dietary administration of dieldrin for a 24-day period. In addition, Nebeker
et al., (1992) also recorded significant concentrations of dieldrin in the tissues of mallard
ducklings that were fed dieldrin-spiked food.
The NOAEL reported by Nebeker et al., (1992) was used to calculate the lexicological
benchmark for birds because it focused on developmental growth as a critical endpoint and
dietary concentrations were administered via oral ingestion during a critical life-stage period.
The study by Smith et al.., (1970) on chicken eggs was not selected for benchmark derivation
because data were not identified on either (1) direct absorption of dieldrin from direct contact
with the eggs or (2) maternal transfer from mother to egg. Without these absorption data, it
is difficult to estimate the internal dose to the hatchling from the egg-injected dose.
The principles for allometric scaling were assumed to apply to birds,' although specific studies
supporting allometric scaling for aviaa species were not identified. Thus, the NOAEL from
the Nebeker et al. (1992) study was scaled for differences between the body size of the test
species and the body size of the species of interest. This cross-species scaling was completed
using the method described by Opresko et al. (1994).
August 1995
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APPENDIX B Dieldrin - 4
Data were available on the reproductive and developmental, effects of dieldrin, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. Laboratory
experiments of similar types were not conducted on a range of avian species and as such,
inter-species differences among wildlife species were not identifiable. Based on the avian
data set for dieldrin, the benchmarks developed from Nebeker et al. (1992) were categorized
as adequate.
Fish and Aquatic Invertebrates: The Final Chronic Value (FCV) of 6.25 E-05 mg/L (U.S.
EPA, 1993c) was selected as the benchmark protective of fish and aquatic invertebrates in the
generic freshwater ecosystem. It should be noted that a Final Residue Value (FRV) of 1.9E-7
mg/1 was reported (57 FR 60911) however, it was not considered appropriate for a benchmark
value because residues and bioaccumulation are already taken into account by the Thomann et
al., (1992) model. Because the benchmark was based on a FCV derived for the Sediment
Quality Critieria Document, this benchmark is categorized as adequate.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECM) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Aquatic plant data was not identified for dieldrin and, therefore, no benchmark was
developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K,,,.) to determine a protective sediment concentration
(Stephan, 1993). The EQP number is the chemical concentration that may be present in
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. The FCV reported in the Sediment Quality Criterion (SQC) document for dieldrin
(U.S. EPA, 1993c) was used to calculate an EQP number of 12.7 mg dieldrin /kg organic
carbon. Assuming a mass fraction of organic carbon for the sediment (f^) of 0.05, the
benchmark for the benthic community is 0.637 mg/kg. Since the EQP number was based on a
FCV established for the SQC, the sediment benchmark is categorized as adequate.
August 1995
-------
APPENDIX B
Dieldrin - 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
RoptMtnUrtlw
SfMCteS
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted
sandpiper
herring gull
kingfisher
BwtottfTMfk
ValiM*
mgfflcg-d
0.010 (a)
0.006 (a)
0.04 (a)
0.04 (a)
0.04 (a)
0.08 (a)
0.05 (a)
0.11 (a)
0.05 (a)
o:08 (a)
Study
rat
rat
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
duckling
mallard
duckling
mallard
ducklings
'. EfnMI
rep
rep
dev
dev
dev
dev
dev
dev
dev
dev
Study
VahM
n^pHQ^fli
0.014
0.014
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0.08
0^*,
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
-
-
-
-
-
-
Original Bourc*
Harret al., 1970
Harretal., .1970
Nebeker et al., 1992
Nebeker etal., 1992
Nebeker et al., 1992
Nebeker et al., 1992
Nebeker et al., 1992
Nebeker et al.. 1992
Nebeker et al., 1992
Nebeker et al.. 1992
'Benchmark categories, a=adequate, p=provisional, i=interim
above the NEL or LEL for other adverse effects.
a '*' indicates that the benchmark value was an order of magnitude or more
August 1995
-------
APPENDIX B
Dieldrin - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftapfMMIttthNI
Spaolw
fish and aquatic
invertebrates
aquatic plants
benttiic
community
MflCJUIIWIT
Vilu»» mgfl.
6.25E-05 (a)
ID
0.072(a)
mg/Vg
sediment
Study SfwciM
aquatic
organisms
aquatic
organisms
Mwcnptfofi
FCV
-
FCVxK^
OriQlnti Sowcc
U.S. EPA, 1993C
-
U.S. EPA, 1993c
II.
Benchmark categories, a=adequate, p=provisional, i=intenm; a '*' indicates that the benchmark value was an
order of magnitude or more above the NEL or LEI for other adverse effects.
ID = Insufficient Data .
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: An additional chronic study was found for mammalian wildlife in which dose-
response data were reported. In this chronic reproductive study, raccoons fed 0.73 or 2.2 ppm
of dieldrin experienced adverse effects on the estrous cycle, decreased incidence of
pregnancy, reduction in litter size, resorption of embryos, and increased fetal death
(Frederickson, 1973 as cited in NIOSH, 1978). Using a mean raccoon body weight of 5.8 kg
(U.S. EPA, 1993e) and an estimated food consumption, a LOEL of 0.036 mg/kg-day was
calculated from the original 0.73 ppm. However, the timing or exposure duration in
Fredrickson's research was not revealed.
The NOAEL in Harr et al.'s (1970) .study was chosen for the derivation of a benchmark for
mammals in the generic terrestrial ecosystem because: (1) it was performed on a surrogate
August 1995
-------
APPENDIX B Dieldrin - 7
species, (2) it focused on reproductive toxicity as a critical endpoint, and (3) chronic
exposures were administered via oral ingestion. Unspecified exposure duration and limited
dose-response information, prevented the Frederickson (1973 as cited in NIOSH, 1978) study
from being used to derive a benchmark. Since Hair et al. (1970) documented reproductive
effects from dieldrin exposure to male and female rats, male and female body weights for
each representative species were used in the cross-species scaling algorithm to obtain
terrestrial benchmarks. Based on the data set for dieldrin the benchmarks developed were
categorized as adequate, as in the aquatic ecosystem..
/
Birds: No additional avian toxicity studies were identified for species representing the
terrestrial ecosystem. Thus, the benchmarks calculated for the avian members of the generic
terrestrial ecosystem were based on the same study value used for the generic freshwater
ecosystem (Nebeker et al., 1992). The cross-species scaling method of Opresko et al. (1994)
was used to adjust the study value for differences in animal body size. Based on the avian
data set for dieldrin, the benchmarks developed from the Nebeker et al. (1992) study were
categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for dieldrin and, as a result, a benchmark
could not be developed.
Soil community: A dataset from which a soil community benchmark could be calculated was
not identified.
August 1995
-------
APPENDIX B
Dieldrin - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RtpfMWIttthtt
SfMCiM
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
. American robin
American
woodcock
plants
soil community
tMnCIUTMrK
Vibw'mgAc*
*y
0.028 (a)
0.029 (a)
0.025 (a)
0.010 (a)
0.007 (a)
0.007 (a)
0.003 (a)
0.05 (a)
0.08 (a)
0.08 (a)
0.09 (a)
0.08 (a)
ID
ID
Study
SfwdM
rat
rat
rat
rat
rat
rat
rat .
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
-
-
EfftCt ,
rep
rep
rep
rep
rep
rep
rep
dev
dev
dev
dev
dev
-
-
Study
Vaiin
rngft*
.- day
0.014
0.014
0.014
0.014
0.014
0.014
0.014
0.08
0.08
0.08
0.08
0.08
-
Doci'lpUon
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
-
8F
-
-
-
-
-
-
-
-
-
-
-
-
-
Harret al., 1970
Harretal., 1970
Harret al., 1970
Harr et al:, 1970
Harretal., 1970
Harretal., 1970
Harretal., 1970
Nebeker et al..
1992
Nebeker et al.,
1992
Nebeker et al.,
1992
Nebeker et al.,
1992
Nebeker et al.,
1992
'Benchmark categories, a=adequate, p=provisional, i=interim; a '" indicates that the benchmark value was an order of magnitude or more
above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Dieldrin - 9
III. Biological Uptake Measures ,
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K^w values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log K^ values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for dieldrin. The
bioconcentration factor for fish, was also estimated from the Thomann models (i.e., log Kow -
dissolved BCF/) and multiplied by the dissolved fraction (/) as defined in Equation 6-21 to
determine the total bioconcentration factor (BCF/). The dissolved bioconcentration factor
(BCF,d ) was .converted to the BCF,1 in order to estimate the acceptable lipid tissue
concentration (TC/) in fish consumed by piscivorous fish (see Equation 5-115). The BCF/
was required in Equation 5-115 because the surface water, benchmark (i.e., FCV or SCV)
represents a total water concentration (C). Mathematically, conversion from BCF/1 to BCF/
was accomplished using the relationship delineated in the Interim Report on Data and
Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S.
EPA, 19931):
BCF,d x fd = BCF,'
August 1995
-------
APPENDIX B Dieldrin-10
Converting the predicted BCF,d of 251,768 L/kg LP to the BCF/ of 143,867 L/kg LP was in
reasonable agreement (i.e., within a factor of 4) of the geometric mean of two measured BCF,'
values presented in the master table on dieldrin (geometric mean = 216,700).
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of several
values with sources in Table 4 (see master table). For earthworms and terrestrial
invertebrates, the bioconcentration factor was estimated as described in Section 5.3.5.2.3.
Briefly, the extrapolation method is applied to hydrophobic organic chemicals assuming that
the partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs
and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for "a
chemical lacking measured data (in this case dieldrin) is compared to the BBF for TCDD and
that ratio (i.e., dieldrin BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial
vertebrates, invertebrates, and earthworms, respectively. For hydrophobic organic
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf
trans location, direct deposition on leaves and grasses, and uptake into the plant through air
diffusion. For metals, empirical data were used to derive the BCF for aboveground forage
grasses and leafy vegetables.
August 1995
-------
APPENDIX B
Dieldrin- 11
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic ,
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
llpld-fcesed or
... w.i W-..A-
wnoie>>ovciy
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
763,219 (d)
618,762 (d)
.1 42,668 (t)
71 3,460 (d).
742,334 (d)
1,322,661 (d)
23
0.0033
0.024
0.029
source
predicted value based on
Thomann, 1989, food chain
model
predicted value based on
Thomann, 1989, food chain
model
predicted value based on
Thomann, 1989 and adjusted to
estimate total BCF
predicted value based on
Thomann et al., 1992, food web
model
predicted value based on
Thomann et al., 1992, food web
model
predicted value based on
Thomann et al., 1992, food web
model
geometric mean of values (e.g.,
Mendenhall et al., 1983 and
Aulerich et al., 1972 as cited in
WHO, 1989)
Cooke, 1972 as cited in WHO,
1989
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Dieldrin-12
References
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APPENDIX B Dieldrin-13
DeWitt, J.B. 1956. Chronic Toxicity to Quail and Pheasants of Some Chlorinated Insecticides.
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August 1995
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APPENDIX B Dieldrin - 14
Nebeker, A.V., W.L. Griffis, T.W. Stutzman, G.S. Schuytema, L.A. Carey, and S.M. Schefer.
1992. Effects of aqueous and dietary exposure of dieldrin on survival, growth, and
bioconcentration in mallard ducklings. Environ. Toxicol. Chem. 11:687-699.
Ottolenghi, A.D., J.K. Haseman, and F. Suggs. 1973. Teratogenic effects of aldrin, dieldrin,
and endrin in hamsters and mice. Teratology 9:11-16.
Opresko, D.M., B.E. Sample, and G.W.. Suter. 1994. lexicological Benchmarks for Wildlife:
1994 Revision. Oak Ridge National Laboratory, ORNL ES/ER/TM-86/R1
Parrish, P.R., J.A. Couch, J. Forrester, J.M. Patrick, Jr., and G.H. Cook. 1974. Dieldrin:
Effects on Several Estuarine Organisms. In: Southeastern Association of Game and Fish
Commissioners, Twenty-Seventh Annual Conference, pp. 427-434. As cited in Stephan,
1993, Derivations of Proposed Human Health and Wildlife Bioaccumulation Factors for
the. Great Lakes Initiative, PB93-154672, Environmental Research Laboratory, Office of
Research and Development, Duluth, MN.
RTECS (Registry of Toxic Effects of Chemical Substances) Database. March 1994. National
Institute for Occupational Safety and Health.
Sanders, H.O. and O.B. Cope. 1966. Toxicities of Several Pesticides to Two Species of
Cladocerans. Trans. Am. Fish. Soc. 95(2): 165-169. As cited in AQUIRE (AOUatic
Toxicity /nformation /?£trieval Database), Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
Santharam, K.R., B. Thayumanavan, and S. Krishnaswamy. 1976. Toxicity of Some
Insecticides to Daphnia carinata King, an Important Link in the Food Chain in the
Freshwater Ecosystems. Indian J. Ecol. As cited in AQUIRE (AOt/atic Toxicity
/nformation /?£trieval Database), Environmental Research Laboratory, Office of Research
and Development, U.S. Environmental Protection Agency, Duluth, MN.
Shubat, P.J., and L.R. Curtis. 1986. Ration and toxicant preexposure influence dieldrin
accumulation by rainbow trout (Salmo gairdneri). Environ. Toxicol. Chem. 5:69-77. As
cited in Stephan, 1993. Derivations of Proposed Human Health and Wildlife
Bioaccumulation Factors for the Great Lakes Initiative, PB93-154672, Environmental
. Research Laboratory, Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Dieldrin-15
Smith, S.I., C.W. Weber, and B.L. Reid. 1970. The effect of injection of chlorinated
hydrocarbon pesticides on hatchability of eggs. Toxicol. Appl. Pharmacol. 16:179-185.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter II, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
Thorpe, E., and A.I.T. Walker. 1973. The toxicology of dieldrin (HEOD). II Comparative
long-term oral ioxicity studies in mice with dieldrin, DDT, phenobarbitone, 3-BHC and
y-BHC. Fd. Cosmet. Toxicol. 11:433-442.
Travis, C.C. and A.D. Arms. 1988. Bjoconcentration of organics in beef, milk, and
vegetation. Environ. Sci. Technol. 22(3):271-274.
i
Treon, J.F., and P.P. Cleveland. 1955. Toxicity of certain chlorinated hydrocarbon
insecticides for laboratory animals, with special reference to aldrin and dieldrin. Agric.
Food Chem. 3(5):402-408.
U.S. Department of Health, Education, and Welfare. 1978. Special Occupational Hazard
Review for Aldrin/Dieldrin. Public Health Service, Center for Disease Control, National
Institute for Occupational Safety and Health, Division of Criteria Documentation and
Standards Development, Rockville, Maryland.
U.S. EPA (U.S. Environmental Protection Agency). 1980. Ambient Water Quality Criteria
for Aldrin/Dieldrin. PB81-117301. Environmental Criteria and Assessment Office, Office
of Water Regulations and Standards, Washington, DC.
August 1995
-------
APPENDIX B Oieldrin - 16
U.S. EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment. Washington, DC. January.
U.S. EPA (U.S. Environmental Protection Agency). 1993b. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Science and Technology, Office of Water, Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1993c. Sediment Quality Criteria for the
Protection of benthic Organisms: Dieldrin. EPA-822-R-93-015. Office of Science and
Technology, Office of Water, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993d. Technical Basis for Deriving
Sediment Quality Criteria for Nonionic Organic Contaminants for the Protection of
Benthic Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of
Water, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993e. Wildlife Exposure Factors Handbook.
EPA/600/R-93/187a. Office of Research and Development, Washington, DC
U.S. EPA (U.S. Environmental Protection Agency). 1993i. Interim Report on Data and
Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Life and
Associated Wildlife. EPA/600/R-93/055. Office of. Research and Development,
Washington, DC.
Virgo, B.B., and G.D. Bellward. 1975. Effects of dietary dieldrin on reproduction in the
Swiss-Vancouver (SWV) mouse. Environ. Physiol. Biochem. 5:440-450.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxicity - Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
mouse
mouse
mice
SWV mice
(females)
mice
Swiss mice
rats
female rats
hamsters
mice
Endpolnt
et
fet
rep
fet.rep
iver
rep
rep
rep
ter
ter
Description
NOAEL
LOAEL
NOAEL
LOAEL
AEL
NOAEL
NOAEL
NOAEL
AEL
AEL
Value
1.5
3
5
t.02
10
3
0.189
0.014
30
15
Units
mg/kg-
day
mg/kg-
day
ppm
mg/kg-
diet
ppm
mg/kg-
diet
mg/kg-
day
mg/kg-
day
mg/kg-
dlet
mg/kg-
diet
Exposure Route
(oral, B.C., l.v.,
l.p., Inlectlon)
gastric intubation
gastric intubation
oral
oral
oral
oral
/
oral
oral (10 two-fold
cone.)
oral
oral
Exposure
Duration/Timing
days 7- 16 of
gestation
days 7- 16 of
gestation
120 days
4 wks prior to their
2nd mating, cont. to
day 28 postpartum
2-year study
6-generation study
(2 liners/
generation)
2 years
lifetime
observations
Given a single dose
on day 7, 8 , or 9 of
gestation.
Given a single dose
on day 9
j
Reference
Chemoff et al., 1975
Chemotf etal., 1975
Good and Ware, 1969
Virgo and Bellward, 1975
Thorpe and Walker, 1973
Keplinger et al., 1970
Treon and Cleveland,
1955
Harr etal., 1970
Ottolenghi etal., 1973
Ottolenqhi et al.. 1973
Comments
No teratogenic or fetotoxic effects
were observed at this dose level.
An increased percentage of '
supernumerary ribs was observed
in mice.
No effect on maternal mortality.
fertility, or fecundity. (Single dose)
'No effect on the incidence of
breeding in parous females, fetal
survival, the duration of gestation,
or parturition.'
Liver enlargement by week 50;
first liver tumours after a 12-
month exposure period.
'no effects on fertility, viability, or
gestation were observed in 6
generations of mice*
The number of pregnancies were
not affected at this dose level.
0.24 ppm is 'the highest dietary
dieldrin level consistent with
normal reproductive values'
Embryocidal and teratogenic
effects were observed in pregnant
hamsters.
Teratogenic effects, but the
frequency and gravity of the
defects produced were less
pronounced.
-------
Freshwater . ^ity - D!e!dr!n
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
Daphnia
carinata
Daphnia pulex
Simocephalus
serrulatus
bluegill
striped bass
aquatic
organisms
aquatic
organisms
rainbow trout
fathead
minnow
Type of
Effect
immob.
immob.
immob.
mort. -
mort.
mort.
mort.
mort.
mort.
Description
EC50
EC50
EC50
LC50
LC50
FCV
FCV
LC50
LC50
Value
130
251
190-240
(213.8)
7.9-17
(10.72)
1-500
(98.86)
0.29
0.0625
1.1 - 10000
(159.59)
18
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ufl/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
NA
NA
NA
NA
Exposure
Duration/
Tlmlna
48 hour
48 hour
48 hour
96 hour
96 hour
NA
NA
96 hour
96 hour
Reference
Santharam et al., 1976 as
cited in AQUIRE, 1995
AQUIRE, 1995
Sanders etal, 1966 as
cited in AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
U.S. EPA, 1980
U.S. EPA, 1993c
AQUIRE. 1995
Henderson et al., 1959 as
cited in AQUIRE, 1995
Comments
NA = Not applicable
-------
Freshwater Biological Uptake Measures - Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
Species
fish
fish
fish
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
Value
467
2091
2245
Measured
or
predicted
(m,p)
P
m
m
Units
NS
NS
NS
Reference
Stephan, ,1993
Shubat andCurtis, 1986 as cited
Stephan, 1993
Parrish et al., 1974 as cited in.
Stephan, 1993
Comments
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
NS = Not specified
-------
APPENDIX B Diethyl phthalate - 1
Toxicological Profile for Selected Ecological Recptors
Diethyl phthaJate
Cas No.:84-66-2
Summary: This profile on diethyl phthalate summarizes the toxicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire toxicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Mammals: Only one subchronic study investigating the effects of oral diethyl phthalate
exposure in mammalian species was identified. Brown et al. (1978) fed female rats diethyl
phthalate at doses of 150, 750 and 3710 mg/kg-day for 16 weeks. Male rats were maintained
on diets containing 150, 770 and 3160 mg/kg-day for the same period of time. No changes in
behavior or other clinical signs of toxicity were observed in either sex. However, at the
highest dose levels, decreases in food consumption and weight gain were observed in both
sexes as well as increases in relative weights of the brain, liver, kidney, stomach, small
intestines and full caecum. Based on these results, a NOAEL of 750 mg/kg-day and a
LOAEL of 3160 mg/kg-day were reported.
The dose levels used in this study were sufficient to establish a dose-response relationship for
pathological effects. However, benchmark values were not derived because the study does
not evaluate reproductive or developmental endpoints.
August 1995
-------
Terrestrial Biological, .ke Measures Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
Species
plants
cattle (beef)
cattle (milk)
B-factor
(BCF, BAF,
BMP)
BCF
BTF
BTF
Value
0.123
0.0079
0.0107
Measured
or
predicted
(m.p)
P
m
m
Units
(ug/g DW
plant)/(ug
/g soil)
NS
NS
Reference
U.S. EPA, 1990e
Travis and Arms, 198B
Travis and Arms, 1988
Comments
BTF = Biotransfer factors
BTF = Biotransfer factors
NS = Not specified
-------
Terrestrial Biological Uptake Measures - Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
cattle
cattle
swine
swine
swine
swine
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
Value
1
2.8
6.5
18
8.7
10.6
3
1.6
1.76
0.8
2.68
1.8
Measured
or
predicted
(m.p)
m
m
m
m
m
m
m
m
m
m
m
m
Unto
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS .
NS
Reference
Nebeker et at., 1992
Nebekeretal., 1992
Nebekeretal, 1992
Nebekeretal., 1992
Nebekeretal., 1992
Nebekeretal., 1992
Clabom, et.al., 1960 as cited
inKenaga, 1980
Clabom, et.al., I960 as cited
in Kenaga, 1 980
Clabom, et.al., 1960 as cited
in Kenaga, 1 980
Clabom, et.al , 1960 as cited
in Kenaga, 1 980
Clabom, et.al., 1960 as cited
inKenaga, 1980
Clabom. et.al., 1956 as cited
in Kenaga, 1 980
Comments
Steady state BCF at mean food
dieldrin concentration of 606 =(-) 16
ug/g.
Steady state BCF at mean food
dieldrin concentration of 272 +(-) 17
ug/g
Steady state BCF at mean food
dieldrin concentration of 155 +(-) 15
ug/g. .
Steady state BCF at mean food
dieldrin concentration of 48 +(-) 5 -
ug/g
Steady state BCF at mean food
dieldrin concentration of 16.4 +(-) .3
ug/g.
Steady state BCF at mean food
dieldrin concentration of .3 +() .03
ug/g.
,
-------
Terrestrial Biological i .ke Measures - Oieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
common frog
common
load
barn owl
short-tailed
shrew
mink
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
mallard
ducklings
B-factor
(BCF, BAF,
BMF)
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
Value
387.5
280
18.8
-\
1.6
8.4
1,124
706
1,085
1,427
1,325
1,995
1,753
Measured
or
predicted
(m,p)
m
m
m
m
m
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Cooke, 1972
Cooke, 1972
Mendenhall et al., 1983
Blus, 1978
Aulerich et al , 1972
Nebeker el al., 1992
Nebeker et al., 1992
Nebeker et at., 1992
Nebeker et al., 1992
Nebeker etal., 1992
Nebeker et al., 1992
Nebeker et al., 1992
Comments
Exposure duration = 2 days; whole
body.
Exposure duration = 2 days; whole
body.
Exposure duration = 2 years;
carcass.
Exposure duration = 17 days;
carcass.
Exposure duration = 4-10 weeks;
fat.
Steady state BCF at mean dieldrin
concentrations in water of .193 +(-)
8mg/L.
Steady state BCF at mean dieldrin
concentrations in water of .177 +(-)
11 mg/L.
Steady state BCF at mean dieldrin
concentrations in water of . 1 1 8 +(-)
11mg/L.
Steady state BCF at mean dieldrin
concentrations In water of .075 +(-)
1-mg/L.
Steady state BCF at mean dieldrin
concentrations in water of .052 +(-)
4 mg/L.
Steady state BCF at mean dieldrin
concentrations in water of .019 +(-)
2 mg/L.
Steady state BCF at mean dieldrin
concentrations in water of .014 +(-)
1 mg/L.
-------
Terrestrial 1 Aty - Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
raccoons
mallard
duckling
mallard
duckling
bam owls
quail
quail
pheasants
embryos and
young
growing
chicks
embryos and
young
growing
chicks
Endpolnt
rep
dev
dev
rep
rep
rep
rep
dvp, rep
dvp, rep
Description
LOEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
0.036
0.08
4.27
0.5
O.B
1
10
2.5
5
Units
mg/kg-d
mg/kg-
day
mg/kg-
day
ppm
mg/kg-
day
ppm
ppm
mg/egg
mg/egg
Exposure Route
(oral, s.c., l.v.,
l.p.. Inlectlon)
oral
oral
oral
oral
oral
oral
oral
injection
injection
Exposure
Duration/Timing
NS
24 days
24 days
2 years
1 54 days
61 days
61 days
injected either prior
to incubation or
after a 7-day
incubation period
injected either prior
to incubation or
after a 7-day
incubation period
Reference
Frederickson, 1973 as
cited in NIOSH. 1978
Nebeker et al., 1992
Nebeker et al., 1992
Mendenhalletal., 1983
DeWitt, 1955
DeWitt, 1956
DeWitt, 1956
Smith etal., 1970
Smith etal.. 1970
Comments
'statistically significant adverse
effects on the estrous cycle and
on the incidence of pregnancy.
fetal death, resorption of embryos
and reduced litter size.
No developmental effects were
observed at this dose level.
Growth and survival were effected
at this dose level.
No reduction in breeding success;
a single dose.
Significant decreases in
hatchability of eggs and viability o
chicks were observed.
No effects on egg production,
percentage fertility, or percentage
hatchability were observed at this
dose level.
Hatchability was decreased and
unusually high mortality of chicks
during the first 2 weeks were
observed at this dose level.
No reproductive effects were
observed.
This result is in conflict with '
(Dunachie and Fletcher, 1966),
which indicated that 10 mg could
be injected with no apparent
deleterious effects.
-------
Terrestrial Toxicity - Oieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
Species
chickens
rat
mouse
dog
monkey
rabbit
pig
guinea pig
hamster
pigeon
chicken
quail
duck
wild bird
canada
goose
fulvous
whistling
duck
Endpolnt
rep
acute
acute
acute
acute
acute *"
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
Description
NOAEL
LD50
LD50
LD50
L050
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
1
38300
38
65
3
45
38
49
60
23700
20
10780
381
13300
<141
100-
200
Units
ppm
ug/kg-
Dody wl
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
bodywt.
mg/kg-
body wt.
mg/kg-
body wt.
ug/kg-
body wt.
mg/kg-
bodywt.
ug/kg-.
body wt.
mg/kg-
body wt.
ug/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
Exposure Route
(oral, s.c., l.v.,
l.p., Inlectlon)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Ouratlon/Tlmlnq
2 years
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
/
Reference
Brown et at., 1965
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
U.S. EPA. 1993b
U.S. EPA, 1993b
Comments
Fertility and hatchability were not
affected at this dose level.
.
Behavioral effects. .
-------
APPENDIX B Diethyl phthalate - 2
Birds: No toxicity studies documenting avain exposure to diethyl phthalate in the fresh water
ecosystem were identified and, therefore benchmarks were not derived.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for diethyl phthalate. Therefore, the Tier n method described in Section 4.3.5 was
used to calculate a Secondary Chronic Value (SCV) of 0.22 mg/L. Tier n values or SCVs
were developed so that aquatic benchmarks could be established for chemicals with data sets
that did not fulfill all the requirements of the National AWQC. Because the benchmark is
based on an SCV, it was categorized as interim.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). A
value of 86 mg/L as reported in the SQC document for diethyl phthalate was selected as the
benchmark value. As described in Section 4.3.6, all benchmarks for aquatic plants were
designated as interim.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or other chronic water quality measures, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community the harmful effects of chemcial
exposure. Because no FCV was available, a Secondary Chronic Value (SCV) was calculated
as described in Section 4.3.5. The SCV reported for diethyl phthalate was used to calculate
an EQ., number of 44.2 mg diethyl phthalate/kg organic carbon. Assuming a mass fraction of
organic carbon for the sediment (f^ of 0.05, the benchmark for the benthic community is 2.2
mg diethyl phthalate/kg of sediment Because the EQp number was set using a SCV derived
using the Tier n method, it was categorized as interim.
August 1995
-------
APPENDIX B
Diethyl phthalate 3
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
£tt^*||M
^fjF^^F^^^
mink
riwr otler
bald eagle
osprey
great bkw heron
mallard
lesser scaup
spotted sandpiper
herring gul
kingfisher
Benchmark
V*tu#ma/*fr
«*
10
ID
ID
ID
ID
ID
ID
ID
«
ID
ID
SJwfr
*»4**
- '
'
-
-.
«««<*
-
-
-
-
-
Study Wh*
*Bft»4*
-
:
-
-
-
-
li^^^^PT^p^^W
-
-
-
-
- .
-
8f
.
.
Origipatdeurc*
* .
-
-
'
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID - Insufficient Data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
^pWNWiiluwwi
Sp*d**
fish and aquatic
invertebrates
aquatic plants
benlhic community
dtnehmarit
Vatae*
B»0ft.
0.22 (i)
86 (i)'
2.2 (i)
Study
gpicttw
aquatic
organisms
aquatic
plants
aquatic
organisms
Deecrtpfion
scv
scv
SCV x «,.
Oriafeal&Mjrea
AQUIRE. 1995
Sutar and Mabrey,
1994
AQUIRE. 1995
'Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value
was an order of magnitude or more above the NEL or LEL (or other adverse effects.
August 1995
-------
APPENDIX B Diethyl phthalate 4
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (_) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned in the freshwater ecosystem discussion, no suitable subchronic or
chronic toxicity studies were found for mammalian wildlife exposure to diethyl phthalate.
Since no additional laboratory mammal studies focusing on reproductive or other critical
endpoints were identified, a mammalian benchmark for terrestrial ecosystems was not derived.
Birds: Adequate toxicity data with which to derive a benchmark protective of the terrestrial
avain community were not available.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for diethyl phthalate and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Diethyl phthalate 5
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with .Terrestrial Ecosystem
ftapnstenwfae
$P4Q|e0
dew mouse
short-tajtod
shrew
meadow vole
Eastern
cottontail
red fox .
raccoon
white-tailed deer
red- tailed hawk
American Kestrel
Northern
bobowhite
DeoflhmMlt
VtJue*
4rt0W(Mtoy
ID
10
ID
ID
ID
ID
ID
ID
ID
ID
Study
, Sp*U~
-
-
-
-
:
-
.-
. -
Etf-ct
-
-
-
-
-
-
-
-
-
Study
V«A»
»*ft«-
*v
-
-
-
V^ataWtfttslkfts^Mh.
»F^W^P^MW%
-
-
-
-
-
-
-'
-
SF
'
-
-
-
-
OrfefelBl $ewe*
.
-
-
.
-
-
American robin
American
plants
soil community
ID
ID
ID
ID
woodcock
-------
APPENDIX B
Diethyl phthalate - 5
Table 3. Toxkological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
flapfaaantinvt
$p4ci*#
deer mouse
short-tailed
shrew
meadow vole
Eastarn
cottontail
red fox
raccoon
white- tailed dear
red- tailed hawk
American kaetnel
Northern
bobowhrte
American robin
American
woodcock
plants
toil community
Banclunafic
V«hW
waftHM**
10
ID
ID
ID
ID
ID
ID
ID '
ID
ID
ID
ID
ID
ID
&wtf
Spade*
-
.-
-
-
. -
-
-
-
-
BfrCt
-
-
-
-
-
.
-
-
-
Study
Vafee
mtfUB-
day
-
-
-
-
-
-
-
|i^^^vY*f*j*iB'9%
.
-
-
-
-
-
-
-
-
l$f
.
. -
-
-
-
-
-
-
<<*&(* $CHWX*
.
-
-
-
-
.
-
-
-
-
-
'Benchmark Category, a a adequate, p = provisional, i * interim; a "*' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID - Insufficient Data
HI. Biological Uptake Measures
This section presents the biological uptake measures (i.e., BCFs, and BAFs) used to derive protective
surface water and soil concnetrations for constituents considered to bioconcnetrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and sources
are presented in Table 4 for ecological receptor categories: tropic level 3 and 4 fish in the limnetic and
littoral ecosystems, general fish (BCF only), aquatic invertebrates, earthworms, other soil invertbrates,
terrestrial vertebrates, and plants. Each value is idenfieid as whole-body or lipid-based and, for the
generic aquatic ecosystems, the biological uptake factors are deignated with a "d" if the value reflects
dissolved water concentrations, and a "t" if the value reflects total surface water concentrations. For
organic chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
August 1995
-------
APPENDIX B Diethyl phthalate - 6
assumed to refer dissolved water concentrations (i.e., dissolved water concentration equals total water
concentration). For organic chemicals with log Kow values above 4, the BCFs were assumed to refer
to total water concentrations and concentrations in fish. The brief discussion proceeding Table 4
describes the rationale for selecting the biological uptake factors and provides the context for
interpretting the biological uptake values.
The bioconcentration factor for fish was estimated from the Veith (1980) equation for phthalates. The
measured BCF from Stephan (1993) was not used because the value may be artifically high
since it is based on uptake of radioactivity with no verification of the parent chemical.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e.; parathion BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Diethyl phthalate - 7
Table 4. Biological Uptake Properties
9QOtOQl00l
receptor
fish
littoral trophic
10*42
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAF.or
8SAF
BCF
-
BAF
BCF
BCF
BCF
JipfcMMMdor
Mftirtt* fcirt Ar
WfKHPnKWy
lipid
-
whoto-body
whote-body
whole-body
whole-plant
»*hie
500 (d)
ID
2.7 E -06
2.6 E - 06
2.1 E-05
1.7
QUKW
predicted; Veith et a).. 1980
calc
calc
calc
U.S. EPA, I990e
d = retere to dissolved surface water concentration
t » retort to total surface water concentration
ID = loiufficieiH D»u
August 1995
-------
APPENDIX B Diethyl phthalate - 8
References
Barrows, M. E., S. R. Petrocelli, and K. J. Macek. 1980. Bioconcentration and elimination
of selected water pollutants by bluegill sunfish (Lepomis macrochirus). In: Dynamics,
Exposure and Hazard Assessment of Toxic Chemicals. R. Haque, Ed. Ann Arbor Science
Pub. Inc., Ann Arbor, MI. pp. 379-392. As cited in Stephan, C.E. 1993. Derivations of
Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative. PB93-154672. Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
Brown, D., K. R. Butterworth, I. F. Gaunt, P. Grasso, and S. D. Gangolli. 1978. Short-term
oral toxicity study of diethyl phthalate in the rat. Food Cosmet. Toxicol. 16:415-422.
Food Research Laboratories, Inc. 1955. Toxicological studies of diethyl phthalate. Laboratory
No. 67567, Celanese Corporation of America, Summit Research Laboratories, Summit,
NJ. As cited in U.S. EPA (Environmental Protection Agency), IRIS (Integrated Risk
Information System). March 1994.
Geiger, D.L., C.E. Northcott, D.J. Call, and L.T. Brooke. 1985. Acute toxicities of organic
chemicals to fathead minnows (Pimephales promelas), Vol. 2. Center for Lake Superior
Environmental Studies, University of Wisconsin, Superior, WI:326 p. As cited in
AQUIRE (AQUatic Toxicity Information ^Etrieval Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
National Library of Medicine. HSDB (Hazardous Substance Database). 1994.
NTP (National Toxicology Program). 1984. Diethyl Phthalate: Reproduction and fertility
assessment in CD-1 mice when administered in the feed. Final report. NTP, Research
Triangle Park, NC. As cited in U.S. EPA (Environmental Protection Agency). IRIS
(Integrated Risk Information System). March 1994.
Singh, A. R., W. H. Lawrence, and J. Autian. 1972. Teratogenicity of phthalate esters in
rats. Journal of Pharmaceutical Sciences 61(l):51-55.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Diethyl phthalate - 9
Suter n, G. W. and J.- B. Mabrey. 1994. Toxicological Benchmarks for Screening of
Potential Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-
AC05-84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction. '
Environmental Toxicology and Chemistry. 11:615-629.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D. C. January.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
Veith, G. D. and K. J. Macek, S. R. Petrocelli and J. Carroll. 1980. An evaluation of using
partition coefficients and water solubility to estimate bioconcentration factors for organic
chemicals in fish. /. Fish. Res. Board Can. ( Prepublication copy) As cited in Lyman,
W. J., W. F. Reehl and D. H. Rosenblatt. 1990. Handbook of Chemical Proper?
Estimation Methods. American Chemical Society, Washington, D. C. p. 5-4.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxicity - Diethyl phthalate Cas No.: 84-66-2
'
Chemical
Name
diethyl
phthalate
diethyl
phthalate
diethyl
phthalate
^
diethyl
phthalate
diethyl
phthalate
diethyl
phthalate
Species
rat
rat
mice
rats
rat
guinea pig
NS - Not Specified
Endpoint
growth
growth
rep
fet
behv. dev _j
behv, dve
Description
LOAEL
Value
3160
LOAEL
NOAEL
LOAEL
LD50
LD50 -
5
2.5
0.506
8600
8600
Units
mg/kg-day
%
%
ml/kg
mg/kg-body wt.
mg/kg-body wt.
Exposure
Route (oral,
S.C., I.V., l.p.,
injection)
oral
Exposure
Duration
/Timing
1 6 weeks
oral
oral
i.p. injection
oral
oral
2 years
1 8 weeks
days 5,10
and 15 of
gestation
NS
NS
;
Reference
Brown etal., 1978
Food Research
Lab.. 1955 as cited
in IRIS, 1994
NTP, 1984 as cited
in IRIS, 1994
Singh etal., 1972
RTECS. 1994
RTECS, 1994
Comments
Decreased growth rate, food
consumption and organ
weights were observed at
this dose level.
Growth of the animals was
retarded throughout the study
at this dose level. (0.5%,
2.5%, and 5.0%)
Reproductive pertomance
was not altered at this dose
level, the highest dose of
three.
Skeletal abnormalities were
noted at this dose level, the
lowest dose level of three
dose levels.
-------
Freshwater Toxicity - Diet. . phthaiate Cas No.:84-66-2
Chemical
Name
diethyl
phthaiate
. Specie?
aquatic
organisms
diethyl
phthaiate
fathead
minnow
NS = Not Specified
Type of
Effect
chronic
mort.
Description
scv
LC50
Value
Units
220 ug/L
31.800
ug/L
Test Type
(Static/Flow
Through)
NS
NS
Exposure
Duration
/Timing
Reference
Suter and Mabrey,
NS J1994
96-hour
Comments
Geiger etal., 1985
as cited in AQUIRE,
1995
-------
Freshwater Biological Uptake Measures - Diethyl phthalate Cas No.: 84-66-2
Chemical
Name
diethyl
phthalate
diethyl
phthalate
diethyl
phthalate
= BCF value
NS = Not Spe
Species
fish
fish
fish
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
s may have come from sing
cified |
Value
73-
B.54
24.38'
e source
Measured
or
Predicted
(m.P)
m
P
m
Units
L/kg
NS
NS
Reference
U.S. EPA, 1992
Slephan, 1993
Barrows et al., 1980 as
cited in Stephan, 1993
Comments
Normalized to 3%
lipjd.
Normalized to 1 .0%
lipids.
Normalized to 1 .0%
lipids.
24.33
-------
Terrestrial Bioiogicai Uptake Measu. Diethy! prtirtaiate Gas No.: B4=66-2
Chemical
Name
diethyl
phthalate
Species
plant
B-factor
(BCF, BAF,
BMP)
BCF
Value
1.7
Measured
or
Predicted
(m.P)
P
units
(ug/g DW
plant)/(ug/g
soil)
Reference
Comments
.
U.S. EPA, 1990e
-------
APPENDIX B Dimethyl phthalate - 1
lexicological Profile for Selected Ecological Receptors
Dimethyl phthalate
CasNo.: 131-11-3
Summary: This profile on dimethyl phthalate summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
/
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (CJ for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Mammals: Plasterer et al. (1985) investigated the effects of oral dimethyl phthalate exposure
in laboratory mammals. Pregnant mice were given 3500 mg/kg-day in corn oil by gavage on
gestation days 7 through 14. In a similar study by Hardin et al. (1987) pregnant mice were
given 3500 mg/kd-day in distilled water or corn oil by gavage on gestation days 6 through 13.
No toxic effects were observed in treated mothers or in their offspring in either study. Since
no adverse effects on reproductive endpoints were identified, benchmark values protective of
the mammalian community were not derived.
Birds: Toxicity data were not identified involving dimethyl phthalate toxicity in avian species
and, therefore benchmarks were not derived.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not available for
dimethyl phthalate. Therefore, the Tier II method described in Section 4.3.5 was used to calculate and
August 1995
-------
APPENDIX B Dimethyl phthalate - 2
Secondary Chronic Value (SCV) of 140 mg/L. Tier II values or Secondary Chronic Values were
developed so that the aquatic benchmarks could be established for chemcials with data sets that did not
fulfill all the requirements of the National AWQC. Because the benchmark is based on an SCV, it
was categorized as interim.
Aquatic plants: The lexicological benchmarks for aquatic plants were either (1) a no observed
effects concentration (NOEQ or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or (2) an effective concentration (EC^) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum capricornutum). Adequate data for the
development of benchmarks for dimethyl phthalate were not identified in Suter and Mabrey (1994) or
in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined using the
Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value (FCV) or SCV,
along with the fraction of organic carbon and the octanol-carbon partition coefficient (K^ to
determine protective sediment concentration (Stephan, 1993). The EQp number is the chemical
concentration that may be present in the sediment while still protecting the benthic community from
the harmful effects of chemical exposure. The SCV for dimethyl phthalate, as calculated from the
AQUIRE database was used to calculate and EQp value of 5.78 mg dimethyl phthalate/kg organic
carbon. Assuming a mass fraction of organic carbon for the sediment (f^) of 0.05, the benchmark for
the benthic community is 0.29 mg/kg sediment. Since the EQp number was based on an SCV, the
sediment benchmark was categorized as interim.
August 1995
-------
APPENDIX B
Dimethyl phthalate 3
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ftofvwwtaiw
Slt*0i*0
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser tcaup
spottad sandpiper
herring guH
kingfisher
BwwhBMrft
V«iu* maty*.
**y
ID
ID
ID
ID
ID
ID
ID
ID
ID
ID
*u*
,ap«ofa«
-
-
.
-
-
-
-
-
itt*ci
- .
-
-
-
-
Study V*faNi
*aftHNr
-
-
-
. ffcuM*|M|tlttft
r^^l^^*^^^*
-
-
-
-
-
-
-
° W '
-
-
-
-. J.
' -
'
-
-
. '
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value-was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and aquatic
invertebrates
aquatic plants
benthic community
140 (i)
No data
0.29 (i)
Stody
scv
AQUIRE. 1995
SCVxK^ AQUIRE. 1995
Benchmark Category, a - w*wn*to. p = provisional, i = interim; a "" indicates that the benchmark value was an
order of magnitude or more above (he NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Dimethyl phthalate - 4
n. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cp^ for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem.
Mammals: As mentioned in the freshwater ecosystem discussion, no suitable subchronic or chronic
lexicological studies were found for mammalian wildlife exposure to dimethyl phthalate. Therefore, a
mammalian benchmark for terrestrial ecosystems was not derived.
Birds: No avian toxicity studies were identified for dimethyl phthalate and therefore, benchmark
values protective of avian species were not derived.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root lengths. As presented in Will and Suter (1994), phytotoxicity benchmarks were selected
by rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or
fewer values for a chemical, the lowest LOEC was used. If there were more than 10 values, the
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or
other effects reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for dimethyl
phthalate and, as a result, a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Dimethyl phthalate - 5
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
R»pr«eenlati»i
%»*de*
dear mouse
short-tailed
shrew
meadow vote
Eastern
cottontail
rod fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern
bobowhite
American robin
American
woodcock
plants
toil community
Banchm**
Vain**
Wgflt9**t
ID
ID
ID
ID
ID
ID
ID .
ID
ID
ID
ID
ID
No data
No data
Study
Specie*'
-
- ,
-
-
-
-
Effect
-
-
-
-
-
-
Study -
Vrt»
% *«**
4*Y
-
/
-
'
-
-
\ .
Description
.
- .
.
. -
-
.
-
' 8F
-
-
-
-
-
-
-'
CMf0nai Soura*
*
-
-
-
-
-
- -
-
-
-
Benchmark Category, a = adequate, p - provisional, i = interim; a "' indicates (hat the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
m. Biological Uptake Measures
This section presents the biological uptake measures (i.e., BCFs, and BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcnetrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and sources
are presented in Table 4 for ecological receptor categories: tropic level 3 and 4 fish in the limnetic and
littoral ecosystems, general fish (BCF only), aquatic invertebrates, earthworms, other soil invertbrates,
terrestrial vertebrates, and plants. Each value is idenfieid as whole-body or lipid-based and, for the
generic aquatic ecosystems, the biological uptake factors are deignated with a "d" if the value reflects
dissolved water concentrations, and a "t" if the value reflects total surface water concentrations. For
organic chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
August 1995
-------
APPENDIX B Dimethyl phthalate - 6
assumed to refer dissolved w^r concentrations (i.e., dissolved water concentration equals total water
concentration). For organic chemicals with log K,,w values above 4, the BCFs were assumed to refer
to total water concentrations and concentrations in fish. The brief discussion proceeding Table 4
describes the rationale for selecting the biological uptake factors and provides the context for
interpreting the biological uptake values.
The bioconcentration factor for fish was estimated from the Veith (1980) equation for phthalates. The
measured BCF from Stephan (1993) was not used because the value may be artifically high
since it is based on uptake of radioactivity with no verification of the parent chemical.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e., parathion BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Dimethyl phthalate - 7
Table 4. Biological Uptake Properties
OTPJOfjhnfll
: tetiepiof
fish
trophic lavol 2
invertebrates
tarrettriai
vertebrates
terrestrial
invertebrates
earthworms
plants
8C£,«AF»ar
BSAF
BCF
BAF
BCF
BCF
BCF
lipkUNMtd or
Mifcii>iat tut iiii
wmwoooy
lipid
whole-body
whole-body
whole-body
whole-plant
vatoe
100 (d)
ID
5.4E-07
5.2E-07 '
4.1E-06
4.4
»
tkounw
predtcted; Veith et a)., 1980
calc
caic .
cak
U.S. EPA. 1990e
d » reiers to dissolved surface water concentration
t = retort to total surface water concentration
ID a Insufficient Data
August 1995
-------
APPENDIX B Dimethyl phthalate - 8
References
Barrows, M. E., S. R, Petrocelli, and K. J. Macek. 1980. Bioconcentration and elimination
of selected water pollutants by bluegill sunfish (Lepomis macrochirus). In: Dynamics,
Exposure and Hazard Assessment of Toxic Chemicals. R. Haque, Ed. Ann Arbor Science
Pub. Inc., Ann Arbor, MI. pp. 379-392. As cited in Stephan, C.E. 1993. Derivations of
Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative. PB93-154672. Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
Geiger, D.L., C.E. Northcott, D.J. Call, and L.T. Brooke. 1985. Acute toxicities of organic
chemicals to fathead minnows (Pimephales promelas), Vol. 2. Center for Lake Superior
Environmental Studies, University of Wisconsin, Superior, WI:326 p. As cited in
AQUIRE (AQUatic Toxicity information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Hardin, B.D., R.L. Schuler, J.R. Burg, G,M. Booth, K.P. Hazelden, K.M. MacKenzie, V.J.
Piccirillo, and K. N. Smith. 1987. Evaluation of 60 chemicals in a preliminary
developmental toxicity test. Teratogenesis, Carcinogenesis, and Mutagenesis. 7:29-48.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
National Library of Medicine. HSDB (Hazardous Substance Database). 1994.
Plasterer, M.R., W.S. Bradshaw, G.M. Booth, and M.W. Carter. 1985. Developmental
toxicity of nine selelcted compounds following prenatal exposure in the mouse:
naphthalene, p-nitrophenol, sodium selenite, dimethyl phthalate, ethylenethiourea, and four
glycol ehter derivatives. Journal of Toxicology and Environmental Health, 15: 25-38.
Singh, A. R., W. H. Lawrence, and J. Autian. 1972. Teratogenicity of phthalate esters in
rats. Journal of Pharmaceutical Sciences 61(l):51-55.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter n, G. W. and J. B. Mabrey 1994. Toxicological Benchmarks for Screening of
Potential Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-
AC05-.84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
August 1995
-------
APPENDIX B , Dimethyl phthalate 9
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615-629.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D. C. January.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
U.S. EPA (Environmental Protection Agency). 1994. Integrated Risk Information System.
March.
Veith, G. D. and K. J. Macek, S. R. Petrocelli and J. Carroll. 1980. An evaluation of using
partition coefficients and water solubility to estimate bioconcentration factors for organic
chemicals in fish. /. Fish. Res. Board Can. ( Prepublication copy) As cited in Lyman,
W. J., W. F. Reehl and D. H. Rosenblatt. 1990. Handbook of Chemical Property
Estimation Methods. American Chemical Society, Washington, D. C. p. 5-4.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial toxicity - Diemthyl phthalate Cas No.: 131-11-3
Chemical
Name
dimethyl
phthalate
dimethyl
phthalate
dimethyl
phthalate
dimethyl
phthalate
dimethyl
phthalate
Species
rats
rats
mouse
rat
mouse
NS = Not Specified
Endpolnt
ter
fet
rep
behv, dev
behv, dev
Description
LOAEL
NOAEL
NOAEL
LD50
LD50
Value
0.338
5000
3500
6800
6800
Units
ml/kg
mg/kg-day
mg/kg-day
mg/kg-body wt.
mg/kg-body wt.
Exposure
Route (oral,
B.C., I.V., I. p.,
Injection)
i.p. injection
oral (gavage)
oral (gavage)
oral
oral
Exposure
Duration /Timing
days 5, 10 and 15
of gestation
gestation days 6-
13
day 7 through
day 14 of
gestation
NS
NS
Reference
Singh etal.. 1972
Hardin et. al., 1987
Plasterer etal., 1985
RTECS, 1994
RTECS, 1994
i Comments
Teratogenic effects were
observed at this dose level, the
lowest of three dose levels.
No toxic effects in the treated
mothers or in their offspring. (2
dose levels)
No effect on maternal weight
gain, litter size, or average pup
weight was observed at this
single dose level.
-------
Freshwater Toxiclty-Dirneth; ^hthalates Gas No.: 131-11-3
Chemical
Name
dimethyl
phthalate
Species
fathead
minnow
NS = Not Specified
Type of
Effect
mort.
Description
LC50
Value
.121,000
Units
ug/L
Test Type
(Static/Flow
Through)
NS
Exposure
Duration
/Timing
96-hour
Reference
Geiger et al., 1990 as cited in
AQUIRE. 1995
Comments
-------
Freshwater Biological Uptake Measures - Dimethyl phthalate Cas No.: 131-11-3
Chemical
Name
dimethyl
phthalate
dimethyl
phthalate
dimethyl
phthalate
Species
fish
fish
fish
NS = Not Specified
B-factor
(BCF, BAF.
BMF)
BCF
BCF
BCF
Value
36"
1.38
11.9*
* = BCF values may have come from a single source
Measured
or
Predicted
(m,p)
m
m
Units
L/kjL
NS
NS
Reference
U.S. EPA, 1992
Stephan, 1993
Barrows et al., 1980 as cited
in Stephan, 1993
Comments
Normalized to 3%
lipjd.
Normalized to 1 .0%
lipids.
Normalized to 1 .0%
lipids.
- -
-------
Terrestrial Biological Uptake Measure. Jimethyl phthalate cas No.: 131-11-3
Chemical Name
dimethyl phthalate
Species
plant
B-factor
(BCF. BAF,
BMP)
BCF
Value
4.4
Measured
or
Predicted
(m,P)
P
units
(ug/g DW plant)/(ug/g
soil)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Endosulfan .
lexicological Profile for Selected Ecological Receptors
Endosulfan
CasNo.: 115-29-7
Summary: This profile on endosulfan summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulatio'n, and biorriagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
KOW between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No subchronic or chronic studies on mammalian wildlife were found in which
dose-response data were reported. However, several chronic and subchronic toxicity studies
involving endosulfan have been conducted using laboratory rats, mice and dogs. Hoechst
(1984) conducted a reproductive sludy in which rals were fed a diet of endosulfan for a
August 1995
-------
APPENDIX B Endosulfan - 2
period of 84 days or two generations. At the highest dose administered, Hoechst (1984)
reported a NOAEL of 3.8 mg/kg-day for reproductive effects. In another reproductive study,
(Hoechst, 1988; NCI, 1978), mice fed dietary concentrations of endosulfan for 2 years
exhibited no reproductive effects at a dose of 2.51 mg/kg-day. As in the earlier mentioned
Hoechst (1984) study, these two studies observed no adverse reproductive effects at the
highest dose levels. Increased mortality was observed in a two-year study in which 25 male
and 25 female rats were fed dietary concentrations of 10, 30, and 100 ppm of endosulfan
(FAO/WHO, 1968). Abnormalities in weight gain and hematological parameters, and effects
on kidney size and function were also observed, but only in the highest dose group. A
NOAEL of 1.5 mg/kg-day (30 ppm) was reported for this study (FAO/WHO, 1968). In a
subchronic developmental study with rabbits, FMC (1981) reported that there were no signs
of developmental toxicity at doses equal to or less than 1.8 mg/kg-day, the highest
administered dose. Gupta and Chandra (1977) reported an NOAEL of 5 mg/kg-day for
changes in testicular weight after their treatment of male albino rats. This study administered
oral doses of 0.0, 5.0 and 10.0 mg/kg-day for a period of 15 days.
The NOAEL in the Gupta and Chandra (1977) study was chosen to derive the lexicological
benchmark because: (1) the study contained sufficient dose-response information, and reported
significant effects on a reproductive endpoint. The studies by Hoechst (1984), Hoechst
(1988), and NCI (1978) were not selected for the derivation of a benchmark because they
lacked dose-response data. In each of these sf.dies, the NOAEL was based on the lack of
effects seen at the highest dose group. The FAO/WHO (1968) study was not selected as the
basis for benchmark derivation because it did not assess developmental or reproductive
endpoints.
The NOAEL from the Gupta and Chandra (1977) study was scaled for species representative
of a freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994)
( bw N1/4
Benchmark. = NOAEL, x __L
August 1995
-------
APPENDIX B Endosulfan-3
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Gupta and Chandra (1977) study documented reproductive effects from endosulfan exposure
to male rats, male body weights for each representative species were used in the scaling
algorithm to obtain the lexicological benchmarks.
Data were available on reproductive and developmental effects as well as on growth or
chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. All of the studies identified
were conducted using laboratory animals and as such, inter-species differences among wildlife
species were not identifiable. Therefore, an inter-species uncertainty factor was not applied.
There were several study values in the data set which were lower than the benchmark value
by at least an order of magnitude. These values corresponded to effects on behavioral and
reproductive endpoints and effects on chronic survival. Based on the data set for endosulfan
and the terrestrial benchmarks developed from Gupta and Chandra (1977) were categorized as
adequate, with a "*" to indicate that adverse effects may occur at the benchmark level.
Birds: No subchronic or chronic studies on representative or surrogate avian species were
located. Sources reviewed for avian toxicity ^formation included: Endosulfan (WHO,
1984); an on-line search of the TOXLIT, RTECS, and DART databases; and an extensive
library search at National Institute for Environmental Health Statistics (NIEHS) library. As a
result, no avian toxicity benchmark was developed.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 5.6E-5 mg/1 reported in
the AWQC document for endosulfan was (U.S. EPA, 1980) selected as the benchmark value
protective of fish and aquatic invertebrates. Because the benchmark is based on an FCV
developed for a AWQC, it was categorized as adequate.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECU) for a species
of freshwater algae, frequently a spec'rSs of green algae (e.g., Selenastrum capricprnutum).
Aquatic plant data was not identified for endosulfan and, therefore, no benchmark was
developed.
August 1995
-------
APPENDIX B
Endosulfan - 4
Benthic Community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the pctanol-carbon partition coefficient (K^.) to determine a protective chemical concentration
(U.S. EPA, 1993c). The EQP number is the chemical concentration that may be present in
sediment while still"protecting the benthic community from the harmful effects of chemical
exposure. The Final Chronic Value (FCV) reported in the document for endosulfan was used
to calculate a EQP number of 0.148 mg endosulfan /kg organic carbon. Assuming a mass
fraction of organic carbon for the sediment (f^) of 0.05, the benchmark for the benthic
community is 7.38E-3 mg/kg. Since the EQP number was based on a FCV established for the
AWQC, the sediment benchmark is categorized as adequate.
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
mink
2.9 (a')
rat
rep
NOAEL
Gupta and Chandra.
1977
river otter
1.8 (a*)
rat
rep
NOAEL
Gupta and Chandra,
1977
bald eagle
10
osprey
ID
great blue heron
ID
mallard
ID
lesser scaup
ID
spotted sandpiper
ID
herring gull
ID
kingfisher
ID
Benchmark Category, a a adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order
'of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B
Endosulfan 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and aquatic
invertebrates
aquatic plants
benthic community
5.6E-05 (a)
ID
7.38E-03 (a)
mg/Vg sediment
*"%;* j:: ^
AWQC species
A WQC species
FCV
FCVxK.
U.S EPA. 1980
U.SEPA, 1980
II.
'Benchmark Category, a = adequate, p = provisional, i a interim; a "' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
ID 3 insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants, and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to endosulfan.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Gupta and Chandra, 1977) was used to derive the endosulfan lexicological benchmark
for mammalian species representing the terrestrial ecosystem. The selected study NOAEL
was scaled to species representative of the terrestrial ecosystem using the inter-species scaling
method of Opresko et al. (1994). ^
Based.on the data set for endosulfan and because the selected study value is not the lowest
August 1995
-------
APPENDIX B Ehdosulfan - 6
NOAEL in the data set,"the mammalian benchmarks were categorized as adequate, with a "*"
to indicate that adverse effects may occur at the benchmark level.
Birds: Although numerous sources were reviewed for toxicity information, no subchronic or
chronic studies were identified for representative or surrogate avian species. Therefore, no
benchmarks were developed for birds in the terrestrial ecosystem.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. Phytotoxicity benchmarks were selected as the lowest
concentration identified for plant growth, yield reductions, or other effects reasonably
assumed to impair the ability of a plant population to sustain itself, such as a reduction in
seed elongation. However, adequate data with which to derive a benchmark protective of the
plant community were not identified. '
Soil community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Endosulfan 7
Table 3. lexicological Benchmarks for Representative Mammals and Birds Associated
with Terrestrial Ecosystem
g>-^j ^-^^^^A^aA.j»
noptOTwrauiw
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail '
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American
plants
soil community
GlMMilllt^Ck*
VriH-mgn*.
,; **,-. .
8.4 (a')
8.7 (a')
6.9 (a«)
3.1 (a')
2.1 (a')
2.0 (a')
1.0 (a')
ID
ID
ID
ID
10
10
ID
i'£3£:-
rat
rat
rat
rat
rat
rat
rat
-
'
-
-
-
"Hhielv'
g. ;:. ; .«,
rep
rep
rep
rep
rep
rep
rep
-
-
-
-
-
-
:'OHgtmt,
V^o»
;."*B(Nfe*»'
5
5
5
5
5
5
5
-
-
-
-
-
-
E-~""
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
-
-
-
-
^ tr
-
-
-
-
.-
-
-
-
fkMnaf firunw
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
Gupta and
Chandra, 1977
-
-
-
- .
'
Benchmark Category, a 3 adequate, p * provisional, i = interim; a "" indicates that the benchmark value was an order
of magnitude or more above the NEL or tSt. for other adverse effects.
.ID = Insufficient data
August 1995
-------
APPENDIX B . . ' N Endosulfan - 8
III. Biological Uptake Measures
This section presents biological uptake measures (i.e. BCFs, BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for selected ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For the generic aquatic ecosystems, the BCF value is
identified as whole-body or lipid-based and designated with a "d" if the value reflects
dissolved water concentrations, and a "t" if the value reflects total surface water
concentrations. For organic chemicals with log K^ values below 4, bioconcentration factors
(BCFs) in fish were always assumed to refer to dissolved water concentrations (i.e., dissolved
water concentration equals total water concentration). The following discussion describes the
rationale for selecting the biological uptake factors and provides the context for interpreting
the biological uptake values presented in Table 4.
The bioconcentration factor for fish was estimated from the Thomann (1989) model (i.e., log
K^ - dissolved BCF/) because: (1) no appropriate measured values were identified, (2) the
BCF was in close agreement with predicted.BCFs based on other methods (i.e., regression
equations), and (3) there were no data (e.g., metabolism) to suggest that the log K,,w = BCF,d
relationship deviates for endosulfan (log Kow --3.48). As stated in section 5.3.2, the dissolved
bioconcentration factor (BCF,d ) for organic chemicals with log K^w below 4 was considered
to be equivalent to the total bioconcentration factor (BCF/) and, therefore, adjusting the BCF,d
by the dissolved fraction (fd) was not necessary.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e., endosulfan BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
August 1995
-------
APPENDIX B
Endosuffan - 9
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
Table 4. Biological Uptake Properties
roccpior
fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAP, or
BSAF
BCF
BAF
BAF
BCF
BCF
. BCF
wffiotewooy
liptd
lipid
whole-body
whole- body
whole-body
whole-plant
VtiM
2,990 (d or t)
-
3.7E-05
3.6E-05
2.9E-04
0.38
predicted value based on
Thomann. 1989
date under review
estimated based on beef
biotransfer ratio with 2.3.7,8-
TCDO
estimated based on beef
biotransfer ratio with 2,3.7,8-
TCOO
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
U.S. EPA. 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Endosulfan - 10
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE (AOUatic Toxicity /nformation /tetrieval Database), 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Chandler, G.T. and G.I. Scott. 1991. Effects of Sediment-Bound Endosulfan on Survival,
Reproduction, and Larval Settlement of Meiobenthic Polychaetes and Copepods.
Environmental Toxicology and Chemistry, Vol. 10, pp. 375-382.
CEC (Commission of European Communities). 1981. Criteria (Dose/Effect Relationships)
for Organochlorine Pesticides. Pergamon Press. Oxford. As cited in WHO (World
Health Organization), 1984, Endosulfan, Environmental Health Criteria 91, Geneva,
Switzerland.
Ernst, W. 1977. Determination of the bioconcentration potential of marine organisms: a
steady state approach. Chemosphere. 6:""" 1-740.
FAO/WHO (Food Agriculture Organization/ World Health Organization). 1968. Evaluations
of Some Pesticides Residues in Food. Food and Agriculture Organization in the United
Nations, Rome.
FMC (Food Machinery and Chemical, Corp.). 1959. Thiodan technical: Repeated oral
administration - dogs. Final report. Conducted for Food Machinery and Chemical
Corporation, Niagara Chemical Division. Hazleton Laboratories, Inc., Falls Church, VA.
As cited in Agency for Toxic Substances and Disease Registry (ATSDR), 1993,
Toxicological Profile for Endosulfan, Public Health Service, U.S. Department of Health
and Human Services, Atlanta, GA.
August 1995
-------
APPENDIX B Endosulfan -11
FMC (Food Machinery and Chemical, Corp.)- 1965. Three-generation reproduction study in
albino rats on thiodan: Results through weaning of Fib litters. Conducted for Food
Machinery and Chemical Corporation, Niagara Chemical Division. Industrial Bio-Test
Laboratories, Inc., Northbrook, IL. As cited in Agency for Toxic Substances and Disease
Registry (ATSDR), 1993, Toxicological Profile for Endosulfan, Public Health Service,
U.S. Department of Health and Human Services, Atlanta, GA.
FMC (Food Machinery and Chemical, Corp.). 1967. Two-year chronic oral toxicity of
thiodan technical - beagle dogs. Conducted for Food Machinery and Chemical
Corporation. Industrial Bio-Test Laboratories, Inc., Northbrook, DL. As cited in Agency
for Toxic Substances and Disease Registry (ATSDR), 1993, Toxicological Profile for
Endosulfan, Public Health Service, U.S. Department of Health and Human Services,
Atlanta, GA.
FMC (Food Machinery and Chemical, Corp.). 1980. Final Report: Teratology Study with
FMC 5462 in Rats. Conducted for Food Machinery and Chemical Corporation. Raltech
Scientific Services, Madison, WI. Raltech study no. 79041. As cited in Agency for Toxic
Substances and Disease Registry (ATSDR), 1993, Toxicological Profile for Endosulfan,
Public Health Service, U.S. Department of Health and Human Services, Atlanta, GA.
FMC (Food Machinery and Chemical, Corp.) 1981. Teratology study with FMC 5462 in
rabbits. Conducted for Food Machinery and Chemical Corporation. Raltech Scientific
Services, Madison, WI. Raltech study no. 80070. As cited in: IRIS (Integrated Risk
Information System). 1994. U.S. Environmental Protection Agency, Office of Research
and Development, Washington, DC
Gupta, P.K., and S.V. Chandra. 1977. Toxicity of endosulfan after repreated oral
administration to rats. Bull Environ. Contam. Toxicol. 18:378-384.
Gupta, P.K. and R.C. Gupta..1977; Effect of Endosulfan Pretreatment on Organ Weights and
on Phenobarbital Hypnosis in Rats. Toxicol. 7:283-288.
August 1995
-------
APPENDIX B Endosulfan -12
Hoechst. 1984. Effect of Endosulfan-Techriical (Code HOE 02671 O I AT209) on
Reproductive Function of Multiple Generations in the Rat. Conducted for Hoechst
Aktiengesellschaft, Frankfurt, Germany. Huntington Research Centre, Cambridgeshire,
England. HST 204/83768. As cited in Agency for Toxic Substances and Disease
Registry (ATSDR), 1993, Toxicological Profile for Endosulfan, Public Health Service,
U.S. Department of Health and Human Services, Atlanta, GA.
*
Hoechst. 1988. Beta-Endosulfan (Code Hoe 052619 Oi Zc99 0001): Testing for Acute Oral
Toxicity in the Male and Female Wistar Rat. Hoechst Aktiengesellschaft, Frankfurt,
Germany. TOXN No. 83.0113. As cited in Agency for Toxic Substances and Disease
Registry (ATSDR), 1993, Toxicological Profile for Endosulfan, Public Health Service,
U.S. Department of Health and Human Services, Atlanta, GA.
Hoechst. 1989. Endosulfan - Substance Technical (Code HOE 02671 OI ZD96 0002):
Testing for Toxicity by repeated oral administration (1-Year Feeding Study to Beagle
Dogs. Conducted for Hoechst Aktiengesellschaft, Frankfurt, Germany. Project no.
87.0643. As cited in Agency for Toxic Substances and Disease Registry (ATSDR), 1993,
Toxicological Profile for Endosulfan, Public Health Service, U.S. Department of Health
and Human Services, Atlanta, GA.
Macek, K.J., M.A. Lindberg, S. Sauter, K.S. jsuxton, and P.A. Costa. 1976. Toxicity of Four
Pesticides to Water Fleas and Fathead Minnows. Acute and Chronic Toxicity to Acrolein,
Heptachlor, Endosulfan, and Trifluralin to the Water Flea (Daphnia Magna) and the
Fathead Minnow (Pimephales Promelas).
NCI (National Cancer Institute). 1978. Bioassay of endosulfan for possible carcinogencity.
DHEW Publication No. NIH 78-1312. NCI Technical Report Series No. 62,
Carcinogenesis Jesting Program, National Cancer Institute, Bethesda, MD.
Opresko, D.M., B.E. Sample, and G.W. Suter El. 1994. Toxicological Benchamrks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1. Oak Ridge National Laboratory,
Environmental Sciences Division: Oak Ridge, TN.
August 1995
-------
APPENDIX B Endosulfan -13
Pickering, Q.H., and C. Henderson. 1966. The acute toxicity of some pesticides to fish.
Ohio J. Sci. 66(5):508-513. As cited in AQUIRE (AQt/atic Toxicityjnformation
ftEtrieval Database), Environmental Research Laboratory, Office of Research and
Development, U.S. Environmental Protection Agency, Duluth, MM.
Roberts, D. 1972. The assimilation and chronic effects of sub-lethal concentrations of
endosulfan on condition and spawning in the common mussel (Mytilus Edulis). Marine
Biology \6:l\%-\25.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health. Washington, DC.
Schimmel, S.C., A.M. Patrick, and A.J. Wilson. 1977. Acute ioxicity to and
bioconcentration of endosulfan by estuarine animals. In: F.L. Mayer and J.L. Hamelink
(eds.), Aquatic Toxicology and Hazard Evaluation, 1st Symposium, ASTM STP 634,
Philadelphia, PA. As cited in AQUIRE (AOUatic Toxicity /nformation fl£trieval
Database), Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency, Duluth, MN.
Schoettger, R.A. 1970= Toxicology ofThiodan in Several Fish and Aquatic Invertebrates.
Invest. Fish Control No. 35, U.S.D.I. As cited in AQUIRE (AQC/atic Toxicity
/nformation REtrieval Database), Environmental Research Laboratory, Office of Research
and Development, U.S. Environmental Protection Agency, Duluth, MN.
Schoettger, R.A. 1970. Toxicology of Thiodon in Several Fish and Aquatic Invertebrates.
U.S. Department of the Interior, Bureau of Sport, Fish and Wildlife, Investigations in Fish
Control. 35:1-31. As cited in WHO (World Health Organization), 1984, Endosulfan,
Environmental Health Criteria 91, Geneva, Switzerland.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Endosulfan -14
Suter II, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model, of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
U.S. EPA (U.S. Environmental Protection Agency). 1980. Ambient Water Quality Criteria
for Endosulfan. 440/5-80/046. Environmental Criteria and Assessment Office, Office of
Water Regulations and Standards, Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, DC. January.
U.S. Environmental Protection Agency, 1993 - Technical Basis for Deriving Sediment Quality
Criteria for Nonionic Organic Contaminants for the Protection of Benthic Organisms by
Using Equilibrium Partitioning. EPA/822-R-93/011. Office of Water, Washington, D.C.
WHO (World Health Organization). 1984. Endosulfan. International Programme on
Chemical Safety. Environmental Health Criteria 40. Geneva, Switzerland.
Yoshioka, Y. and Y. Ose. 1993. A quantitative Structure-Activity Relationship Study arid
Ecotoxicological Risk Quotient for the Protection from Chemical Pollution.
Environmental Toxicology and Water Quality, Vol. 8, 87-101.
=*».
August 1995
-------
Terrestrial To y - Endosulfan
Cas No. 115-29-7
Chemical
Name
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
Specie*
rat
rat
rat
rat
rat
rats
(togs
rat
rat
Endpolnt
dev
rep
j
rep
systemic
kidney
chronic
chronic
rep
rep
Description
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Value
5
5
10
5
10
1.5
0.75
5.4
2.5
Unit*
mg/Kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mo/kg-day
mg/kg-day
mo/ko-dav
Exposure
Route (oral,
».c., l.v.. l:p.,
Injection)
oral
oral
oral
oral
oral
oral
oral (by
capsule)
ore)
oral
Exposure
Duration/Timing
gestation days 6-
14(1x/day)
15 days, once/day
15 days, once/day
15 days, once/day
15 days, once/day
104 weeks
6 days/week for 10
months
2-generabon study
170 da vs. ad lib
Reference
Gupta and Gupta, 1977
Gupta and Chandra,
1977
Gupta and Chandra,
1977
Gupta and Chandra,
1977
Gupta and Chandra,
1977
FAO/WHO, 1968 as
cited In WHO, 1984
CEC, 1981 as cited in
WHO. 1984
Hoechst, 1984 as died
InATSDR, 1993
FMC. 1965 as cited In
ATSDR. 1993
Comments
No change in ovarian weight
was observed at this dose
level.
No significant change in
body weight and absolute
and relative weights of testes
was observed.
Increased lestes weight and
tubule degeneration were
observed in male rats.
No systemic effects were
observed at (his dose level.
Histopathological alterations
were observed in the
kidneys.
No lexicological effects were
observed.
No lexicological effects were
observed.
No evidence of reproductive
toxiclty was found at any of
the dose levels tested.
(NOAEL = 6.6 mg/kg-day for
emales)
-------
Terrestrial Toxicity - Endosulfan
Cas No. 115-20-7
Chemical
Name
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosulfan
endosultan
endosulfan
endosultan
v
Species
rat
rabbit
mica
male dogs
female dogs
rat
mouse
dog
cat
rabbit
Endoolnt
rep
dev
reP,l
J
rep
rep
acute
acute
acute
acute
acute
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LD50
LD50
LD50
LD50
LD50
Value
3.8
1.8
2.51
2
18
18
7360
76700
2
-
26
Units
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (orar.
8.C., I.V., l.p.,
Inlectton)
oral
oral
oral
mg/kg-day. oral
mg/kg-day
mg/kg-body
wt.
ug/kg-body
wt.
ug/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
84 days (two-
generation study)
gestation days 6-
28
2 years
1 - 2 years
1 - 2 years
NS
NS
NS
NS
NS
Reference
Hoechst. 1984 as cited
inATSDR. 1993
FMC. 1981 as cited in
IRIS, 1994
Hoechst, 1988 as cited
in ATSDR. 1993; NCI,
1978.
FMC. 1959, 1967 and
Hoechst, 1989 as died
in ATSDR, 1993
FMC. 1959, 1967 and
Hoechst, 1989 as died
inATSDR. 1993
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
Comments
No effect on the size.
mortality, or sex ratio of the
litters. ;
No developmental effects
were observed. (NOAEL is
the highest dose level).
No toxic effects on the
reproductive organs.
No toxic effects on the
reproductive organs.
No toxic effects on the
reproductive organs.
-------
Terrestrial Tc -y - Endosulfan
Cas No. 115-29-7
Chemical
Name
endosultan
endosultan
endosulfan
endosultan
Species
hamster
duck
domestic
animal
wild bird
Endpolnt
acute
acute
acute
A
acute
Description
LD50
LD50
LD50
LO50
Value
118
33
26
35 '
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral.
S.C.. I.V.. l.p.,
Inlection)
oral
oral
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
Reference.
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
Comments
' *
NS = Not specified
-------
Freshwater Toxicity - Endosulfan
Cas No. 115-29-7
Chemical
Name
.
endosulfan
endosulfan
endosulfan
endosullan
endosullan
endosultan
endosullan
endosullan
endosultan
endosultan
Species
aquatic
organisms
Daphnia
carinata
Daphnia
longispina
Daphnia
magna
Daphnia
magna
blu'egill
striped. bass
rainbow trout
fathead
minnow
tathead
minnow
Endpolnt
chronic-
immob.
mort
immob.
mow
mort.
mort.
mort.
acute
chron
Description
AWQC
EC50
LC50
EC50
LC50
LC50
LC50
LC50
LC50
CV
Value
0.056
180
0.3
158-720
(345.14)
62.0 - 740
(279.33)
3.3 - 4.4
(3.81)
0.1
0.17-2.43
(0.69)
0.29 - 3.45
(1.29)
0.20-0.40
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ »ow
through)
NA
NA
NA '
NA
NA
NA
NA
NA
NA
partial life
cycle test
Exposure
Duration/
Timing
NS
48 hour
48 hour
48 hour
48 hour
4 days
96 hour
96 hour
96 hour
NS
Reference
U.S. EPA, 1980
Santharam et al.. 1976 as
cited in AQUIRE, 1995
Magadza et al., 1983 as
cited in AQUIRE, 1995
AQUIRE. 1995
AQUIRE. 1995
Pickering and Henderson.
1966 as cited in AQUIRE.
1995
Korn et al.. 1974 as cited in
AQUIRE. 1995
AQUIRE, 1995
AQUIRE. 1995
Maceketal, 1976
Comments
]
«
Critical life stage end points
embryo, larval, and early
juvenile; hatchattlity.
NS = Not specified
-------
Freshwater Biological Ut .8 Measures - Endosuifsn
Cas No. 115-29-7
Chemical
Nam*
endosultan
endosullan
endosulfan
endosullan
endosullan
endosultan
endosullan
endosultan
endosultan
endosultan
endosultan
endosullan
Species
ish
mussel
whole
body)
mussel
whole
body)
mussel
whole
body)
mussel
(whole
body)
goldtish
(liver)
goldtish
(muscle)
white
sucker
(muscle)
white
sucker
(muscle)
white
sucker
(liver)
whHe
sucker
(liver)
striped
mullet
B-factor
(BCF, BAF,
BMF)
BCF
BCF
BCF
BCF
J
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
55.6
t7
It
8.1
600
781
314
65
55
550
695
2755
Measured
or
predicted
(m,p)
P
m
m
m
m
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Slephan.1993
Roberts, 1972
Roberts. 1972
Roberts. 1972
Ernst, 1977
! ftoettger. 1970 as cited in WHO.
1984
Schoettger. 1970 as cited in WHO.
1984
Schoetlger . 1970 as cited in WHO.
1984
Schoetlger. 1970 as cited in WHO.
1984
Schoettger. 1970 as cited in WHO,
1984
Schoettger, 1970 as cited in WHO.
1984
Schimmel el al. , 1 977 as cited in
AQUIRE, 1994
Comments
Normalized to 1 .0% lipid; value derived tor
alpha and beta isomers ol endosullan
Exposure time =112 days; dose (ug/L) = 100.
Exposure time =112 days; dose (ug/L) = 500
Exposure time =112 days; dose (ug/L) =
1000.
Exposure time =112 days; dose (ug/L) =
0.14.
Exposure time = 1 1-20 days; dose (ug/L) = 7.
Exposure time = 5-20 days; dose (uq/L) - 7.
Exposure time =12 hours; dose (ug/L) = 20;
temperature at 19 C.
Exposure time = 9 hours; dose (ug/L) = 20;
temperature at 19 C.
Exposure time = 12 hours; dose (ug/L) = 20;
temperature at 19C.
Exposure time = 9 hours; dose (ug/L) = 20;
temperature at 19 C.
Lite stage = 25.6 MM; 28 day lest
NS - Not speeded
-------
Terrestrial Biological Uptake Measures - Endosulfan
Cas No. 115-29-7
Cnamlcal
Nam*
endosullan
Spaclaa
plants
B-factor
(BCF, BAF.
BMF)
BCF
Value
15
Maaaurad 01
pradlctad
(m.p)
P
UnlU
(ug/gWW
plant)/(ug/g soil
water)
Rafaranca
U.S. EPA. 19906
Comments
-------
APPENDIX B Endrin
Toxicological Profile for Selected Ecological Receptors
Endrin
Cas No.: 72-20-8
Summary: This profile on endrin summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e.. Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs. for the
freshwater ecosystem were calculated for organic constituents with log Kow between 4 and
6.5. For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire lexicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from data presented in the technical support document for
the Hazardous Waste I (identification Rule (HWIR): Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (CLJ for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found which reported dose-
response data for mammalian wildlife. However, lexicological studies involving endrin
exposure to mammals have been conducted using laboratory mice and other rodents. In a
chronic study designed to test the effects of endrin on reproduction, Good and Ware (1969)
fed CFW Swiss mice 5 mg/kg of dietary endrin for 120 days, starting 30 days prior to
mating. The PEL (frank effecis level) of 5 mg/kg produced significant parent mortality, and
smaller litters. The reported 5 mg/kg dietary dose corresponded to a daily dose of 0.93
mg/kg-day based on the geometric mean body weight of 0.0297 kg for laboratory mice (U.S.
EPA, 19881) and the derived food consumption rate of 0.0055 kg/day (Nagy, 1987). It was
noted that reproductive effects observed at the 0.93 mg/kg-day dose level were directly
August 1995
-------
APPENDIX B Endrin - 2
attributed to parent mortality due to exposure to endrin. In a study by Ottolcnghi et al.
(1973), pregnant Syrian golden hamsters and GDI mice were administered, via oral
intubation, single doses of endrin in corn oil. Hamsters were administered 5 mg/kg-day, on
days 7, 8, or 9 of gestation and mice were given 2.5 mg/kg-day on day 9 of gestion.
Statistically significant increases in fetal deaths were observed in hamsters treated on day 7 or
8 (5 mg/kg inferred as PEL). Teratogenic effects were observed in both hamsters (fused ribs)
and mice (open eye and cleft palate), with the frequency and gravity of the defects being less
pronounced in the mice. Fetal death and weight reduction were not observed in the mice but
the 2.5 mg/kg dose was inferred as a PEL based on teratogenic effects. Kavlock et al. (1981)
examined the development of fetal mice in a study in which mice were administered endrin
via gastric intubation in doses of 0.5, 1.0, 1.5 and 2.0 mg/kg/day. Endrin was noted to be
fetotoxic in the mouse, as evidenced by dose related decreases in fetal weight and skeletal
and visceral maturation. A LOAEL of 1.0 mg/kg-day and a NOAEL of 0.5 mg/kg-day was
inferred based on these fetotoxic effects.
The NOAEL for fetotoxic effects from the Kavlock et al. (1981) study was chosen to derive
the lexicological benchmark because (1) chronic exposures were administered via oral
intubation, (2) the study focused on longterm reproductive success as a critical endpoint, (3)
the study contained dose response information, and (4) the study contained the lowest toxicity
value for a critical endpoint. The Good and Ware (1969) and Ottolenghi et al. (1974) studies
were not chosen for the derivation of the benchmark primarily because they did not contain
sufficient dose response information. Therefore, the NOAEL of 0.5 mg/kg-day from the
Kavlock et al. (1981) study was chosen for the derivation of a mammalian benchmark value.
The study value from Kavlock et al. (1981) was scaled for species representative of a
. freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogeniciry assessments and reponable quantity
( bw >/4
. Benchmark = NOAEL. x L
VKJ
documents for adjusting animal data to^an equivalent human dose (57 FR 24152). Since the
Kavlock et al. (1981) study documented fetotoxic effects from endrin exposure to female rats,
the mean female body weight of representative species was used in the scaling algorithm to
obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of endrin, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. Based on the data set
for endrin, the benchmarks developed from the Kavlock et al. (1981) study were categorized
as adequate.
August 1995
-------
APPENDIX B Endrin - 3
Birds: Only two chrenic studies were identified that investigated the effects of endrin toxicity
on avian species. In a study examining the reproduction of mallard ducks conducted by
Roylance et al. (1985), duck pairs were fed diets containing 0.5 and 3.0 ppm endrin for
approximately 20 weeks. Although egg production, fertility, and hatchability were not
affected, there was a 9.6% drop in embryo survival in the 3.0 ppm treatment group. This
endpoint resulted in an inferred NOAEL of 0.5 ppm and a LOAEL of 3.0 ppm. These dietary
doses correspond to daily doses of 0.028 and 0.17 mg/kg-day, respectively. The ppm doses
were converted to daily doses using the male and female mean body weight of 1.162 kg and
the reference food intake rate of 0.064 kg/day (U.S. EPA, 1993g) for mallard ducks. In a
similar study, Spann et al. (1986) administered doses of 1.0 and 3.0 ppm endrin to mallards
for approximatly 10 weeks. These doses corresponded to 0.057 and 0.17 mg/kg/day based on
the geometric mean of 1.126 kg for the male and female body weights taken from the study,
and the mean food intake rate of 0.064 kg/day for male and female mallards (U.S. EPA,
1993g). While the authors noted that the birds receiving the 3 ppm dose appeared to
reproduce more poorly than controls, any differences were not demonstrated to be statistically
significant. .
The NOAEL of 0.028 mg/kg-day from the Roylance et al. (1985) study was selected to derive
the avian benchmark value for the freshwater ecosystem. This study was chosen because (1)
chronic exposures were administered via oral ingestion, (2) reproductive toxicity was one of
the primary endpoints examined, and (3) the study contained sufficient dose-response
information. The Spann et al. (1986) study was not chosen due to the lack of adequate dose-
response information.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian
species representative of a freshwater ecosystem, the NOAEL value of 0.028 mg/kg-day from
the Roylance et al. (1985) study was scaled using the cross-species scaling method of
Opresko et al. (1994). Since Roylance et al. (1985) administered dietary doses of endrin to
both male and female mallards the mean of the male and female body weights for each
representative species was used in the scaling algorithm to obtain the lexicological
benchmarks.
Data were available on reproductive and developmental effects of endrin as well as on growth
and survival. In addition, the data set contained studies that were conducted over chronic and
subchronic durations as well as durin^a sensitive life stage. There were no other values in
the data set that were at least an order of magnitude below the benchmark value. Based on
the avian data set for endrin, the benchmarks developed from the .NOAEL in the Roylance et
al. (1985) study were categorized as adequate.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 6.1E-05 mg/L for endrin
was selected as the benchmark protective of fish and aquatic invertebrates (U.S. EPA,
1993m). Since the FCV was derived in the sediment quality criteria document, the
benchmark was categorized as adequate.
August 1995
-------
APPENDIX B Endrin - 4
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEQ for vascular aquatic
plants (e.g. duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). For endrin there
was insufficient data for the development of a benchmark value.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The FCV, taken from the sediment quality criteria,'for endrin was used to calculate
an EQp number of 7.8 mg endrin per kg organic carbon. Assuming a mass fraction of
organic carbon for the sediment (f^.) of 0.05, the benchmark for the benthic community is
3.9E-01 mg endrin per kg of sediment. Because the EQ, number was set using a SCV
derived from sediment quality criteria, it was categorized as adequate.
<*»
August 1995
-------
APPENDIX B
Endrin 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Reprwseelftttv*
mink
river otter
bald eagle
o spray
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benotunarit
Value" mg*^
4*t ^ ;
0.23 (a)
0.13(a)
0.021 (a)
0.025 (a)
0.023 (a)
0.028 (a)
0.031 (a)
\
0.063 (a)
0.028 (a)
0.046 (a)
Slujfr
\8ped**
mouse
mouse
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
fitted
feto
feto
rep
rep
reP.
rep
rep
rep
rep
rep
Study Value
wo/fcfday
o.s
0.5
0.028
0.028
0.028
0.028
0.028
0.028
0.028
0.028
Description
- -.H
NOAEL
NQAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
\'"9f 'V
#VV^*:
^i^^^'^^i
^Otflwiif tt&tmia; ' -
Kavtock et al., 1981
Kavtock et ).. 1981
Roytance et al., 1985
Roytance et al.. 1986
Roytance et al., 1986
Roytanee et al., 1985
Roytance et al., 1985
Roytance et al.. 1985
Roytance et «i. 1985
Roytance et al., 1985
Benchmark Category, a « adequate, p provisional, i » interim; a "' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL tor other adverse effects.
Table 2. Torieotogical Benchmarks tor Representative Hah
Associated with Freshwater Ecosystem
yJRepnieenlatfae
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark
*: * **%
wfljf?^.
6. IE-OS (a)
10
3.9E-01 (a)
Study
Specie*
aquatic
organisms
aquatic'
organisms
^
FCV
'
FCV x K^
5 \ , "
SQC
-
SQC
Benchmark Category, a - adequate, p » provisional, i * interim; a '*' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
ID - insufficient data
August 1995
-------
APPENDIX B Endrin. 6
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ^ for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to endrin.
Because of the lack of additional mammalian toxicity studies, the same surrogate species
study (Kavlock et al., 1981) was used to derive the endrin lexicological benchmark for
mammalian species representing the terrestrial ecosystem. The study NOAEL of 0.5 mg/kg-
day was scaled for species in the terrestrial ecosystem using a cross-species scaling algorithm
developed by Opresko et al. (1994). Since the Kavlock et al. (1981) study documented
reproductive effects from endrin exposure to female rats, the female body weight of each
representative species was used in the scaling algorithm to obtain the lexicological
benchmarks.
Based on the data set for endrin the benchmarks developed from the Kavlock et al. (1981)
study were categorized as adequate.
Birds: As in the freshwater ecosystem, the study by Roylance et al. (1985) was used to
calculate the benchmarks for birds in the generic terrestrial ecosystem. The study NOAEL of
0.5 ppm (0.028 mg/kg-day) was scaled for the representative species by using the cross-
species scaling algorithm developed by Opresko et al. (1994). Since Roylance et al. (1985)
administered dietary doses of endrin to both male and female mallards the mean of the male
and female body weights for each representative species was used in the scaling algorithm to
obtain the lexicological benchmarks. Based on the avian data set for endrin, the benchmarks
developed from the Roylance et al., (1985) study were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LQEC values and then approximating the 10 percentile.
If there were 10 or fewer values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, studies were not identified for benchmark development for endrin.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Endrin - 7
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
R«pTB*efltath/*
8p«ch»,
deer mouse
short-tailed
shrew
meadow vote
Eastern
cottorttu
red fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
Benchmark.
Value*
moftQxtt*
0.56 (a)
0.58 (a)
0.47 (a)
0.20 (a)
0.15 (a)
0.14 (a)
0.07 (a)
0.028 (a)
0.049 (a)
0.044 (a)
0.054 (a)
0.045 (a)
10
ID
atopy .
Spade*
mouse
mouse
mouse
mouse
mouse
mouse
mouse
mallard
duck
malard
duck
mallard
duck
malard
duck
mallard
duck
-
Effect
(eto
teto
feto
feio
feto
feto
feto
rep
rep
rep
rep
rep
Study
yafcat
mtfig.
day
0.5
0.5
0.5
0.5
0.5
. 0.5
0.5
0.028
0.028
0.028
0.028
0.028
Description
NOAEL
NOAEL
NOAEL
NOAEL .
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
-
-
QriglMtSMnw
Kavtock et at..
. 1981
Kavtock M !..
1981
Kavtock »t al.,
1981
Kavtock at al..
1981
Kavtock at al..
1981
Kavtock at al..
1981
Kavtock at al..
1981
Roytance at al.,
1985
Roytance at al.,
1985
Roytance at al.,
. . 1985
Roylance at al.,
1985
Roylance at al..
1985
'Benchmark Category, a adequaat. p ^(gyisional, i interim: a
magnitude or more above the NEL or LEL foe other adverse affects
10 - insufficient dad
"' indicates that the benchmark value was an order of
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
August 1995
-------
APPENDIX B Endrin - g
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K,,w values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration)., For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between'dissolved water concentrations and concentrations
in fish. The following brief discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values.
As stated in section 5.3.2, the BAFls for consituents of concern were generally
estimated using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for
the littoral ecosystem; these models were considered appropriate to estimate BAFls for endrin.
The bioconcentration factor for fish was also estimated from the Thomann models (i.e., log
Kow - dissolved BCF1) and multiplied by the dissolved fraction (fd) as defined in Equation 6-
21 to determine the total bioconcentration factor (BCF,'). The dissolved bioconcentration
factor (BCF|d ) was converted to the BCF,' in order to estimate the acceptable lipid tissue
concentration (TCI) in fish consumed by piscivorous fish (see Equation 5-115). The BCF,'
was required in Equation 5-115 because the surface water benchmark (i.e., FCV or SCV)
represents a total water concentration (C1). Mathematically, conversion from BCF,d to BCF,'
was accomplished using the relationship delineated in the Interim Report on Data and
Methods for Assessment of23,7,8'Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S.
EPA, 1993i):
BCF,d x fd = BCF,'
Converting the predicted BCF,d of 156,675 Ukg LP to the BCF,' of 105,881 L/kg LP was in
reasonable agreement (i.e., within a factor of about 2) with the geometric mean of five
measured BCF/ values presented uv«fee master table on endrin (geometric mean = 49,600).
The bioaccumulation factor for terrestrial vertebrates, and the bioconcentration factors for
earthworms and invertebrates were estimated as described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobic organic chemicals assuming that the
partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs and
BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data (in this case endrin) is compared to the BBF for TCDD and
that ratio (i.e., endrin BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial
vertebrates, invertebrates, and earthworms, respectively. For hydrophobic organic
-------
APPENDIX B Endrin - 9
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses. The BCF is based on route-co-leaf
translocation, direct deposition on leaves and grasses, and uptake inta the plant through air
diffusion. For metals, empirical data were used to derive the BCF for aboveground forage
grasses and leafy vegetables.
August 1995
-------
APPENDIX B
Endrin 10
Table 4. Biological Uptake Properties
cQlOQiCBi
receptor
limnetic trophic
lovel 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrate*
earthworms
plants
BCF, SAP, o*
BSAF
BAF
BAF
8CF
BAF
BAF
BAF
BAF
BCF
8CF
BCF
Ifptf-bM** or
who4ebooy
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
valu*
318,304 (d)
290.126 (d)
105,881 (t)
298,094 (d)
322.806 (d)
643,213 (d)
0.0019
0.0018
0.01S
0.038
ourc*
4- '
predicted value based on
Thomann.. 1989, food chain
modal
predicted value based on
Thomann. 1989, food chain
model
predicted value based on
Thomann, 1989 and adjusted to
estimate total BCF
predicted valu* based on
Thomann et a!.. 1992. food web
model
predicted value based on
Thomann et at., 1992. food web
model
predicted value based on
Thomann et at., 1992. food web
model
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCOO
estimated based on beef
biotranster ratio with 2.3,7,8-
TCDO
, estimated based on beef
biotransfer ratio with 2,3.7,8-
TCDD
USEPA. 1992C
d - refers to dissolved surface water.concentrauon
t » refers to total surface water concentration
-------
APPENDIX B Endrin-H
References
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Technical Support Document for the Proposed Land Application Rule. Prepared for U.S.
Environmental Protection Agency. Contract No. 68-DO-0020.
Anderson, R.L. and D.L. Defoe. 1980. Toxicity and bioaccurnulation of endrin and
methoxychlor in aquatic invertebrates and fish. Environ. Pollut. 22A(2):111-121.
(Author Communication Used). As cited in AQUIRE (AQC/atic Toxicity Ynformation
/?Etrieval Database), Environmental Research Laboratory, Office of Research and
Development, U.S. Environmental Protection Agency, Duluth, MN.
AQUIRE ( AQt/atic Toxicity Mormation fl£trieval Database). 1994. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN. .
Blus, L. J. 1978. Short-tailed shrews: toxicity and residue relationships of DDT, dieldrin,
and endrin. Archives of Environmental Contamination and Toxicology. 7:83-98.
Chemoff, N., R. J. Kavlock, R. C. Hanisch, D. A. Whitehouse, J. A. Gray, L E. Gray, Jr. and
G. W. Sovocool. 1979. Perinatal toxicity of endrin in rodents. I. Fetotoxic effects of
prenatal exposure in hamsters. Toxicology. 13:155-165.
Deichmann, W. B;, W. E. MacDonald, E. Blum, M. Bevilacqua, J. Radomski, M. Keplinger,
M. Balkus. 1970. Tumorigenicity of aldrin, dieldrin,and endrin in the albino rat.
Toxicology. 39:37-45. .
Eisenlord, G., G. S. Loquvam, S. Leung. 1968. Results of Reproduction Study of Rats Fed
Diets Containig Endrin Over Three Generations. Shell Chemical Company and Velsicol
Chemical Corporation. The Hine Laboratories, Inc., San Francisco, CA.
57 FR 24152. June 5, 1992. U.S. Emnvironmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogenic Risk Assessment Based on
Equivalence of mg/kg^/day.
'*»
Goldenthal, E. I. 1978. Teratology Study in Hamsters. International Research and
Development Corporation: prepared for the Vesicol Chemical Corporation.
Good, E.E. and G.W. Ware. 1969. Effects of insecticides on reproduction in the laboratory
mouse: endrin and dieldrin. Toxicology and Applied Pharmacology. 14:201-203.
August 1995
-------
APPENDIX B Endrin - 12
Gray, L. E., R. J. Kavlock, N. Chemoff, J. A. Gray, J. McLamb. 1981. Perinatal toxicity of
endrin in rodents. III. Alterations of behavioral ontogeny. Toxicology. 21:187-202.
Hermanutz, R. 1978. Endrin and malathiontoxicity to flagfish (Jordanella floridae). Arch
Environ. Contam. Toxicol. 7: 159-168. As cited in Rand, G.M. and S'.R. Petrocelli, 1985,
Fundamentals of Aquatic Toxicology, Hemisphere Publishing Corporation, New York.
Howard, P. 1991. Handbook of Environmental Fate and Exposure Data for Organic
Chemicals: Volume III Pesticides. Lewis Publishers.
Jarvinen, A.W., and R.M. Tyo. 1978. Toxicity to fathead minnows of endrin in food and
water. Arch. Environ. Contam. Toxicol. 7(4): 409-421. As cited in AQUIRE (AQUztic
Toxicity /nformation /?£trieval Database), Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
Jarvinen, A.W., and R.M. Tyo. 1978. Toxicity to fathead minnows of endrin in food and
water. Arch. Environ. Contam. Toxicol. 7(4): 409-421. As cited in Rand, G.M. and S.R.
Petrocelli, 1985, Fundamentals of Aquatic Toxicology, Hemisphere Publishing
Corporation, New York.
Kavlock, R. J., N. Chemoff, E. H. Rogers. 1 °85. The effect of acute maternal toxicity on
fetal development in the mouse. Teratogenesis, Carcinogenesis, and Mutagenesis.
5:3-13. .
Kavlock, R. J., N. Chemoff, R. C. Hanisch, J. Gray, E. Rogers and L. E^ Gray, Jr. 1981.
Perinatal toxicity of endrin in rodents. II. Fetotbxic effects of prenatal exposure in rats and
mice. Toxicology. 21:141-150.
Kettering Laboratory. 1971. The Reproductive Capacity Amoung Dogs Fed Diets Containing
Endrin. Department of Environmental Health, College of Medicine, University of
Cincinnati, Cincinnati, Ohio.
NTIS. 1993. Sediment Quality Criteria for the Protection of Benthic Organisms: Endrin.
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Technology, Washington, D.C.
Nagy, K. A., 1987. Field metabolic rate and food requirement scaling in mammals and birds.
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Opresko, D. M., B. E. Sample, and G. W. Suter. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
-------
APPENDIX B Endrin.13
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August 1995
-------
APPENDIX B Endrin - 14
Thurston, R.V., T.A. Gilfoil, E.L. Meyn, R.K. Zajdel, T.L. Aoki, andG.D. Veith. 1985.
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s.
U.S. EPA (Environmental Protection Agency). 1993g. Wildlife Exposure Factors Handbook:
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Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for the U. S. Department of Energy.
-------
Terrestrial"i .city - Endrin
Cas No! 72-20-8
Chemical
endun
endrin
endrin
Ullllllll
.
hamsters
mice
CFW Swiss
mice
CD mice
CD mice
Type of
Effect
teratogenic
J
teratogenic
rep
letotoxic
lelotoxic
Description
PEL
PEL
PEL
NOAEL
LOAEL
Value
5
2.5
5
0.5
1
Units
mo/kg-day
mg/kg-d?
ppm
mg/kg-day
mg/kg-day
ExpotJre Route
(oral, B.C., l.v.. l.p..
Inlectlon)
Pesticides were
dissolved in 1 5 ml/kg
corn oil immediately
before administration
by oral Intubation.
Pesticides were
dissolved in 1 .5 ml/kg
com oil immediately
before administration
by oral intubation.
oral
dissolved in com oil
and administered via
gastric Intubation (0 1
ml/day)
dissolved in corn oil
and administered via
gastric intubation (0. 1
nil/day)
Exposure
Duration/
Timing
Given a
single dose
on day 7. 8 .
or 9 of
gestation.
Given a
single dose
on day 9 of
gestation.
120 days;
beginning 30
days before
mating
treated on
gestation
days 7- 17
treated on
gestation
days? 17
Reference
Ottolenghi, 1974
Cntolenghi, 1974
Good and Ware.
1969
Kavlock el al..
1981
Kavlock el al..
1981
Comments
Malformations weie highest after
treatment on day 8 , Embryocidal
action and lelotoxicily were more
pronounced when treatment
occurred on day 7 or 8 than 9 of
gestation.
In mice, endrin was teratogenic, but
the frequency and gravity of the
defects produced were less
pronounced than in the hamsters.
There was a direct influence of
endrin on reproduction in mice, due
to fetal mortality.
Petal weight and skeletal and
visceral maturity were adversely
affected at this dose. Teratogenic
effects and embryo lethality not
evident even al levels of maternal
lethality.
-------
Terrestrial Toxicity - Endrin
Cas No. 72-20-8
Name
endrin
endrin
endrin
endrin
endrin
endrin
Species
rat
rat
hamster
hamster
hamster
CO rat .
TvDA Of
Effect
letotoxic
rep
letotoxic
letotoxic
behavioral
behavioral
Description
NOAEL
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
Value
045
2
0.75
1.5
'1.5
0.15
Unite
mg/kg-day
b»
ppm
mg/kg-day .
mg/kg-day
mg/kg-day
mg/kg-day
Exposure Route
(oral, B.C., l.w., l.p.,
Inlectlon)
dissolved in com oil
and administered via
gastric intubation (0.2
ml/day)
oral (through the diet)
adminstered by oral
gavage In com oil
administered by oral
gavage In com oil
administered by
gastric Intubation
administered via
gastric Intubation
Exposure
Duration/
Timing
treated on
gestation
days 7- 20
weanling rats
were
maintained
on the diet
for 79 days
treated on
days 5- Hot
gestation
treated on
days 5-14 ol
gestation
days 5-14 ol
gestation
days 7- 15 of
lactation
Reference
Kavlock et al..
1981
Eisenlord et al..
1968
Chernoff et al..
1979
Chernottetal..
1979
Gray et al., 1981
Grayetal., 1981
Comments
No dose related effects on fetal
mortality, weight, degree of skeletal
and visceral maturation and
incidences ol skeletal and visceral
anomalies.
There was no difference in
behavior, weight, the number of
litters, or the percent survival- of
pups at this dose.
Maternal toxicity and fetal toxicity
were noted at doses above 0 75
mg/kgday
Significant maternal lethality and
fetal toxicity were noted at this
dose.
This dosage produced a persistent
elevation in the locomotor activitv
This dosage produced an elevation
in locomotor activity in rat pups that
was attenuated by 3 months of aqe
-------
Terrestrial . .city - Eneirin
Cas No. 72-20-8
Chemical
endrin
endrin
endiin
endrin
endrin
uniiiin
hamster
mouse
dog
mallard
ducks
mallard
ducks
inallaid
ducks
_ .
lype 01
leratogenic
letotoxic
.1
rep
rep
rep
tup
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
NOAEL
Value
25
7
2
3
0.5
3
Units
mg/kg-day
mg/kg-day
ppm
ppm
ppm
ppm
Exposure Rout*
(oral, s.c., l.v., l.p.,
Injection)
administered via
gastric intubation
administered as
solution in com oil
oral; once per day
oral
oral
01 at
Exposure
Duration/
Timing
days 4-1 3 ol
gestation
day 8 ol
gestation
1 5 months
20 weeks
20 weeks
>2(X) days
* Reference
Goldenlhal el al..
1978
Kavlock el al..
1985
The Kettering
Laboratory, 1971
Roylance et al..
1985
Roylance et al..
-1985
Spaim el al.. 1966
Comments
Although scoliosis with or without
fused ribs was observed in
hamsters in this study, it is
attributed to heredity^in laboratory
animals. In elfect. endrin given .
orally to hamsters in this study at
doses up to 2.5 mg/kg-day is not
considered a teratogen.
Fetal weight was reduced at this
level.
Based on morphological
observations and numbers ol pups.
no effects could be attributed to
endrin in the daily die! al 2 0 ppm
(0.051 mg/kg-day), or less.
A 9.6% drop in embryo survival was
observed at this dose. However, no
effects on egg production, fertility.
and hatcnability were reported.
No effect on egg production, fertility
and hatchability were reported.
. No significant effects on
reproduction were observed at this
dose level. A later hatching date
and poorer hatching success were
reported at this dose level.
-------
Terrestrial Toxicity - Endrin
Cas No. 72-20-8
Chemical
Name
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
endrin
oinJun
Species
fertile hen
eggs
ral
mouse
monkey
rabbit
guinea pig
hamsler
pigeon
quail
duck
wild bird
mallard
sharp tailed
grouse
California
quail
Tvoe of
Effect
dvp
acute
acute . '
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
Description
NOAEL
LD50
L050
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LO50
LD50
LD50
LD50
Value
0.2
3
1370
3
7
16
10
5600
4210
5330
1780
5.64
1.06
1.19
Unite
mo/egg
mo/kg
uflfcg
mg/kg
mfl/kg
mo/kg
mg/kg
Ug/kfl
u&*g
ugfcg
ug/kg
mg/kg
mg/kg
mg/kg
Exposure Route
(orsl, s.c., l.v., l.p..
Injection)
injection via com
carrier
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
NS
NS
NS
Exposure
Duration/
Timing
injected.
either prior to
incubation or
alter a 7-day
incubation
period
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Smith et al , 1970
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
RTECS, 1994
RTECS, 1994
US EPA, 1993h
US EPA, 1993h
US EPA. 1993H
Comments
Forty percent hatchabihty occurred
in comparison with 86.2% lor the
controls.
'
-------
Terrestrial . city - Endrin
Cas No. 72-20-8
Chemical
Name
endrin
endrin
endrin
endrin
Species
'pheasant
rock dove
mule deei
domestic
goat
Type ol
Effect
acute
acute
- acute
acute
Description
LD50
LD50
LD50
LD50
Value
1 78
2.0-5.0
&25-12.5
25.0-50.0
Units
mg/kg
mg/kg
mg/kg
mg/kg
' Exposure Route
(oral. s.c.. l.v., l.p.,
Inlectlon)
NS
NS
NS
NS
Exposure
Duration/
Timing
NS
NS
NS
NS
Reference
US EPA. 1993h
U.S. EPA. 1993H
U.S. EPA, 1993h
U.S EPA, 1993h
Comments
NS = Not specified
-------
Freshwater Toxicity - Endrin
Cas No. 72-20-8
Chemical
Nam*
endrin
endnrt
endnn
endrin
endrin
Species
lathead
minnow
llaglish
Ceriodaphnia
. reticulala
Daphnia
magna
Daphnia
magna
Type of
Effect
chron
. Chron
immob.
immob.
morl
Description
MATC
MATC
EC50
EC50
LC50
Value
<0 14
022-30
24
59
88 230
(141.25)
Units .
uo/L
ug/L
UQ/L
ug/L
UQ/L
Test Type
(static/ flow
through)
complete lite
cycle lest
complete lite
cyde test
NS
NS*
NS
Exposure
Duration/
Timing
NS ,
NS
48 hour
48 hour
48 hour
Reference
Jarvinen and Tyo,
1978 as cited in
Rand and PetroceUi.
1985
Hermanulz, 1978 as
cited in Rand and
PetroceUi, 1985
AQUIRE, 1994
AQUIRE. 1994
AQUIRE, 1994
Comments
Critical life stage end
points: embryo, larval.
and early juvenile;
mortality
Critical lite stage end
points: embryo, larval.
and early juvenile;
growth
-------
Freshwater icily - Erjdrirs
Cas No. 72-20-8
Chemical
Name
endrin
endrin
endrin
endrin
endfin
endrin
Species
Simocephalus
serrulatus
cattish
channel cattish
brook trout
bluegill
largemouth
bass
Type of
Effect
immob. .
mort
mort
mort
mort
mort
Description
EC50
LC50
LC50
LC50
LC50
LC50
1
Value
26 - 45 (33 9)
2800
041 -0.8
(0.52)
0.36 0.59
(0.46)
0.19-061
(0.39J
027
Units
ug/L
ug/L
ug/L
ug/L
ua/L
ug/L
Test Type
(static/ flow
through)
NS
NS
NS
NS
NS
NS
Exposure
Duration/
Timing
48 hour
96 hour
96 hour
96 hour
96 hour
48 hour
Reference
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AOUIRE. 1994
AQUIRE, 1994
AQUIRE. 1994
Comments
-------
Freshwater Toxicity - Endrin
Cas No. 72-20-8
Chemical
Name
endrin
endrin
Species
rainbow trout
fathead
minnow
Type of
Effect
mort
rfiort
Description
LC50
LC50
Value
0.33 - 2.5
(O.B2)
0.26 - 3.8
(0.84)
Units
ug/L
ug/L
Test Type
(static/ flow
through)
NS
NS
Exposure
Duration/
Timing
96 hour
96 hour
Reference
AQUIRE, 1994
AQUIRE, 1994
Comments
NS = Not specified.
-------
Freshwater Biological . «ake Measures - Endrin
Cas No. 72-20-8
Chemical
name
endiin
endrin
endrin
endrin
endrin
endrin
endrin
Species
fish
fish
black
bullhead
black
bullhead
lathead
minnow
channel
catfish
fathead
minnow
B-taclor
(BCF, BAF.
BMF)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
168.6
1130
3700
6200
13000
1640 - 2000
300
Measured or
predicted
(m,p)
P
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
. Reference
U.S. EPA. 1993h
Jarvinen and Tyo, 1978
as cited in U.S. EPA,
1993h
Anderson and Oefoe,
1 980 as cited in
AQUIRE, 1994
Anderson and Defoe,
1980 as cited in
AQUIHE, 1994
Jarvinen and Tyo, 1978
as cited in AQUIRE,
1994
U.S. EPA, 1980 as cited
in Roward, 1991
Jarvinen and Tyo, 1978
as cited in AQUIRE.
1994
Comments
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
NS = Not specified.
-------
Terrestrial Biological Uptake Measure - Endrin
Cas No. 72-20-8
Chemical
name
endrin
Species
plants
= Not specified.
B-factor
(BCF, BAF,
BMP)
BCF
Value
0058
Measured or
predicted
(m.p)
P
Units
NS
Reference
U.S EPA.
19906
Comments
-------
APPENDIX B Fluoranthene - 1
Toxicological Profile for Selected Ecological Receptors
. Fluoranthene
Cas No.: 206-44-0
Summary: This profile on fluoranthene summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e.. Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) arid, if available.
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwaler ecosystem. Table 1 coniains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and litloral ecosystems, including
aqualic planis, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Adequate toxicity data measuring reproductive or developmental endpoints pertinent to
population sustainability were not identified. Thus, no benchmark values protective of the mammalian
community in a freshwater ecosystem were derived.
Birds: 'No toxicity studies documenting terrestrial avain exposure to fluoranthene were identified.
August 1995
-------
APPENDIX B Fluoranthene - 2
Fish and aquatic invertebrates: A Final Chronic Value (FCV) of 0.0062 mg/L as reported in the SQC
document for fluoranthene was selected as the benchmark value protective of fish and aquatic
invertebrates. Because the benchmark is based on FCV from SQC, it was categorized as adequate.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no observed effects
concnetration (NOEC) or a lowest observed concnetration (LOEC) for vascular aquatic plants (e.g.,
duckweed) or (2) an effective concentration (EC,,) for a species of freshwater algae, frequently a
species of green algae (e.g., Selenastrum capricornutum). The aquatic plant benchmark for
fluoranthene is 5.44 E+07 mg/L (Suter and Mabrey, 1994). As described in Section 4.3.6, all
benchmarks for aquatic plants were designated as interim.
Benthic community: Benchmarks for the protection of benthic organisms were determined using the
Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value (FCV) or Secondary
Chronic Value (SCV), along with the fraction of organic carbon and the octanol-carbon partition
coefficient (K,,,.) to determine protective sediment concentration (Stephan, 1993). The EQpl number is
the chemical concentration that may be present in the sediment while still protecting the benthic
community from the harmful effects of chemical exposure. The FCV reported in the SQC document
for fluroanthene was used to calculate an EQp value of 720 mg fluoranthene/kg organic carbon.
Assuming a mass fraction of organic carbon for the sediment (f,,,.) of 0.05, the benchmark for the
benthic community is 36 mg/kg. Since the EQp value was based on a FCV from the SQC, the
sediment benchmark is categorized as adequate.
August 1995
-------
APPENDIX B
Fluoranthene - 3
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative
Specie*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark Value*
mg/Kg-day
ID
, ID
ID
ID
ID
ID
ID
ID
ID
ID
Study
Spedea
-
-
-
-
-
. . .
-
-
-
Effect
-
-
-
-
-
-
:
Study Value
mg/kg-day
-
-
'
-
-
-
Deacrfptfofi
-
-
-
-
-
-
-
-
SF
-
-
-
-
-
-
-
Original Source
-
'Benchmark Category, a = adequate, p = provisional, i = iniehm; a '*' indicates that the benchmark value was an order of magnitude or
more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and
aquatic
invertebrates
aquatic plants
benthic
community
DeMCIVMnl
mgft
0.0062 (a)
5.44 E +07
(i)
24.8 mg/kg
sediment (a)
Study Sp«cta*
aquatic
organisms
Pimephalas
promelas and
Daphnia magna
benthic
community
o~*«»
FCV
CV .
FCV x K,* '
SMVO*
U.S.EPA,
1993k
Suter and
Mabrey, 1994
SQC
Benchmark Category, a = adequate, p = provisional, i = interim; a '' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Fluoranthene - 4
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As discussed previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to fluoranthene.
Thus, no benchmark values protective of the mammalian community were derived. -
Birds: No toxicity studies documenting avain exposure to fluoranthene were identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10*
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were .not identified for fluoranthene and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Fluoranthene - 5
Table 3. Toricological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RcprVMvitstfvv Sptcwt
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
American robin
American woodcock
plants
soil community
B«nctunark VthW
mgrt(p"d«y
ID
ID
ID
10
ID
ID
ID
ID
ID
ID
ID
No data
No data
Study Spwiv
-
-
-
' -
-
-
-
-
en«ct
-
-
'
-
Study VMM
ma/kpday
-
-
-
-
-
-
-
OMcHptton
.-
-
-
#
-
-
-
-
UP
- .
-
-
-
-
-
-
-
OrigbMl Soura
-
.
'Benchmark Category, a = adequate, p = provisional, i = interim: a '" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEI for other adverse effects.
ID = Insufficient Data
III. Biological Uptake Measures
This section presents the biological uptake measures (i.e., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcnetrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: tropic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertbrates, terrestrial vertebrates, and plants. Each
value is idenfieid as whole-body or llpid-based and, for the generic aquatic ecosystems, the
biological uptake factors are deignated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K^ values below 4, bioconcentration factors (BCFs) in fish were always
August 1995
-------
APPENDIX B Fluoranthene - 6
assumed to refer dissolved water concentrations (i.e., dissolved water concentration equals
total, water concentration). For organic chemicals with log K^, values above 4, the BCFs
were assumed to refer to total water concentrations and concentrations in fish. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values presented in Table 4.
As stated in section 5.3.2, the BAP/s for constituents of concern were generally estimated
, using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. However, these models were considered inappropriate to estimate BAF/s for
fluoranthene because they fail to consider metabolism in fish. A number of studies have
demonstrated that polycyclic aromatic hydrocarbons (PAHs) such as fluoranthene are readily
metabolized in the tissue of fish (see Polycyclic Aromatic Hydrocarbon Hazards to Fish,
Wildlife, and Invertebrates: A Synoptic Review. U. S. Fish and Wildlife Service Biol. Rep.
85[1.11]. The BAF/s selected for fish in the limnetic and littoral ecosystems for fluoranthene
are from Stephan (1993). This document contains unpublished field data by Burkard with a
geometric mean BAP of 96 qbtained from fish at 5% lipids. Steady-state measured data on
biological uptake of fluoranthene (and most PAHs) are very limited at this time and should be
interpreted with caution. Since no measured fish BCf values were identified, the fish BAF
reported by Stephan (1993) was used for bioconcentration factor for fish.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, earthworms and
terrestrial invertebrates were estimated asu described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobia organic chemicals assuming that the
partitioning to tissue is dominated by lipids. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Fluoranthene - 7
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF.
-
BAF
BCF
BCF
BCF
llplcMMMdor
whoto body
lipid
lipid
lipid
lipid
lipid
-
whole-body
whole-body
whole-body
whole-plant
vatwi
1 ,900 (t) .
1 ,900 (t)
1 ,900 (t)
1 ,900 (t)
1,900(t)
10
1.8E-03
1.7E-03
1.3E-02
4.1 E -02
ourca
measured; Stephan. 1993
measured; Stephan, 1993
measured; Stephan. 1993
measured; Stephan, 1993
measured; Stephan. 1993
calc
calc
calc
U.S. EPA, I990e
August 1995
-------
APPENDIX B Fluoranthene - 8
References
Brooke, L., 1991. Memorandum to Walter Berry. Summary' of Results of Acute and Chronic
Exposures of Fluoranthene Without and With Ultraviolet (UV) Light to Various
Freshwater Organisms. December 3. 5pp. As cited in U.S. Environmental Protection
Agency, 1993k. Sediment Quality Criteria for the Protection of Benthic Organisms:
Fluoranthene. Office of Water, Office of Research and Development, Office of Science
and Technology, Health and Ecological Criteria Division, Washington, D. C, EPA-822-R-
93-012.
Buccafusco, R. J., S. J. Elis and G.A. LeBlanc. 1981. Acute Toxicity of Priority Pollutants to
Bluegill (Lepomis macrochirus). Bull. Environ. Contam. Toxicoi, 26:446-452. As cited in
U.S. Environmental Protection Agency, 1993k. Sediment Quality'Criteria for the
Protection of Benthic Organisms: Fluoranthene. Office of Water, Office of Research and
Development, Office of Science and Technology, Health and Ecological Criteria Division,.
Washington, D. C., EPA-822-R-93-012.
Carlson R. M. et al., 1979. Implications to the aquatic environment of polynuclear aromatic
hydrocarbons liberated from Northern Great Plains Coal, USEPA-600/3-79-093. As cited
in Hazardous Substance Database (HSDB), National Library of Medicine, 1994.
Clements, W.H., J.T. Oris, and T.E. Wissing, 1994. Accumulation and food chain transfer of
fluoranthene and benzo(a)pyrene in Chironomus riparius and Lepomis macrochirus. Arch.
Environ. Contam. Toxicoi. 26:261-266.
Gendusa, A. C., 1990. Toxicity of Chromium and Fluoranthene from Aqueous and Sediment
Sources to Selected Freshwater Fish. Ph.D. Thesis, University of North Texas. U.M.I.
300 N. Zeeb Rd., Ann Arbor, MI 48106. 138pp, As cited in U.S. Environmental
Protection Agency, 1993k. Sediment Quality Criteria for the Protection of Benthic
Organisms: Fluoranthene. Office of Water, Office of Research and Development, Office
of Science and Technology, Health and Ecological Criteria Division, Washington, D. C.,
EPA-822-R-93-012.
Gerhart, E. H. and R. M. Carslon, 1978. Hepatic mixed-function oxidase activity in
rainbow trout exposed to several polycyclic aromatic compounds. Environmental
Research, 17:284-295.
August 1995
-------
APPENDIX B Fluoranthene - 9
Home. J. D. and B.R. Oblad, 1983. Aquatic Toxicity Studies of Six Priority Pollutants. Final
Report Task II. U.S. EPA. Contract No. 68-01-6201. As cited in U.S. Environmental
Protection Agency, 1993k. Sediment Quality Criteria for the Protection of Benthic
Organisms: Fluoranthene. Office of Water, Office of Research and Development, Office
of-Science and Technology, Health and Ecological Criteria Division, Washington, D. C.
EPA-822-R-93-012.
LeBlanc, G. A. 1980. Acute Toxicity of Priority Pollutants to Water Flea (Daphnia
magna). Bull. Environm. Contam. Toxicoi., 24:684-691.
Newsted, J. L. and J. P. Giesy. 1987. Predictive models for photoinduced acute toxicity of
polycyclic aromatic hydrocarbons to Daphnia Magna, Strauss (Cladocera, Crustacea).
Environmental Toxicology and Chemistry, Vol. 6, pp. 445-461.
Oris, J.T., R.W. Winner, and M.V. Moore, 1991. A Four-Day Survival and Reproduction
Toxicity Test for Ceriodaphnia dubia. Environ. Toxicoi. Chem., 10:217-224. As cited in
U.S. Environmental Protection Agency, 1993k. Sediment Quality Criteria for the
Protection of Benthic Organisms: Fluoranthene. Office of Water, Office of Research and
Development, Office of Science and Technology, Health and Ecological Criteria Division,
Washington, D. C., EPA-822-R-93-012.
RTECS (Registry of Toxic Effects of Chemical Substance), March 1994. National Institute
for Occupational Safety and Health.
Spehar, RL et al., 1980. J Water Pollut Control Fed 52:1703-74. As cited in Hazardous
Substance Database (HSDB), National Library of Medicine, 1994.
Stephan, C.E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suedel, B. C.., J. H. Rodgers, Jr. and P. A. Clifford, 1993. Bioavailability of fluoranthene
in freshwatersSediment toxicity tests. Environmental Toxicology and Chemistry, Vol. 12,
pp. 155-165.
August 1995
-------
APPENDIX B Fluoranthene - 10
Suter II, G.W., M.A. Futrell, and G.A. Kerchner, 1992. Toxicological Benchmarks
forScrening of Potential Contaminants of Concern for Effects on Aquatic Biota on theOak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, D. C.
Suter, G.W. II and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management.
Thomann, R. V. 1989. Bioaccumulation model of organic chmeical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6): 699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulationin aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615 - 629.
U.S. Environmental Protection Agency. 1978. In-depth studies on health and environmental
impacts of selected water pollutants. U.S. EPA. Contract No. 68-01-4646. As cited in
U.S. Environmental Protection Agency. 1980. Ambient Water Quality Criteria for
Fluoranthene. Criteria and Standards Division, Washington, D. C... October 1980, 86p.
U.S. Environmental Protection Agency. 1988. 13-week Mouse Oral Subchronic Toxicity
Study. Prepared by Toxicity Research Laboratories, Ltd., Muskegon, Michigan for the
Office of Solid Waste, Washington, D, C. As cited in IRIS Database, 1994.
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment. Washington, D.C. January.
U.S. Environmental Protection Agency. 1993k. Sediment Quality Criteria for the Protection
of Benthic Organisms: Fluoranthene. Office of Water, Office of Research and
Development, Office of Science and Technology, Health and Ecological Criteria Division.
Washington, D.C., EPA-822-R-93-012.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxic. , - Fluoranthene
Cas No.: 206-44-0
Chemical
Name
lluoranthene
tluoranthene
tluoranthene
Species
mouse
iat
mouse
-
Type of Effect
acute
acute
sub-chronic
Description
LD50
LD50
NOAEL
Value
100
2,000
125
Units
mg/kg
mg/kg-
body wt.
mg/kg-
body wt.
Exposure Route
(oral, s.c., i.v.. i.p..
injection)
i.v.
oral
oral
Exposure Duration
/ Timing
NS
NS
1 3-weeks
Reference
RTECS, 1994
U.S.EPA, 1968 as cited in
IRIS, 1994
Comments
Doses of the study were 0. 125. 250 and 500
mg/kg-d. Critical effects observed at 250 and
500 mg/kg-d dose were nephropathy,
increased liver weights, hematological
alterations and clinical effects.
NS = Not Specified
I luoi.iiithe.ne Page 6
-------
Freshwater Toxlclty - Fluoranthene
Cos No!: 206-44-0
Chemical Name
lluoranlhene
fluoranthene
Iluoranthene
ftuoranthene
fluoranthene
S
fluoranthene
fluoranthene
fluoranthene
Species
aquatic
organisms
aquatic
organisms
channel
cattish
rainbow
trout
rainbow
trout
blueglll
bluegill
fathead
minnow
Type of
Effect
chion
chron
acute
acute
acute
acute
acute
acute
Description
NAWQC
FCV
LC50/EC50
LC50/EC50
LC50/EC50
LC50
LC50/EC50
LC50/EC50
Value
1.7
6.16
36
7.7
187
3.980
4.000
7.71
Uunlts
ug/l
ug/l
ug/l
UQ/I
ug/l
ug/l
ug/l
ug/l
Test Type
(static/ flow
through)
NS
NS
static
flow-
through
static
static
static
static
Exposure
Duration /
Timing
NS
NS
NS
NS
NS
96-hour
NS
NS
Reference '.
Suter et al.. 1992
U.S. EPA. 1993k
Gendusa. 1990 as
cited In U.S EPA.
1993k
Brooke. 1991 as cited
in U.S. EPA. 1993k
Home and Oblad.
1983 as cited in
U.S.EPA. 1993k
U.S.EPA. 1980
Buccafusco et al..
198 las cited In U.S.
EPA. 1993k
Gendusa. 1990 as
ctted in U.S. EPA.
1993k
Comments
freshwater (daik) FCV is basud
on FAV=33.6 ug/l. and the tina
ACK=5.45
I luorr
-------
Freshwater Tox. / - Fluoranthene
Cos No.: 206-44-0
Chemical Name
fluofcinlhene
(luoranthene
lluoianlhene
fluoranthene
fluoranthene
fluoranthene
fluoranlhene
lluoianthene
Species
faiheud
1 Ilil II IOW
fathead
minnow
fathead
minnow
fathead
minnow
daphnla
magna
daphnla
magna
daphnla
magna
daphnia
magna .
Type ol
Ettect
acute
acute
chronic
chronic
acute
acute
acute
acute
Description
LC60/EC60
LC507EC50
/
cv
cv
'LC50/EC50
LC50/EC50
LC50/EC50
LCSO
Value
12.22
95
15.02
2.59
45
102.8
0.97
325.000
Uunlts-
ug/r
ug/l
ug/i
ug/l
ug/l
ug/l
ug/l
ug/l
Test Type
(slctic/ flow
through)
(low-
through
static
NS
NS
static
static
flow-
through
static
Exposure
Duration /
Timing
NS
NS
NS
NS
NS
NS
NS
48 hour
Reference
Brooke. 1991 as cited
inUS.EPA. 1993k
Home and Oblad.
1983 as cited in
U.S.EPA. 1993k
U.S.EPA. 1993k
U.S.EPA. 1993k
Orisetal.. 1991 as
cited in U.S. EPA.
1993k
Brooke; 1991 as cited
In U.S. EPA. 1993k
Brooke. 1991 as cited
In U.S. EPA. 1993k
U.S.EPA. 1978 as
cited in U.S.EPA. 1980
Comments
^
test performed in dark
environment
test performed in lighted (UV)
environment
luornf ithune - Page 8
-------
Freshwater Toxlclty - Fluoranthene
CasNo.: 206-44-0
Chemical Name
fluoranthene
(luoranthene
fluoranthene
fluoranthene
Species
daphnia
maqna
daphnia
magna
daphnia
magna
daphnia
magna
Type of
Effect
acute
acute
chronic
chronic
Description
LC50
EC50
CV
CV
Value
320.000
102.6
30.37
0.92
Uunlts
uq/l
ug/l
ug/l
ug/l
Test Type
(static/ flow
through)
static
static
NS
NS
Exposure
Duration /
Timing
48-hour
10-day
NS
NS
Reference
. LeBlanc. 1980
Suedeletal.. 1993
U.S.EPA. 1993k
U.S.EPA, 1993k
Comments
test performed in dark
environment
test performed in lighted (UV)
environment
NS=Not Specified
I luuir ',-ne - Page 9
-------
Freshwater Biological Upi. > Measures - Ffuoranth.ene
Cos No.: 206-44-0
Chemical Name
lluoranlhene
lluoranthene
lluoranlhene
(luoranlhene
lluoranthene
lluoranthene
Species
rainbow trout
rainbow trout
fathead minnow
aquatic organisms
Oaphnia magna
Chironomus riparius
B-factor (BCF,
BAF, BMP)
BCF
BCF
BCF
BCF
BCF
BAF
Value
378
380
398
570
1742
31
Measured 01
Predicted
(m,p)
m
m
m
P
P
m
Units
LAg
NS
NS
NS
NS
(chironomtds
ug/kg) / (sediments
uo/kg)
Reference
Gerhart & Carlson. 1978
Spehar el al , 1980 as
cited in HSDB. 1994
Carslon et al., 1979 as
cited in HSDB. 1994
U.S.EPA, 1993b
Newsted & Geisy, 1987
Clements et a! . 1994
Comments
. Rainbow trout were exposed to both
lluoranthene and pyrene in one experiment
The How-through exposure was lor 21-
days. BCF value was measured in.lhe livei
tissue ot Ihe rainbow Iroul.
21 -day biocoricentralion test in a How
through tank.
28-day experiment in flow-through lank log
BCF 3.60 peak after 7 days: depuration
occurs in two days
BCF normalized to 1% lipid
Reported BAF is measured from the
sediments with the highest concentration ol
fluoranthene (4 040 uq/kq)
NS = Not Specified
IUOK inthune Cage 10
-------
Terrestrial Biological Uptake Measures - Fluoranthene
CasNo.: 206-44-0
Chemical Name
fluoranthene
Species
plant
B-factor
(BCF. BAF.
BMR
BCF
Value
0045
Measured
or
Predicted
(m.o)
P
Units
(ug/g DW
plant)/(ug/g
soil)
Reference
U.S. EPA. 1990E
Comments
Plant uptake from soil pertains to
foraqed plants
I IIK ire ' v;f>e - Page 1 I
-------
APPENDIX B Heptachlor. 1
Toxicological Profile for Selected Ecological Receptors
Heptachlor
Cas No.: 76-44-8
Summary: This profile on heptachlor summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, arid soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 coniains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in ihe limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several lexicological studies involving heptachlor exposure to mammals have
been conducted using laboratory mice and rais. Green et al. (1970) fed male and female
Sprague-Dawley rats diets containing 0.25 mg/kg-day heptachlor for 60 days (females
continued to receive test diet through gestation). A LOAEL of 0.25 mg/kg-day was
established based on decreased fertility in females. In another study, Akay and Alp (1981, as
cited in ATSDR, 1989) fed male and female mice 7.5, 15, or 30 mg/kg-day heptachlor. Over
a 10 week period an overall failure to reproduce at all dose levels was observed.
In a similar investigation, rats of both sex were exposed to concentrations of 1.5, 3.0, 5.0, 7.0
and 1.0.0 ppm heptachlor for 7 weeks (Witherup et al., 1955). The introduction of heptachlor
August 1995
-------
APPENDIX B Heptachlor-2
into the diet of rats in concentrations greater than 7.0 ppm increased the probability of
fatalities in offspring, reflecting some weakness in the reproductive function. Since no
information was provided on daily food consumption or body weight, conversion from mg/kg-
diet to mg/kg-day required the use of an allometric equation:
Food consumption = 0.056(W°-6611) where W is body weight in kg (Nagy, 1987).
Assuming a body weight of 0.346 kg, a NOAEL of 5.0 ppm was converted to 0.429mg/kg-
day and a LOAEL of 7.0 ppm was converted to 0.600 mg/kg-day.
The Witherup et al. (1955) study measures chronic reproductive effects that may impair the
fecundity of a wildlife population. Therefore, the study NOAEL of 0.429 mg/kg-day was
chosen for derivation of a benchmark value. The subchronic study conducted by Green et al.
(1970) was not selected to extrapolate a benchmark value due to lack of an adequate dose-
response regime. The data from the Akay and Alp (1981) study also was not used in the
development of benchmarks due to limited details and statistical analysis.
The NOAEL value from the Witherup et al., (1955) was then scaled for species representative
of a freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
Benchmark = NOAEL. x
KJ
where NOAELj is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Witherup et al. (1955) study documented reproductive effects from heptachlor expoure to both
male and female rats, representative body weights of both genders were used in the scaling
algorithm to obtain lexicological benchmarks.
Data were available on reproductive, developmental, growth and survival endpoints for
heptachlor exposure. In addition, the data set contained acute and chronic toxicity studies that
were conducted during sensitive life stages. Based on the data set for heptachlor, the
benchmarks developed from the Witherup et al. (1955) study were categorized as adequate.
Birds: Adequate toxicity studies documenting avain exposure to heptachlor were not
identified and therefore, no benchmarks were developed.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for heptachlor. Therefore, the Tier n method described in Section 4.3.5 was used to
calculate a Secondary Chronic Value (SCV) of 6.9E - 03 mg/L. Because the benchmark is
-------
APPENDIX B Heptachlor-3
based on a SCV and there were no lower toxicity values in the data set, it was categorized as
interim.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascualr aquatic plants (e.g., duckweed) or (2) an effective concentration (ECxx) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwri).
The aquatic plant benchmark for heptachlor is 26.7 mg/L (Suter and Mabrey, 1994). As
described in Section 4.3.6, all benchmarkds for aquatic plants were designated as interim.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQ method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
octanol-carbon partition coefficient (K^ to determine protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from the harmful effects. of chemical
exposure. Because no FCV was available, a SCV value of 0.584 mg heptachlor/kg organic
carbon was used to calculate an EC" value. Assuming a mass fraction of organic carbon for
the sediment (f^ of 0.05, the benchmark for the benthic community is 2.92.E-02 mg/kg
sediment. Since the EQp number was based on a SCV, the sediment benchmark was
categorized as interim.
August 1995
-------
APPENDIX B
HepUchlor - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
W^^ppB^^fWdw^^P-
mink
river otter
baJdeagto
osprey
great biua haron
mallard
lesser scaup
spotted sandpiper
herring gul
kingfisher
ftttwhmartt
Vahtt* m4ko>
VWV^mgm^
**
0.33 (a)
, 0.20(8)
ID
10
ID
ID
ID
ID
ID
ID
»«dy
SMeiM
rat
rat.
-
-
-
Cnwl
«p
rep
-
iMiittir VhfeM
»9»*+9
0.43
0.43
-
-
. -
-
-
, .
D««ct»*»
NOAEL
NOAEL
1
-
-
-
^ * ^
-
'
-
-
-
* rnhjih*n.>"i'«
, s^ < *
WHhvup at al., 1955
Wittwrup at al., 1955
-
-
- '
*B0nchma/fc Catagory, a - adequate, p « provisional, i - interim; a "' indicates that ins banchmark valua was an order of
magnitude or more above the NEL or LEL for otter adverse effects.
ID - Insufficient Data
August 1995
-------
APPENDIX B
Heptachlor - 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftepreeantrfv*
Species^
fish and aquatic
invertebrates
aquatic plants
benlhic community
Benchmark
«*.
6 9E-03 (i)
0.0267 (i)
2.9E-02 (i)
Study
aquatic
organisms
aquatic
plant*
aquatic
organisms
^
SCV
cv
SCVxK,,.
OUalwHwitw
Gil. 1992
Suter and Mabrey,
1994
GLI, 1992
Benchmark Category, a - adequate, p a provisional, i * interim; a "" indicate* that the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverte effects.
August 1995
-------
APPENDIX B HepUchlor - 6
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^) for the generic terrestrial ecosystem. Table 3 contains
benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial
ecosystem.
Mammals: Becasue of the lack of additional mammalian toxicity studies, the same surrogate-
species study (Witherup et al., 1955) was used to derive the heptachlor lexicological
benchmark for mammalian species representing the general terrestrial ecosystem. The study
value was scaled for species in the terrestrial ecosystem using the cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the (Witherup et al., 1955) documented
reproductive effects from heptachlor exposure to both male and female rats, the representative
body weights of both sexes were used in the scaling algorithm to obtain lexicological
benchmarks. Based on the data set for heptachlor, the benchmarks developed for the
terrestrial ecosystem were categorized as adequate.
Biros: As mentioned in the freshwater ecosystem discussion, adequate data with which to
derive a benchmark protective of the avian community were not identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for heptachlor and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Heptachlor 7
Table 3 lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Raprm«u*tfv»
Stoefile^:
deer mouee
short-tailed
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
white-tailed dear
red-tailed hawk
American kestrel
Northern
bobowhile
American robin
American
woodcock
planU
toil community
Vahi*
mgrKa-cte|r
0.89 (a)
0.91 (a)
0.77 (a)
0.31 (a)
0.23 (a)
0.21 (a)
0.11 (a)
ID
>P
ID
ID
ID
No data
No data
Study
SpedM
rat
rat
rat
rat
rat
rat
rat
-
StfM*
rep
rep
rep
rep
rep
rep
rep
-
.-
Vtfu*
«**a~
**
0.43
0.43
0,43
0.43
0.43
0.43
0.43
-
-
-
~
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
'
&
0»^iB^O«»»
WBhenjp et aJ..
1955
WHherup et al.,
1955
Wrtherup et al..
1955
Wrthenjp etal..
1955
Wrthwupetal..
1955
WHherup et al.. .
1955
Wrthenjp et at..
. 1955
'Benchmark Category, a > adequate, p = provisional, i = interim; a *" indicate* that the benchmark value was an order of
magnibdB or more above the NEL or LEL for other adverse effects.
ID - Insufficient Data
August 1995
-------
APPENDIX B Heptachlor. 8
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial invertebrates, and plants. Each
value is identified as whole-boy or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). The following discussion describes the rationale for selecting the
biological uptake factors and provides the context for interpreting the biological uptake values
presented in Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for heptachlor. The
bioconcentration factor for fish was also estimated from the Thomann models (i.e., log Kow -
dissolved BCF/) and multiplied by the dissolved fraction (/j) as defined in Equation 6-21 to
determine the total bioconcentration factor (BCF/). The dissolved bioconcentration factor
(BCF/1) was convened to the BCF/ in order to estimate the acceptable lipid tissue
concentration (TC/) in fish consumed by piscivorous fish (see Equation 5-115). The BCF/
was required in Equation 5-115 because the surface water benchmark (i.e., FCV or SCV)
represents a total water concentration (C1). Mathematically, conversion from BCF/1 to BCF/
is accomplished using the relationship delineated in the Interim Report on Data and Methods
for Assessment of 2 S,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S. EPA,
1993i):
BCF/1 x fd = BCF/
Converting the predicted BCF,d of 102,329 L/kg LP to the BCF/ of 77,781 L/kg LP was in
reasonable agreement (i.e., within a factor of 4) of the geometric mean of three measured
BCF/ values presented in the master table on heptachlor (geometric mean = 146,900).
The bioaccumulation factor for terrestrial vertebrates; invertebrates, and earthworms were
estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation method is applied to
hydrophobic organic chemicals assuming that the partitioning to tissue is dominated by lipids.
Further, the method assumes that the BAFs and BCFs for terrestrial wildlife developed for
2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and
TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient quality to serve as the standard.
August 1995
-------
APPENDIX B Heptachlor - 9
The beef biotransfer factor CBBFs) for a chemical lacking measured data (in this case
heptachlor) is compared to the BBF for TCDD and that ratio (i.e., heptachlor BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion:
August 1995
-------
APPENDIX B
Heptachlor - 10
Table 4. Biological Uptake Properties
ootogicai
limnalic frophic
leveMNsh
limnetic trophic
Ievel3fish
fish
littoral trophic
level 4 fish
littoral tophic
level 3 fish
littoral topfrc
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF,
-------
APPENDIX B Heptachlor-11
References
Agency of Toxic Substances and Disease Registry (ATSDR). 1993. Toxicological Profile for
Heptachlor/Heptachlor Epoxide. Washington, D.C. U.S. Public Health Service (USPHS).
Akay, M.T. and U. Alp. 1981. The effects of BHC and heptachlor on mice. Hacettepe Bull
NatSciEng 10:11-22.
AQUIRE (AQUatic Toxicity Information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U. S. Environmental Protection Agency,
Duluth, MN, June 1995.
Arnold, D.W., G.L. Jr. Kennedy, M.L. Keplinger, et al. 1977. Dominant lethal studies with
technical chlordane, HCS-3260, and heptachlor: heptachlor epoxideJ. Toxicol. Environ.
Health 2:547-555. .
Blus, L. J., C. J. Henny, D.J. Heptachlor: Toxicolgy and safety evaluation. Industrial
Medicine & Surgery, p. 840 -844.
Lenhart, and T.E. Kaiser. 1984. Effects of hepachlor- and lindane-treated seed on Canada
geese. J. Wild!. Manage. 48(4): 1097 - 1111.
Eisler, M. 1968. Heptachlor: Toxicology and safety evaluation. Industrial Medicine &
Surgery, pp. 840 -844.
Garten, C. T. and J. R. Trabalka. 1983. Evaluation of models for predicting terrestrial food
chain behavior of xenobiotics. Environ. Sci. Technol. 17(10): 5 90-595.
Gossett, R.W., D.A. Brown, and D.R. Young. 1983. Predicting the bioaccumulation of
organic compounds in marine organisms using octanol/water partition coefficients.
Marine Pollution Bulletin, Vol. 14, No. 10 pp. 387 - 392.
s
Great Lakes Water Quality Initiative GLI, 1992. Tier D Water Quality Values for Protection
of Aquatic Life in Ambient Water - Support Documents. 11/23/92.
Green, V.A. 1970. Effects of pesticides on rat and chick embryo. In: Hemphill D. ed. Trace
substance environmental health 3rd. Proc. 3rd Ann Conf, University of Missouri, 183-209.
Kenaga, E. E. 1980. Correlation of bioconcentration factors of chemicals in aquatic and
terrestrial organisms with their physical and chemical properties. American Chemical
Society, 14(5): 553- 556.
August 1995
-------
APPENDIX B Heptachlor - 12
Lu. P., R.L. Metcalf, A. SI Hirwe, and J.W. Williams. 1975. Evaluation of environmental
distribution and fate of hexachlorocyclopentadiene, chlordene, heptachlor, and heptachlor
epoxide in a laboratory model ecosystem. /. Agric. Food Chem. 23(5): 967 - 973.
Mestitzova, M. 1966. On reproduction studies and the occurrence of cataracts in rats after
Rand, G.M., and S:R. long-term feeding of the insecticide heptachlor. Experientia.
23/1:42-43.
Nagy, K. A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:11-128.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
Opresko, D.M., B.E. Sample, G.W. Suter II, 1994. lexicological Benchmarks for Wildlife
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratoy, Oak Ridge, Tennessee.
Rand, G.M. and S.R. Petrocelli. 1985. Fundamentals of Aquatic Toxicology: Methods and
Applications. Hemisphere Publishing Corporation, New York.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter n, G. W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615-629.
Tiara, M. C. and L. De Viale. 1980. Porphyrinogen carboxy-lyase from chick embryo liver
in vivo effect of heptachlor and lindane. Int. J. Biochem. 12: 1033 - 1038.
\
U.S. EPA (Environmental Protection Agency). 1985. Drinking water criteria document for
heptachlor, heptachlor epoxide and chlordane (Final Draft) .S.B. Wilbur, et al..
Environmental Protection Agency, Cincinnati, OH. Environmental Criteria and
Assessment Office. March 1985. EPAy600/X-84/197-l.
-------
APPENDIX B Heptachlor - 13
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D.C. January.
U.S. EPA (Environmental Protection Agency). Integrated Risk Information System.
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region TV.
U.S. EPA (Environmental Protection Agency). 1993. Derivations of Proposed Human
Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative. PB93-
154672. Environmental Research Laboratory, Office of Research and Development,
Duluth, MR
Veith, G.D., D.L. DeFoe and B.V. Befgstedt 1979. Measuring and estimating the
bioconcentration factor of chemicals in fish; J. Fish. Res. Bd. of Canada. 36: 1040 -
1048.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
Witherup, S., F. P. Cleveland, F.E. Shaffer. 1955. The physiological effects of the
introduction of heptachlor into the diet of experimental animals in varying levels of
concentration: report of experiment No. 2. (unpublished study).
Witherup, S., K. Stemmer, P. Taylor. 1967. The effectts exerted upon the fertility of rats
and upon the viability of their offspring by the introduction of heptachlor into their daily
diets. Unpublished study prepared by University of Cincinnati.
August 1995
-------
terrestrial Toxicity - Heptachlor
Cas No. 76-44-8
Chemical
Heptachlor
Heptachlor
heptachlor
heptachlor
heptachlor
heptachlor
heplachlor
hepiachlor
8 male CD-I
mice
Male 'and
female
Sprague-
Dawtey rats
mouse
rat
mouse
rat
fat
rat
Type of
Elf eel
rep
rep
rep
rep
rep
rep
path
patti
Description
NOAEL
FEL
LOAEL
LOAEL
NOAEL
PEL
NOAEL
LOAEL
Value
15
0.25
6.5
2.6
10
6
0.15
0.25
Units
mg/kg/day
mo/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kgBW
mg/kg/day
mg/kg/day
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
gavage
-
oral
orai
oral
gavage
oral
oral
oral
Exposure
Duration/
Timing
single dose
60 days
1 0 weeks
80 weeks
5 successive
weekdays
18 months
2 years
2 years
Reference
Arnold et al., 1977
Green, 1970 in .
Hemphill, 1970
Akay and Alp, 1981
NCI. 1977 as cited in
ASTDR, 1993
Epstein et al., 1972
Mestitzova. 1967
Velsicol Chemical
Corporation, 1955aas
cited in IRIS, 1994
Velsicol Chemical
Corporation. 1955a as
cited in IRIS. 1994
Comments
The dose given was a
heptachlorheplachlor epoxide
mixture (25%:75%). No
adverse effect on the
reproductive capacity of the
male mice was noted.
The number of pregnancies.
the number of embryos, and
the mean litter size were all
reduced in (he F1 generation
and completely reduced in the
F2 generation.
100% infertility
Vaginal bleeding was reported
at this dose level.
No early fetal deaths or
preimplantation losses were
reported outside the control
limits.
A decrease in litter size and a
shortening of the life span of
the sucklings was observed at
this dose level.
-
Liver-to-body weight increases
were recorded in males only.
-------
Terrestrial To., .ty - Heptachlor
Cas No. 76-44-8
Name
heptachlor
heptachlor
heptachlor
heplachlor
heplachlor
Species
rat
rat
mouse
guinea pig
hamster
Tvoe of
Effect
rep
acute
acute
acute
acute
Description
NOEL
LD50
LD50
LD50
LD50
Value
0.5
40
68
116
100
Units
mg/kg/day
mg/kg
mg/kg
mg/ka
mg/kfl
Exposure
Route (oral,
s.c., l.v., l.p.,
Injection)
NS
oral
oral
oral
oral
Exposure
Duration/
Timing
3-generation
study
NS
NS
NS
NS
Reference
Velsicol Chemical,
1967 as cited in IRIS.
1993
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
Comments
No adverse effects were
reported at (his dose.
Page '2.
-------
Freshwater Toxicity - Heptachlor
CasNo. 76-44-8
Chemical
name
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Species
Aquatic
organisms
Fathead
minnow
Fish
Daphnids
Fish
Oaphnia
magna
Daphnia
magna
Simocephalus
serrulatus
Striped bass
Bluegill
Rainbow trout
Fathead
minnow
NS = Not Specified.
NA - Not Applicable.
Type of
effect
chronic
chron
chron
chron
chron
mort
mod
immob.
mort
mort
mort
mort
-
Description
NAWQC
MATC
CV
cv
EC20
LC50
EC50
EC50
LC50
LC50
LC50
LC50
Value
1.00E-02
0.86-1.84
1.26
3.18
0.86
78 - 120
(105.32)
42
47-80
(60.95)
3
18.0-220
(19.65)
7.0-19.4
(9.35)
78
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/flow
through)
NA
complete life
cycle test
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
Exposure
Ouratlon/T
Imlng
NS
NS
NS
NS
NS
48 hour
48 hour
48 hour
96 hour '
96 hour
96 hour
96 hour
Reference
Suter. 1992
Maceketal. 1976c
as cited in Rand
an'd Petrocelli.
1985
Suleretal, 1992
Suter el al., 1992
Suter etal. 1992
AQUIRE. 1994
AQUIRE. 1994
AQUIRE, 1994
AQUIRE. 1994
AQUIRE. 1994
AQUIRE. 1994
Henderson et al.,.
1959 as cited in
AQUIRE. 1994
Comments
Critical life stage end
points: embryo; larval, and
early juvenile; mortality
-------
Freshwater Biological U^ .^.ice Measures - Heptachlor
Cas No. 76-44-8
Chemical
name
Heptachlor
Heptachlor
Heptachlor
Heptachlor
Heptachlor
lieplachloi
heotachlor
heplachlor
Species
Fathead
minnow
Fathead
minnow
Three
species ot
marine fish
Five species
ot marine fish
Red pointer
crab
lish
fish
fish
B-lactor
(BCF, BAF,
BMP)
BCF
BCF
BSAF
BSAF
BSAF
BAF
BA.F
BCF
Value.
9500
17,000-
23.814
(20,773)
6
0
0
13804
3802
484
Measured or
predicted
(m,p)
m
m
P
P
P
NS
NS
c
Units
NS
NS
ug/g
ug/g
ug/g
LAg
L/kg
NS
Reference
Veithetal. 1979
Macek et al., 1976 as cited in
AQUIRE, 1994
Columboet. al., 1990
Cosset et al.. 1983
Gossetet. at., 1983
Garten and Trabalka, 1 983
Garten and Trabaika. 1983
Stephan 1993
Comments
Adult lifestage
2 generations
»
Flowing water; All estimates were
calculated based on published
data, the type ot studies from
which the data were, taken were
not specified. '
Microcosm; All estimates were
calculated based on published
data, the type of studies trom
which the data were taken were
not specified.
Assuming 1 .0% lipid
Page I
-------
Freshwater Biological Uptake Measures - Heptachlor
Cas No. 76-44-8
Chemical
name
neptachlor
Heptachlor
heptachlor
Species
fish
fish
f
fish
NS = Not Specified.
B-tactor
(BCF. BAF,
BMP)
BCF
BCF
BCF
Value
2726
931
1250
Measured or
predicted
(m.p)
m
m
m
Units
NS
NS
NS
Reference
Schimmel et al., 1976 as cited
in Stephan 1993
Goodman et al.. 1978 as cited
in Stephan 1993
Veilh et al., 1979b as cited in
Stephan 1993
Comments
Assuming 1 .0% lipid
Assuming 1 .0% lipid
0
Assuming 1 .0% lipid
-------
Terrestrial Biological Uptake Me*_ures - Heptachior Cas No.: 76-44-8
Chemical
Name
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Heptachior
Species
Cattle
Cattle
Swine
Swine
Cattle
(beef)
Cattle (milk)
sheep
poultry
cow
swine
B-factor
(BCF. BAF.
BMP)
BCF
BCF
BCF
BCF
BTF
BTF
BAF
BAF
BAF
BAF
Value
0.4
0.6
6.55
6.4
0.0154
0.00323
-----
Measured
or
Predicted
(m.p)
M
M
M
M'
M
M
NS
NS
NS
NS
Units
NS
NS
NS
NS
NS
NS "
kg fat/ kg
diet
kg fat/ kg
diet
kg fat/ kg
diet
kg fat/ kg
diet
Comments
BTF =
Biotransfer
factors
BTF =
Biotransfer
factors
Reference
(Claborn, et.al.. 1960 as
cited in Kenaga, 1980)
(Claborn, et.al., 1960 as
cited in Kenaga, 1980)
(Claborn, et.al., 1956 as
cited in Kenaga, 1980)
(Claborn, et.al.. 1956 as
cited in Kenaga, 1 980)
(Travis and Arms, 1 988)
(Travis and Arms, 1 988)
(Garten and Trabalka,
1983)
(Garten and Trabalka,
1983)
(Garten and Trabalka,
1983)
(Garten and Trabalka,
1983)
Page I
-------
Terrestrial Biological Uptake Measures - Heptachlor Cas No.: 76-44-8
heptachlor
heptachlor
NS "= Not S
P'9
earthworm
pecifjed
BAF
m
. .
NS
Pigs were
led 5
mg/kg/day,
resulting in a
fat
heptachlor
level of .37
ppm
.
Halackaelal . 1974 (as
cited in Toxicological Profile
for Heptachlor/Heptachlor
Epoxide. ASTDR, 1993)
. -
-»2
-------
APPENDIX B Hepuchlor epoxide - I
Toxicological Profile for Selected Ecological Receptors
Heptachlor epoxide
Cas No.: (1024-57-3)
Summary: This profile on heptachlor epoxide summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwaterecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive orother effects reasonably assumed to
impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish were
generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria).
Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs)" are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with,the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for which reported
dose-response data for mammalian wildlife. However, lexicological studies involving
heptachlor epoxide exposure to mammals have been conducted using laboratory mice. Arnold
et al. (1977) administered single oral doses of 75 percent hepaichlor epoxide (25 percent
hepatchlor) lo eight male Charles River CD-I mice at 7.5 and 15 mg/kg-day dose levels.
These mice were bred with ihree untreated females for 6 weeks with no effecis on ihe
reproductive capacity noted in the male mice, therefore the NOAEL was set at 15 mg/kg-day.
In anoiher sludy by Epsiein et al. (1972), heptachlor epoxide exposure at 8 mg/kg-day in mice
did not produce early fetal deaths or preimplantation losses outside the control Limits (inferred
as a NOEL).
August 1995
-------
APPENDIX B Heptachlor epoxide - 2
The dose levels used In the aforementioned studies were not sufficient for establishing a
dose-response relationship. Since no adverse effects on reproductive endpoints were identified,
benchmark values protective of the mammalian community in a freshwater ecosystem were
not derived.
Birds: No studies were identified concerning heptachlor epoxide toxicity in avian speciesand
therefore, no benchmarks were developed.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is
available for heptachlor epoxide. Therefore, the Secondary Chronic Value (SCV) of 0.51
mg/L as reported in AQUIRE for heptachlor epoxide was selected as the benchmark value
protective of fish and aquatic invertebrates. Because the benchmark is based on an SCV and
there were no lower toxicity values in the data set, it was categorized as interim.
Aquatic plants: The lexicological benchmarks for aquatic plants were either (1) a no
observed effects concentration (NOEQ or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Adequate data sufficient for the development of benchmark values were not identified in
Suter and Mabrey (1994) or in AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQJ method. The EQ method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
octanol-carbon partition coefficient (K^ to determine protective sediment concentration
(Stephan, 1993). The EQ_ number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The SCV, for heptachlor epoxide was used to calculate an EQp value of 0.5128 mg
heptachlor epoxide/kg organic carbon. Assuming a mass fraction of organic carbon for the
sediment (f,^) of O.OS.rthe benchmark for the benthic community is 1.2 mg/kg sediment.
Since the EQp number was based on an SCV, the sediment benchmark was categorized as
interim.
August 1995
-------
APPENPIX B
Hepitachlor epoxide 3
Table 1. Toxicoiogical Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
fefWMMMN*
Speeiea
rnv)K
river otter
baldeagia
osprey
great blue heron
malard
lesser scaup
spotted sandpiper
herring gul
kingfiehar
9afvoMMii(
Yakwrna**-
**
ID
10
10
ID
ID
ID
ID
ID
ID
ID
9tMaY
-
-
-
-
-
ittact
-
-
-
9****
-
-
DMOt»fci>
-
-
-
-
-
tt»
-
-
OrtyMtSMm
'-
.
-
*Band>maffc Category, a » adaquata. p provisional, i > intorim; a "" IndicatM tut th* banchrruirk valua was an ordar, of
magnituda or mora above the NEL or LEL for other adverse effects.
ID > Insufficient Data
August 1995
-------
APPENDIX B
Heptachlor epoxide - 4
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
SvprMwrtrtw
Sped** -
fish and aquatic
invertebrate*
aquatic plants
bentftic community
Benchmark
Vato*
wflflL. .
5.1E-01 (i)
No data
1.25E+00(i)
8fuo₯
aquatic
organisms
-
aquatic
organisms
OMOfeto*
scv
-
SCVxK^
' OrfeinrtScHK?
. AQUIRE
-
AQUIRE
'Benchmark Category, a » adequate, p = provisional, i = interim; a "" indicate* lhat the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverse effects.
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (C^ for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned in the freshwater ecosystem discussion, no suitable subchronic or
chronic studies were found for mammalian wildlife exposure to heptachlor epoxide. Since no
additional laboratory mammal studies focusing on reproductive or other critical endpoints
were identified, a mammalian benchmark for terrestrial ecosystems was not calculated.
Birds: Toxicity studies documenting terrestrial avain exposure to heptachlor epoxide were not
identified and thus, no benchmarks were derived.
August 1995
-------
.APPENDIX B
Heptachlor epoxide S
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
deer mouse
short-taitod shrew
meadow vote
Eastern cottontail
red fox
raccoon
white- tailed deer
red- tailed haw*
American kestrel
Northern bobwhite
American robin
American woodcock
lants
U soil community
«0*0Hfcy
10
10
ID
ID
10
10
10
ID
ID
ID
ID
ID
ID
10
Hudyt»»eiM
-
-
'
-
-
-
!.:!.:***:
-
-
-
-
-
-
SktfyVUw
r««*»Hh»
-
-
-
DMCripdM
-
.
:
-
-
-
-
«F
.
'
-
-
-
-
(MQlnriSNMt-
--
-
-
-
-
'Benchmark Category, a * adequate, p > provisional, i interim; a indcatet (hat (ho benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID > Insufficient Data
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for heptachlor epoxide and, as a result, a
benchmark could not be developed. .
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
August 1995
-------
APPENDIX B Heptachlor epoxide - 6
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified' as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using '
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for heptachlor
epoxide. The bioconcentration factor for fish was also estimated from the Thomann models
(i.e., log Kow - dissolved BCF/) and multiplied by the dissolved fraction (/j) as defined in
Equation 6-21 to determine the total bioconcentration factor (BCF,1). The dissolved
bioconcentration factor (BCF/1) was converted to the BCF/ in order to estimate the
acceptable lipid tissue concentration (TCI) in fish consumed by piscivorous fish (see Equation
5-115). The BCF/ was required in Equation 5-115 because the surface water benchmark (i.e.,
FCV or SCV) represents a total water concentration (C1). Mathematically, conversion from
BCF;d to BCF/ is accomplished using the relationship delineated in the Interim Report on
Data and Methods for Assessment of 2 3,7,8~Tetrachlorodibenzo-p-dioxin Risks to Aquatic
Wildlife (U:S. EPA, 1993i):
BCF,d x fd = BCF/
Converting the predicted BCF/1 of 56,234 L/kg LP to the BCF/ of 47,850 L/kg LP was
considerably lower than the single measured BCF/ value of 189,500 cited in Stephan (1993).
However, the predicted value was used because: (1) no information was identified that
suggested that heptachlor epoxide should deviate appreciably from the log KOW/BCF
relationship, and (2) given the variability in measured BCF values, a single measured value
was considered an insufficient (and possibly overconservative) basis for the bioconcentration
factor.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates and
August 1995
-------
APPENDIX B
Heptachlor epoxide - 7
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. For hydrophobic organic constituents, the bioconcentration factor for
plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, driect depostion on leaves
and grasses, and uptake into the plant through air diffusion.
Table 4. Biological Uptake Properties
ES?'
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophK
level 2
invertebrate*
terrestrial
vertebrates
terrestrial
invertebrates
plants
BCF, BAF, «r
i8AF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BAF
BAF
BAF
Itpid baa«d or
lipid
lipid
lipid
lipid
lipid
lipid .
whole-body
whole-body
whole-body
whole- plant
. 70, 966 ( d)
70,31 3 (d)
47.850 (t)
64,836 (d)
70,553 (d)
146,934 (d)
1.1 E-04
'l.l E-04
8.6E-04
2.0 E -01
owe*
predicted value; Thomann,
1986 .
predicted value; Thomann,
1989
predicted valun based on
Thomann, 1989 and adjusted to
estimate tonal BCF
predicted value: Thomann «t
al.. 1992
predicted value ;Thomann at al.,
1992
predicted value: Thomann 0t
al., 1992
estimated based on beef
biotransfer ratio with 2.3,7,8-
TCDO
estimated based on beef
biotransfer ratio with 2.3,7.8-
TCDD
estimated baaed on beef
biotransfer ratio wft 2,3,7.8-
TCDO
U.S. EPA. 1990e
d = refers to dissolved surface water concentration
t » refers to total surface water concnetralion
ID - insufficient data
August 1995
-------
APPENDIX B HeptachJor epoxide - 8
References
AQUIRE (AQUatic Toxicityjnformation REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U. S. Environmental Protection Agency,
Duluth, MN, June 1995.
Arnold, D.W., G.L. Jr. Kennedy, M.L. Keplinger, et al. 1977. Dominant lethal studies with
technical chlordane, HCS-3260, and heptachlor: heptachlor epoxide. /. Toxicol . Environ.
Health 2:547-555.
Agency of Toxic Substances and Disease Registry (ATSDR). 1993. Toxicological Profile for
Heptachlor/Heptachlor Epoxide. Washington, D.C U.S. Public Health Service (USPHS).
Epstein, S.S., E. Arnold, J. Andrea, et al. 1972. Detection of chemical mutagens by the
dominant lethal assay in the mouse. Toxicol. Appl. Pharmacol. 23:288-325.
Garten, C. T., Jr., and J.R. Trabalka. 1983. Evaluation of models for predicting terrestrial
food chain behavior of xenobiotics. Environ. Sci. Technol. 17 (10): 590 -595.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. .
Suter n, G. W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D. C.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D.C. January.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connolly, and T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615-629.
August 1995
-------
Terrestrial Toxicity - Heptachior epoxide
Cas No. 1024-57-3
Chemical
Name
heptachlor
epoxide
heptachlor
epoxide
heptachlor
epoxide
heptachlor
epoxide
heptachlor
epoxide
Species
mice
mice
rat
mouse
rabbit
NS =- Not Specified
Type of
Effect
rep
rep
acute
acute
acute
Description
NOAEL
NOAEL
LD50
LD50
LD50
Value
15
8
15
39
144
Units
mg/kg-day
mg/kg-day
mg/kg
mg/kg
mo/kg
Exposure
Route (oral,
S.C., I.V., I. p.,
Infection)
gavage
gavage
oral
oral
oral
Exposure
Duration/
Timing
single dose
5 successive
workdays
NS
NS
NS
Reference
Arnold el al.,
1977
Epstein et al..
1972
RTECS^1994
RTECS. 1994
RTECS. 1994
Comments
The dose given was a
heptachlorheptachlor
epoxide mixture
(25%:75%). No adverse
effect on the reproductive
capacity of the male mice
was noted.
No early fetal deaths or
preimplantation losses
were reported outside of
the control limits.
-------
APPENDIX B Heptachlor epoxide - 9
U.S. EPA (Environmental Protection Agency). 1992. 304(a) Criteria and Related
Information for Toxic Pollutants. Water Management Division, Region IV.
U.S. EPA (Environmental Protection Agency). 1993i. Interim Report on Data and Methods
for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin rishks to Aquatic Wildlife.
EPA/600/R-93/055. Office of Research and Development, Washington, D.C.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Freshwater Bioiogica! uptake measures - Heptachtor epoxide
Cas No. 1024-57-3
Chemical
name'
leplachlor
epoxide
heptachlor
epoxide
leptachlor
epoxide
rteplaclilor
epoxide
heptachlor
epoxide .
Species
fathead
minnow
fish
fish
fish
fish
B-factor
(BCF. BAF,
BMP)
BCF
BAF
BAF
BCF
BCF
Value
14400
14454
4898
89.2
1895
Measured or
Predicted
(m,p)
m
NS
NS
P
m
Units
NS
M
-------
Freshwater ToxicU, Heptachlor epoxide
Cas No. 1024-57-3
Chemical
Name
heptachlor
epoxide
heptachlor
epoxide
Species
Daphnia
magna
Daphnia
magna
NA = Not Applicable.
Type of
Effect
mort
mort
Description
LC50
LC50
Value
120
240
Units
ug/L
ug/L
Test type
(static/flow
through)
NA
NA
Exposure
Duration/
Timing
1.08 day
48 hour
Reference
AQUIRE, 1994
AQUIRE, 1994
Comments
-------
APPENDIX B Hexachlorobenzene 1
lexicological Profile for Selected Ecological Receptors
Hexachlorobenzene
Cas No.: 118-74-1
Summary: This profile on hexachlorobenzene summarizes the lexicological benchmarks and
biological uptake .measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biornagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the data presented in the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found which reported dose-
response data for mammalian wildlife. However, lexicological studies involving
hexachlorobenzene exposure to laboratory rats have been conducted. Grant et al. (1977)
conducted a two-year, four-generation rat study involving the dietary administration of 0, 10,
20, 40, 80, 160, 320, or 640 ppm hexachlorobenzene. For the Fl and F3 generations, the
pups from dams fed 40 ppm developed increased liver weights and increased aniline
hydroxylase activities. Rats fed diets containing less than 20 ppm hexachlorobenzene
displayed no reproductive effects. To convert these levels inio daily doses, a Sprague-Dawley
reference body weight of 458 grams and a food consumption rate of 0.033 kg/day was used
(U.S. EPA, 1988). The resulting NOEL and LOEL based on ihe aforementioned
August 1995
-------
APPENDIX B Hexachlorobenzene - 2
j>iBBBaMMPaaa«BMiiaBff^
developmental effects were calculated as 1.46 mg/kg-d and 2.92 mg/kg-d. In another study,
Khera (1974) administered 0, 10, 20, 40, 60, 80, or 120 mg/kg-day hexachlorobenzene to rats
via gavage on gestation days 6 to 21. Maternal toxicity and reduced fetal weights were
observed at the 80 and 120 mg/kg-day dose levels, and, therefore, a NOEL of 60 mg/kg-day
was interpreted from this study.
The developmental NOEL of 1.46 mg/kg-d from the Grant et al. (1977) study was selected to
derive the lexicological, benchmark for aquatic mammals because: (1) it was a four-
generation study which focused on developmental toxicity as the critical endpoint, (2) the
study contained sufficient dose-response information, and (3) the study resulted in the lowest
NOEL among identified studies with sufficient dose-response information. In the other
studies reviewed, dose-response curves were established for a single generation and/or
reproductive effects were observed at higher dose levels.
The study value from the Grant et al. (1977) study was scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994)
.
Benchmark = NOAEL, x _ L
I bw...
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Grant et al. (1977) study documented reproductive effects from hexachlorobenzene exposure
to female and male rats, the mean body weight of both genders of representative species were
used in the scaling algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental, effects of hexachlorobenzene, as
well as chronic survival. All of the studies identified were conducted using laboratory rats
and mice and as such, inter-species differences among wildlife species were not identifiable.
Therefore, an inter-species uncertainty factor was not applied. There were several short-term
studies of oral exposure to laboratory mammals which reported histopathological toxicity of
hexachlorobenzene at levels more than a magnitude lower than the benchmark value. Based
on the data set for hexachlorobenzene, the benchmarks developed from the Grant et al. (1977)
study were categorized as adequate, with a "*" to indicate that adverse effects may occur at
the benchmark level.
Birds: No subchronic or chronic studies, with adequate dose-response regimes, were
identified for hexachlorobenzene exposure to the representative avian species. However, two
studies were identified which documented the reproductive effects of hexachlorobenzene on
Japanese quail. Vos et al. (1971) administered HCB to Japanese quail for 90 days at dietary
-. ,- TiCiC
-------
APPENDIX B Hexachlorobenzene - 3
concentrations of 1, 5, 20, and 80 ppm. Increased liver weights and slight liver damage were
detected in adult quails dosed at 5 ppm. Since it is unclear whether or not these liver effects
would exhibit an adverse impact at the population level, reproductive endppinits relevant to
population sustainability were extracted from the Vos et al. (1971) study. Reproductive
effects, in terms of reduced hatchability and significantly reduced volume of eggs, were
observed in the quails exposed to the 20 ppm dietary dose. Because no information on daily
food consumption rates were provided, the use of an allometric equation was required to
convert the dose from dietary mg/kg to mg/kg-d:
Food consumption = 0.648(W°'651), where W is body weight in g (Nagy, 1987)
A calculated mean body weight of 127 g for male and female Japanese quail (Vos et al.,
1971; Schwetz et al., 1974) was used to convert the 5 ppm NOEL for reproductive effects to
a daily dose of 0.60 mg/kg-d. Schwetz et al. (1974) reported that hexachlorobenzene
decreased the survival of Japanese quail chicks following the administration 20 mg/kg (diet)
for 90 days. Using the same food consumption equation as above and a mean body weight of
121 g (Schwetz et al., 1974), the Adverse Effects Level (AEL) of 20 mg/kg diet was
converted to a daily dose of 2.4 mg/kg-d.
The NOEL reported by Vos et al. (1971) was selected as the toxicological benchmark
representative of avian species because it was the lowest toxicity value in the data set, had
sufficient dose response data, and focused on reproductive toxicity as a critical endpoint. The
study by Schwetz et al. (1974) was not considered suitable for derivation of an avian
benchmark value based on the lack of dose-response information.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian
species representative of a freshwater ecosystem, the NOAEL of 0.60 mg/kg-day from Vos et
al. (1971) was scaled using the cross-species scaling method of Opresko et al. (1994).
Data were identified on the reproductive arid mortality effects of hexachlorobenzene exposure
to avian species. Laboratory experiments of similar types were not conducted on a range of
avian species and as such, inter-species differences among wildlife species were not
identifiable. There were no other values in the data set which were lower than the benchmark
value. Since the avian data set for hexachlorobenzene contained a sufficient set of
endpoints for population sustainability, as discussed in 4.3.2, the benchmarks developed from
the Vos et al. (1971) study were categorized as adequate.
Fish and aquatic invertebrates: A Final Chronic Value (FCV) of 6.0E-3 for
hexachlorobenzene was inferred from the Ambient Water Quality Criteria document (U.S.
EPA, 1980). The available data in the AW.QC indicate that "hexachlorobenzene does not
cause significant adverse effects of freshwater aquatic life at or below 6 ug/1." Therefore, a
FCV of 6E-03 mg/1 was selected as the benchmark protective of fish and aquatic
invertebrates. This benchmark was categorized as adequate since it was inferred by the
AWQC document.
August 1995
-------
APPENDIX B Hexachlorobenzene 4
Aquatic Plants: The "toxicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (EC^) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwn),
Aquatic plant data was not identified for hexachlorobenzene and, therefore, no benchmark
was developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (Koc) to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. The FCV interpreted from the AWQC document for hexachlorobenzene (U.S.
EPA, 1980) was used to calculate an EQp number of 1,530 mg hexachlorobenzene/kg
organic carbon.. Assuming a mass fraction of organic carbon for the sediment (f^) of 0.05, .
the benchmark for the benthic community is 76.5 mg/kg. Since the EQp number was based
on a FCV established for the AWQC, the sediment benchmark is categorized as adequate.
August 1995
-------
APPENDIX B
Hexachlorobenzene 5
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative.
Specie*.
mink
river ottar
bald eagle
ospray
great blue heron
mallard
lessor scaup
spotted sandpiper
herring guU
Kingfisher
Benchmark
Value* raoAg-d
1.2 (a')
0.71 (a')
0.26 (a)
0.32 (a)
0.29 (a)
0.34 (a)
0.38 (a)
0.79 (a) '
0.35 (a)
0.67 (a)
Study
Specie*
rat
rat
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Etfect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
: Study
Value
mg/kg-day
1.5
1.5
0.6
0.6 .
0.6
0.6
0.6
0.6
0.6
0.6
Description
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
SF
OriQln*! Saute*
Grant etal., 1977
Grant et aJ.. 1977
Vos et aJ.. 1971
Vos etal., 1971
Vos etal., 1971
Vos et.al... 1971
Vos etal.. 1971
Vos et ai., 1971
Vos et ai.. 1971
Vos etal., 1971
"Benchmark Category, a » adequate, p » provisional, i » interim; a "" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Hexachlorobenzene 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Repre««ntaUv*
Specie* ,/:
fish and aquatic
invertebrates
aquatic plants
benlhic
community
Benchmark
:.-,.; jvvaiu«»:..y.
;: "":.V:_,_lt
": n>yi»
3.68 E-03 (a)
10 .
76.5 (a) mg/kg
sediment
Study Sp«cl**
aquatic
organisms
-
aquatic
organisms
Description
FCV (»)
FCVxK^.
Original
Sourc*
. U.S.EPA, 1980
-
U.S.EPA, 1980
'Benchmark Category, a * adequate, p provisional, i = interim; a "" indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
* = from AWQC 'the available data indicate that HC8 does not cause significant adverse effects of freshwater
aquatic life at or below 6 ug/f
10 = Insufficient Data
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to
hexachlorobenzene. Because of the lack of additional mammalian toxicity studies, the same
surrogate-species study (Grant et al., 1971) was used to derive the hexachlorobenzene
lexicological benchmark for mammalian species representing the terrestrial ecosystem. The
study value from the Grant et al. (1971) study was scaled for species representative of a
terrestrial ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994). Since the Grant et al. (1977) study documented reproductive effects from
hexachlorobenzene exposure to female and male rats, the mean body weight of both genders
of representative species were used in the scaling algorithm to obtain the lexicological
benchmarks. Based on the data set for hexachlorobenzene, the benchmarks developed were
categorized as adequate, with a "*" to indicate that adverse effects may occur at the
benchmark level.
Birds: No additional avian toxiciiy siudies were identified for species representing the
terrestrial ecosystem. Thus, for the avian species representative of a terrestrial ecosystem, the
NOAEL of 0.60 mg/kg-day from ihe Vos el al. (1971) study was used as the benchmark
August 1995
-------
APPENDIX B Hexachlorobenzene - 7
value. This value was then scaled for species representative of a terrestrial ecosystem using a
cross-species scaling algorithm adapted from Opresko et al. (1994). Since the avian data set
for hexachlorobenzene contained a sufficient set of endpoims for population sustainability, as
discussed in 4.3.2, the benchmarks developed from the Vos et al. (1971) study were
categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for hexachlorobenzene and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Hexachlorobenzene g
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Representative
Specie*
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plants
soil community
Benchmark Value'
mg/kg-d
3.2 (a')
3.3 (a')
2.8 (a')
1.1 (a')
0.82 (a*)
0.78 (a')
0.35 (a)
0.61 (a)
0.55 (a)
0.67 (a)
0.56 (a)
ID
ID
Study
Specie*
rat
rat
rat
rat
rat
rat
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
-
Effect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study
Value
mg/kg-day
1.5
1.5
1.5
1.5
1.5
1.5
0.60
0.60
0.60
0.60
0.60
Description;.
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
NOEL
SF
-
-
-
Origin*) Souroe-
Grant etal., 1977
Grant etal.. 1977
Grant etal.. 1977
Grant et a!.'. 1977
Grant etal., 1977
Grant et al., 1977
Vos etal., 1971
Vos etaJ.. 1971
Vo» etal.. 1971
Vos etal., 1971
Vos etaJ.. 1971
Benchmark Category, a > adequate, p - provisional, i * interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEI for other adverse effects.
ID - Insufficient Data
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
August 1995
-------
APPENDIX B Hexachlorobenzene - 9
chemicals with log K-ow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above: 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for
hexachlorobenzene. The predicted BAF/1 for trophic level 4 fish in both the limnetic and
littoral ecosystems is in reasonable agreement (i.e., within a factor of 2) with the geometric
mean BAF/ (2,085,900) of the three measured values presented in Derivation of Proposed
Human Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative (Stephan,
1993). The geometric mean of the measured values was based on data from Oliver and Nicol
(1982) and Oliver and Niimi (1983 and 1988) for trout and salmonids. The bioconcentration .
factor for fish was estimated as the geometric mean of 7 measured BCF/ values presented in
Stephan (1993). Although the predicted value of 160,115 did not differ significantly from the
geometric mean of measured values (i.e., within a factor of approximately 2), the high quality
and number .of values in the data set was considered sufficient rationale for using the
geometric mean.
The bioaccumulation factor for terrestrial vertebrates and invertebrates was estimated as
described in Section 5.3.5.2.3. Briefly, the extrapolation method is applied no hydrophobic
organic chemicals assuming that the partitioning to tissue is dominated by lipids. Further, the
method assumes that the BAFs and BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD
in the Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp
and Paper Sludge (Abt, 1993) are of sufficient quality to serve as the standard. The beef
biotransfer factor (BBFs) for a chemical lacking measured data is compared to the BBF for
TCDD and that ratio (i.e., hexachlorobenzene BBFyTCDD BBF) is multiplied by the TCDD
standard for terrestrial vertebrates, invertebrates, and earthworms, respectively. The BCFl for
earthworms was a measured value identified in a study by Belfroid et al. (1994) on
earthworm exposure to chlorobenzenes in soil. Assuming a lipid fraction for earthworms of
0.01 (Belfroid et al., 1993), the measured value was converted to a whole-body BCF by
multiplying the lipid-based BCF/ by the lipid fraction, resulting in a whole-body BCF of 2.15.
For hydrophobic organic constituents, the bioconcentration factor for plants was estimated as
described in Section 6.6.1 for above ground leafy vegetables and forage grasses. The BCF is
based on route-tb-leaf translocation, direct deposition on leaves and grasses, and uptake into
the plant through air diffusion.
August 1995
-------
APPENDIX B
Hexachlorobenzene 10
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 Jish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF '
BAF
BAF
BAF
BAF
BCF
BCF
BCF
8pid*b«eed or
whole-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole- body
lipid
whole-plant
value
1,201. 943 (d)
905,176 (d)
336,600 (t)
1,142,641 (d)
1, 160,307 (d)
1,918.811 (d)
0.0039
0.0037
4,100
0.026
ourc*
predicted value based on Thomann, 1989,
food chain model
predicted value based on Thomann, 1989,
food chain model
predicted value based on Thomann, 1989
and adjusted to estimate total BCF
predicted value based on Thomann et aJ.,
1992. food web model
predicted value based on Thomann el al. .
1992. food web model .
predicted value based on Thomann et al.,
1992, food web model
estimated based on beef biotransfer ratio
with 2,3,7.8- TCDD
estimated based on beef biotransfer ratio
with 2,3.7,8- TCDO
measured value in g soM/g lipid from Betfroid
et al., 1994
U.S. EPA, 1992e
d » refers to dissolved surface water concentration
t » refers to total surface water concentration
August 1995
-------
APPENDIX B Hexachlorobenzene.il
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
Agency of Toxic Substances and Disease Registry (ATSDR). 1989. Toxicological Profile for
Hexachlorobenzerte Epoxide. Washington, D.C. U.S. Public Health Service (USPHS).
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Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Bailey, J., V. Knauf, W. Mueller and W. Hobson. 1980.. Transfer of hexachlorobenzene and
polychlorinated biphenyls to nursing infant rhesus monkeys: Enhanced toxicity. Environ.
Res. 21(1):190-196. As cited in U.S. EPA (Environmental Protection Agency. 1984.
Health Effects Assessment for Hexachlorobenzene. Environmental Criteria and
Assessment Office, Cincinnati, OH.
Belfroid, A., A. Van Wezel, M. Sikkenk, W. Seinen, K. Van Gestel, and J. Hermens. 1994. ,
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andrei): experiments
in soil. Environmental Toxicology and Chemistry. 13:93-99.
Belfroid, A., A. Van Wezel, M. Sikkenk, K. Van Gestel, W. Seinen, and J. Hermens. 1993.
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andrei): experiments
in water. Ecotox. and.Environ. Safety. 25:154-165.
Boger, A., G. Koss, W. Koransky, R. Naumann and H. Frenzel. 1979. Rat liver'alterations
after chronic treatment with hexachlorobenzene. Virchows Arch. [Path Anat.J 832(2): 127-
137. As cited in U.S. EPA (Environmental Protection Agency. 1984. Health Effects
Assessment for Hexachlorobenzene. Environmental Criteria and-Assessment Office,
Cincinnati, OH.
Carlson, A.R. and P.A. Kosian. 1987. Toxicity of Chlorinated Benzenes to Fathead
Minnows (Pimephales promelas). Arch. Environ. Contam. Toxicoi., 16:129-135. As cited
in Stephan. 1993. Derivations of Proposed Human Health and Wildlife-Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
den Tonkelaar, E.M., H.G. Verschuuren, J. Bankovska, et al.. 1978. Hexachlorobenzene
toxicity in pigs. Toxicoi. Appl. Pharmacol. 43:1370. As cited in U.S. EPA
(Environmental Protection Agency. 1984. Health Effects Assessment for
Hexachlorobenzene. Environmental Criteria and Assessment Office, Cincinnati, OH.
August 1995
-------
APPENDIX B , Hexachlorobenzene 12
57 FR 24152. June 5; 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-species Scaling Factor for Carcinogen Risk Assessment Based oh
Equivalence of mg/kg3/4/day.
Grant, D.L., F. Iverson, G.U. Hatina and D.C. Villeneuve. 1974. Effects of
hexachlorobenzene on hiver porphyrin levels and microsomal enzymes in rats. Environ.
Physiol. Biochem. 4:159. As cited in U.S. EPA (Environmental Protection Agency.
1984. Health Effects Assessment for Hexachlorobenzene. Environmental Criteria and
. Assessment Office, Cincinnati, OH.
Grant D.L., W.E. Phillips, and G.V. Hatina. 1977. Effect of hexachlorobenzene on
reproduction in the rat. Arch Environ Contam Toxicol 5:207-216.
Hansen, L.G., S.B. Dorn, S.M. Sundlof, and R.S. Vogel. 1978. No'title provided. J. Agric.
Food Chem. 26(6): 1369. As cited in U.S. EPA (Environmental Protection Agency. 1984.
Health Effects Assessment for Hexachlorobenzene. Environmental Criteria and
Assessment Office, Cincinnati, OH.
Howard, P.H. 1990. Handbook of Environmental Fate and Exposure Data for Organic
Chemicals. Volume I. Large Production arid Priority Pollutants. Lewis Publishers. Chelsea,
Michigan.
IARC (International Agency for Research of Cancer). 1979. I ARC Monographs on the
Evaluation of the Carcinogenic Risk of Chemicals to Humans: Some Halogenated
Hydrocarbons - Vol. 20.
Khera, K.S. 1974. Teratogenicity and dominant lethal studies on hexachlorobenzene in rats.
Food Cosmet. Toxicol. 12:471-477.
Kitchin,'K.T., R.E. Linder, T.M. Scotti, et al. 1982. Offspring mortality and maternal lung
pathology in female rats fed hexachlorobenzene. Toxicology. 23:33-39.
Konemann, H., and K. van Leeuwen. 1979. Toxicokinetics in Fish: Accumulation and
Elimination of Six Chlorobenzenes by Guppies. In: Quantitative Structure-Activity
Relationships for Kinetics and Toxicitv of Aquatic Pollutants and Their Mixtures in Fish,
H. Konemann, Ed. pp. 19-31. As cited in Stephan, 1993. Derivations of Proposed
Human Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative.
PB93-154672. Environmental Research Laboratory, Office of Research and Development,
Duluth, MN. " ' ,
August 1995
-------
APPENDIX B Hexachlorotenzene - 13
Konemann, H. and K-. van Leeuwen. 1980. -Toxicokineti.es in Fish: Accumulation and
Elimination of Six Chlorobenzenes by Guppies. Chemosphere* 9: 3-19. As cited in .
Stephan, 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Kosian, P., A. Lemke, K. Studders, and G. Vieth. 1981. The Precision of the ASTM
Bioconcentration Test EPA 600/3-81-022, National Technical Information Service,
Spirngfield VA. As cited in Stephan, 1993. Derivations of Proposed Human Health and
Wildlife Bioaccumulation Factors for the Great Lakes Initiative. PB93-154672.
Environmental Research Laboratory, Office of Research and Development, Duluth, MN.
Kosian, P., A. Lemke, K. Studders, and G. Vieth. 1981. The Precision of the ASTM
Bioconcentration Test EPA 600/3-81-022, U.S.EPA, Duluth MN: 20 p. As cited in
AQUIRE (AQt/atic Toxicity Mormation /?Etrieval Database). 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Kuiper-Goodman, T., D.L. Grant, C.A. Moodie, G.O. Korsrud and I.C. Munro. 1977.
Subacute Toxicity of Hexachlorobenzene in the Rat. Toxicol. Appl. Pharmacol.
40(3):529-549. As cited in U.S. EPA (Environmental Protection Agency. 1984. Health
Effects Assessment for Hexachlorobenzene. Environmental Criteria and Assessment
Office, Cincinnati, OH.
Mendoza, C.E., B.T. Collins, J.B. Shields and G.W. Laver. 1978. Effects of
Hexachlorobenzene or Hexabromobenzene on body and organ weights of preweanling rats
after a reciprocal transfer between the treated and control dams. J. Agric. Food Chem.
26(4): 941-945. As cited in U.S. EPA (Environmental Protection Agency. 1984. Health
Effects Assessment for Hexachlorobenzene. Environmental Criteria and Assessment
Office, Cincinnati, OH.
Mendoza, C.E., J.B. Shields and G.W. Laver. 1979. Comparison of the porphyrinogenic
activity of hexabromobenzene and hexachlorobenzene in primiparous Wistar rats. Bull.
Environ. Contam. Toxicol. 21(3):358-364. 26(4): 941-945. As cited in U.S. EPA
(Environmental Protection Agency. 1984. Health Effects Assessment for
Hexachlorobenzene. Environmental Criteria and Assessment Office, Cincinnati, OH.
Murty., A.S. and P.D. Hansen. 1983. Influence of the Carrier Solvent on Aquatic Toxicity
Tests In: K. Christiansen (ed.), Chemicals in the Environment, Proc. Int Symp., Lyngby,
Denmark, 1982, Publ. West Germany: 334-342. As cited in AQUIRE (,40t/atic Toxicity >
/nformation /?£trieval Database). 1995. Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
August 1995
-------
APPENDIX B Hexachlorobenzene - 14
i i mil
Nagy, K.A. 1987. Feild metabolism rate and food requirement scaling in mammals and
birds. Ecol. Mono. 57:111-128.
Nebeker, A.V., W.L. Griffis, CM. Wise, E. Hopkins, and J.A. Barbitta. 1989. Survival,
reproduction and bioconcentration in invertebrates and fish exposed to hexachlorobenzene.
Environmental Toxicology and Chemistry. 8:601-611.
Oliver, B.G. 1987. Biouptake of chlorinated hydrocarbons from laboratory-spiked and field
sediments by oligochaete worms. Environ. Sci. Technol. 21:785-790.
Oliver, B.G. and K.D. Nicol. 1982. Chlorobenzenes in Sediments, Water, and Selected Fish
from Lakes Superior, Huron, Erie, and Ontario. Environmental Science and Technology,
16:532:536. -
Oliver, B.G. and A.J. Niimi. 1983. Bioconcentration of Chlorobenzenes from water by
Rainbow Trout: Correlations with partition coefficients and environmental residues.
Environ. Sci. Technol. 17:287-291.
/
Oliver, B.G. and A.J. Niimi. 1988. Trophodynamic analysis of polychlorinated biphenyl
congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ.
Sci. Technol. 22:388-397. As cited in Stephan, 1993. Derivations of Proposed Human
Health and Wildlife Bioaccwnulation Factors for the Great Lakes Initiative.
PB93-154672. Environmental Research Labpratory, Office of Research and Development,
Duluth, MN.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
.Schrap, S.M. and A. Opperhuizen. 1990. Relationship between bioavailability and
hydrophobicity: Reduction of the uptake of organic chemicals by fish due to the sorption
on panicles. Environ. Toxicol. Chem., 9:715-724. As cited in Stephan, 1993.
Derivations of Proposed Human Health and Wildlife Bioaccwnulation Factors for the
Great Lakes Initiative. PB93-154672. Environmental Research Laboratory, Office of
Research and Development, Duluth, MN.
Schwetz, B.A.y J.M. Norris, R.J. Kociba, P.A. Keeler, R.F. Cornier, and P.J. Gehring. 1974.
Reproduction study in Japanese quail fed hexachlorobenzene for 90 days. Toxicol. appl.
Pharmacol. 30:255-265.
Stephan, C. E. 1993. Derivations of proposed human health and wildlife bioaccumulation
factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Hexachloroberuene - 15
Suter H, G.W. and J.B. Mabrey. 1994. lexicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
840R21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
v
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
Chlorinated Benzenes. EPA-440/5-80-028. Criteria and Standards Division, Washington,
DC.
U.S. EPA (Environmental Protection Agency. 1984. Health Effects Assessment for
Hexachlorobenzene. Environmental Criteria and Assessment Office, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. P338-179874,
Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990e Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (Environmental Protection Agency). 1993b. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Water, Office of Science and Technology, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993c. Technical Basis for Deriving
Sediment Quality Criteria for Nonionic Organic Contaminants for the Protection of
Benthic Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of
Water, Washington, DC.
Vieth, G.D., D.L. Defoe, and B.V. Bergstedt, 1979. Measuring and Estimating the
Bioconcentration Factor of Chemicals in Fish. J. Fish. Res. Board Can. 36(9): 1040-
1048. As cited in U.S. EPA (Environmental Protection Agency). 1993a. Derivations of
Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative. PB93-154672. Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
August 1995
-------
APPENDIX B Hexaehlorobenzene - 16
Vos, J.G., H.L. Van Der Maas, A. Musch and E. Ram. 1971. Toxicity of
Hexaehlorobenzene in Japanese Quail with Special Reference to Porphyria, Liver Damage,
Reproduction, and Tissue Residues. Toxicology and Applied Pharmacology, 18:944-957.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxlcity - Kexachlorobenzene
Cas No. 118-74-1
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hb^acl iloioUei izene
Specie*
pig
pig
rtiesus
monKey
Japanese '
quail
Japanese
quail
Japanese
quail
tat
Endpolnt
hepatic
NS
leto
rep
rep
rep
acute
Description
NOAEL
NOAEL .
AEL
NOAEL
LOAEL
AEL
LD50
Value
0.05
0.025
64
053
2.11
2.12
100
Units
mg/Kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
ppm
Exposure
Route (oral,
S.C., I.V., l.p.,
Infection)
oral
oral
oral
oral
oral
oral
oral
Exposure
Duratlon/TI
mlnq
90 days
gestation
and nursing
(-5-6
months)
-
60 days
90 days
90 days
90 days
96 days
Reference
den Torikelaar elal.,
1978-as cited U.S
EPA, 1984
Hansen el at.. 1978
as cited U.S. EPA,
1984
Bailey elal., 1980 as
cited in U.S. EPA,
1984
Vos et at., 1971
Voselal., 1971
Schwetz el at., 1974
Kilchin et al , 1982
Comments
No effects were observed at this level.
No effects were observed al this level.
s
Hypoaclivity arid lethargy, progressing to
ataxia and death as well as clinical signs ol
toxcity were observed al this dosage lovel
Egg production, percent halchability.
eggshell thickness, and volume of eggs
were not affected at this dosage level.
Egg volume and percent halchability were
significantly reduced al this dosage level.
Al this dosage level, a decreased survival
ot chicks hatched during the study and
increase in the liver weigh! of adult birds at
the end of the study was observed.
-------
Terrestrial Toxiciti .exachlorobenzene
Cas No. 118-74-1
Chemical Name
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Species
rat
rat
rat
rat
rat
rat
rat
Endoolnt
feto
toxlclty
ler
NS
NS
rep
rep
Description
NOEL
NOAEL
NOEL
NOAEL
NOAEL
AEL
NOAEL
Value
1.42
0.50
60.00
0.025
0.007
6.5
6.5
Unite
mg/kg-dav
mg/kg-day
mg/kg-dav
mg/kg-day
mg/kg-dav
mg/kg-day
mq/kg-dav
Exposure
Route (oral,
Infection)
oral
oral
gavage
oral
oral
oral
oral
Exposure
Duration/TI
mirm
1 .5 years
9-10
months
days 6-21
of gestation
15 weeks
29 weeks
135 days
149 days
Reference
'Grant et al., 1977
Grant etal.. 1974 as
cited In U.S. EPA,
1964
Khera, 1974
Kulper-Goodman et
al., 1977 as cited In
U.S. EPA, 1984
Bogeretal., 1979
U.S. EPA. 1984
Mendoza et al.. 1978
as died In U.S. EPA,
1984
Mendoza el al., 1979
as cited in U.S. EPA,
1984
Comments
There was no effect on the number of
liners, pups per liner, and total liner weight
at this dose level.
No effects were observed at this level.
There was no significant increase in the
incidence of unilateral or bilateral 14th rib in
lifters from dams exposed to this dose level
No effects were observed at this level. No
dose-response information.
No effects were observed at this level. No
dose response information
Reduced survival al weaning was observed
. at this dose level.
There were no differences in the number of
liners, average number of pups/liner,
average number of pups at birth or
qestation index.
-------
Terrestrial Biological Uptake . asures - Hexachlorobenzene
CAS No. 118-41-1
Chemical Name
hexachlorobenzene
hexachlorobenzene
hexachlorobenzene
Species
plants
oligochaete
worms
earthworm
B-factor
(BCF, BAF,
BMR
BCF
BCF
BCF
Value
0.026
24.000
4085
Measured
or
predicted
(m.p)
P
m
m
Units
(ug/g DW
plant)/(ug/g soil)
Ukg
ml/glipld
Reference
U.S. EPA; 19900
Oliver, 1987
Belfroid el al., 1993
Comments
worm BCF = chemical cone, (ng/lg) in
worm dry weight / pore water cone. (nq/L)
value based on total body weight
-------
Freshwater Toxiclty - Hexachlorobenzene
CasNo. 118-74-1,
Chemical Name
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Species
aquatic
organisms
Daphnla
maqna
Fathead
minnow
Endpolnt
chronic
rep
dvp. rep
Description
FCV
EC50
NOEC
Value
6
16
5
Units
ug/l
ug/L
uoA
Teat type
(static/ flow
through)
NS
NA
flow through
Exposure
Duration/
Timing
NS
14 day
2 to 68 days
Reference
U.S. EPA, 1980
AQUIRE. 1995
Nebekeretal ,
1989
Comments
Irom AWQC 'the available data
indicate thai HCB does not cause
significant adverse effects of
freshwater aquatic life at or below 6
ug/r
NS = No) specified
-------
Terrestrial Toxlcit\ .exachlorobenzene
Cas No. 118-74-1
Chemical Name
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
-
Species
ral
mallard
pheasant
EndooJnt
acute
acute
acute
Description
LD50
LO50
LD50
Value
140
71414
118.00
Unit*
ppm
mg/kg-
body wt.
mg/kg-
bodv wt.
Exposure
Route (oral,
8.C.. I.V., l.p.,
Inlectlon)
oral
NS
NS
Exposure
Duration/TI
mlnfl
96 days
NS
NS
Reference
Kltchin et al , 1982
U.S. EPA. 1993b
U.S. EPA. 1993b
Comments
NS = Not Specified
-------
Freshwater Biological Uptake Measures - Kexachlorobertzene
Cos No. 309-00-2
Chemical name
hexachlorobenzene
hexachlorobenz ene
hexachlorobenzene
hexachlorobenzene
hexachlofobenzene
hexachlorobenzene
hexachlorobenzene
Species
fish
fish
fish
fish
fish
rainbow troul
salmon
B-factor
(BCF, BAF,
BMF)
BCF
BCF
BCF
BCF
BCF
BAF
BAF
Value
5702
1833
5805
2116
5400
68987
23030
Measured or
predicted (m,p)
m
m
m
m
m
m
m
Units (17kg,
NS. other)
NS
NS
NS
NS
NS
NS
NS
Reference
Kosianelal . 1981 as
cited in Slephan, 1993
Oliver and Niimi, 1983
Carlson and Koslan,
1987 as cited in
Slephan, 1993
Nebekerelal , 1989
Schrap and
Opperhuizen. 1990 as
cited in Slephan, 1993
Oliver and Niiml, 1983
Oliver and Niimi, 1988
as cited in Slephan,
1993
Comments
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
Normalized to 1.0% lipid
Normalized, to 1.0% lipid
Normalized to 1.0% lipid
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
NS = Not specified
-------
Freshwater Biological Uptake, jasures - Hexachlorobenzene
Cos No. 309-00-2
Chemical name
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
Hexachlorobenzene
hexachlorobenzene
hexachlorobenzene
hexachlorobenzene
Species
fathead
minnow (whole
body)
rainbow trout
rainbow trout
rainbow trout
talhead
minnow
fathead
minnow
fish
fish
fish
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
22000
7800
6202
4645
12600
39000
1668
2434
2900
Measured or
predicted (m,p)
NS
NS
m
m
m
m
P
m
m
Units (L/kg,
NS, other)
NS
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1980
U.S. EPA, 1980
Murty and Hansen,
1983 as cited in
AQUIRE. 1995
Murty and Hansen,
1983 as cited In
AQUIRE. 1995
Nebekeretal.. 1989
Koslanelal , 1981 as
cited In AQUIRE. 199S
Stephan. 1993
Veithetal., 1979 as
cited In Stephan. 1993
Konemann and van
Leeuwen. 1979, 1980
as cited in Stephan,
1993
Comments
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
-------
APPENDIX B Hexachlorocyclopentadiene - 1
Toxicological Profile for Selected Ecological Receptors
Hexachlorocyclopentadiene
Cas No.: 77-47-4
Summary: This profile on hexachlorocyclopentadiene summarizes the lexicological
benchmarks and biological uptake measures (i.e., bioconcentration, bibaccumulation, and
biomagnification factors) for birds, mammals, daphnids and fish, aquatic plants and benthic
organisms representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
K<,w between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including .
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No subchronic or chronic studies were identified for mammalian wildlife species,
exposure to hexachlorocyclopentadiene. However, several studies have documented
subchronic and critical-lifetime exposure of hexachlorocyclopentadiene (HEX) to laboratory
animals. Naishlein and Lisovskaya (1965 as cited in U.S. EPA, 1984) reported a subchronic
NOEL of 0.2 mg/kg for rats exposed orally lo HEX over 216 days. Abdo et -al. (1984) dosed
rats by gavage with 0, 10, 19, 38, 75, or 150 mg HEX/kg body weight at 5 days/week for 13
August 1995
-------
APPENDIX B Hexachlorocyclopentadiene - 2
weeks. Abdo et al. (1984) observed lesions in the forestomach at dose levels of 19 mg/kg
and dose-related depression in mean body weight was noticed in male rats at dose levels of
38 mg/kg and above. Murray et al. (1980) attempted to determine the teratogenic potential of
hexachlorocyclopentadiene in mice and New Zealand rabbits. Mice were given 0, 5, 25, or
75 mg HEX/kg/day by gavage from days 6 to 15 of gestation; rabbits were given the same
dose levels of HEX by gavage from days 6 to 18 of gestation. No maternal toxicity or
teratogenic effects were observed in mice given HEX. Unlike mice, rabbits given 75 mg/kg-
day experienced diarrhea, weight loss, and mortality. Teratogenic effects at this dose were
limited to only one minor skeletal variation in the offspring. At the 75 mg/kg dose, thirteen
ribs were seen more frequently among the fetuses of the rabbits dosed (the normal number of
pairs of ribs in the rabbit is 12 or 13). A NOAEL of 25 mg/kg-day and a LOAEL of
75mg/kg-day were inferred from the data set.
The study by Murray was selected for developing a mammalian benchmark value because: (1)
the dose was orally administered, (2) it contained sufficient dose-response data and (3) it
reported endpoints that may impair population sustainability. The study by Naishtein and
Lisovskaya (1965 as cited in U.S. EPA, 1984) was not used to develop benchmarks because
the dosing-range and lexicological endpoints could not be determined. Similarly, while Abdo
et al. (1984) reported a low NOAEL for rats, this study was not used because reproductive or
developmental endpoints that could reasonably cause adverse effects to populations in the
wild were not assessed.
The NOAEL value from Murray et al., (1980) was then scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
bw 4
Benchmark,, = NOAEL, x
'
where NOAEL, is the NOAEL (or LOAEL/ 10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Murray et al. (1980) study documented reproductive effects from hexachlorocyclopentadiene
expoure to female rabbits, the representative body weights of the female species were used in
the scaling algorithm to obtain lexicological benchmarks.
August 1995
-------
APPENDIX B Hexachlorocyclopentadiene - 3
Data were available on reproductive, developmental, growth and survival endpoints for
heptachlor exposure. In addition, the data set contained acute and chronic toxicity studies that
were conducted during sensitive life stages. Given the data set for
hexachlorocyclopentadiene, the benchmarks developed from Witherup et al. (1955) were
categorized as adequate.
Birds: Adequate toxicity studies documenting avain exposure to hexachlorocyclopentadiene
were not identified and therefore, no benchmarks were developed.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for hexachlorocyclopentadiene. Therefore, the Tier II method described in Section
4.3.5 was used to estimate a Secondary Chronic Value (SCV) of 6.9E - 03 mg/L as reported
in AQUIRE. Because the benchmark is based on a SCV and there were no lower toxicity
values in the data set, it was categorized as interim.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascualr aquatic plants (e.g., duckweed) or (2) an effective concentration (ECxx) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). No
Adequate data sufficient for the development of benchmark values were not identified and
therefore, benchmarks were not derived.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or Secondary Chronic Value (SCV), along with the fraction of organic carbon and the
octanol-carbon partition coefficient (K,,,.) to determine protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in the
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. Because no FCV was available, a SCV value of 50.3 mg hexachlorocyclopentadiene
/kg organic carbon was used to calculate an EQp value. Assuming a mass fraction of organic
carbon for the sediment (f^) of 0.05, the benchmark for the benthic community is 2.5 mg/kg
sediment. Since the EQp number was based on a SCV, the sediment benchmark was
categorized as interim.
August 1995
-------
APPENDIX B
Hexachlorocyclopentadiene - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
RapraMntthm
SpodoQ
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
V«hw* mgfte-
day
38.9 (a)
21. 6 (a)
ID
ID
ID
ID
ID
ID
ID
ID
Study
SpMta
rabbit
rabbit
-
-
-
-
-
-
-
-
Eftoet
« -
. rep
rep
-
-
-
-
-
-
-
'- .
Study Vah»
mg/kg-diy
25
25
-
-
-
-
-
-
-
> '^ -»mtmittntu
UMCflpWMt
NOAEL
NOAEL
'
'
-
-
-
-
1
SF
-
-
-
/
-
-
-
-
-
Original Sourc*
Murray et at., 1979
Murray et al., 1979
-
.
-
- '
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August1995
-------
APPENDIX B
Hexachlorocyclopentadiene - 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
R«prM«ntattv«
Spacfes
fish and aquatic
invertebrates
aquatic plants
benthic community
otndiiwfK
VtilM'
Rtyl*.
7.5E-04 (i)
No data
2.5E+00 (i)
Study
SfMCiM
aquatic
organisms
aquatic
organisms
Qofcnptfoit
scv
'
SCVx K.
Original Sown
AQUIRE
.
AQUIRE
II.
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value
was an order ofmagnitude or more above the NEL or LEU for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Mammals: Becasue of the lack of additional mammalian toxicity studies, the same surrogate-
species study (Murray et al., 1980) was used to derive the hexachlorocyclopentadiene
toxicological benchmark for mammalian species representing the general terrestrial ecosystem.
The study value was scaled for species in the terrestrial ecosystem using the cross-species
scaling algorithm adapted from Opresko et al. (1994). Since Murray et al. (1980) documented
reproductive effects from hexachlorocyclopentadiene exposure to female rabbits,
representative body weights of the female species were used in the scaling algorithm to obtain
toxicological benchmarks. Based on the data set for hexachlorocyclopentadiene, the
benchmarks developed for the terrestrial ecosystem were categorized as provisional.
Birds: Adequate toxicity data with which to derive a benchmark protective of the avian
community were not identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
August 1995
-------
APPENDIX B Hexachlorocyclopentadiene - 6
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for hexachlorocyclopentadiene and, as a
result, a benchmark could not be developed.
Soil Community. Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Hexachlorocyclopentadiene - 7
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RopreMntBtive
Sp*cto»
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox .
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobowhite
American robin
American
woodcock
plants
soil comrhnity
Bmctimarfc
VtflM*
mg/kg
-------
APPENDIX B Hexachlorocyclopentadiene - 8
III. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: general fish
(BCF only), aquatic invertebrates, earthworms, other soil invertebrates, terrestrial
invertebrates, and plants. Each value is identified as whole-boy or lipid-based and, for the
generic aquatic ecosystems, the biological uptake factors are designated with a "d" if the
value reflects dissolved water concentrations, and a "t" if the value reflects total surface water
concentrations. For organic chemicals with log K^ values below 4r bioconcentration factors
(BCFs) in fish were always assumed to refer to dissolved water concentrations (i.e., dissolved
water concentration equals total water concentration). The following discussion describes the
rationale for selecting the biological uptake factors and provides the context for interpreting
the biological uptake values presented in Table 4,
Although the log Kow value for HEX (4.9) suggests that this chemical will bioaccumulate
appreciably in the aquatic ecosystem, studies have demonstrated that HEX is readily
metabolized by fish. Consequently, the measured value reported in Stephan (1993) was
selected as the bioconcentration factor for HEX. It should be noted that the BCF in Stephan
(1993) was converted to a lipid-based bioconcentration factor (i.e., BCF/) using the lipid
fraction reported in the study.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemical assuming that the partitioning to tissue is
dominated by lipids. For hydrophobic organic constituents, the bioconcentration factor for
plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Mexachlorocyclopentadiene - 9
Table 4. Biological Uptake Properties
ecological
recaptor
fish
terrestrial
vertebrates
terrestrial
invertebrates
' earthworms
plants
BCF, BAF, or
BSAF
BCF
BAF
BAF
BAF
BAF
llpld-basad or
wttoJa-body
lipid
whole-body
whole-body
whole-body
whole-plant
valua
400 (t)
3.2 E-04
3.0 E-04
- 2.4 E-03
1.1 £-01
aouroe
measured; metabolized by fish
(Stephan, 1.993)
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated basecl on beef
biotransfer' ratio with 2,3,7,8-
TCDD
estimated basecl on beef
biotransfer ratio with 2,3,7,8-
TCDD
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = Insufficient Data
August 1995
-------
APPENDIX B ' Hexachlorocyclopentadiene -10
References
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Toxicity of Hexachlorocyclopentadiene: Subchronic (13-week) Administration by Gavage
to F344 Rats and B6C3F1 Mice. Journal of Applied Toxicology, Vol. 4, No. 2.
Applegate, V.C. efal. 1957. Toxicity of 4,346 Chemicals to Larval Lamprey and Fishes.
U.S. Fish Wild. Serv. Spec. Rep. -- Fish. No. 207. Washington, DC., U.S. Dept. of Inter.
As cited in U.S. Environmental Protection Agency, 1980. Ambient Water Quality Criteria
for Hexachlorocyclopentadiene^ Criteria and Standards Division, EPA-440/5-80-055.
AOUIRE (AOUatic Toxicity Information REtrieval Database). 1995 Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency.
Duluth, MN.
Buccafusco, R.J. and G.A. LeBlanc. 1977. Acute Toxicity of Hexachlorocyclopentadiene to
Bluegill (Lepomis macrochirus). Channel Catfish (Ictalurus punctatus). Fathead Minnow
(Pimephales promelas). and the Water Flea (Daphnia magna). Unpublished report
prepared for Velsicol Chemical Corporation, Chicago, IL. As cited in U.S. Environmental
Protection Agency. 1988. Health and Environmental Effects Document for Chlorinated
Cyclopentadienes. Environmental Criteria and Assessment Office, Office of Health and
Environmental Assessment. ECAO-CIN-G029.
EG & G, Bionomics, 1977. Acute Toxicity of Hexachlorocyclopentadiene to bluegill
(Lepomis macrochirus), channel catfish (Ictalurus punctatus), fathead minnow (Pimephales
promelas) and the water flea (Daphnia magna). Toxicity Test Report submitted to
Velsicol Chemical Corporation, Chicago, Illinois. As cited in U.S. Environmental
Protection Agency, 1980. Ambient Water Quality Criteria for Hexachlorocyclopentadiene.
Criteria and Standards Division, EPA-440/5-80-055.
Freitag, D., L. Ballhorn, H. Geyer, and F. Korte. 1985. Environmental Hazard Profile of
Organic Chemicals. Chemosphere 14(10): 1589-1616.
Henderson, D. 1956. Bioassay investigations for International Joint Commission. Hooker
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A. Taft Sanitary Engineering Center, Cinn., Ohio. As cited in U.S. Environmental
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Criteria and Standards Division, EPA-440/5-80-055.
August 1995
-------
APPENDIX B , Hexachlorocyclopentadiene - 11
IRDC (International Research and Development Corp.) 1978. Pilot Teratology Study in Rats.
Submitted to Velsicol Chemical Corp. (Unpublished). As cited in U.S. Environmental
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Criteria and Stanards Division, lOOp.
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Kuhn, R., M. Pattard, K. Pernak, and A. Winter. '1989. Results of the Harmful Effects of
Water Pollutants to Daphnia magna in the 21 Day Reproduction Test. Water Res., 23(4):
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Environmental Research Laboratory, Office ;of Research and Development, U.S.
Environmental Protection Agency. Duluth, MN.
Lu, Po-Yung, Robert L. Metcalf, Asha S. Hirwe, and John W. Williams. 1975. Evaluation
of Environmental Distribution and Fate of Hexachlorocyclopentadiene, Chlordene,
Heptachlor, and Heptachlor Epoxide in a Laboratory Model Ecosystem. J. Agric. Food
Chem., Vol. 23, No. 5.
Murray, F.J., B.A. Schwetz, M.F. Balmer, and R.E. Staples. 1980. Teratogenic Potential of
Hexachlorocyclopentadiene in Mice and Rabbits. Toxicology and Applied Pharmacology,
53, 497-500.
Nagy, K. A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecoi Mono. 57:11-128.
Naishtein, S.Y. and E.V. Lisovskaya. 1965. Maximum Permissible Concentration of
Hexachlorocyclopentadiene in water bodies. Hyg. Sanit., 30: 177-182. (Trans. Rus.). As
cited in U.S. Environmental Protection Agency, 1984. Health Assessment Document for
Hexachlorocyclopentadiene. Office of Health and Environmental Assessment.
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Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife
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Laboratoy, Oak Ridge, Tennessee.
August 1995
-------
APPENDIX B Hexachlorocyclopentadiene 12
Root, M.S., D.E. Rodwell, and E.I. Goldenthal. Teratogenic Potential of
Hexachlorocyclopentadiene in Rats. Velsicol Chemical Corporation. As cited in The
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SRI (Southern Research Institute). 1980a. Acute Toxicity Report on
Hexachlorocyclopentadiene (C53607) in Fischer-344 Rats and B6C3F1 Mice.
Unplublished Report for NTP. 44p. As cited in U.S. Environmental Protection Agency,
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Environmental Assessment. Washington/DC. EPA-600/8-84-001F..
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Hexachlorocyclopentadiene (C53607) in Fischer-344 Rats and B6C3F1 Mice.
Unplublished Report for NTP. 33p. As cited in U.S. Environmental Protection Agency,
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Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN. .
Suter II, G; W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
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Thomann, R. V. 1989. Bioaccuraulation model of organic chemical distribution in aquatic
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Environmental Toxicology and Chemistry. 11:615-629.
August 1995
-------
APPENDIX B Hexachlorocyclopentadiene 13
Treon, J.F., P.P. Cleveland and J. Cappel. 1955. The Toxicity of
Hexachlorocyclopentadiene. Arch. Ind. Health, 11: 459-472. As cited in U.S.
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Hexachlorocyclopentadiene. Criteria and Standards Div. EPA-440/5-80-055.
U.S. Environmental Protection Agency. 1982. Symposium: Carcinogenic Polynuclear
Aromatic Hydrocarbons in the Marine Environment held at Pensacola Beach, Florida on
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for Chlorinated Cyclopentadienes. Office of Solid Waste and Emergency Response,
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August 1995
-------
APPENDIX B Hexachlorocyclopentadiene -14
U.S. Environmental Protection Agency. 1991'. Health Assessment Document for
Hexachlorocyclopentadiene. Office of Health and Environmental Assessment,
Washington, D.C.
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Bioconcentration Factor of Chemicals in Fish. J. Fish Res. Board Can. 26, 1040-1048. As
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Vilkas, A.G. 1977. The Acute Toxiciry of Hexachlorocyclopentadiene to the Water Flea,
Daphnia magna Straus. Union Carbide Environmental Services. Prepared for Velsicol
Chemical Corp., Chicago, IL. As cited in U.S. Environmental Protection Agency. 1988.
Health and Environmental Effects Document for Chlorinated Cyclopentadienes.
Environmental Criteria and Assessment Office, Office of Health and Environmental
Assessment. ECAO-CIN-G029.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Biological Uptake Measures - Kexachioi0cyctcpe>ntQdier»«
Cos No.: 77-47-4
Chemical Norn*
hexachlorocydopentadiene
Specie*
plant
B-tactor
(BCF.BAf,
BMF)
BCF
Value
5.60E-02
Measured
or
Ptedteled
(m,p)
P
Untts
NS
Reference
U.S. EPA. 1990
Comments
Plant uptake from soil pertains to
leafy vegetables.
NS = not specified
-------
Freshwater Biological Uptak* Me^^jres - H«xcrchlprocyclop«ntacU«n«
Cos No.: 77-47-4
Chemical Name
hexachlorocydopentadiene
hexochlofocvdopentadiene
hexachlorocydopentadiene
hexachlorocydopentadiene
hexachlorocydopentadiene
hexachlorocydopentadiene
hexachlorocydopentadiene
Specie*
fathead minnow
fathead minnow
Golden ide
fish
fish
mosquito fish
algae to snail .
B-fador
(BCF. BAF,
BMF)
BCF
BCF
BCF
BCF
BCF
BAF
BMF
Value
-------
TerrestrksS Toxictty - Hexcechiorocyclopentadlene
Cos. No.: 77-47-4
Chemical Name
hexachlorocyclopentadiene
hexachlorocyclopentadiene
hexachlorocyclopentadiene
hexachlorocyclopentadiene
hexachlorocyclopentadiene
Species
mouse
mouse
mouse
CM mice
New Zealand
rabbits
EEndpoint
subchronic
subchronic
subchronic
sub chronic.
terat
sub chronic,
terat
Description
NOAEL
FEL
LOEL
NOAEL
LOAEL
Value
19
SO
38
75
75
Unto
mg/kg-body wt
ma/kg
mg/kg-body wt.
mg/kg-day
mg/kg-day
Exposure
Route (oral.
B.C.. I.V., i.p.,
InjactJon)
oral
oral
gavage
gavage
gavage
Exposure
Duration Aiming
5 days/week (or
13 weeks
12 day
exposure
duration .
90-day feeding
study
days 6 15
days of .
gestation
days 6- 18
days of
gestation
Reference
Abdoetal.. 1984
SRI. 1980aas
cited in U.S. EPA.
1984
Abdo et al., 1984
Murray et al.,
1980
Murray et al..
1980
Comments
Dose levels of 0. 19, 38. 75, 150. and
300mg/kg. Dose was 94.3% - 97.4%.
HEX In com oil. Dose levels of 38 mg/k(
and above caused lesions In the
forestomach
Study doses were 10.19.38,75,150. and
300mg/kg. Lesions of forestomach In
female rats at 38 mg/kg.
No maternal toxidty. no embryotoxlc nor
teratogenic effects observed In offspring
Sub-chronic maternal toxidty Included
diarrhea, weight loss, and mortality.
Teratogenic effect - increase of minor
skeletal variation In offspring.
-------
Terrestrial Toxictty - He^uchlorocyclopentadien*
Cos. No.: 77-47-4
Chemical Name
hexachlorocyclopentadiene
hexachkxocydopentadiene
he xachlorocY dopentadiene
he xachtorocy dopentadiene
hexachkxocydopentadiene
hexachlorocyclopentadiene
hexachkxocydopentadiene
Species
rat
rat
rat
rat
rat
rat
rat
Eindpoint
subchronic
subchronic
subchronic
subchronic
subchronic
subchronic
terat
Description
NOEL
NOAEL
NOEL
NOAEL
LOEL
LOAEL
NOAEL
Value
0.2
10
10
25
19
30
30
Unto
mg/kg
mg/ka-booV wt.
mg/kg
mg/kg
mg/kg-bodywt
mg/Ka-day
mg/kg-day
Exposure
Route (oral.
S.C.. I.V., l.p ,
Hectton)
oral
oral
oral
oral
gavage
gavage
gavage
-Expoeur*
Duration /timing
216 day
exposure
duration
5 days/Week tor
13 weeks
10 day
exposure
duration
12 day
exposure
duration
90-day feeding
study
days 6-15 of
gestation
days 6-15 days
Reference
Natshtein and
Lteovskaya, 1965
as cited in US
EPA, 1984
Abdoetal. 1984
IROC, 1978 as
cited In U.S. EPA.
1984
SRI. 1980bas
cited In US. EPA.
1984
Abdoetal.. 1984
International
Research and
Development
Corp., 1978 as
cited In U.S. EPA.
1980
RootetaL.as
cited In The
lexicologist,
March 1983.
Convnonto
i
Dose levels of 0, 10. 19. 38, 75. and 150
mg/kg Dose was 94 3% - 97.4% HEX Ir
corn oil. Dose levels of 19 mg/Kg and
above caused lesions in the
forestomach. Dose-related depression
in mean body weight was noticed In male
rats at 38 mg/Kfl
Study was used to extrapolate human
RfD. study doses were 10.19.38.75.150.
and 300 mg/kg. Lesions of forestomach
In female rats at 19 mg/kg.
Study doses were 3.10.30.100 and 300
mg/kg-d. Rats receiving 30 mg/kg-d
showed reduced body weight gains and
staining of the anogeniteJ area
Dosage levels of 0 3 10 30mg/kg d No
embryotoxicrty nor teratogenicity was
observed at highest dose.
-------
freshwater Toxteity - HexachSorocycSopentadiens
Cos. No.: 77-47-4
hexachlofocyclopentadiene
hexachlofocydopentadiene
hexachlofocydopentadiene
hexachlorocydopentadiene
hexachlordcydopentadiene
"»exachlorocydopentadiene
hexachlofocydopentadiene
hexachlofocydopentadiene
hexach!GfGcydopeniadi6r.s
fathead minnow
fathead minnow
fathead minnow
Daphnia magna
Daphnia magna
Daphnia magna
channel catfish
bluegill
blucgi!!
chronic
acute
acute
acute
acute
acute
acute
acute
bshav
CV
NOEC
NOEC
NOEC
NOEC
NOEC
NOEC
NOEC
PEL
5.2
3.7
87
32
18
9
56
65
1.000
ua/l
ua/l
ua/l
ua/l
ug/l
ug/l
UQ/I
ug/l
ug/l
flow
through
flow-
through
static
static
static
MS
static
static
NS
4 day
30-day
NS
48-hour
48-hour
21 day
NS
NS
24-hour
Spehar et al..
1979
Spehar et al..
1979
Buccafusco and
LeBlanc. 1977 as
cited in U.S. EPA.
1988
Vilkas. 1977 as
cited in EPA HEED
1988
Buccafusco and
LeBlanc, 1977 as
cited in U.S. EPA.
1988
Kuhnetal.. 1989
as dted in
AQUIRE, 1994
Buccafusco and
LeBlanc. 1977 as
cited in U.S. EPA.
1988
Buccafusco and
LeBlanc. 1977 as
cited In U.S. EPA.
1988
Applegate et al..
1957 as dted h
U.S. EPA. 1980
CV was calculated by
geomean of NOEC and
LOEC (CV presented in
AWQC document)
water temp « 25 C. soft
water. NOEC based on
lethal toxidty and growth
data.
water temp = 22 C. soft
water
water temp = 1 7 C. soft
water
water temp = 22 C. soft
water
water temp = 22 C. soft
water
water temp = 22 C. soft '
water
distress was obsevered in
1/2 hour
-------
Freshwater Toxtetty - H»~achlorocyclop«ntaclien«
Cos. No.: 77-47-4
Chemical Name
lexachlorocyclopentadiene
hexachlofocydopentadiene
hexachlofocyclopentadiene
hexachlofocydopentadiene
hexachlofocydopentadiene
vexachlococydopentadiene
hexachlofocydopentadiene
hexachlofocydopentadiene
hexacNotocydopentadiene
hexachlofocydopentadiene
spades
daphnia magna
daphnia magna
fathead minnow
fathead minnow
fathead minnow
fathead minnow
fathead minnow
fathead minnow
channel catfish
bluegill
Type of Effect
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
Description
LC50/EC60
LC50/EC60
LC60
LC50
LC50
LC60
LC60
LC50
LC50
LC50
Value
39
52
7
6.7
180
104
-
78
69
97
130
Units
ufl/l
ua/l
ua/l
ua/l
UQ/I
ua/l
ug/l
ua/l
ua/l
ua/l
Test Type
(it otto/ flow
through)
static
static
flow-
through
flow-
thfouah
static
static
static
static
static
static
Exposure
Duration/
Timing
MS
MS
96-hour
30-day
4 day
4 day
4 day
4 day
4 day
4 day
Reference
EG&G Bionomics.
1977 as cited in
U.S.EPA. 1980
Union Carbide
Environmental
Services. 1977 as
cited in U.S.EPA.
1980
Spehar et al..
1979
Spehar et al..
1979
EG&G Bionomics.
1977 as cited In
U.S.EPA. 1980
Henderson. 1956
as dted in
U.S.EPA. 1980
Henderson. 1956
as cited in
U.S.EPA. 1980
Henderson. 1956
as cited in
U.S.EPA. 1980
EG&G Bionomics.
1977 as dted In
U.S.EPA. 1980
EG&G Bionomics.
1977 as cited in
U S.EPA. 1980
Comments
study duration was not
specified
study duration was not
specified
-------
APPENDIX B Hexachlorophene -1
Toxicological Profile for Selected Ecological Receptors
Hexachlorophene
Cas No.: 70-30-4
\
Summary: This profile on hexachlorophene summarizes the toxicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulatioh, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from exiting regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also '
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire toxicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the information presented in the
technical support document for the "Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rational behind toxicological benchmarks used to derive protective
media concentrations (C^) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies reporting adequate dose-response data
for mammalian wildlife were identified. However, several toxicological studies involving
hexachlorophene exposure to mammals have been performed with laboratory rats. Gaines et
al. (1973) conducted a study on rats fed 0, 20 and 100 ppm hexachlorophene. Although no
August 1995
-------
APPENDIX B Hexachlorophene - 2
reproductive effects were observed at levels of 20 ppm, a reduction in pup survival rate
occurred at dietary levels of 100 ppm. Since no information was provided on daily food
consumption or body weight, conversion from mg/kg-diet to mg/kg-day required the use of an
allometric equation:
Food consumption = 0.056CVV066") where W is body weight in kg (Nagy, 1987).
Assuming a body weight of 0.330 kg, a NOAEL of 1.63 mg/kg-day and a LOAEL of 8.18
mg/kg-day could be inferred for reproductive effects. In other studies, Kennedy et al. (1976)
observed that hexachlorophene administered to rats at 10 mg/kg-day reduced the survival of
12 and 21 day-old offspring. Kennedy et al., 1975b demonstrated that rabbit dams had
reduced rates of body weight gain and possibly increased resorption at 6 mg/kg-day (6-18d)
gestion. Hexachlorophene administered orally to mature male dogs and rats at >3 mg/kg-day
for 9 weeks resulted in a rapid, but transitory reduction in spermatogenesis and degeneration
of the germinal epithelium of the testis (Thorpe, 1967; James et al., 1980; as cited in U.S.
EPA, 1986). A LOAEL of 3 mg/kg-day was inferred from these studies.
Selection of the Gaines et al. (1973) study for the derivation of protective benchmarks was
based on demonstrated reproductive effects at the lowest dose of all the studies reviewed. The
values in the Thorpe (1967) and James et al. (1980) (as cited in U.S. EPA, 1986) studies were
not used in deriving benchmark values as the toxicity data were insufficient to infer
differences in sensitivity between surrogate species and representative species. Furthermore,
the values from the other studies would require the use of a NOAEL extrapolation safety
factor of 10 to account for differences between the original LOAEL reproductive endpoint
and a NOAEL reproductive endpoint.
The NOAEL value from Gaines et al. (1973) was used to extrapolate a benchmark value
representative of a freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994):
bw
Benchmark^ = NOAEL, x I '.
bWw
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
August 1995
-------
APPENDIX B Hexachlorophene - 3
Gaines et al. (1973) study documented reproductive effects from hexachlorophene exposure to
female mice, the representative body weights of female species were used in the scaling
algorithm to obtain toxicological benchmarks.
The Gaines et al. (1973) study was selected for derivation of benchmark values because it
provided data on reproductive, developmental, growth and survival endpoints for
hexchlorophene exposure. The data set on hexachlorophene also contains information on
acute and chronic toxicity studies conducted during sensistive life stages. In addition, it
contained a study value for neurological endpoints (Robinson et al., as cited in HEPA, 1986)
that was approximately an order of magnitude lower than the benchmark value. Based on the
data set for hexachlorophene, the benchmarks developed from the Gaines et al. (1973) study
were categorized as adequate, with a "*" to indicate that some adverse effects have been
observed at the benchmark level. .
Birds: No studies were identified concerning hexachlorophene toxicity in avian species and
therefore, no benchmarks were developed.
Fish and aquatic invertebrates: No AWQC, Final Chronic Value (FCV) or Secondary
Chronic Value (SCV) data were available on hexachlorophene for the development of
protective benchmarks for the fish and aquatic community.
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concnetration (NOEC) or a lowest observed effects concnetration (LOEC) for vascular aquatic
plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g.,. Selenastrum capricomutum). Adequate data
for the development of benchmarks were not identified in Suter and Mabrey (1994) or in
AQUIRE.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanbl-carbon partition coefficient (K,,,.) to determine a protective sediment concnetration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community from harmful chemical exposure. No
FCV or Secondary Chronic Value (SCV) data were identified for hexachlorpohene and,
therefore, no benchmark was developed.
August 1995
-------
APPENDIX B
Hexachlorophene - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Representative
Specie*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark Value*
mp/kg-day
1.35 (a')
0.75 (a*)
ID
ID
ID
ID
ID
ID
ID
ID
Study
Species
rat
rat
-
-
-
-
.
-
-
-
Effect
rep
rep
.
-
-
-
-
-
-
Study Value
mg/kg-day
1.53
1.53
'
-
-
-
-
-
-
-
Description
NOAEL
NOAEL
-
- '
-
-
-
-
SF
-
'
-
-
-
-
-
.
-
Original Source
Gaines et al, 1973
Gaineset al, 1973
-
-
-
-
-
-
-
-
Benchmark Category, a = adequate, p = provisional, i = interim; a
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
indicates mat the benchmark value was an order of.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Representative
Species
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
Value
mg/L
ID
No data
ID
Study Species
-
-
p\..M ! tMiui
uescnpoon
-
, -
-
Original Source
-
.
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a '*' indicates that the benchmark value was an order
of magnitude or more above the 'NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Hexachlorophene - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Because of the lack of additional mammalian toxicity studies, the same surrogate
species study (Gaines et ah, 1973) was used to calculate benchmark values for mammalian
species representing the general terrestrial ecosystem. The NOAEL from the Gaines et al. .
(1973) study was scaled for species in the terrestrial ecosystem using the cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the Gaines et al. (1973) study
documented reproductive effects from hexachlorophene exposure to female rates, the
representative body weights for female species were used in the scaling algorithm to obtain
the lexicological benchmarks. Based on the data set for hexachlorophene, the benchmarks
developed from the Gaines et al. (1973) study were categorized as adequate.
Birds: Adequate toxicity data on hexachlorophene with which to derive a benchmark
protective of the avian community were not identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for hexachlorophene and, as a result, a
benchmark could not be developed.
Soil community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Hexachlorophene 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Representative
Species
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plants
soil community
Benchmark Value*
mgrtcg-day
3.33 (a*)
3.42 (a*)
2.78 (a')
. 1.17 (a*)
0.87 (a')
0.84 (a*)
ID
ID
ID
ID
ID
ID
ID
Study
Species
rat
rat
rat
rat
rat
rat
-
-
-
-
- '
-
-
Effect
rep
rep
rep
rep
. rep
rep
-
-
-
-
- '
-
-
Study
Value
mg/kg-day
1.53
1.53
1.53
1.53
1.53
1.53
-
-
-
-
-
-
-
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL.
'
-
-
-
'
-
-
SF
-
-
-
-'
-
-
-
-
-
-
-
Original Source
Gaines et al., 1973
Gaines et al., 1973
Gaines et al.. 1973
Gaines etal., 1973
Gaines et al., 1973
Gaines etal., 1973
'
. -
-
'
-
-
-
Benchmark Category, a = adequate, p = provisional, i = interim; a indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
III. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcehtrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. Each value is identified as whole-body or lipid-based and,
for the generic aquatic ecosystems, the biological uptake factors are designated with a "d" if
the value reflects dissolved water concentrations, and a "t" if the value reflects total surface
August 1995
-------
APPENDIX B Hexachlorophene 7
water concentrations. For organic chemicals with log K^, values below 4, bioconcentration
factors (BCFs) in fish were always assumed to refer to dissolved water concentrations (i.e.,
dissolved water concentration equals total water concentration). For organic chemicals with
log K,,w values above 4, the BCFs were assumed to refer to total water concentrations. The
brief discussion preceding Table 4 describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values.
No data were identified oh the bioconcentration/bioaccumulation of hexachloirophene in fish.
The bioaccumulation factor for terrestrial vertebrates and invertebrates was estimated as
described in Section 5.3.5.2.3. Briefly, the extrapolation method is applied to hydrophobic
organic chemicals assuming that the partitioning to tissue is dominated by lipids. For
hydrophobic organic constituents, the bioconcentration factor for plants was estimated as
described in Section 6.6.1 for above ground leafy vegetables and forage grasses. The BCF is
based on rout-to-leaf translocation, direct depostion on leaves and grasses, and uptake into the
plant through air diffusion.
August 1995
-------
APPENDIX B
Hexachlorophene - 8
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
-
'-
-
-
-
BAF
BCF
BCF
BCF
llpM-bamd or
...*- i^. » -i
WHOM OOGy
-
. .
-
-
-' . .
-
whole-body
whole-body
whole- body
whole-plant
value
No data
No data
ID
10
ID
ID
0.32
0.31
2.5
2.0E-03
source
.
-
-
- .
-
calc
calc
calc
U.S. EPA, 1990e
ID = Insufficient Data
August 1995
-------
APPENDIX B Hexachlorophene - 9
References
AQUIRE (AQUatic Toxicity information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U. S. Environmental Protection Agency,
Duluth, MN, June 1995.
\
Gaines, T.B., R. D. Kinbrough, and R.E. Linder. 1973. Theoral and dermal toxicity of
hexacholophene in rats. Toxicology and applied pharmacology. 25: 332 -343.
Gellert, R. J., C. A. Wallace, E.M. Wiesmeier and R. M. Shuman. 1978. Topical exposure
of Neonates to hexachlorphene: longstanding effects on mating behavior and prostatic
development in rats. Toxicology and Applied Pharmacology. 43: 339 - 349.
IRDC (International Research and Development Corporation). 1974. Unpublished study.
IRDC# 281-013, 7/9/74. As cited in U. S. EPA (Environmental Protection Agency).
1986 Health and Environmental Effects Profile for Hexachlorophene. March 1986.
IRDC (International Research and Development Corporation). 1979. Unpublished study.
IRDC# 380-002, 8/9/79. As cited in U. S. EPA (Environmental Protection Agency).
1986 Health and Environmental Effects Profile for Hexachlorophene. March 1986.
James, R.W. R. Hey wood, and D: Crook. ;1980. Quatnitiative aspects of spermatogenesis
in rats and dogs after repeated hexachlorophene treatment. Toxicol. Lett. 5(6): 405-412.
Kennedy, G. L. Jr., S.H. Smith, M. L. Keplinger, and J.C. Calandra. 1975a. Effect of
hexachlorophene on reproduction in rats. J. Agric. Food Chem., Vol. 23, No. 5. :
866-868.
Kennedy, G. L. Jr., S.H. Smith, M. L. Keplinger, and J. C. Calandra. 1975b. Evaluation of
the teratological potential of hexachlorophene in Rabbits and rats. Teratology, 12: 83 -88.
Kennedy, G. L. Jr., S.H. Smith, M. L. Keplinger, and J. C. Calandra. 1976. Reproductive
and peri- and postnatal studies with hexachlorophene. Fd. Cosmet. Toxicol. 14: 421-423.
Kimmel, C.A., W. Moore, O.K. Hysell, and J. F. Stara: 1974. Teratogenicity of
hexachlorophene in rats. Arch Environ Health. 28: 43-48.
August 1995
-------
APPENDIX B Hexachlorophene-10
Nagy, K. A. 1987. Field metabolic rate and food requiremnt scaling in mammals and birds.
Ecol. Mono. 57: 111 - 128. ' .
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemcial Substances) Database. March 1994.
Oakley, G. P. and T.H. Shepard. 1972. Possible teratogenicity of hexachlophene in rats.
Teratology. 5(2): 264. As cited in U. S. EPA (Environmental Protection Agency). 1986
Health and Environmental Effects Profile for Hexachlorophene. March 1986.
Opresko, D. M., B. E. Sample, and G. W. Suter II. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revisions. ES/ER/TM-86/R1. U. S. Department of Energy, Oak Ridge
NationalLaboratory, Oak Ridge, Tennessee.
Robinson, G. R.,.D. J. Wagstaff, J. J. Colaianne and A.G. Ulsamer. 1975. Experimental
hexachlorophene intoxication in young swine. Am. J. Vet. Res. 36 (11): 1615 - 1618.
As cited in U. S. EPA (Environmental Protection Agency). 1986 Health and
Environmental Effects Profile for Hexachlorophene. March 1986.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. Environmental Research Laboratory, Office of
Research and Development, Duluth, MM. PB93-154672.
Suter II, G. W. and J. B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for the Effects on Aquatic Biota: 1994 Revision. DE- AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U. S.
Department of Energy, Washington, D. C.
Thorpe, E. 1967. Some pathological effects of hexachlorophene in the rat. J, Comp. Pathol.
77(2): 137-142. As cited in U. S. EPA (Environmental Protection Agency). 1986
Health and Environmental Effects Profile for Hexachlorophene. March 1986.
Thomann, R. V. 1989. Bioaccumulation model of organic chmeical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6): 699-707.
Thomann, R. V., J. P. Connolly, arid T. F. Parkerton. 1992. An equilibrium model of
organic chemical accumulation in aquatic food webs with sediment interaction.
Environmental Toxicology and Chemistry. 11:615 - 629.
August 1995
-------
APPENDIX B Hoxachlorophene - 11
U. S. Environmental Protection Agency. 1986 Health and Environmental Effects Profile for
Hexachlorophene. EPA/600/22. U.S. EPA, Cincinnati, OH..
U.S. Environmental Protection Agency. 1990e. Methodology for Assessing Health Rishks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment. Washington, D. C. January.
Weiss, L. R., J. T. Williams, and S. Krop. 1973. Behavioral toxicity of hexachlorophene in
rats. Toxicol. Appl. Pharmocol. 25(3): 439. As cited in U. S. EPA (Environmental
Protection Agency). 1986 Health and Environmental Effects Profile for
Hexachlorophene. March 1986.
Will, M. E. and G. W. Suter II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S.> Department of Energy.
August 1995
-------
Freshwater Toxicity - Hexachlorophene Cas No.: 70-30-4
Chemical Name
hexachlorophene
hexachlorophene
Species
Pimephales
promelas
Pimephales
promelas
Type of
Effect
chronic
chronic
Description
LC50
EC50
Value
21
260
Units
ygfl-
ug/L
Test Type
(Static/Flow
Through)
NS
NS
Exposure
Duration
/Timing
4
1
Reference
AQUIRE. 1995
AQUIRE, 1995
Comments
-------
Terrestrial Toxicity - HexachSorophene
CAS No. 70-30-4
Chemical N«me
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Specie*
rats
rabbits
rabbits
rats
rats
Type of
rep
ter
ter
ter
ter
Description
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
3.54
3
6
15
30
Unite
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
Exposure
Route (orel,
WedtJM)?
oral
oral
oral
gavage
gavage
-'"; Expoeunj , .-
beginning at 21 d
of age until
sacrifice of the
parental animals
(3 generations)
d-18 of gestation
d 6-18 of gestation
d 6-15 of gestation
d 6- 15 of gestation
ft.*--
Kennedy et
al.. 1975a
Kennedy et
al., 1975b
Kennedy et
al., 1975b
Kennedy et
al.. 1975b
Kennedy et
al.. 1975b
Comments
Administration of
this dose 'did not
produce any <
.changes with .
respect to mating,
fertility, length of
gestation, and the
number of
deliveries.
No teratological
response to HCP
could be detected
at this level.
Small but not
statistically
significant increase
in incidence of
acrania and minor
skeletal
malformation,
fetotoxicity
(resorptions) and
. maternal toxicity
(reduced rate of
body weight gain).
No teratological
response to HCP
could be detected
at this level.
At this dose level.
there was a
reduced rate of
body weight gain in
the mothers and
reduced fetal body
weights.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Nam*
Hexachlorophene
?
Hexachlorophene
Hexachlorophene
*
Hexachlorophene
Hexachlorophene
Hexachlorophene
SpodM
rats
rats
rats
rats
3-4 Wistar rats
rats
Typ»of
Eft**
rep
. rep
rep
rep
ter, rep
ter
' .;
DMCriptJion
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
1.63
7.08
10
20
1,000
25
Units
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
ppm
mg/kg/d
Exposure
ROIIIB (On),
fMX, I.V., Lp.,
Injection)
oral
oral
gavage
gavage
diet
oral
Exposure
DufAUonffitnlng
began at 4 to 5
wks; continued
through 2
generations
began at 4 to 5
wks; continued
through 2
generations
d 7-15 of gestation
d 7-15 of gestation
throughout
pregnancy
. d 7-20 of gestation
nGMfWIOt
Gaines et
al.. 1973
Gaines et
al.. 1973
-
Gaines et
al., 1973
Gaines et
al., 1973
Thorpe,
1967 as .
cited in
HEPA, 1986
Oakley and
Shepard,
1972 as
cited in
HEPA, 1986
Commant*
Study conducted
over F(o), F1a,
F1b, F2a ,
generations
At this dose level,
the rate of survival
of pups to weaning
was reduced in the
Flaand F1b
generations, with
the latter
generation having
the greater effect.
The dosage level
did not affect
reproduction.
This dose level
caused a reduction
in the number of
rats bom and in
the body weight of
the pups at
weaning.
No adverse effects
were reported on
the fetus or on
fertility.
Small fetuses with
cleft palate were
evident at this
dose level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
«*
Chemical Nona
Hexachlorophene
Hexachlorophene
SpeclM
rats
rats
Type of
Effect
ter
rep
Description
LOAEL
NOAEL
Value
80
50
Unto
mg/kg/d
ppm
Exposure
Route (oral,
Ix;, l.v,, Ip,,
Injection)
introduced
into the
vagina of
pregant rats
oral
Exposure
Duration/Timing
d 7-10 of
gestation
3-generation
study
Rcferoncw
Kimmel et
al., 1974
Plank et
al., 1973
as cited
in HEPA,
1986
Comments
A significant
increase over
the controls
was produced at
this level. The
following
malformations
were observed:
hydrocephaly,
anophthalmia,
microphthalmia,
wavy ribs, and
urogenilal
defects.
There was no
indication of
impaired
fertility or
evidence of
increased fetal
mortality
resulting from
prenatal
exposure within
any generation
of this study.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Name
Hexachlorophene
Hexachlorophene
u
Hexachlorophene
Speder
rats
rats
rats
Type of
ftt* nt
CTTeCI
rep, let
rep, fet
ter
Description
NOAEL
LOAEL
NOAEL
Value
15
30
20
Unto
mg/kg/d
mg/kg/d
mg/kg/d
Exposure
Route (oral,
t,c.i l.v., tp.,
Injection)
oral
oral
introduced into
the vagina of
pregnant rats
Expoture
Dwation/riining
d 15 of gestation
and throughout
lactation
d 15 of gestation
and throughout
lactation
d 7- 10 of gestation
Reference
Kennedy et
al., 1976
Kennedy et
al., 1976
Kimmel et
al., 1974
Comments
Animals treated at
this dose level did
not display any ,
adverse reactions
during the dosing
period. Maternal
toxicity was not
observed at this
dose level.
Maternal toxicity, in
the form of weight
loss and abnormal
neurological signs,
was observed at
this dose level.
The number of
viable pups
delivered
decreased and the
number of stillborn
pups increased at
this dose level.
No teratogenic
effects were
observed at
this level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Ww
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
SpackM
rats
rats
rats
newborn pigs
fyp»of
Effort
rep
behv
behv
neuro
. Dotcffption
LOAEL
NOAEL
LOAEL
LOAEL
VpkM
60
10
25
0.5
.
Unto
ppm
.
mg/kg/d
mg/kg/d
mg/kg/d
Expo*ure
' Mi^ivMi,' '
Infection)
oral
oral
oral
oral
' '" '- ' - '. '
.;'-; ExfXMur* :.,;
- . Duration/Timing '
3 generations
.
30 d
30 d
36 d
, ."'.:'*
IRDC, 1979
as cited
in HEPA,
1986
Weiss et
al., 1973,
1978 as
cited in
HEPA, 1986
Weiss et
al., 1973,
1978 as
cited in
HEPA, 1986
Robinson
et al..
1975 as
-cited in
HEPA, 1986
Comnwnt* ,
The following
effects were
observed: <
decreased
number of
corpora lutea
and
implantations;
decreased
number of F1a
pups surviving
until lactation
d-4; decreased
body weight of
pups.
No effects on
behavior
occurred at
tfiis dose
level.
Deterioration
of avoidance
responding
behavior
occurred at
this level.
No neurological
signs or
lesions were
evident at this
dosage level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Name
Hexachlorophene
Hexachlorophene
Spaces
rats
rats
Type of
Effed
rep
rep
Description
NOAEL
NOAEL
Value
5
""
20
Unto
s-; *..»
mg/kg
ppm
Exposure
Route (oral,
A, U, Ip.,
Injection)
oral
oral
Expcwur*
Duration/Timing
critical life
stage
3 generations
ReforancA
Plank et
al., 1973
as cited
in HEPA,
1986
IRDC. 1979
as cited
in HEPA,
1986
Comments
No unusual
behavioral
reactions were >
observed among
pups from any
group. There
was no evidence
of increased
fetal mortality
resulting from
prenatal
exposure in
this study.
However, pup
survival,
during the
lactation
period, was
reduced at this
level.
No effects were
observed at
this dosage
level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Name
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
finnflai
"r J"r * ,
rats
rabbits
rabbits
rats
rats
Type of
rep
ter
ter
ter
ter
Description
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
3.54
3
6
15
30
Unto:
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
Exposure
ex., I.V., Ip.,
Injection)
oral
oral
oral
gavage
gavage
Exposure
Duration/Timing
beginning at 21 d
of age until
sacrifice of the
parental animals
(3 generations)
d 6-18 of gestation
d 6-18 of gestation
d 6-15 of gestation
d 6- 15 of gestation
Reference
Kennedy et
al., 1975a
Kennedy et
al., 1975b
Kennedy et
al., 1975b
Kennedy et
al., 1975b
Kennedy et
al.. -1975b
Comments
Administration of
this dose did not
produce any ,
changes with
respect to mating,
fertility, length of
gestation, and the
number of
deliveries.
No teratological
response to HCP
could be detected
at this level.
Small but not
statistically
significant increase
in incidence of
acrania and minor
skeletal
malformation,
fetotoxicity
(resorptions) and
maternal toxicity
(reduced rate of
body weight gain).
No teratological
response to HCP
could be detected
at this level.
At this dose level,
there was a
reduced rate of
body weight gain in
the mothers and
reduced fetal body
weights.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Nam*
Hexachlorophene
Hexachlorophene
Hexachlorophene
Specie*
rats
rats
rats
Type of
Effect
rep, fet
rep, fet
ter
DMCriutluii
NOAEL
LOAEL
NOAEL
Value
15
30
20
Unto
mg/kg/d
mg/kg/d
0
mg/kg/d
Exposure
Route (oral,
A, P.v, tp.,
Injection)
oral
oral
introduced into
the vagina of
pregnant rats
Exposure ;
Duretion/nmlng
d 1 5 of gestation
and throughout
lactation
d 1 5 of gestation
and throughput
lactation
d 7- 10 of gestation
Reference
Kennedy et
al., 1976
Kennedy et
al., 1976
Kimmel et
al.. 1974
Comments
Animals treated at
this dose level did
not display any >
adverse reactions
during the dosing
period. Maternal
loxicity was not
observed at this
dose level.
Maternal loxicity, in
the form of weight
loss and abnormal
neurological signs,
was observed at
this dose level.
The number of
viable pups
delivered
decreased and the
number of stillborn
pups increased at
this dose level.
No teratogenic
effects were
observed at
this level.
-------
Terrestrial Toxicity - Hexachlorophene
GAS No. 70-30-4
ChwnteaiNanra
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
^T^!?T- .
rats
rats
rats
rats
3-4 Wistar rats
rats
Type of
'Eftet
rep
rep
rep
rep
ter, rep
ter
Description
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
1;63
7.08
10
20
1,000
25
>ii
Unto
mg/kg/d
mg/kg/d
mg/kg/d
mg/kg/d
ppm
mg/kg/d
Exposure
Route (oral,
-------
Terrestrial 1 .ity - Dieldrin
Cas No. 60-57-1
Chemical
Name
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
dieldrin
-
Species
mallard
California
quail
Japanese
quail
pheasant
chukar
gray
partridge
rock dove
house
sparrow
mule deer
domestic
goat
Endpoint
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
381
8.78
69.7
79
25.3
8.84
26.6
47.6
75 - 150
100.-
200
Units
mg/kg-
body wl.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
bodywt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
bodywt.
mg/kg-.
bodywl.
mg/kg-
body wt.
Exposure Route
(oral, ».c., l.v.,
l.p., Inlectlon)
oral
oral
oral
oral
oral '
oral
oral
oral
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1993b
U.S. EPA. 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA. 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
NS = Not specified
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Name
Hexachlorophene
Hexachlorophene
Spades
rats
rats
Type of
Effect
ter
rep
Description
LOAEL
NOAEL
Value
80
50
Unto
mg/kg/d
ppm
*
exposure
Route (oral,
e.0, l.v., Ip.,
Injection)
introduced
into the
vagina ot
pregant rats
oral
Exposure
Duration/Timing
d 7-10 of
gestation
3-generadon
study
Reference
Kimmel et
al., 1974
Plank et
al., 1973
as cited
in HEPA,
1986
Comments
A significant
increase over
the controls ,
was produced at
this level. The
following
malformations
were observed:
hydrocephaly.
anophthalmia,
microphthalmia,
wavy ribs, and
urogenital
defects.
There was no
indication of
impaired
fertility or
evidence of
increased fetal
mortality
resulting from
prenatal
exposure within
any generation
of this study.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chomleti N«m«
Hexachlorophene
t
Hexachlorophene
8p*%-
rats
rats
Typaof
Effect
rep
rep
OMeripUon
NOAEL
NOAEL
VlIlM
5
20
Unto
mg/kg
ppm
Exposure
Route (owl.
.c,, l.v., Ip.,
InJMitlon)
oral
oral
Expowjro
Punrttonfllmbifi
critical life
stage
3 generations
fwfOT0flC9
Plank et
al., 1973
. as cited
in HERA,
1986
IRDC, 1979
as cited
in HEPA,
1986
COWMTWtlftl
No unusual .
behavioral
reactions were ,
observed among
pups from any
group. There
was no evidence
of increased
fetal mortality .
resulting from
prenatal
exposure in
this study.
However, pup
survival,
during the
lactation
period, was
.reduced at this
level.
No effects were
observed at
this dosage
level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
CtemtealNam
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
SfMClM
',': -V "-.-);
newborn pigs
dog/beagle
rat *
mouse
rabbit
guinea pig
Type of
Effect
neuro
path
acute
acute
acute
acute
Description
NOAEL
PEL
LD50
LD50
LD50
LD50
Value
'-''>. i '-'.
0.1
0.75
56
67
40690
60
Unto
mg/kg/d
mg/kg/d
mg/kg
mg/kg
M9/kg
mg/kg
Exposure
RoufMoral,
§;«, l.v, tp.,
r'}Mn|eca«!)-.
oral
oral
oral
oral
oral
oral
Evpowm
Duration/Timing
36 d
1 3 weeks
NS
NS
NS
NS
Rcfwanc*
Robinson
et al ,
1975 as
cited in
HEPA. 1986
IRDC, 1974
as cited
in HEPA,
1986
RTECS,
1994
RTECS.
1994
RTECS,
1994
RTECS.
1994
ConvMnti
. J'41' '''
Neurological
signs were
evident in half >
of the piglets
at this dosage
level.
The effects
reported at
this dose level
were swollen
salivary gland
and status
spongiosus in
brain and optic
nerves.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical N«M
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
-***,;/' :
rats
_
rats
rats
newborn pigs
Type of
Effect
rep
behv
behv
neuro
Dfiertpiion
' -,'- '!. " '
.- '-t- '
LOAEL
NOAEL
LOAEL
LOAEL
Value
60
10
25
0.5
Unto
ppm
mg/kg/d
mg/kg/d
mg/kg/d
Exposure
*?«««««,
i«4 Mr, (.p.,
Injection)
oral
oral
oral
oral
/:.jxpoavr».
Durttton/Tlmlng
' :s>: f.'fy.- * ' ' ! . .
3 generations
30 d
30 d
36 d
Refenwice
IRDC, 1979
as cited
in HEPA,
1986
Weiss et
al., 1973,
1978 as
cited in
HEPA, 1986
Weiss et
al., 1973.
1978 as
cited in
HEPA, 1986
Robinson
etal.,
1975 as
cited in
HEPA. 1986
COAWMOtS
The following
effects were
observed:
decreased
number of
corpora lutea
and
implantations;
decreased
number of F1 a
pups surviving
until lactation
d-4; decreased
body weight of
pups.
No effects on
behavior
occurred at .
this dose
level.
Deterioration
of avoidance
responding
behavior
occurred at
this level.
No neurological
signs or
lesions were
evident at this
dosage level.
-------
Terrestrial Toxicity - Hexachlorophene
CAS No. 70-30-4
Chemical Nome
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Hexachlorophene
Species
newborn pigs
dog/beagle
rat
mouse
rabbit
guinea pig
Type oi
T'tt lit
Cllvul
neuro
path
acute
acute
acute
acute
*#!&
NOAEL
PEL
LD50
LD50
LD50
LD50
:***;
0.1
0.75
56
67
40690
60
Unto
mg/kg/d
mg/kg/d
mg/kg
mg/kg
*»
mg/kg
Exposure
5S-
oral
oral
oral
oral
oral
oral
Duration/Timing
C';; .? '' ' -" '- :'.*'; '; ' ''
36 d
13 weeks
NS
NS
NS
NS
Reference
Robinson
etal.,
1975 as
cited in
HEPA, 1986
IRDC, 1974
as cited
in HEPA,
1986
RTECS,
1994
RTECS,
1994
RTECS.
1994
RTECS.
1994
Comment*
Neurological
signs were
evident in half >
of the piglets
at this dosage
level.
The effects
reported at
this dose level
were swollen
salivary gland
and status
spongiosus in
brain and optic
nerves.
------- |