&EPA
United States
Environmental Protection
Agency
Office of Research and Development
Washington, DC 20460
EPA-600-R-02-009
www.epa.gov
Procedures for the Derivation of
Equilibrium Partitioning
Sediment Benchmarks (ESBs)
for the Protection of Benthic
Organisms: Endrin
I
: . '--
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EPA/600/R-02/009
August 2003
for the of
for the
Walter! Bew*
Robert M. Burgess**
National Health and Environmental Effects Research Laboratory
Atlantic Ecology Division
Narragansett, RI
David J. Hansen
HydroQual, Inc., Mahwah, NJ
Great Lakes Environmental Center, Traverse City. MI (formerly with U.S. EPA)
Dominic M. DiToro
Manhattan College, Riverdale, NY; HydroQual. Inc..
Mahwah, NJ
Laurie D.DeRosa
HydroQual. Inc., Mahwah, NJ
Heidi E. Bell*
Mary C. Rciley
Office of Water, Washington, DC
Frank E. Stancil, Jr.
National Exposure Research Laboratory
Ecosystems Research Division
Athens, GA
Christopher S. Zarba
Office of Research and Development, Washington, DC
David R. Mount
Robert L. Spchar
National Health and Environmental Effects Research Laboratory
Mid-Continent Ecology Division
Duluth, MN
* Principle U.S. EPA Contacts
** Document Editor
U.S. Environmental Protection Agency
Office of Research and Development
National Health and Environmental Effects Research Laboratory
Atlantic Ecology Division, Narragansett, RI
Mid-Continent Ecology Division, Duluth, MN
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Notice
The Office of Research and Development (ORD) has produced this document to provide procedures for the
derivation of equilibrium partitioning sediment benchmarks (ESBs) for the insecticide endrin. ESBs maybe
useful as a complement to existing sediment assessment tools. This document should be cited as:
U.S. EPA. 2003. Procedures for the Derivation of Equilibrium Partitioning Sediment
Benchmarks (ESBs) forthe Protection of Benthic Organisms: Endrin. EPA-600-R-02-009.
Office of Research and Development. Washington, DC 20460
The information in this document has been funded wholly by the U.S. Environmental Protection Agency. It
has been subject to the Agency's peer and administrative review, and it has been approved for publication as
an EPA document.
Mention of trade names or commercial products does not constitute endorsement or recommendation for use.
Abstract
This equilibrium partitioning sediment benchmark (ESB) document describes procedures to derive
concentrations of the insecticide endrin in sediment which are protective of the presence of benthic
organisms. The equilibrium partitioning (EqP) approach was chosen because it accounts for the varying
biological availability of chemicals in different sediments and allows for the incorporation of the appropriate
biological effects concentration. This provides for the derivation of benchmarks that are causally linked to
the specific chemical, applicable across sediments, and appropriately protective of benthic organisms.
EqP can be used to calculate ESBs for any toxicity endpoint for which there are water-only toxicity data; it is
not limited to any single effect endpoint. For the purposes of this document, the Final Chronic Value (FCV)
from the Water Quality Criterion (WQC) for endrin was used as the toxicity benchmark. This value is
intended to be the concentration of a chemical in water that is protective of the presence of aquatic life. The
ESB is derived by multiplying the FCV by the chemical's KQC, yielding the concentration in sediment
(normalized to organic carbon) that should provide the same level of protection in sediment that the FCV
provides in water. For endrin, this concentration is 5.4 |j,g endrin/goc for freshwater sediments and 0.99 |j,g/
goc for saltwater sediments. Confidence limits of 2.4 to 12 |J.g/goc forfreshwater sediments and 0.44 to 2.2 |j,g/
goc for saltwater sediments were calculated using the uncertainty associated with the degree to which toxicity
could be predicted by multiplying the KQC and the water-only effects concentration. The ESBWQCs should be
interpreted as chemical concentrations below which adverse effects are not expected. At concentrations
above the ESBWQCs, effects may occur with increasing severity as the degree of exceedance increases.
The ESBs do not consider the antagonistic, additive or synergistic effects of other sediment contaminants in
combination with endrin or the potential for bioaccumulation and trophic transfer of endrin to aquatic life,
wildlife or humans.
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Foreword
Under the Clean Water Act (CWA), the U. S. Environmental Protection Agency (EPA) and the
States develop programs for protecting the chemical, physical, and biological integrity of the
nation's waters. To support the scientific and technical foundations of the programs, EPA's Office
of Research and Development has conducted efforts to develop and publish equilibrium
partitioning sediment benchmarks (ESBs) for some of the 65 toxic pollutants or toxic pollutant
categories. Toxic contaminants in bottom sediments of the nation's lakes, rivers, wetlands, and
coastal waters create the potential for continued environmental degradation even where water
column contaminant levels meet applicable water quality standards. In addition, contaminated
sediments can lead to water quality impacts, even when direct discharges to the receiving water
have ceased.
The ESBs and associated methodology presented in this document provide a means to estimate the
concentrations of a substance that may be present in sediment while still protecting benthic
organisms from the effects of that substance. These benchmarks are applicable to a variety of
freshwater and marine sediments because they are based on the biologically available
concentration of the substance in the sediments. These ESBs are intended to provide protection to
benthic organisms from direct toxicity due to this substance. In some cases, the additive toxicity
for specific classes of toxicants (e.g., metal mixtures or polycyclic aromatic hydrocarbon
mixtures) is addressed. The ESBs do not consider the antagonistic, additive or synergistic effects
of other sediment contaminants in combination with endrin or the potential for bioaccumulation
and trophic transfer of endrin to aquatic life, wildlife or humans.
ESBs may be useful as a complement to existing sediment assessment tools, to help assess the
extent of sediment contamination, to help identify chemicals causing toxicity, and to serve as
targets for pollutant loading control measures.
This document provides technical information to EPA Regions, States, the regulated community,
and the public. It does not substitute for the CWA or EPA's regulations, nor is it a regulation
itself. Thus, it cannot impose legally binding requirements on EPA, States, or the regulated
community. EPA and State decisionmakers retain the discretion to adopt approaches on a case-by-
case basis that differ from this technical information where appropriate. EPA may change this
technical information in the future. This document has been reviewed by EPA's Office of Research
and Development (Mid-Continent Ecology Division, Duluth, MN; Atlantic Ecology Division,
Narragansett, RI), and approved for publication.
Mention of trade names or commercial products does not constitute endorsement or
recommendation of use.
This is contribution AED-02-046 of the Office of Research and Development National Health and
Environmental Effects Research Laboratory's Atlantic Ecology Division.
Front cover image provided by Wayne R. Davis and Virginia Lee.
ill
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IV
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Contents
Notice [[[ i
Abstract [[[ i
Forward [[[ iii
Acknowledgments [[[ ix
Executive Summary [[[ xi
Glossary [[[ xiii
Section 1
Introduction [[[ 1-1
1.1 General Information [[[ 1-1
1.2 General Information: Endrin [[[ 1-2
1.3 Applications of Sediment Benchmarks [[[ 1-3
1.4 Overview [[[ 14
Section 2
Partitioning [[[ 2-1
2.1 Description of EqP Methodology [[[ 2-1
2.2 Determination of ^Tow for Endrin [[[ 2-2
2.3 Derivation of KQC from Adsorption Studies [[[ 2-2
2.3.1 KQC from Particle Suspension Studies [[[ 2-2
2.3.2 Koc from Sediment Toxicity Tests [[[ 2-3
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Contents
Section 5
Derivation of Endrin ESBWQCs 5-1
5.1 Derivation of ESBWQCS 5-1
5.2 Uncertainty Analysis 5-2
5.3 Comparison of Endrin ESB WQCs and Uncertainty Concentrations to Sediment
Concentrations that are Toxic or Predicted to be Chronically Acceptable 5-3
5.4 Comparison of Endrin ESB s to STORET and Corps of Engineers,
San Francisco Bay Databases for Sediment Endrin 5-6
5.5 Limitations to the Applicability of ESB s 5-9
Section 6
Sediment Benchmark Values: Application and Interpretation 6-1
6.1 Benchmarks 6.1
6.2 Considerations in the Application and Interpretation ESB 6-1
6.2.1 Relationship of ESB WQC to Expected Effects 6-1
6.2.2 Use of EqP to Develop Alternative Benchmarks 6-1
6.2.3 Influence of Unusual Forms of Sediment Organic Carbon 6-2
6.2.4 Relationship to Risks Mediated through
Bioaccumulation and Trophic Transfer 6-2
6.2.5 Exposures to Chemical Mixtures 6-2
6.2.6 Interpreting ESBs in Combinations with Toxicity Tests 6-2
6.3 Summary 6-3
Section 7
References 7-1
Appendix A A-I
Appendix B B-I
Appendix C c-i
VI
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Tables
Table 2-1. Endrin measured and estimated log10^Tow values 2-2
Table 3-1. Test-specific data for chronic sensitivity of freshwater and saltwater
organisms to endrin 34
Table 3-2. Summary of freshwater and saltwater acute and chronic values, acute-chronic ratios, and
derivation of final acute values, final acute-chronic ratios, and final chronic values for endrin 3-5
Table 3-3. Results of approximate randomization (AR) test for the equality of the freshwater and saltwater
FAV distributions for endrin and AR test for the equality of benthic and combined benthic
and water column WQC FAV distributions 3-6
Table 4-1. Summary of tests with endrin-spiked sediment 4-2
Table 4-2. Water-only and sediment LC50 values used to test the applicability of the EqP theory for endrin.... 4-7
Table 5-1. Equilibrium partitioning sediment benchmarks (ESB s) for endrin using the WQC FCV
as the effect concentration 5-1
Table 5-2. Analysis of variance for derivation of confidence limits of the ESB ss for endrin 5-3
Table 5-3. Confidence limits of the ESB s for endrin 5-3
Figures
Figure 1-1. Chemical structure and physical-chemical properties of endrin
Figure 2-1. Observed versus predicted partition coefficients for nonionic organic chemicals
using Equation 2-4 2-3
Figure 2-2. Organic carbon-normalized sorption isotherm for endrin and probability plot
of KQC from sediment toxicity tests 24
Figure 3-1. Genus mean acute values from water-only acute toxicity tests using freshwater species
versus percentage rank of their sensitivity 3-2
Figure 3-2. Genus mean acute values from water-only acute toxicity tests using saltwater species
versus percentage rank of their sensitivity 3-3
Figure 3 -3. Probability distribution of FAV difference statistics to compare water-only data from freshwater
versus saltwater, benthic versus WQC freshwater, and benthic versus
WQC saltwater data 3-7
Figure 4-1. Percent mortality of amphipods in sediments spiked with acenaphthene or phenanthrene, endrin,
or fluoranthene, and midge in sediments spiked with kepone relative to interstitial
water toxic units 4-5
Figure 4-2. Percent mortality of amphipods in sediments spiked with acenaphthene or phenanthrene,
dieldrin, endrin, or fluoranthene, and midge in sediments spiked with dieldrin relative
to predicted sediment toxic units 4-8
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Contents
Figure 5-1. Predicted genus mean chronic values (PGMCV) calculated from water-only toxicity values
using freshwater species versus percentage rank of their sensitivity 54
Figure 5-2. Predicted genus mean chronic values (PGMCV) calculated from water-only toxicity values
using saltwater species versus percentage rank of their sensitivity 5-5
Figure 5-3. Probability distribution of concentrations of endrin in sediments from streams, lakes,
and estuaries in the United States from 1986 to 1990 from the STORET database compared
with the endrin ESBs values 5-7
Figure 5 -4. Probability distribution of organic carbon-normalized sediment endrin concentrations from
the U.S. Army Corps of Engineers (1991) monitoring program of SanFrancisco Bay 5-8
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Acknowledgments
Coauthors
Walter J. Berry* U.S. EPA, NHEERL, Atlantic Ecology Division,
Narragansett, RI
Robert M. Burgess U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
David J. Hansen HydroQual, Inc., Mahwah, NJ; Great Lakes Environmental
Center, Traverse City, MI (formerly with U.S. EPA)
Dominic M. Di Toro Manhattan College, Riverdale, NY; HydroQual, Inc.,
Mahwah, NJ
Laurie D. De Rosa HydroQual, Inc., Mahwah, NJ
Heidi E. Bell* U.S. EPA, Office of Water, Washington, DC
Mary C. Reiley U.S. EPA, Office of Water, Washington, DC
Frank E. Stancil, Jr. U.S. EPA, NERL, Ecosystems Research Division, Athens, GA
Christopher S. Zarba U.S. EPA, Office of Research and Development, Washington, DC
David R. Mount U.S. EPA, NHEERL, Mid-Continent Ecology Division, Duluth, MN
Robert L. Spehar U.S. EPA, NHEERL, Mid-Continent Ecology Division,
Duluth, MN
Significant Contributors to the Development of the Approach and Supporting Science
Herbert E. Allen University of Delaware, Newark, DE
Gerald T. Ankley U.S. EPA, NHEERL, Mid-Continent Ecology Division,
Duluth, MN
Christina E. Cowan The Procter & Gamble Co., Cincinnati, OH
Dominic M. Di Toro Manhattan College, Riverdale, NY; HydroQual, Inc.,
Mahwah, NJ
David J. Hansen HydroQual, Inc., Mahwah, NJ; Great Lakes Environmental
Center, Traverse City, MI (formerly with U.S. EPA)
Paul R. Paquin HydroQual, Inc., Mahwah, NJ
Spyros P. Pavlou Ebasco Environmental, Bellevue, WA
Richard C. Swartz Environmental consultant (formerly with U.S. EPA)
Nelson A. Thomas U.S. EPA, NHEERL, Mid-Continent Ecology Division,
Duluth, MN (retired)
Christopher S. Zarba U.S. EPA, Office of Research and Development, Washington, DC
Technical Support and Document Review
Patricia DeCastro Computer Sciences Corporation, Narragansett, RI
Robert A. Hoke E.I. DuPont deNemours and Company, Newark, DE
Scott D. Ireland U. S. EPA, Office of Water, Washington, DC
Heinz P. Kollig U.S. EPA, NERL, Ecosystems Research Division, Athens, GA
Tyler K. Linton Great Lakes Environmental Center, Columbus, OH
Robert L. Spehar U.S. EPA, NHEERL, Mid-Continent Ecology Division, Duluth, MN
* Principal U. S. EPA contact ix
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Executive Summary
This equilibrium partitioning sediment benchmark (ESB) document describes procedures to derive
concentrations of the insecticide endrin in sediment which are protective of the presence of
benthic organisms. The equilibrium partitioning (EqP) approach was chosen because it accounts
for the varying biological availability of chemicals in different sediments and allows for the
incorporation of the appropriate biological effects concentration. This provides for the derivation
of benchmarks that are causally linked to the specific chemical, applicable across sediments, and
appropriately protective of benthic organisms.
EqP theory holds that a nonionic chemical in sediment partitions between sediment organic
carbon, interstitial (pore) water and benthic organisms. At equilibrium, if the concentration in
any one phase is known, then the concentrations in the others can be predicted. The ratio of the
concentration in water to the concentration in organic carbon is termed the organic carbon
partition coefficient (K^, wnicn is a constant for each chemical. The ESB Technical Basis
Document (U.S. EPA, 2003a) demonstrates that biological responses of benthic organisms to
nonionic organic chemicals in sediments are different across sediments when the sediment
concentrations are expressed on a dry weight basis, but similar when expressed on a ,wg
chemical/g organic carbon basis G«g/goc). Similar responses were also observed across sediments
when interstitial water concentrations were used to normalize biological availability. The
Technical Basis Document further demonstrates that if the effect concentration in water is known,
the effect concentration in sediments on a ,wg/goc basis can be accurately predicted by multiplying
the effect concentration in water by the chemical's ^Toc.
EqP can be used to calculate ESBs for any toxicity endpoint for which there are water-only
toxicity data; it is not limited to any single effect endpoint. For the purposes of this document,
the Final Chronic Value (FCV) from the Water Quality Criterion (WQC) for endrin was used as
the toxicity benchmark. This value is intended to be the concentration of a chemical in water that
is protective of the presence of aquatic life. If an FCV is not available, a secondary chronic value
(SCV) can be substituted. The ESBWQC is derived by multiplying the FCV by the chemical's KQC,
yielding the concentration in sediment (normalized to organic carbon) that should provide the
same level of protection in sediment that the FCV provides in water. Ancillary analyses
conducted as part of this derivation suggest that the sensitivity of benthic/epibenthic organisms is
not significantly different from pelagic organisms; for this reason, the FCV and the resulting
ESBWQC should be fully applicable to benthic organisms. For endrin, this concentration is 5.4 ^g
endrin/goc for freshwater sediments and 0.99 ^g/goc for saltwater sediments. Confidence limits
of 2.4 to 12 ^g/goc for freshwater sediments and 0.44 to 2.2 ,wg/goc for saltwater sediments were
calculated using the uncertainty associated with the degree to which toxicity could be predicted by
multiplying the KQC and the water-only effects concentration. The ESB cs should be interpreted
as chemical concentrations below which adverse effects are not expected. At concentrations
above the ESB cs, effects may occur with increasing severity as the degree of exceedance
increases. In principle, above the upper confidence limit effects are expected if the chemical is
bioavailable as predicted by EqP theory. A sediment-specific site assessment would provide
further information on chemical bioavailability and the expectation of toxicity relative to the
ESB cs and associated uncertainty limits.
As discussed, while this document uses the WQC value, the EqP methodology can be used by
environmental managers to derive a benchmark with any desired level of protection, so long as
the water-only concentration affording that level of protection is known. Therefore, the resulting
benchmark can be species or site-specific if the corresponding water-only information is available.
For example, if a certain water-only effects concentration is known to be protective for an
XI
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
economically important benthic species, the organic carbon-normalized sediment concentration
protective for that benthic species could be derived using the effects concentration and the
partition coefficient. Such a benchmark might be considered as providing "site-specific
protection" for a species or endpoint, if the goal is to derive a benchmark for that particular site
or species. Another way to make an ESB site-specific would be to incorporate information on
unusual partitioning, if suspected, at the site (see U.S. EPA 2003b).
The ESBs do not consider the antagonistic, additive or synergistic effects of other sediment
contaminants in combination with endrin or the potential forbioaccumulation and trophic transfer
of endrinto aquatic life, wildlife or humans. Consistent with the recommendations of EPA's
Science Advisory Board, publication of these documents does not imply the use of ESBs as stand-
alone, pass-fail criteria for all applications; rather, ESB exceedances could be used to trigger the
collection of additional assessment data. ESBs apply only to sediments having >0.2% organic
carbon by dry weight.
Tier 1 and Tier 2 ESB values were developed to reflect differing degrees of data availability and
uncertainty. Tier 1 ESBs have been derived for endrin in this document, and for the nonionic
organic insecticide dieldrin, metal mixtures, and polycyclic aromatic hydrocarbon (PAH) mixtures
inU.S. EPA(2003c, d, e). Tier2 ESBs are reported inU.S. EPA(2003f).
xn
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Glossary of Abbreviations
ACR
ANOVA
AR
CFR
CWA
DOC
EC50
EPA
EqP
ESB
ESBOC
ESBWQC
pen
WQCOC
FACR
FAV
FCV
FDA
Joe
FRY
GMAV
SoC
HECD
HMAV
Acute-chronic ratio
Analysis of variance
Approximate randomization
Code of Federal Regulations
Clean Water Act
Dissolved organic carbon
Chemical concentration estimated to cause adverse effects to 50% of the test
organisms within a specified time period
United States Environmental Protection Agency
Equilibrium partitioning
Equilibrium partitioning sediment benchmark; for nonionic organics, this
term usually refers to a value that is organic carbon-normalized (more formally
ESBOC) unless otherwise specified
Organic carbon-normalized equilibrium partitioning sediment benchmark
Equilibrium partitioning sediment benchmark derived based on the Water
Quality Criteria for a specific chemical
Dry weight-normalized equilibrium partitioning sediment benchmark derived
based on the Water Quality Criteria for a specific chemical
Organic carbon normalized equilibrium partitioning sediment benchmark
derived based on the Water Quality Ctieria for a specific chemical
First progeny generation
Final acute-chronic ratio
Final acute value
Final chronic value
U.S. Food and Drug Administration
Fraction of organic carbon in sediment
Final residue value
Genus mean acute value
Gram organic carbon
U.S. EPA, Health and Ecological Criteria Division
Habitat mean acute value
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Glossary
IUPAC
IWTU
LC50
LC50S>OC
LC50W
NAS
NERL
NHEERL
NOEC
NTIS
OC
OEC
OST
PAH
PGMCV
PSTU
SCV
SE
SMACR
STORE!
TOC
TU
WQC
International Union of Pure and Applied Chemistry
Interstitial water toxic unit
Organic carbon-water partition coefficient
Octanol-water partition coefficient
Sediment-water partition coefficient
The concentration estimated to be lethal to 50% of the test organisms
within a specified time period
Organic carbon-normalized LC50 from sediment exposure
LC50 from water-only exposure
National Academy of Sciences
U.S. EPA, National Exposure Research Laboratory
U.S. EPA, National Health and Environmental Effects Research
Laboratory
No observed effect concentration
National Technical Information Service
Organic carbon
Observed effect concentration
U.S. EPA, Office of Science and Technology
Polycyclic aromatic hydrocarbon
Predicted genus mean chronic value
Predicted sediment toxic unit
Secondary chronic value
Standard error
Species mean acute-chronic ratio
EPA's computerized database for STOrage and RETrieval of
water-related data
Total organic carbon
Toxic unit
Water quality criteria
xiv
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 1
Introduction
1.1 General Information
Toxic pollutants in bottom sediments of the
nation's lakes, rivers, wetlands, estuaries, and marine
coastal waters create the potential for continued
environmental degradation even where water column
concentrations comply with established WQC. In
addition, contaminated sediments can be a significant
pollutant source that may cause water quality
degradation to persist, even when other pollutant
sources are stopped. The absence of defensible ESBs
make it difficult to accurately assess the extent of the
ecological risks of contaminated sediments and to
identify, prioritize, and implement appropriate cleanup
activities and source controls.
As a result of the need for a procedure to assist
regulatory agencies in making decisions concerning
contaminated sediment problems, the EPA Office of
Science and Technology, Health and Ecological
Criteria Division (OST/HECD) and Office of Research
and Development National Health and Environmental
Effects Research Laboratory (ORD/NHEERL)
established a research team to review alternative
approaches (Chapman, 1987). All of the
approaches reviewed had both strengths and
weaknesses, and no single approach was found to be
applicable for the derivation of benchmarks in all
situations (U.S. EPA, 1989a). The EqP approach was
selected for nonionic organic chemicals because it
presented the greatest promise for generating
defensible, national, numeric chemical-specific
benchmarks applicable across a broad range of
sediment types. The three principal observations that
underlie the EqP approach to establishing sediment
benchmarks are as follows:
1. The concentrations of nonionic organic chemicals
in sediments, expressed on an organic carbon
basis, and in interstitial waters correlate to
observed biological effects on sediment-dwelling
organisms across a range of sediments.
2. Partitioning models can relate sediment
concentrations for nonionic organic chemicals on
an organic carbon basis to freely-dissolved
concentrations in interstitial water.
3. The distribution of sensitivities of benthic
organisms to chemicals is similar to that of water
column organisms; thus, the currently established
WQC FCV or SCV can be used to define the
acceptable effects concentration of a chemical
freely-dissolved in interstitial water.
The EqP approach, therefore, assumes that (1) the
partitioning of the chemical between sediment organic
carbon and interstitial water is at or near equilibrium;
(2) the concentration in either phase can be predicted
using appropriate partition coefficients and the
measured concentration in the other phase (assuming
the freely-dissolved interstitial water concentration can
be accurately measured); (3) organisms receive
equivalent exposure from water-only exposures or from
any equilibrated phase: either from interstitial water
via respiration, from sediment via ingestion or other
sediment-integument exchange, or from a mixture of
exposure routes; (4) for nonionic chemicals, effect
concentrations in sediments on an organic carbon basis
can be predicted using the organic carbon partition
coefficient (^Toc) and effects concentrations in water;
(5) the FCV concentration is an appropriate effects
concentration for freely-dissolved chemical in
interstitial water; and (6) ESBs derived as the product
of the ^Toc and FCV are protective of benthic
organisms. ESB concentrations presented in this
document are expressed as /j,g chemical/g sediment
organic carbon (,wg/goc) anc^not on an interstitial water
basis because (1) interstitial water is difficult to sample
and (2) significant amounts of the dissolved chemical
may be associated with dissolved organic carbon; thus,
total concentrations in interstitial water may
overestimate exposure.
Sediment benchmarks generated using the EqP
approach are suitable for use in providing technical
information to regulatory agencies because they are:
1-1
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Introduction
1. Numeric values
2. Chemical specific
3. Applicable to most sediments
4. Predictive of biological effects
5. Protective of benthic organisms
ESBs are derived using the available scientific data
to assess the likelihood of significant environmental
effects to benthic organisms from chemicals in
sediments in the same way that the WQC are derived
using the available scientific data to assess the
likelihood of significant environmental effects to
organisms in the water column. As such, ESBs are
intended to protect benthic organisms from the effects
of chemicals associated with sediments and, therefore,
only apply to sediments permanently inundated with
water, to intertidal sediment, and to sediments
inundated periodically for durations sufficient to permit
development of benthic assemblages. ESBs should not
be applied to occasionally inundated soils containing
terrestrial organisms, nor should they be used to
address the question of possible contamination of upper
trophic level organisms or the synergistic, additive, or
antagonistic effects of multiple chemicals. The
application of ESBs under these conditions may result
in values lower or higher than those presented in this
document.
ESB values presented herein are the concentrations
of endrin in sediment that will not adversely affect
most benthic organisms. It is recognized that these
ESB values may need to be adjusted to account for
future data. They may also need to be adjusted because
of site-specific considerations. For example, in spill
situations, where chemical equilibrium between water
and sediments has not yet been reached, sediment
chemical concentrations less than an ESB may pose
risks to benthic organisms. This is because for spills,
disequilibrium concentrations in interstitial and
overlying water may be proportionally higher relative to
sediment concentrations. Research has shown that the
source or "quality" of total organic carbon (TOC) in
the sediment does not affect chemical binding (DeWitt
et al., 1992). However, the physical form of the
chemical in the sediment may have an effect. At some
sites, concentrations in excess of an ESB may not pose
risks to benthic organisms because the compound may
be a component of a particulate such as coal or soot, or
exceed solubility such as undissolved oil or chemical.
In these situations, an ESB would be overly protective
of benthic organisms and should not be used unless
modified using the procedures outlined in "Procedures
for the Derivation of Site-Specific Equilibrium
Partitioning Sediment Benchmarks (ESBs) for the
Protection of Benthic Organisms" (U.S. EPA, 2003b).
If the organic carbon has a low capacity (e.g., hair,
sawdust, hide), an ESB would be underprotective. An
ESB may also be underprotective where the toxicity of
other chemicals are additive with an ESB chemical or
where species of unusual sensitivity occur at the site.
This document presents the theoretical basis and
the supporting data relevant to the derivation of ESBs
for endrin. The data that support the EqP approach for
deriving ESBs for nonionic organic chemicals are
reviewed by Di Toro et al. (1991) and EPA (U.S. EPA,
2003a). Before proceeding through the following text,
tables, and calculations, the reader should consider
reviewing Stephan et al. (1985), U.S. EPA (1985) and
U.S.EPA(2003a).
1.2 General Information: Endrin
Endrin is the common name of a "broad spectrum"
organochlorine insecticide/rodenticide. It was
formulated for use as an emulsifiable concentrate, as a
wettable or dustable powder, or as a granular product.
It has been used with a variety of crops including
cotton, tobacco, sugarcane, rice, and ornamentals.
One of its major uses in the United States was for
control of Lepidoptera larvae on cotton. During the
1970's and early 1980's its use was increasingly
restricted until it was banned on October 10, 1984, in
part as a result of its observed toxicity to non-target
organisms, bioaccumulation potential, and persistence
[49 CFR 42792 (October 24, 1984)].
Structurally, endrin is a cyclic hydrocarbon having a
chlorine substituted methanobridge structure (Figure 1-
1). It is similar to dieldrin, an endo-endo stereoisomer,
and has similar physicochemical properties, except that
it is more easily degraded in the environment (Wang,
1988). Endrin is a colorless crystalline solid at room
temperature, having a melting point of about 235 °C and
specific gravity of 1.7 g/cc at 20°C. It has a vapor
pressure of 0.026 mPa (25°C) (Hartley and Kidd, 1987).
Endrin is toxic to non-target aquatic organisms,
birds, bees, and mammals (Hartley and Kidd, 1987).
The acute toxicity of endrin ranges from genus mean
acute values (GMAVs) of 0.15 to 716.88 /^g/L for
freshwater organisms and 0.037 to 790 /-ig/L for
saltwater organisms (Appendix A). There is little
difference between the acute and chronic toxicity of
1-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
endrinto aquatic species; acute-chronic ratios (ACRs)
range from 1.881 to 4.720 for three species (see Table
3-2 in Section 3.3). Endrin bioconcentrates in aquatic
animals from 1,450 to 10,000 times the concentration in
water (U.S. EPA, 1980). The WQC for endrin (U.S.
EPA, 1980) was derived using a Final Residue Value
(FRV) calculated using bioconcentration data and the
Food and Drug Administration (FDA) action level to
protect marketability of fish and shellfish; therefore,
the WQC is not "effects based." In contrast, the ESB
for endrin is effects based. It is calculated from the
FCV derived in Section 3.
1.3 Applications of Sediment Benchmarks
ESBs are meant to be used with direct toxicity
testing of sediments as a method of evaluation
assuming the toxicity testing species is sensitive to the
chemical of interest. They provide a chemical-by-
chemical specification of what sediment concentrations
are protective of benthic aquatic life. The EqP method
should be applicable to nonionic organic chemicals
with a Kow above 3.0. Examples of other chemicals to
which this methodology applies include dieldrin, metal
mixtures (Cd, Cu, Pb, Ni, Ag, Zn), andpolycyclic
aromatic hydrocarbon (PAH) mixtures.
MOLECULAR FORMULA
MOLECULAR WEIGHT
DENSITY
MELTING POINT
PHYSICAL FORM
VAPOR PRESSURE
380.93
1.70 g/cc (20 °C)
235 °C
Colorless crystal
0.026 mPa (25 °C)
CAS NUMBER:
TSL NUMBER:
COMMON NAME:
TRADE NAME:
CHEMICAL NAME:
72-20-8
IO15750
Endrin (also endrine and nendrin)
Endrex (Shell); Hexadrin
1,2,3,4,10,10, heiachloro-lR, 4S, 4aS, 5nS, 6,7R, 8R, 8aR-
oct»hydro-6t 7-epoxy-l, 4:5, 8-dimethanonaphthalene (IUPAC)
or Hexachloroepoxy-octahydru-endo-endo-di methanonaphthalene
Figure 1-1. Chemical structure and physical-chemical properties of endrin (from Hartley and Kidd, 1987).
1-3
-------
Introduction
For the toxic chemicals addressed by the ESB
documents Tier 1 (U.S. EPA, 2003c, d, e, and this
document) and Tier 2 (U. S. EPA, 2003f) values were
developed to reflect the differing degrees of data
availability and uncertainty. Tier 1 ESBs are more
scientifically rigorous and data intensive than Tier 2
ESBs. The minimum requirements to derive a Tier 1
ESB include: (1) Each chemical's organic carbon-water
partition coefficient (KQC) is derived from the octanol-
water partition coefficient (KQW) obtained using the
SPARC (SPARC Performs Automated Reasoning in
Chemistry) model (Karickhoff et al., 1991) and the KQW-
Koc relationship from DiToroetal. (1991). This KQC
has been demonstrated to predict the toxic sediment
concentration from the toxic water concentration with
less uncertainty than KQC values derived using other
methods. (2) The FCV is updated using the most
recent toxicological information and is based on the
National WQC Guidelines (Stephanetal., 1985). (3)
EqP confirmation tests are conducted to demonstrate
the accuracy of the EqP prediction that the KQC
multiplied by the effect concentration from a water-
only toxicity test predicts the effect concentration
from sediment tests (Swartz, 1991; De Witt etal., 1992).
Using these specifications, Tier 1 ESBs have been
derived for the insecticide endrin in this document, the
nonionic organic insecticide dieldrin(U.S. EPA, 2003c),
metals mixtures (U.S. EPA, 2003d), andpolycyclic
aromatic hydrocarbon (PAH) mixtures (U.S. EPA,
2003e). In comparison, the minimum requirements for a
Tier2 ESB (U.S. EPA, 2003f) are less rigorous: (1) The
Kow for the chemical that is used to derive the Koc can
be from slow-stir, generator column, shake flask,
SPARC or other sources. (2) FCVs can be from
published or draft WQC documents, the Great Lakes
Initiative or developed from AQUIRE. Secondary
chronic values (SCV) from Suter and Mabrey (1994) or
other effects concentrations from water-only tests can
be also used. (3) EqP confirmation tests are
recommended, but are not required for the development
of Tier 2 ESBs. Because of these lesser requirements,
there is greater uncertainty in the EqP prediction of the
sediment effect concentration from the water-only
effect concentration, and in the level of protection
affordedby Tier2ESBs. Examples of Tier 2 ESBs for
nonionic organic chemicals are found in U.S. EPA
(20031).
1.4 Overview
Section 1 provides a brief review of the EqP
methodology and a summary of the physical-chemical
properties and aquatic toxicity of endrin. Section 2
reviews a variety of methods and data useful in deriving
partition coefficients for endrin and includes the ^Toc
recommended for use in deriving endrin ESBWQCs.
Section 3 reviews aquatic toxicity data contained in the
endrin WQC document (U.S. EPA, 1980) and new data
that were used to calculate the FCV used in this
document to derive ESBWQC concentrations. In
addition, the comparative sensitivity of benthic and
water column species is examined, and justification is
provided for use of the FCV for endrin in the derivation
of ESBWQCs. Section 4 reviews data on the toxicity of
endrin in sediments, the need for organic carbon
normalization of endrin sediment concentrations, and
the accuracy of the EqP prediction of sediment toxicity
using ^Toc and an effect concentration in water. Data
from Sections 2, 3, and 4 were used in Section 5 as the
basis for the derivation of the ESBWQCs for endrin and
its uncertainty. ESBWQCs for endrin are then compared
with two databases on endrin's environmental
occurrence in sediments. Section 6 concludes with the
sediment benchmarks for endrin and their application
and interpretation. The references cited in this
document are listed in Section 7.
1-4
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 2
Partitioning
2.1 Description of EqP Methodology
ESBs are the numeric concentrations of individual
chemicals that are intended to be predictive of
biological effects, protective of the presence of benthic
organisms, and applicable to the range of natural
sediments from lakes, streams, estuaries, and near-
coastal marine waters. For nonionic organic
chemicals, ESBs are expressed as /j,g chemical/goc and
apply to sediments having >0.2% organic carbon by dry
weight. A brief overview follows of the concepts that
underlie the EqP methodology for deriving ESBs. The
methodology is discussed in detail in "Technical Basis
for the Derivation of Equilibrium Partitioning Sediment
Benchmarks (ESBs) for the Protection of Benthic
Organisms: Nonionic Organics" (U.S. EPA, 2003a),
hereafter referred to as the ESB Technical Basis
Document.
Bioavailability of a chemical at a particular
sediment concentration often differs from one sediment
type to another. Therefore, a method is necessary to
determine ESBs based on the bioavailable chemical
fraction in a sediment. For nonionic organic
chemicals, the concentration-response relationship for
the biological effect of concern can most often be
correlated with the interstitial water (i.e., pore water)
concentration (jj.g chemical/L interstitial water) and
not with the sediment chemical concentration (/j,g
chemical/g sediment) (Di Toro et al., 1991). From a
purely practical point of view, this correlation suggests
that if it were possible to measure the interstitial
water chemical concentration, or predict it from the
total sediment concentration and the relevant sediment
properties, then that concentration could be used to
quantify the exposure concentration for an organism.
Thus, knowledge of the partitioning of chemicals
between the solid and liquid phases in a sediment is a
necessary component for establishing ESBs. For this
reason, the methodology described below is called the
EqP method. As stated above, an ESB can be derived
using any given level of protection, in the following
example the FCV from the endrin WQC is used.
The ESB Technical Basis Document shows that
benthic species, as a group, have sensitivities similar
to all benthic and water column species tested (taken as
a group) to derive the WQC concentration for a wide
range of chemicals. The data showing this for endrin
are presented in Section 3.4. Thus, an ESB can be
established using the FCV, calculated based on the
WQC Guidelines (Stephan et al. , 1985), as the
acceptable effect concentration in interstitial or
overlying water (see Section 5). The partition
coefficient can then be used to relate the interstitial
water concentration (i.e., the calculated FCV) to the
sediment concentration via the partitioning equation.
This acceptable concentration in sediment is an
An ESB is calculated as follows. Let FCV
be the acceptable concentration in water for the
chemical of interest, then compute an ESB using the
partition coefficient, Kf (L/kgsediment), between sediment
and water
ESBWQC
FCV
(2-1)
This is the fundamental equation used to generate an
ESBWQC. Its utility depends on the existence of a
methodology for quantifying Kp.
Organic carbon appears to be the dominant sorption
phase for nonionic organic chemicals in naturally
occurring sediments and, thus, controls the
bioavailability of these compounds in sediments.
Evidence for this can be found in numerous toxicity
tests, bioaccumulation studies, and chemical analyses
of interstitial water and sediments (Di Toro et
al., 1991). The evidence for endrin is discussed in
this section and in Section 4. The organic carbon
binding of a chemical in sediment is a function of
that chemical's KQC and the weight fraction of organic
carbon (foc) in the sediment. The relationship is as
follows
^-p foe KO
It follows that
Aoc FCV
(2-2)
(2-3)
2-1
-------
Partitioning
where ESBWQCOC is an ESBWQC on a sediment organic
carbon basis. Fornonionic organics, "ESBWQC" usually
refers to a value that is organic carbon-normalized
(more formally ESBWQCOC) unless otherwise specified.
KQC is not usually measured directly (although it
can be done; see Section 2.3). Fortunately, KQC is
closely related to the octanol-water partition coefficient
(KOff) (Equation 2-5), which has been measured for
many compounds and can be measured very accurately.
The next section reviews the available information on
the KQVf for endrin.
2.2
Determination of Kow for Endrin
Several approaches have been used to determine
KQVf for the derivation of an ESB, as discussed in the
ESB Technical Basis Document. In an examination of
the literature, primary references were found listing
measured log10^Tow values for endrin ranging from 4.40
to 5.19 and estimated log^^ values ranging from 3.54
to 5.60 (Table 2-1). Karickhoff and Long (1995, 1996)
established a protocol for recommending KQVf values for
uncharged organic chemicals based on the best
available measured, calculated, and estimated data.
The recommended log lQKOVf value of 5.06 for endrin
from Karickhoff and Long (1995) is used to derive ESBs
for endrin.
2.3 Derivation of Koc from Adsorption
Studies
Two types of experimental measurements of KQC
are available. The first type involves experiments
designed to measure the partition coefficient in particle
suspensions. The second type is from sediment toxicity
tests in which sediment endrin, sediment organic
carbon (OC) and freely-dissolved endrin in interstitial
water were used to compute KQC; endrin associated
with dissolved organic carbon (DOC) was not included.
2.3.1 Kocfrom Particle Suspension Studies
Laboratory studies to characterize sorption are
generally conducted using particle suspensions. The
high concentrations of solids and turbulent conditions
necessary to keep the mixture in suspension make data
interpretation difficult as a result of the particle
interaction effect. This effect suppresses the partition
coefficient relative to that observed for undisturbed
sediments (Di Toro, 1985; Mackay and Powers, 1987).
Based on analysis of an extensive body of
experimental data for a wide range of compound types
and experimental conditions, the particle interaction
model (Di Toro, 1985) yields the following relationship
for estimating K
focK(
OC "OC
(2-4)
mfoc K
OC "OC ' ux
where m is the particle concentration in the suspension
(kg/L) and i>x, an empirical constant, is 1.4. The KQC
is given by
= 0.00028 + 0.
(2-5)
Figure 2-1 compares observed partition coefficient
data for the reversible component with predicted values
estimated with the particle interaction model
(Equations 2-4 and 2-5) for a wide range of compounds
Table 2-1. Endrin measured and estimated log10A"ow values
Method
Measured
Measured
Measured
Measured
Estimated
Estimated
Estimated
Log10^ow
4.40
4.92
5.01
5.19
3.54
5.40
5.60
Reference
Rapaport and Eisenreich, 1 984
Ellington and Stancil, 1988
Eadsforth, 1986
DeBruijnetal., 1989
Mabeyetal., 1982
Karickhoff etal., 1989
Neeleyetal.,1974
2-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
(DiToro, 1985). The observed partition coefficient for
endrin using adsorption data (Sharom et al. , 1980) is
highlighted on this plot . The observed log IQK of 2 . 04
reflects significant particle interaction effects. The
observed partition coefficient is about nine times lower
than the value expected in the absence of particle
effects (i.e., Iog10^ = 2.98 fmmfocK
QC = 958 L/kg).
is related to K
In the absence of particle effects, KQC
via Equation 2-5 . For log10^Tow = 5 .06 (see Section
2.2), this expression results in an estimate of log10^Toc
= 4.97.
2.3.2 Kocfrom Sediment Toxicity Tests
Measurements of KQC were available from the
sediment toxicity tests using endrin (Nebeker et al.,
1989; Schuytemaetal., 1989; Stehly, 1992). These tests
used different freshwater sediments having a range of
organic carbon contents of 0.07% to 11.2% (see Table
4-1; Appendix B). Endrin concentrations were
measured in the sediment and interstitial waters,
providing the data necessary to calculate the partition
coefficient for an undisturbed bedded sediment. In the
case of the data reported by Schuytema et al. (1989),
the concentration of endrin in the overlying water at the
end of the 10-day experiment was used. Nebeker et al.
(1989) demonstrated in their experiments, which were
static and run in the same way as those of Schuytema et
al. (1989), that overlying water and interstitial water
endrin concentrations were similar. Figure 2-2A is a
plot of the organic carbon-normalized sorption
isotherm for endrin, where the sediment endrin
concentration (/wg/goc) is plotted versus interstitial
water concentration (,wg/L). The data used to make this
plot are included in Appendix B. The line of unity
slope corresponding to the log10^Toc = 4.97 derived
from the endrin log10^Tow of 5.06 from Karickhoff and
Long (1995) is compared with the data. A probability
plot of the observed experimental log10^Toc values is
shown in Figure 2-2B. The log10^Toc values were
approximately normally distributed, with a mean of
loglQKQC = 4.67 and a standard error of the mean (SE)
of 0.04. This value agrees with the Iog10 KQC = 4.97,
which was computed using the endrin log ^K^ of 5.06
from Karickhoff and Long (1995) using Equation 2-5.
H
M
O
I
ENDRIN
O
O
Predicted
(L/kg)
Figure 2-1. Observed versus predicted partition coefficients for nonionic organic chemicals using Equation 2-4
(figure from DiToro, 1985). Endrin datum is highlighted (Sharom et al., 1980).
2-3
-------
Partitioning
2.4 Summary of Derivation of KQC
Endrin
for
The KQC selected to calculate ESBs for endrin
were based on the regression of log10^Toc to logmKOff
(Equation 2-5) using the endrin log^^ of 5.06 from
Karickhoff and Long (1995). This approach, rather than
use of the KQC from the toxicity tests, was adopted
because the regression equation is based on the most
robust dataset available that spans a broad range of
chemicals and particle types, thus encompassing a wide
range of ^Tow and/oc values. The regression equation
yielded a log IQKQC of 4.97. This value is comparable to
the logloKQC of 4.67 measured in the sediment toxicity
tests.
e
o
0
ii
a
e
U
union
1000
100
10
p i rr~Tntr"~r"~nTTTTn'
E A
Nebeker et al., 1989
Schuytema et al.. 1989
T Stehly, 1992
O.I
/
.1. L : I'Jit _..i..1.11 j.jJJL
ft.ftl 0.1 1 10 100
Interstitial Water Concentration (^g/L)
-TTTITB
10OT
1
u
o
4.0
6-0 "T [ fllH'l J TTTTTTn ""T'T""T 1 I T
5.5
0.1
_LL1J1U...J I. .J...-1
10 20
"7~m n i iiniFr
« "
_L
SO 80
Probability
jjiji j j mini i
99 99.9
Figure 2-2. Organic carbon-normalized sorption isotherm for endrin (A) and probability plot of
Koc (B) from sediment toxicity tests (Nebeker et al., 1989; Schuytema et al., 1989;
Stehly, 1992). The solid line represents the relationship predicted with a Iog10tf
of 4.97.
2-4
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 3
Toxicity of Endrin in
Water Exposures
3.1 Derivation of Endrin WQC
The example used in Section 2 for the EqP method
for derivation of ESBWQCs for endrin uses the WQC
FCV and KQC to estimate the maximum concentration
of nonionic organic chemical in sediments, expressed
on an organic carbon basis, that will not cause adverse
effects to benthic organisms. For this document, life-
stages of species classified as benthic are either
species that live in the sediment (infaunal) or on the
sediment surface (epibenthic) and obtain their food
from either the sediment or water column (U.S. EPA,
2003a). In this section, the FCV from the endrin WQC
document (U.S. EPA, 1980) is revised using new
aquatic toxicity test data, and the use of this FCV is
justified as the appropriate effects concentration for
the derivation of endrin ESBWQCs.
3.2 Acute Toxicity in Water Exposures
A total of 104 standard acute toxicity tests with
endrin have been conducted on 42 freshwater species
from 34 genera (Figure 3-1; Appendix A). Overall
GMAVs ranged from 0.15 to 180 ,wg/L. Fishes,
stoneflies, caddisflies, dipterans, mayflies, glass
shrimp, isopods, ostracods, amphipods, and damselflies
were most sensitive; overall GMAVs for the most
sensitive genera of these taxa range from 0.15 to 4.6
/j,g/L. This database contains 39 tests on the benthic
life-stages of 25 species from 22 genera (Figure 3-1;
Appendix A). Benthic organisms were both among the
most sensitive and the most resistant freshwater
species to endrin. Of the epibenthic species,
stoneflies, caddisflies, fish, mayflies, glass shrimp,
damselflies, amphipods, and dipterans were most
sensitive; GMAVs ranged from > 0.18 to 12 ^g/L.
Infaunal species tested included stoneflies, mayflies,
dipterans, a midge, an oligochaete worm, and an
ostracod; GMAVs ranged from 0.83 /j,g/L for the
midge, Tanytarsus, to > 165 /j,g/L for the oligochaete,
Lumbriculus.
A total of 37 acute toxicity tests were conducted
on 21 saltwater species from 19 genera (Figure 3-2;
Appendix A). Overall GMAVs ranged from 0.037 to
790 ,ug/L. Fishes and a penaeid shrimp were most
sensitive; however, only 7 of the 21 species tested were
invertebrates. Results from 25 tests on benthic life-
stages of 13 species from 11 genera are contained in
this database (Figure 3-2; Appendix A). Benthic
organisms were among both the most sensitive and most
resistant saltwater genera to endrin. The most
sensitive benthic species was the commercially
important pink shrimp, Penaeus duorarum, with a
measured flow-through 96-hour LC50 of 0.037 ,wg/L.
The LC50 represents the chemical concentrations
estimated to be lethal to 50 % of the test organisms
within a specified time period. Other benthic species
for which there are data appeared less sensitive, with
GMAVs ranging from 0.094 to 12 ^g/L.
3.3 Chronic Toxicity in Water Exposures
Life-cycle toxicity tests have been conducted with
the freshwater flagfish (Jordanellafloridae) and fathead
minnow (Pimephales promelas) and with the saltwater
sheepshead minnow (Cyprinodon variegatus) and grass
shrimp (Palaemonetes pugio). Each of these species,
except for P. promelas, has one or more benthic life-
stages.
Two life-cycle toxicity tests have been conducted
with J. floridae (Table 3-1). The concentration-
response relationships were almost identical among the
tests. Hermanutz (1978) observed an 8% reduction in
growth (length) and a 79 % reduction in number of eggs
spawned per female in 0.30 ,wg/L endrin relative to
response of control fish; progeny were unaffected (Table
3-1). Neither parental nor progeny (F t) generation J.
floridae were significantly affected when exposed to
endrin concentrations from 0.051 to 0.22 ,wg/L. The
chronic value from this test was 0.2569. Combined
with the 96-hour companion acute value of 0.85 /-ig/L
(Hermanutz et al., 1985), the acute-chronic ratio (ACR)
for this test is 3.309 (Table 3-2).
In the second life-cycle test, Hermanutz et al.
(1985) observed a 51 % decrease in reproduction in
parental fish exposed to 0.29 ,wg/L endrin, and
3-1
-------
Toxicity of Endrin in Water Exposures
1000
100
o
09
I
o
0.1
0.01
Arthropods
Other Invertebrates
Fish and Amphibians
Pmedoeris (L)
Lumbriculus* (A)
Bufo (L) *
Hexagenia (J)
Dapknia (L) A
Simocephalus (X) A
A OrconKtfS (J)
*
Rana (L)
Jordanella (J)
Asfllui(A)
Parcilia (X),
Atherix (J)
* Gammartis (.4)
A * lschnuni(Jj
Cypridopsis (A)
Tanylarsus (L) A o Palaemoat-tes (A)
Gambwia (/)Q Carassius (J)
Pimephales (J) Q A Baetix (J)
I ° Ptfronarcfila (L)
A *
Claassenia (A)
Lepamis (J)
Acmofuria* (L)
- PerrafJ)
IctaJurua (J) '
n"««j ^/)
ys (A)
20 40 60 80
Percentage Rank of Freshwater Genera
100
Figure 3-1. Genus mean acute values from water-only acute toxicity tests using freshwater species versus
percentage rank of their sensitivity. Symbols representing benthic species are solid; those
representing water column species are open. Asterisks indicate greater than values. A = adult,
J = juvenile, L = larvae, X = unspecified life-stage.
reductions of 73% in survival, 18% in (growth) length,
and 92% in numbers of eggs per female in 0.39 ^g/L.
No significant effects were detected in parental or
progeny generation flagfish in 0.21 /j,g/L. The chronic
value from this test was 0.2468. Combined with the 96-
hour companion acute value of 0.85 /j,g/L (Hermanutz et
al., 1985), the ACR for this test is 3.444. The
geometric mean of these two ACRs is 3.376.
The effect of endrin on P. promelas in a life-cycle
test was only marginally enhanced when exposure was
via water and diet versus water-only exposures
(JarvinenandTyo, 1978). Parentalfishin0.25/j,g/Lin
water-only exposures exhibited about 60% mortality
relative to controls. Mortality of Ft progeny was 70%
in 0.14 ,wg/L, the lowest concentration tested, and 85 %
in 0.25 /-ig/L. Tissue concentrations increased
marginally in fish exposed to the water and diet
treatment relative to fish in water-only exposures.
Effects were observed at all concentrations tested, so
the chronic value for this test is considered to be
< 0.14 ,wg/L. No ACR from this test can be calculated
because no acute value from matching dilution water is
available.
One saltwater invertebrate species, P. pugio, has
been exposed to endrin in a partial life-cycle toxicity
test (Tyler-Schroeder, 1979). Mortality of parental
generation shrimp generally increased as endrin
concentrations increased from 0.11 to 0.79 /-ig/L. Onset
3-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
WOO
100
J3
c;
c
0.11 ,wg/L. The chronic value from this test was
0.01416. Combined with the 96-hour companion acute
value of 0.35 ,wg/L (Tyler-Schroeder, 1979), the ACR
for this test is 4.720.
C. variegatus exposed to endrin in a life-cycle
toxicity test (Hansen et al., 1977) were affected at
endrin concentrations similar to those affecting the two
freshwater fishes described above. Embryos exposed to
0.31 and 0.72 ^g/L endrin hatched early, and all fry
exposed to 0.72 ,wg/L and about half of those exposed to
0.31 ,ug/L died. Females died during spawning, fewer
eggs were fertile, and survival of exposed progeny
decreased in 0.31 ,wg/L. No significant effects were
observed on survival, growth, or reproduction in fish
exposed to 0.027 to 0.12 p-g/L endrin. The chronic
value from this test was 0.1929. Combined with the 96-
hour companion acute value of 0.3629 ,wg/L (Hansen et
al., 1977; Schimmel et al., 1975), the ACR for this test
is 1.881.
The difference between acute and chronic toxicity
of endrin was small (Table 3-2). ACR values were
3.309 and 3.444 forJ.floridae, 4.720 for P. pugio, and
1.881 for C. variegatus. The final ACR (FACR) was
3.106 for both freshwater and saltwater species. Long-
term exposures, not classed as "chronic" in the
National WQC Guidelines (Stephan et al., 1985), also
3-3
-------
Toxicity of Endrin in Water Exposures
Table 3-1. Test-specific data for chronic sensitivity of freshwater and saltwater organisms to endrin
Common
Name,
Scientific
Name
h
Habitat
(life-
Test stage)
Duration
(days)
NOECs°
(Mg/L)
OECs°
(Mg/L)
Observed
Effects
(relative to
controls)
Chronic
Value
(Mg/L)
Reference
Freshwater Species
Flagfish, LC
Jordanella
floridae
Flagfish,
Jordanella
floridae
Fathead
minnow,
Pimephales
promelas
Saltwater Species
LC
E (E,L)
W (J,A)
E(E,L)
W(J,A)
110
140
0.051-
0.22
0.21
0.30
0.29,
0.39
LC
W
(E,L,J,A)
300
<0.14
0.14-
0.25
8% reduction in
growth,
79% reduction
in reproduction
51-92%
reduction in
reproduction,
73% decrease in
survival,
18% reduction
in growth
60% decrease in
adult survival,
70-85%
decrease in
progeny
survival
0.2569
0.2468
Hermanutz,
1978
Hermanutz
etal, 1985
<0.14 Jarvinen and
Tyo, 1978
Grass shrimp,
Palaemonetes
pugio
Sheepshead
minnow,
Cyprinodon
variegatus
PLC W (L) 145
E,W
(E,.LA)
LC E(E) 175
E,W (J,A)
0.03, 0.11- 38-100%
0.05 0.79 decrease in
adult survival,
26-94%
reduction in
progeny growth
0.027- 0.31, 48-100%
0.12 0.72 decrease in
survival;
1 5% reduction
in growth and in
adult
reproduction;
87% decrease in
progeny
survival
0.07416 Tyler-
Schroeder,
1979
0.1929 Hansenet
al., 1977
aTest: LC = life-cycle, PLC = partial life-cycle, ELS = early life-stage.
Habitat: I = infauna, E = epibenthic, W = water column. Life-stage: E = embryo, L = larval, J = juvenile, A = adult.
°NOECs = no observed effect concentrations; OECs = observed effect concentrations.
indicated little difference between acute and chronic
toxicity of endrin. These include tests with the
caddisfly, Brachycentms americanus; stone fly,
Pteronarcys dorsata (Anderson and DeFoe, 1980);
bluntnose minnow, Pimephales notatus (Mount, 1962);
fathead minnow, P. promelas (Jarvinen et al., 1988);
brown bullhead, Ictalurus melas (Anderson and DeFoe,
1980); largemouthbass, Microptems salmoides
(Fabacher, 1976); spot, Leiostomus xanthurus (Lowe,
1966); and mummichog, Fundulus heteroclitus (Eisler,
1970a).
The final acute value (FAV) derived from the
overall GMAVs (Stephan et al., 1985) for freshwater
organisms was 0.1803 /j,g/L. The FAV for saltwater
species was 0.03282 Mg/L (Table 3-2). The FCVs were
used as the effect concentrations for calculating
ESBWQCs for protection of benthic species. The FCV of
3-4
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Table 3-2. Summary of freshwater and saltwater acute and chronic values, acute-chronic ratios, and derivation of
final acute values, final acute-chronic ratios, and final chronic values for endrin
Common Name,
Scientific Name
Acute Value
Og/L)
Species Mean
Chronic Value Acute-Chronic Ratio Acute-Chronic Ratio
Og/L) (ACR) (SMACR)
Freshwater Species
Flagfish,
Jordanellafloridae
Flagfish,
Jordanellafloridae
Fathead minnow,
Pimephales promelas
Saltwater Species
Grass shrimp,
Palaemonetespugio
Sheepshead minnow,
Cyprinodon
variegatus
0.85
0.85
0.35
0.3629
0.2569
0.2468
0.07416
0.1929
3.309
3.444
3.376
4.720
1.881
4.720
1.881
Not used in calculation of SMACR or FACR because acute value from matching dilution water is not available.
Geometric mean of 96-hour LC50 values from three flow-through measured tests (0.34, 0.37, 0.38 ,ug/L) on fry or juvenile fish from
Hansen et al. (1977) and Schimmel et al. (1975). These tests were performed in the same dilution water as the chronic test.
Freshwater:
Final acute value = 0.1803 //g/L
Final acute-chronic ratio = 3.106
Final chronic value = 0.05805 //g/L
Saltwater:
Final acute value = 0.03282 ,ug/L
Final acute-chronic ratio = 3.106
Final chronic value = 0.01057 ,ug/L
0.05805 ,wg/L for freshwater organisms is the quotient
of the FAV of 0.1803 /^g/L and the FACR of 3.106.
Similarly, the FCV for saltwater organisms of 0.01057
A
-------
Toxicity of Endrin in Water Exposures
FAV, computed from the freshwater (combined water
column and benthic) species LC50 values, and the
saltwater FAV, computed from the saltwater (combined
water column and benthic) species LC50 values (Table
3-3). In the AR method, the freshwater LC50 values
and the saltwater LC50 values (see Appendix A) were
combined into one dataset. The dataset was shuffled,
then separated back so that randomly generated
"freshwater" and "saltwater" FAVs could be computed.
The LC50 values were separated back such that the
number of LC50 values used to calculate the sample
FAVs was the same as the number used to calculate the
original FAVs. These two FAVs were subtracted and the
difference used as the sample statistic. This was done
many times so that the sample statistics formed a
distribution representative of the population of FAV
differences (Figure 3-3A). The test statistic was
compared with this distribution to determine its level
of significance. The null hypothesis was that the LC50
values composing the saltwater and freshwater
databases were not different. If this were true, the
difference between the actual freshwater and saltwater
FAVs should be common to the majority of randomly
generated FAV differences. For endrin, the test
statistic occurred at the 99th percentile of the
generated FAV differences. Because the probability
was greater than 95 %, the hypothesis of no significant
difference in sensitivity for freshwater and saltwater
species was rejected (Table 3-3). Note that greater
than (>) values for GMAVs (see Appendix A) were
omitted from the AR analyses for both freshwater
versus saltwater and benthic versus combined water
column and benthic organisms. This resulted in two
endrin freshwater benthic organisms being omitted.
Because freshwater and saltwater species did not
show similar sensitivity, separate tests were conducted
for freshwater and saltwater benthic species. For the
species from each water type, a test of difference in
sensitivity for benthic and all (benthic and water
column species combined, hereafter referred to as
"WQC") organisms, was performed using the AR
method. For this purpose, each life-stage of each test
organism was assigned a habitat (Appendix A) using the
criteria described in U.S. EPA (2003a). The test
statistic in this case was the difference between the
WQC FAV, computed from the WQC LC50 values, and
the benthic FAV, computed from the benthic organism
LC50 values. This was slightly different from the
previous test for saltwater and freshwater species in
that saltwater and freshwater species represented two
separate groups. In this test, the benthic organisms
were a subset of the WQC organisms set. In the AR
method for this test, the number of data points
coinciding with the number of benthic organisms was
selected from the WQC dataset and the "benthic" FAV
was computed. The original WQC FAV and the
"benthic" FAV were then used to compute the
difference statistic. This was done many times, and the
resulting distribution was representative of the
population of FAV difference statistics. The test
statistic was compared with this distribution to
determine its level of significance. The probability
distribution of the computed FAV differences is shown
in Figures 3-3B and 3-3C. The test statistic for this
analysis occurred at the 7th percentile for freshwater
organisms and the 68th percentile for saltwater
organisms, and the hypothesis of no difference in
sensitivity was accepted (Table 3-3). This analysis
suggests that the FCV for endrin based on data from all
tested species was an appropriate effects concentration
for benthic organisms.
Table 3-3. Results of approximate randomization (AR) test for the equality of the freshwater and saltwater FAV
distributions for endrin and AR test for the equality of benthic and combined benthic and water
column WQC FAV distributions
Comparison
Habitat
or Water Type'
a,b
AR Statistic
Saltwater: Benthic vs Water
Column + Benthic (WQC)
Benthic (11)
WQC (19)
0.012
Probability
Freshwater vs Saltwater
Freshwater: Benthic vs Water
Column + Benthic (WQC)
Fresh (32)
Benthic (21)
Salt (19)
WQC (32)
0.149
0.042
99
7
aValues in parentheses are the number of LC50 values used in the comparison.
Note that in both the freshwater vs. saltwater and benthic vs. WQC comparisons, greater than (>) values in Appendix A were omitted.
This resulted in two endrin freshwater benthic organisms being omitted from the AR analysis.
CAR statistic = FAV difference between original compared groups.
Probability that the theoretical AR statistic < the observed AR statistic, given that the samples came from the same population.
3-6
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
05
0.4
OJ
OJ
I -0.1
^ -0,2
^ -0.3
-0.4
-0,5
: A
Freshwater vs Saltwater
_O
05
0.4
OJ
3 °J
8 °-i
| OJ
15 -o.i
O -0.2
jj< -OJ
-0.4
-0.5
TT
TT
B
Benthic vs WQC
Freshwater
at
o
a
B
I
I
U.3
0.4
OJ
02
0.1
OJ
-0.1
-0.2
-0,3
-0,4
-11 C
i i iiFiin i i i iinn i i r
: c
- Benthk vs WQC
Saltwater
-
o
10°
-
_
! 1 HIIIII i i i 1 1 1 111 1 1 I
i i i nun r i r mnn i r
-
-
-
rtiiiiii 1 1 f yyf iTmoEJUDuyL^ A^/fCJO noo Q ^
_
-
-
i i i in ii i ii i mill! i i
o.i
10 20 SO SO 90
Probability
99 99,9
Figure 3-3. Probability distribution of FAV difference statistics to compare water-only data from freshwater
versus saltwater (A), benthic versus WQC freshwater (B), and benthic versus WQC saltwater
(C) data. The solid lines in the figure correspond to the FAV differences measured for endrin.
3-7
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 4
Actual and Predicted Toxicity of
Endrin in Sediment Exposures
4.1 Toxicity of Endrin in Sediments
The toxicity of endrin-spiked sediments was tested
with four freshwater species (two oligochaetesa
lumbriculid worm and a tubificid wormand two
amphipods) and two saltwater species (a polychaete and
the sand shrimp) (Table 4-1). The most common
endpoint measured was mortality; however, impacts
have been reported on sublethal endpoints such as
growth, sediment avoidance, and sediment reworking
rate. All concentrations of endrin in sediments or
interstitial water where effects were observed were
greater than ESBWQC or FCV concentrations reported
in this document. Details about exposure methodology
are provided because sediment testing methodologies
have not been standardized in the way that water-only
toxicity test methodologies have. Generalizations
across species or sediments are limited because of the
limited number of experiments.
Keilty et al. (1988a,b) and Keilty and Stehly (1989)
studied the effects on oligochaete worms of Lake
Michigan sediments spiked with endrin. For all tests,
sediments were dried, passed through a 0.25 mm sieve,
reconstituted with lake water, spiked with endrin
dissolved in acetone, and stirred for 24 hours. The
water (containing the carrier) was aspirated off, new
overlying water added, and sediments placed into
individual beakers for 72 hours before the worms were
added.
Keilty et al. (1988a) examined the effects of
endrin-spiked sediment on sediment avoidance and
mortality of two species of oligochaete worms in
replicate 4-day exposures (Table 4-1). Four-day LC50
values for four tests with Stylodrilus heringianus
averaged 2,110 ,wg endrin/g dry weight sediment and
ranged from 1,050 to 5,400 ,wg endrin/g dry weight
sediment. Four-day LC50 values for three tests with
limnodrilus hoffmeisteri averaged 3,390 ,wg/g dry
weight sediment and ranged from 2,050 to 5,600 /-ig/g
dry weight sediment. Four-day LC50 values from these
tests averaged 194,000 ^g/goc for L. hoffmeisteri and
121,000 ,wg/goc for S. heringianus. Data using this test
method have demonstrated laboratory variabilities by a
factor of 3 to 5 for the same sediment. Sediment
avoidance was seen at much lower concentrations. Over
all tests, burrowing was markedly reduced at > 11.5 /-ig/g
dry weight sediment and possibly at >0.54 /-ig/g dry
weight sediment. EC50s, based on sediment avoidance,
were 59.0 ,wg/g dry weight (3,371 ,wg/goc) for L.
hoffmeisteri and 15.3 and 19.0^g/gdry weight (874 and
1,086 ,wg/goc) sediment for two tests using S.
heringianus. The EC50 represents the chemical
concentration estimated to cause effects to 50% of the
test organisms within a specified time period. Keilty et
al. (1988b) observed 18% mortality of S. heringianus in
11.5 /-ig/g dry weight sediment after a 54-day exposure
and 26% mortality in 42.0 /-ig/g dry weight sediment.
The sediment reworking rate was reported to be
significantly different from the control in sediments
containing >0.54 ,wg/g dry weight sediment. Dry
weights of worms in > 2.33 /-ig/g dry weight sediment
were reduced after 54 days. Keilty and Stehly (1989)
observed no effect of a single, nominal concentration of
50 ,wg/g dry weight sediment on protein utilization by S.
heringianus over a 69-day exposure period. However,
dry weights of worms were significantly reduced.
Nebeker et al. (1989) and Schuytema et al. (1989)
exposed the amphipod Hyalella azteca to two endrin-
spiked sediments, one with a TOC of 11 % and the other
a TOC of 3 %. Nebeker et al. (1989) mixed these two
sediments to obtain a third sediment with a TOC of
6.1%. Sediments were shaken for 7 days in endrin-
coated flasks, and subsequently for 62 days in clean
flasks. The 10-day LC50 values for amphipods in the
three sediments tested by Nebeker et al. (1989) did not
differ when endrin concentration was on a dry weight
basis. The LC50 values decreased with increase in
organic carbon when the concentration was on an
organic carbon basis (Table 4-1). The authors
concluded that endrin data do not support equilibrium
partitioning theory. LC50 values normalized to dry
weight (4.4 to 6.0 Mg/g) or wet weight (0.9 to 1.
differed by less than a factor of 1.5 over a 3.7 fold
range of TOC. In contrast, the organic carbon-
normalized LC50 values ranged from 53.6 to 147
Mg/goc, a factor of 2.7 (Table 4-1).
4-1
-------
Actual and Predicted Toxicity of Endrin in Sediment Exposures
Table 4-1. Summary of tests with endrin-spiked sediment
Sediment Endrin LC50 Interstitial
Common Name,
Scientific Name
Freshwater Species
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Lumbriculid worm,
Stylodrilus
heringianus
Tubificid worm,
Limnodrilus
hoffineisteri
Tubificid worm,
Limnodrilus
hoffineisteri
Tubificid worm,
Limnodrilus
hoffineisteri
Tubificid worm,
Limnodrilus
hoffineisteri
Sediment Source
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
Lake Michigan;
0.25mm sieved
TOC
(%)
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
1.75b
Method,3
Duration
(days)
S,M/4
S,M/4
S,M/4
S,M/4
S,M/4
S,M/4
S,M/54
S,M/54
S,M/54
S,M/54
S,N/69
S,M/4
S,M/4
S,M/4
S,M/4
Dry wt
Response (Mg/g)
LC50 1,400
LC50 1,050
LC50 2,500
LC50 5,400
EC50 19.0
sediment
avoidance
EC50 15.3
sediment
avoidance
26% 42.0
mortality
18% 11.5
mortality
Weight 2.33
loss
Decreased 0.54
sediment
reworking
rate
Weight 50.0
loss
LC50 2,050
LC50 3,400
LC50 5,600°
EC50 59.0
sediment
avoidance
Water
OC LC50
Og/g) Og/L) Reference
80,000 Keilty et al,
1988a
60,000 Keilty et al.,
1988a
143,000 Keilty et al.,
1988a
309,000 Keilty et al.,
1988a
1,086 Keilty etal,
1988a
874 Keilty et al.,
1988a
2,400 Keilty et al.,
1988b
657 Keilty et al.,
1988b
133 Keilty etal.,
1988b
30.8 Keilty etal.,
1988b
2,860 Keilty and Stehly,
1989
117,000 Keilty etal.,
1988a
194,000 Keilty etal.,
1988a
320,000° Keilty et al.,
1988a
3,371 Keilty et al.,
1988a
4-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Table 4-1. Summary of tests with endrin-spiked sediment (continued)
Sediment Endrin LC50
Common Name,
Scientific Name
Amphipod,
Diporeia sp.
Amphipod,
Diporeia sp.
Amphipod,
Diporeia sp.
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Amphipod,
Hyalella azteca
Saltwater Species
Polychaete worm,
Nereis virens
Sand shrimp,
Crangon
septemspinosa
Sediment Source
Lake Michigan;
depth 29m
Lake Michigan;
depth 45m
Lake Michigan;
depth 100m
Soap Creek
Pond No. 7, OR
1 : 1 mixture of
Soap Creek and
Mercer Lake, OR
Mercer Lake, OR
Soap Creek Pond
No. 7, OR;
refrigerated
Soap Creek Pond
No. 7, OR; frozen
Mercer Lake, OR;
refrigerated
Mercer Lake, OR;
frozen
Mercer Lake, OR;
refrigerated
Mercer Lake, OR;
frozen
17% sand, 83%
silt and clay
Sand, wet-
sieved
between l-2mm
sieves
TOC
(%)
0.07
0.55
1.75
3.0
6.1
11.2
3
3
11
11
11
11
2
0.28
Method, a
Duration
(days)
S,M/4
S,M/4
S,M/4
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
R,M/12
R,M/4
Response
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
2 of 5
worms
died
LC50
Dry wt OC
Og/g) Og/g)
0.012 17.0
0.172 31.3
0.224 12.8
4.4 147
4.8 78.7
6.0 53.6
5.1 170
7.7 257
19.6 178
21.7 197
10.3 93.6
9.8 89.1
28 1,400
0.047 16.8
Interstitial
Water
LC50
(,ug/L) Reference
1.07 Stehly, 1992
2.2 Stehly, 1992
0.63 Stehly, 1992
2.1 Nebekeretal.,
1989
1.9 Nebekeretal.,
1989
1.8 Nebekeretal.,
1989
Schuytema et al,
1989
Schuytema et al.,
1989
Schuytema et al.,
1989
Schuytema et al.,
1989
Schuytema et al.,
1989
Schuytema et al.,
1989
McLeeseetal.,
1982
McLeese and
Metcalfe, 1980
aS = static, R = renewal, M = measured, N = nominal.
bValue from Landrum (1991).
CL. hoffmeisteri and S. heringianus tested together.
Clean sediment placed in endrin-coated beakers at beginning of exposure.
4-3
-------
Actual and Predicted Toxicity of Endrin in Sediment Exposures
Schuytema et al. (1989) stored an aliquot of
sediments dosed by Nebeker et al. (1989) for an average
of 9 months and then froze one-half for 2 weeks; the
other half was stored at 4°C for 2 weeks. The toxicity
of endrin to H. azteca did not differ in refrigerated and
frozen sediments from Mercer Lake, OR, and differed
minimally (LC50 = 5.1 vs 7.7 ,wg/g dry weight) in
sediments from Soap Creek Pond. In contrast to the
findings of Nebeker et al. (1989), Schuytema et al.
(1989) used the same test sediments and observed
higher LC50 values in four tests with Mercer Lake
sediments (9.8, 10.3, 19.6, and 21.7 ,wg/g dry weight),
which had a TOC of 11 %, than LC50 values from two
tests using Soap Creek sediments (5.1 and 7.7 ,wg/g dry
weight) where TOC was 3 %.
The only saltwater experiments that tested endrin-
spiked sediments were conducted by McLeese et al.
(1982) and McLeese and Metcalfe (1980). These began
with clean sediments that were added to endrin-coated
beakers just before addition of test organisms. This
was in marked contrast to tests using freshwater
sediments spiked with endrin days or weeks before test
initiation (Nebeker et al., 1989; Schuytema et al.,
1989). As a result, the endrin concentrations in the
sediment and overlying water varied greatly over the
course of these experiments. In addition, the transfer of
test organisms to freshly prepared beakers every 48
hours adds to the uncertainty associated with the
exposure conditions and complicates interpretation of
the results of McLeese et al. (1982).
McLeese et al. (1982) tested the effects of endrin
on the polychaete worm, Nereis virens, in sediment
with 2% TOC (17% sand and 83% silt and clay) in 12-
day toxicity tests. Only two of five worms died at the
highest concentration tested, 28 ^g endrin/g dry weight
sediment or 1,400 ^g endrin/goc. McLeese and
Metcalfe (1980) tested the effects of endrin in sand
with a TOC content of 0.28 % on the sand shrimp,
Crangon septemspinosa. The 4-day LC50 was 0.047 ^g/
g dry weight sediment or 16.8 ,wg/goc. Concentrations
of endrin in water overlying the sediment were
sufficient to explain the observed mortalities of sand
shrimp in sediments.
The need for organic carbon normalization of the
concentrations of nonionic organic chemicals in
sediments is presented in the ESB Technical Basis
Document. For endrin, this need is supported by the
results of the spiked-sediment toxicity tests described
above. When examined individually, experiments in
which H. azteca were exposed to the same sediments by
both Nebeker et al. (1989) and Schuytema et al. (1989)
provide contradictory data concerning the need for
organic carbon normalization (Table 4-1). Nebeker et
al. (1989) observed no change in toxicity with
increasing TOC when endrin was expressed on a dry
weight basis, whereas Schuytema et al. (1989) observed
a decrease in toxicity with increasing TOC when endrin
was expressed on a dry weight basis. However, mean
LC50 values calculated for individual experiments from
both studies were similar when concentrations were
normalized by organic carbon content. The mean
(geometric) LC50 values were 109 ,wg/goc (5 tests) for
sediments from Mercer Lake having a TOC of 11 % and
186 ,wg/goc (3 tests) for sediments from Soap Creek
Pond having 3 % organic carbon. The lack of consistent
evidence supporting organic carbon normalization in
the individual tests reported by Nebeker et al. (1989) is
in contrast with evidence supporting normalization
overall for tests with other nonionic chemicals. The
results for sediments spiked with endrin were most
likely observed because organic carbon concentrations
differed by less than a factor of four and variability
inherent in these tests limited the capacity for
discrimination. Additional tests by Stehly (1992)
provide further support for the need to normalize endrin
concentrations in sediments (Table 4-1). The organic
carbon concentrations for these sediments ranged from
0.07% to 1.75% (a factor of 25). Onadry weight
basis, 4-day LC50 values for Diporeia sp. ranged from
0.012 to 0.224 /j.g/g (a factor of 18.7). The organic
carbon-normalized LC50 values were within a factor of
2.4 and ranged from 12.8 to 31.3 ,wg/goc.
Although it is important to demonstrate that
organic carbon normalization is necessary if ESBWQCs
are to be derived using the EqP approach, it is
fundamentally more important to demonstrate that KQC
and water-only effects concentrations can be used to
predict the effects concentration for endrin and other
nonionic organic chemicals on an organic carbon basis
for a range of sediments. Evidence supporting this
prediction for endrin and other nonionic organic
chemicals follows in Section 4.3.
4.2 Correlation Between Organism
Response and Interstitial Water
Concentration
One corollary of the EqP theory is that freely-
dissolved interstitial water LC50 values for a given
organism should be constant across sediments of
varying organic carbon contents (U.S. EPA, 2003a).
Appropriate interstitial water LC50 values are available
from two studies using endrin (Table 4-1). Nebeker et
al. (1989) found 10-day LC50 values for endrin, based on
4-4
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
interstitial water concentrations, ranged from 1.8 to 2.1
A
-------
Actual and Predicted Toxicity of Endrin in Sediment Exposures
chemical versus interstitial water TUs (IWTUs) for all
sediments. IWTUs are the concentration of the
chemical in interstitial water (,wg/L) divided by the
water-only LC50 (wg/L). Theoretically, 50% mortality
should occur at 1IWTU. At concentrations below 1
IWTU, there should be less than 50% mortality, and at
concentrations above 1 IWTU there should be greater
than 50% mortality. Figure 4-1 shows that, at
concentrations below 1 IWTU, mortality was generally
low and increased sharply at approximately 1 IWTU.
Therefore, this comparison supports the concept that
interstitial water concentrations can be used to make a
prediction that is not sediment-specific of the response
of an organism to a chemical.
4.3 Tests of the Equilibrium Partitioning
Prediction of Sediment Toxicity
Sediment benchmarks derived using the EqP
approach utilize partition coefficients and FCVs from
updated or final WQC documents to derive ESBWQC
concentrations that are protective of benthic organisms.
The partition coefficient KQC is used to normalize
sediment concentrations and predict biologically
available concentrations across sediment types. The
data required to test the organic carbon normalization
for endrin in sediments were available for only one
benthic species. Data from tests with water column
species were not included in this analysis. Testing of
this component of ESBWQC derivation required three
elements: (1) a water-only effects concentration, such
as a 10-day LC50 value, in /j,g/L; (2) an identical
sediment effect concentration on an organic carbon
basis, in ,wg/goc; and (3) a partition coefficient for the
chemical, KQC, in L/kgoc. This section presents
evidence that the observed effect concentration in
sediments (2) can be predicted utilizing the water-only
effect concentration (1) and the partition coefficient (3).
Predicted sediment 10-day LC50 values from
endrin-spiked sediment tests with H. azteca (Nebeker
et al., 1989; Schuytema et al., 1989) were calculated
(Table 4-2) using the Iog10^oc value of 4.97 from
Section 2 of this document and the geometric mean of
the water-only LC50 value (4.1 ^g/L). Overall, ratios
of actual to predicted sediment LC50 values for endrin
averaged 0.33 (range 0.13 to 0.67) in nine tests with
three sediments.
A more detailed evaluation of the accuracy and
precision of the EqP prediction of the response of
benthic organisms can be made using the results of
toxicity tests with amphipods exposed to sediments
spiked with acenaphthene, phenanthrene, dieldrin,
endrin, or fluoranthene. The data included in this
analysis were from tests conducted at EPA laboratories
or from tests that utilized designs at least as rigorous
as those conducted at EPA laboratories. Data from the
kepone experiments were not included because the
recommended KQVf for kepone obtained from Karickhoff
and Long (1995) was evaluated using only one
laboratory measured value, whereas the remaining
chemical KQVf values are recommended based on
several laboratory measured values. Swartz (1991)
exposed the saltwater amphipods E. estuarius and L.
plumulosus to acenaphthene in three marine sediments
having organic carbon contents ranging from 0.82% to
4.2% and to phenanthrene in three marine sediments
having organic carbon contents ranging from 0.82% to
3.6%. Swartz et al. (1990) exposed the saltwater
amphipod R. abronius to fluoranthene in three marine
sediments having 0.18%, 0.31%, and 0.48% organic
carbon. Hoke et al. (1995) exposed the amphipod H.
azteca to three dieldrin-spiked freshwater sediments
having 1.7 %, 2.9 %, and 8.7% organic carbon, and also
exposed the midge C. tentans to two freshwater
dieldrin-spiked sediments having 2.0% and 1.5 %
organic carbon. Nebeker et al. (1989) and Schuytema
et al. (1989) exposed H. azteca to three endrin-spiked
sediments having 3.0 %, 6.1 %, and 11.2% organic
carbon. Figure 4-2 presents the percent mortalities of
amphipods in individual treatments of each chemical
versus predicted sediment TUs (PSTUs) for each
sediment treatment. PSTUs are the concentration of
the chemical in sediment (,wg/goc) divided by the
predicted sediment LC50 (i.e., the product of KQC and
the 10-day water-only LC50 expressed in ,wg/goc). In
this normalization, 50% mortality should occur at 1
PSTU. Figure 4-2 shows that, at concentrations below
1 PSTU, mortality was generally low and increased
sharply at 1 PSTU. Therefore, this comparison
supports the concept that PSTU values also can be used
to make a prediction, that is not sediment-specific, of
the response of an organism to a chemical. The means
of the LC50 values for these tests, calculated on a
PSTU basis, were 1.55 for acenaphthene, 0.73 for
dieldrin, 0.33 for endrin, 0.75 for fluoranthene, and
1.19 for phenanthrene. The mean value for the five
chemicals was 0.80. The fact that this value is so close
to the theoretical value of 1.0 illustrates that the EqP
method can account for the effects of different sediment
properties and properly predict the effects
concentration in sediments using effects concentrations
from water-only exposures.
Data variations in Figure 4-2 reflect inherent
variability in these experiments and phenomena that
4-6
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
have not been accounted for in the EqP model. The
uncertainty of the model is calculated in Section 5.2
of this document. There is an uncertainty of
approximately +2. The error bars shown in Figure 4-2
are computed as +1.96 X (ESBWQC uncertainty). The
value of 1.96 is the t statistic, which provides a 95 %
confidence interval around the ESBWQCs .
Table 4-2. Water-only and sediment LC50 values used to test the applicability of the EqP theory for endrin
Common
Name,
Scientific
Name
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Amphipod,
Hyalella
azteca
Geometric
Mean
Method,a
Duration
(days)
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
S,M/10
Water-
Only
LC50
(,ug/L)
4.2°
3.8°
4.3°
4.1d
4.1d
4.1d
4.1d
4.1d
4.1d
4.1d
Overlying Interstitial
Water Water
LC50 LC50 TOC
(Mg/L) (Mg/L) (%)
1.3° 2.1° 3.0
1.1° 1.9° 6.1
1.2° 1.8° 11.2
1.8° - 3
3.6° 3
3.6° 11
3.9° 11
1.4° 11
1.8° 11
1.9° 1.9C -
Endrin
Sediment
LCSOs
Dry
Wt.
Og/g)
4.4
4.8
6.0
5.1
7.7
19.6
21.7
10.3
9.8
OC
Og/g)
147
78.7
53.6
170
257
178
197
93.6
89.1
125.8
,
Predicted
LC50
(Mg/goc)
392
355
401
383
383
383
383
383
383
383
Ratio:
Actual/
Predicted
LC50
0.38
0.22
0.13
0.44
0.67
0.46
0.51
0.24
0.23
0.33
Reference
Nebeker et al.,
1989
Nebeker et al.,
1989
Nebeker et al.,
1989
Schuytema et
al., 1989
Schuytema et
al., 1989
Schuytema et
al., 1989
Schuytema et
al., 1989
Schuytema et
al., 1989
Schuytema et
al., 1989
S = static; M=measured.
bPredicted LC50 Og/gOc) = water-only LC50 Og/L) x KQC (L/kgoc) x 1 kgoc/1000 goc; where KQC = 104'97.
°Soluble endrin; samples centrifuged prior to analysis.
dMean 10-day water-only LC50 from 3 tests in Nebeker et al. (1989).
4-7
-------
Actual and Predicted Toxicity of Endrin in Sediment Exposures
10(1
SO
3* «
* 4ft
s^
20
0
+ Endrin
O Diddrin
O Plienanlhrene
£\ Kliiorosiithene
V Accnnphthcnc
0
^ *"|
.
° ° A *
^ r4P
» res ftkF'l.
4>. j-^^ ^ wg
, J| °^$ ^tA ^Tgtrf^ljl)^
f i
/<&««;
* * vJ
* J, a
n
in A
"A° ^
g'-1
£ n
S7 \7
I . ydl
]* y0
-
o
^
V
3
"
0.01 0.1 1 II) Mil)
Predicted Sediment Toxic Units witli Uncertainty Bars
Figure 4-2. Percent mortality of amphipods in sediments spiked with acenaphthene or phenanthrene (Swartz,
1991), dieldrin (Hoke et al., 1995), endrin (Nebeker et al., 1989; Schuytema et al., 1989), or
fluoranthene (Swartz et al., 1990; DeWitt et al., 1992), and midge in sediments spiked with dieldrin
(Hoke et al., 1995) relative to predicted sediment toxic units.
4-8
-------
Section 5
Derivation of Endrin ESBWQCs
5.1 Derivation of ESBWQCs
The WQC FCV (see Section 3), without an
averaging period or return frequency, can be used to
calculate the ESBWQCs because the concentration of
contaminants in sediments is probably relatively stable
over time. Thus, exposure to sedentary benthic
species should be chronic and relatively constant.
This contrasts to the situation in the water column,
where a rapid change in exposure and exposures of
limited durations can occur from fluctuations in
effluent concentrations, from dilutions in receiving
waters, or from the free-swimming or planktonic
nature of water column organisms. For some
particular uses of ESBWQCs, it may be appropriate to
use the areal extent and vertical stratification of
contamination at a sediment site hi much the same
way that averaging periods or mixing zones are used
with WQC.
The FCV is the value that should protect 95% of
the tested species included in the calculation of the
WQC from chronic effects of the substance. The FCV
is the quotient of the FAV and the FACR for the
substance. The FAV is an estimate of the acute LC50
or EC50 concentration of the substance corresponding
to a cumulative probability of 0.05 from eight or more
families for the genera for which acceptable acute
tests have been conducted on the substance. The ACR
is the mean ratio of acute to chronic toxicity for three
or more species exposed to the substance that meets
minimum database requirements. For more
information on the calculation of ACRs, FAVs, and
FCVs, see Section 3 of this document and the
National Water Quality Criteria Guidelines (Stephan
et al., 1985). The FCV used in this document differs
from the FCV in the endrin WQC document (U.S.
EPA, 1980) because it incorporates recent data not
included in that document and omits some data that do
not meet the data requirements established in the 1985
WQC Guidelines.
The EqP method for calculating ESBWQCs are
based on the following procedure (also described hi
Section
2-1). If the FCV Gug/L) is the chronic concentration
from the WQC for the chemical of interest, then the
ESBWQCs C"g/g sediment) are computed using the
partition coefficient, Kp (L/g sediment), between
sediment and interstitial water
ESBWQCs = Kp FCV
(5-1)
The organic carbon partition coefficient, #oc, can
be substituted for Kp, because organic carbon is the
predominant sorption phase for nonionic organic
chemicals in naturally occurring sediments (salinity,
grain size, and other sediment parameters have
inconsequential roles in sorption; see Sections 2. 1 and
4.3). Therefore, on a sediment organic carbon basis,
the organic carbon-normalized ESBWQCs (i.e.,
ESBW
QCOC
n
are
ESB
'WQCOC
(5-2)
And because KQ^. is presumably independent of
sediment type for nonionic organic chemicals, so too
are the ESBWQCOC. Table 5-1 contains the calculated
values of the endrin ESBs.
WQC
The ESBWQCOCs are applicable to sediments with
;>0.2%. For sediments with/oc <0.2%, organic
Table 5-1. Equilibrium partitioning sediment benchmarks (ESBWQCs) for endrin using the
WQC FCV as the effect concentration
Type of Water Body
Freshwater
Saltwater
BSBwQcoc-OO^L/kgo,
Log AOW
(L/kg)
5.06
5.06
,) x (ID'3 kgoc/goc)
Log^oc
(L/kg)
4.97
4.97
1 x (0.05805 ^g endrin/L) = 5.4 ,ug ei
FCV
0.05805
0.01057
ESBWQCOC
(Mg/goc)
5.4a
0.99b
bESBWQCOC = (104-
x (KT3
x (0.01057 ^g endrin/L) = 0.99 ,ug endrin/goc.
5-1
-------
carbon normalization and the resulting ESBWQCs do
not apply.
Because organic carbon is the factor controlling
the bioavailability of nonionic organic compounds in
sediments, ESBWQCs have been developed on an
organic carbon basis, not on a dry weight basis.
When the chemical concentrations hi sediments are
reported as dry weight concentrations and organic
carbon data are available, it is best to convert the
sediment concentrations to /^g chemical/g^. These
concentrations can then be directly compared with the
ESBWQCs values. This facilitates comparisons between
ESBWQCs and field concentrations relative to
identification of hot spots and the degree to which
sediment concentrations do or do not exceed ESBWQC
values. The conversion from dry weight to organic
carbon-normalized concentration can be done using
the following formula
Mg chemical/goc = ^g chemical/g^ wt -s- (% TOC -s- 100)
= Mg chemical/g^ wt X 100 + % TOC
For example, a freshwater sediment with a
concentration of 0. 1 ^ug endrin/g^ wt and 0.5% TOC
has an organic carbon-normalized concentration of 20
Mg/goc (= O'1 ^g/Sdryw, X 100 - 0.5), Which
exceeds the freshwater endrin ESBWQC of 5.
Another freshwater sediment with the same
concentration of endrin (0. 1 ^g/g&y wt) but a TOC
concentration of 5.0% would have an organic carbon-
normalized concentration of 2.0 Mg/goc (=0.1 jug/
x 100 -5- 5.0), which is below the freshwater
ESBWQC for endrin.
In situations where TOC values for particular
sediments are not available, a range of TOC values
may be used in a "worst case" or "best case" analysis.
In this case, ESBWQCOC values may be "converted" to
dry weight-normalized ESBWQC values (ESBWQCdry wt).
This "conversion" for each level of TOC is
ESBWQCdrywt
100)
x (% TOC +
For example, the ESBWQCdry w, value for freshwater
sediments with 1 % organic carbon is 0.054
ESBWQCdry wt = 5'
§dry wt
X 1% TOC -r 100 = 0.054 Mg/
This method is used hi the analysis of the STORET
data hi Section 5.4.
5.2 Uncertainty Analysis
Some of the uncertainty of the endrin ESBWQCs
can be estimated from the degree to which the
available sediment toxicity data are predicted using the
EqP model, which serves as the basis for the ESBs.
In its assertion, the EqP model holds that (1) the
bioavailability of nonionic organic chemicals across
sediments is equal on an organic carbon basis and (2)
the effects concentration hi sediment 0"g/goc) can be
estimated from the product of the effects
concentrations from water-only exposures (e.g., FCV
Owg/L)) and the partition coefficient, £oc (L/kg). The
uncertainty associated with the ESBWQCs can be
obtained from a quantitative estimate of the degree to
which the available data support these assertions.
The data used hi the uncertainty analysis are from
the water-only and sediment toxicity tests that were
conducted to fulfill the minimum database
requirements for development of the ESBWQCs (see
Section 4.3 and the ESB Technical Basis Document).
These freshwater and saltwater tests span a range of
chemicals and organisms, they include exposures using
water-only and a number of sediments and are
replicated within each chemical-organism-exposure
media treatment. These data are analyzed using an
analysis of variance (ANOVA) to estimate the
uncertainty (i.e., the variance) associated with the
varying exposure media and that associated with
experimental error. If the EqP model were perfect,
then there would be experimental error only.
Therefore, the uncertainty associated with the use of
EqP is the variance associated with varying exposure
media.
The data used hi the uncertainty analysis are
illustrated hi Figure 4-2. The data for endrin are
summarized hi Appendix B. LC50 values for
sediment and water-only tests were computed from
these data. The EqP model can be used to normalize
the data in order to put it on a common basis. The
LC50 values from water-only exposures (LC50W; ^g/
L) are related to the organic carbon-normalized LC50
values from sediment exposures (LC50S ^ /ug/goc)
via the partitioning equation
(5-3)
As mentioned above, one of the assertions of the EqP
model is that the toxicity of sediments expressed on an
5-2
-------
organic carbon basis equals the toxicity in water tests
multiplied by the X^. Therefore, both LC50S ^ and
KQC x LC50W are estimates of the true LOSO^, for
each chemical-organism pair. In this analysis, the
uncertainty of KQ^, is not treated separately. Any error
associated with KQC will be reflected in the
uncertainty attributed to varying the exposure media.
In order to perform an analysis of variance, a
model of the random variations is required. As
discussed above, experiments that seek to validate
Equation 5-3 are subject to various sources of random
variations. A number of chemicals and organisms
have been tested. Each chemical-organism pair was
tested in water-only exposures and in different
sediments. Let a represent the random variation due
to this source. Also, each experiment was replicated.
Let e represent the random variation due to this
source. If the model were perfect, there would be no
random variations other than those from experimental
error, which is reflected in the replications. Hence, a
represents the uncertainty due to the approximations
inherent in the model and e represents the
experimental error. Let (oa)2 and (o6)2 be the
variances of these random variables. Let i index a
specific chemical-organism pair. Let j index the
exposure media, water-only, or the individual
sediments. Let k index the replication of the
experiment. Then the equation that describes this
relationship is
where ln(LC50j j k) is either ln(LC50w) or
ln(LC50s QC), corresponding to a water-only or
sediment exposure, and ^ is the population ln(LC50)
for chemical-organism pair i. The error structure is
assumed to be log normal which corresponds to
assuming that the errors are proportional to the means
(e.g., 20%), rather than absolute quantities (e.g., 1
/ug/goc). The statistical problem is to estimate ^.,
(oa)2, and (oe)2. The maximum likelihood method is
used to make these estimates (U.S. EPA, 2003a). The
results are shown in Table 5-2. The last line of Table
5-2 is the uncertainty associated with the ESBWQCs;
i.e., the variance associated with the exposure media
variability.
The confidence limits for the ESBWQCs are
computed using this estimate of uncertainty for the
ESBWQCs. For the 95% confidence interval limits, the
significance level is 1.96 for normally distributed
errors.
Hence,
ln(ESBWQCOC)UpPER= ln(ESBWQCOC) +
1.96o,
(5-5)
ln(LC50ijjik) -
(5-4)
ESBWQC
ln(ESBWQCOC)LOWER = ln(ESBWQCOC) -
1-960^^(5-6)
The confidence limits are given in Table 5-3.
The ESB^yococS are applicable to sediments with
/oc 5:0.2%. For sediments with/oc <0.2%, organic
Table 5-2. Analysis of variance for derivation of confidence limits of the ESBWQCsfor endrin
Source of Uncertainty
Parameter
Value
Exposure media
Replication
ESB sediment benchmark
WQC
OESB
WQC
0.41
0.29
0.41
Table 5-3. Confidence limits of the ESBWQCs for endrin
ESB
95% Confidence Limits (/wg/goc)
Type of Water Body
\VQCOC
Lower
Upper
Freshwater
Saltwater
5.4
0.99
2.4
0.44
12
2.2
5-3
-------
carbon normalization and ESBWQCs do not apply.
5.3 Comparison of Endrin ESBWQCs and
Uncertainty Concentrations to
Sediment Concentrations that are
Toxic or Predicted to be Chronically
Acceptable
Insight into the magnitude of protection afforded
to benthic species by ESBWQC concentrations and 95 %
confidence intervals can be inferred using effect
concentrations from toxicity tests with benthic species
exposed to sediments spiked with endrin and sediment
concentrations predicted to be chronically safe to
organisms tested in water-only exposures (Figures 5-1
and 5-2). The effect concentrations are predicted
from water-only toxicity data and £oc values (see
1000000
100000
10000
IQOO
100
10
: PQMCV
A Arthropods
Other Invertebrates
Fish and Amphibians
* DtfMreia sp.
range 3 taste -12,8 to 3U
* H. osteai - 123 pg/gpc < 10d)
range 9 tests - 53.6 to 257
O & ftertttgiamis - 12 1000 tigfeoc
range 5 tests » 60000 to 309000
» ktffflmisteri - 1940QQ iigfgg;
range 3 tests - 117000 to 320000
t
-****
: 5.4
lower: 2.
20 40 60 80
Percentage Rank of Freshwater Genera
100
Figure 5-1. Predicted genus mean chronic values (PGMCV)calculated from water-only toxicity values
(Equation 5-7; Appendix A) using freshwater species versus percentage rank of their sensitivity.
Lines Indicate the freshwater endrin ESBWQC ± 95% confidence limits. Solid symbols are
benthic genera; open symbols are water column ge nera. Arrows indicate greater than values.
Sediment 4-day LC50S oc values (calculated from Keilty et al., 1988a; and Keilty and Stehly,
1989; Nebeker et al. 1989; Schuytema et al., 1989; Stehly, 1992; see lable 4-1) for the amphipods
Diporeia sp. (*) and H. azteca (#) and lumbriculid worm (5. heringianus; 0) and tubificid worm
(L. hoffineisteri; +) are provided for comparison. Error bars around sediment LC50S oc values
indicate observed range of LCSOs.
5-4
-------
Section 4). Chronically acceptable concentrations are
extrapolated from GMAVs from water-only, 96-hour
lethality tests using the FACR. These two predictive
values are used to estimate chronically acceptable
sediment concentrations (predicted genus mean
chronic values, PGMCV) for endrin from GMAVs
(Appendix A), the FACR (Table 3-2), and the K^
(Table 5-1)
PGMCV = (GMAV H- ACR^oc (5-7)
Each PGMCV for fishes and amphibians,
arthropods or other invertebrates tested in water was
plotted against the percentage rank of its sensitivity.
Results from toxicity tests with benthic organisms
exposed to sediments spiked with endrin (Table 4-1;
Appendix B) were placed in the PGMCV rank
appropriate to the test-specific effect concentration.
For example, the mean 10-day LC50S QC for H. azteca,
126 /ug/goc, was placed between the PGMCV of 92
Mg/goc for the amphipod, Gammarus, and the
PGMCV of 138 Mg/goc for me dipteran, Atherix.
Therefore, the LC50 or other effect concentrations are
intermingled in this figure with concentrations
predicted to be chronically safe. Care should be taken
by the reader hi interpreting these data with dissimilar
endpoints. The following discussion of ESBWQCs,
100000
1
u
I
o
10000
1000
100
10
a.
O.I
Water-poly tests:.POMCY
A Arthropods
Other Invertebrates
* Fish mid Amphibians
* A
0*0
* O
o o
upper 2.2
ESBWQC:
lower: 0.44
20
40
60
80
100
Percentage Rank of Saltwater Genera
Figure 5-2. Predicted genus mean chronic values (PGMCV) calculated from water-only toxicity values
(Equation 5-7; Appendix A) using saltwater species versus percentage rank of their sensitivity.
Solid symbols are benthic genera; open symbols are water column genera.
5-5
-------
organism sensitivities, and PGMCVs is not intended to
provide accurate predictions of the responses of taxa
or communities of benthic organisms relative to
specific concentrations of endrin in sediments in the
field. It is, however, intended to guide scientists and
managers through the complexity of available data
relative to potential risks to benthic taxa posed by
sediments contaminated with endrin.
Figures 5-1 and 5-2 are recreations of Figures 3-1
and 3-2, respectively, with GMAVs taken from
Appendix A to calculate PGMCVs using Equation 5-
7. The freshwater ESBWQC for endrin (5.4 ^g/goc) is
less than 33 of the 34 PGMCVs and all of the LC50
values from spiked-sediment toxicity tests (Figure 5-
1). The PGMCV for the fish Perca (4.5 /ug/g^) is
less than the ESBWQC. PGMCVs for 26 of 34
freshwater genera are greater than the upper 95%
confidence interval of the ESBWQC(12 jug/g^). The
PGMCVs for eight genera, including four water
column fish and four benthic arthropod genera, are
below the ESBWQC upper 95% confidence interval.
This illustrates why the slope of the species sensitivity
distribution is important. It also suggests that if the
extrapolation from water-only acute lethality tests to
chronically acceptable sediment concentrations is
accurate, these or similarly sensitive genera may be
chronically affected by sediment concentrations
marginally less than the ESBWQC and possibly less
than the 95% upper confidence interval. For endrin,
the PGMCVs ranged over three orders of magnitude
from the most sensitive to the most tolerant genus. A
sediment concentration 10 times the ESBWQC would
exceed the PGMCVs of 10 of the 22 benthic genera
tested including stoneflies, caddisflies, mayflies,
dipterans, isopods, and fish. Tolerant benthic genera
such as the annelid Lumbriculus may not be
chronically affected in sediments with endrin
concentrations almost 1,000 tunes the ESBWQC. Data
from lethality tests with two freshwater amphipods and
two freshwater annelids exposed to endrin-spiked
sediments substantiate this range of sensitivity. The
LC50 values from these tests range from 2.4 to 59,000
times the ESBWQC of 5.4 y^g/g^.
The saltwater ESBWQC for endrin (0.99 //g/goc) is
less than any of the PGMCVs for saltwater genera
(Figure 5-2). The PGMCVs for the penaeid shrimp
Penaeus (1.1 ^g/gQc) and the fishes Oncorhynchus
(1.44 Mg/goc) and Menidia (1.50 jug/goc) are lower
than the upper 95% confidence interval for the
ESBWQC (2.2
For endrin, PGMCVs from
the most sensitive to the most tolerant saltwater genus
range over two orders of magnitude. A sediment
concentration 20 times the ESBWQC would exceed the
PGMCVs of 6 of the 11 benthic genera tested
including 1 arthropod and 5 fish genera. The hermit
crab Pagurus is less sensitive and might not be
expected to be chronically affected in sediments with
endrin concentrations 300 times the ESBWQC.
5.4 Comparison of Endrin ESBWQCsto
STORET and Corps of Engineers,
San Francisco Bay Databases for
Sediment Endrin
Endrin is frequently measured when samples are
taken to measure sediment contamination, and endrin
values are frequently reported hi databases of sediment
contamination. This means that it is possible that
many of the sediments from the nation's waterways
might exceed the endrin benchmarks. In order to
investigate this possibility, the endrin benchmarks
were compared with data from several available
databases of sediment chemistry.
The following description of endrin distributions
in Figure 5-3 is somewhat misleading because it
includes data from most samples in which the endrin
concentration was below the detection limit. These
data are indicated on the plot as "less than" symbols
(<), but are plotted at the reported detection limits.
Because these values represent upper bounds, not
measured values, the percentage of samples hi which
the ESBWQC values were actually exceeded may be less
than the reported percentage. Very few of the
measured values from either of the databases exceeded
the ESBWQCs.
A STORET (U.S. EPA, 1989b) data retrieval was
performed to obtain a preliminary assessment of the
concentrations of endrin hi the sediments of the
nation's water bodies. Log probability plots of endrin
concentrations on a dry weight basis hi sediments are
shown in Figure 5-3. Endrin was found at significant
concentrations hi sediments from rivers, lakes, and
near-coastal water bodies in the United States. This is
because of its widespread use and the quantity applied
during the 1970s and 1980s. It was banned on
October 10, 1984. Median concentrations were
generally at or near detection limits hi most water
bodies. There is significant variability hi endrin
concentrations in sediments throughout the country.
Lake samples in EPA Region 9 appear to have had
relatively high endrin levels (median = 0.030 Mg/g)
prior to 1986. The upper 10% of the concentrations
were disproportionally found in streams, rivers, and
5-6
-------
Ilndrin
Total Samples: 2*77
ples; 87
1
1
Tbtal Samples: 4 78
Metunircd Sample*: 12
3
1
<*
.1
.9
*C
o
Total Suxnpies: ISO
Meiuured Sumpfcs: 0
50 «ll
Probability
99
99.9
Figure 5-3. Probability distribution of concentrations of endrin in sediments from streams (A), lakes (B),
and estuaries (C) in the United States from 1986 to 1990 from the STORET (U.S. EPA, 1989b)
database compared with the endrin ESBWgC values. Sediment endrin concentrations below the
detection limits are shown as less than symbols (<); measured concentrations are shown as solid
circles (). The upper dashed line on each figure represents the ESBWQC value when TOC=10%,
the lower dashed line represents the ESBWQC when TOC=1%.
5-7
-------
Derivation
lakes in EPA Region 7 and in streams, rivers, lakes,
and estuaries in Region 9 prior to 1986. In some
streams and rivers in Region 7, concentrations
remained high after 1986 (Figure 5-3).
The ESBWQCs for endrin can be compared to
existing concentrations of endrin in sediments of
natural water systems in the United States as contained
in the STORE! database (U.S. EPA, 1989b). These
data were generally reported on a dry weight basis
rather than an organic carbon-normalized basis.
Therefore, ESBWQC concentrations corresponding to
sediment organic carbon levels of 1 % to 10% were
compared with endrin's distribution in sediments as
examples only. For freshwater sediments, ESBWQC
concentrations were 0.054 /ug/g dry weight in
sediments having 1 % organic carbon and 0.54 ,ug/g
dry weight in sediments having 10% organic carbon;
for marine sediments, the ESBWQCs were 0.0099 ,ug/g
dry weight and 0.099 //g/g dry weight, respectively.
Figure 5-3 presents the comparisons of these ESBWQCs
with probability distributions of observed sediment
endrin levels for streams and lakes (freshwater
systems, A and B) and estuaries (marine systems, C).
For streams (n = 2,677), the ESBWQCs of 0.054
//g/g dry weight for 1 % organic carbon sediments and
0.54 yug/g dry weight for 10% organic carbon
freshwater sediments were exceeded in less than 1 % of
the samples. For lakes (n = 478), the ESBWQC of
0.054 /^g/g dry weight for 1 % organic carbon
sediment was exceeded in about 2% of the samples,
and the ESBWQC of 0.54 Mg/g dry weight for 10%
organic carbon freshwater sediments was exceeded in
less than 1 % of the samples. In estuaries, the data (n
= 150) indicate that the ESBWQC of 0.0099 /ug/g dry
weight sediment for 1 % organic carbon sediments was
exceeded in about 8% of the samples, and the ESBWQC
of 0.099 /ug/g dry weight for 10% organic carbon
freshwater sediments was not exceeded by any of the
samples.
A second set of data was analyzed, from the U.S.
Army Corps of Engineers (1991) monitoring program,
for a number of locations in various parts of San
Francisco Bay. For a listing of locations sampled, the
number of observations at each site, and the period
during which the results were obtained, see U.S.EPA
(2003a). These data were collected to examine the
10' I .....Ml
"fib
.8
S
I/I
0.1
10 0 50 80 90
Probability
99
99.9
Figure 5-4. Probability distribution of organic carbon-normalized sediment endrin concentrations from the
U.S. Army Corps of Engineers (1991) monitoring program of San Francisco Bay. Sediment endrin
concentrations less than the detection limits are shown as open triangles (v); measured
concentrations are shown as solid circles (). The solid line is an estimate of the distribution
developed by accounting for nondetected observations.
5-8
-------
quality of dredged sediments in order to determine
their suitability for open water disposal. The database
did not indicate what determinations were made
concerning their acceptability for this purpose.
Investigators compared the frequency of
occurrence of a given sediment endrin concentration
(in individual samples, not dredge sites) with the
ESBWQCs developed using the EqP methodology. A
major portion (93%) of the samples analyzed had/^
>0.2%, for which the ESBWQC concentrations are
applicable. The concentrations of endrin measured in
sediments were normalized by the organic carbon
content, and the results are displayed as a probability
plot in Figure 5-4 to illustrate the frequency at which
different levels are observed. Nearly all of the
samples were less than the varying detection limits of
the analytical tests. Each of the samples for which an
actual measurement was obtained was at least an order
of magnitude lower than the ESBWQCs. An estimate of
the possible frequency distribution of sediment
concentrations of endrin was developed by the
application of an analysis technique that accounts for
the varying detection limits and the presence of
nondetected observations (El-Sharrawi and Dolan,
1989). The results are illustrated by the straight line,
which suggests that no appreciable number of
exceedences is expected. However, the virtual absence
of detected concentrations makes the distribution
estimates unreliable They are presented only to
suggest the probable relationship between the levels of
these two pesticides hi relation to sediment
benchmarks.
Regional-specific differences hi endrin
concentrations may affect the above conclusions
concerning expected example benchmarks
exceedences. This analysis also does not consider
other factors such as the type of samples collected
(i.e, whether samples were from surficial grab
samples or vertical core profiles) or the relative
frequencies and intensities of sampling in different
study areas. It is presented as an aid in assessing the
range of reported endrin sediment concentrations and
the extent to which they may exceed the ESBWQCs.
5.5 Limitations to the Applicability of
ESBs
Rarely, if ever, are contaminants found alone hi
naturally occurring sediments. Obviously, the feet
that the concentration of a particular contaminant does
not exceed the ESBs does not mean that other
chemicals, for which there are no ESBs available, are
not present in concentrations sufficient to cause
harmful effects. Furthermore, even if ESBs were
available for all of the contaminants hi a particular
sediment, there might be additive or synergistic effects
that the benchmarks do not address. In this sense, the
ESBs represent a "best case" benchmark.
It is theoretically possible that antagonistic
reactions between chemicals could reduce the toxicity
of a given chemical such that it might not cause
unacceptable effects on benthic organisms at
concentrations above the ESBs when it occurs with the
antagonistic chemical. However, antagonism has
rarely been demonstrated. More common would be
instances where toxic effects occur at concentrations
below the ESBs because of the additive toxicity of
many common contaminants such as heavy metals and
polycyclic aromatic hydrocarbons (PAHs) (Alabaster
and Lloyd, 1982), and instances where other toxic
compounds for which no ESBs exist occur along with
ESB chemicals.
Care must be used hi the application of
benchmarks hi disequilibrium conditions. In some
instances, site-specific ESBs may be required to
address disequilibrium. The ESBs assume that
nonionic organic chemicals are hi equilibrium with the
sediment and interstitial water and are associated with
sediment primarily through adsorption to sediment
organic carbon. In order for these assumptions to be
valid, the chemical must be dissolved hi interstitial
water and partitioned into sediment organic carbon.
Therefore, the chemical must be associated with the
sediment for a sufficient length of time for equilibrium
to be reached. In sediments where particles of
undissolved endrin occur, disequilibrium exists and the
benchmarks are overprotective. In liquid chemical
spill situations, disequilibrium concentrations hi
interstitial and overlying water may be proportionately
higher relative to sediment concentrations. In this
case, the benchmarks may be underprotective.
Note that the ^Toc values used in the EqP
calculations described hi this document assume that
the organic carbon hi sediments is similar hi
partitioning properties to "natural" organic carbon
found in most sediments. While this has proven true
for most sediments EPA has studied, it is possible that
some sites may have components of sediment organic
carbon with different properties. This might be
associated with sediments whose composition has been
highly modified by industrial activity, resulting in high
5-9
-------
percentages of atypical organic carbon such as rubber,
animal processing waste (e.g., hair or hide fragments),
coal particles, or wood processing wastes (bark, wood
fiber, or chips). Relatively undegraded woody debris
or plant matter (e.g., roots, leaves) may also contribute
organic carbon that partitions differently from typical
organic carbon (e.g., Iglesias-Jimenez et al., 1997;
Grathwohl, 1990; Xing et al., 1994). Sediments with
substantial amounts of these materials may exhibit
higher concentrations of chemicals in interstitial water
than would be predicted using generic K^ values,
thereby making the ESBs underprotective. If such a
situation is encountered, the applicability of literature
KQC values can be evaluated by analyzing for the
chemical of interest in both sediment and interstitial
water. If the measured concentration in interstitial
water is markedly greater (e.g., more than twofold)
than that predicted using the KQ^. values recommended
herein (after accounting for DOC binding in the
interstitial water), then the ESBs would be
underprotective and calculation of a site-specific ESB
should be considered (see U.S. EPA, 2003b).
The presence of organic carbon hi large particles
may also influence the apparent partitioning. Large
particles may artificially inflate the effect of the
organic carbon because of their large mass, but
comparatively small surface area; they may also
increase variability in TOC measurements by causing
sample heterogeneity. The effect of these particles on
partitioning can be evaluated by analysis of interstitial
water as described above, and site-specific ESBs may
be used if required. It may be possible to screen large
particles from sediment prior to analysis to reduce
their influence on the interpretation of sediment
chemistry relative to ESBs.
In very dynamic areas, with highly erosional or
depositional bedded sediments, equilibrium may not
be attained with contaminants. However, even high
KOW nonionic organic compounds come to equilibrium
hi clean sediment in a period of days, weeks, or
months. Equilibrium times are shorter for mixtures of
two sediments that each have previously been at
equilibrium. This is particularly relevant in tidal
situations where large volumes of sediments are
eroded and deposited, even though near equilibrium
conditions may predominate over large areas. Except
for spills and participate chemical, near equilibrium is
the rule and disequilibrium is less common. In
instances where it is suspected that EqP does not apply
for a particular sediment because of the disequilibrium
discussed above, site-specific methodologies may be
applied (U.S. EPA, 2003b). .
5-10
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 6
Sediment Benchmark Values:
Application and Interpretation
6.1 Benchmarks
Based on the level of protection provided by WQC,
the procedures described in this document indicate that
benthic organisms should be comparably protected from
adverse effects of endrin where endrin concentrations
in sediment are below the ESBWQC values of 5.4 (j^g
endrin/goc for freshwater sediments and 0.99 |j,g
endrin/goc for marine/estuarine sediments, except
possibly where a locally important species is very
sensitive or sediment organic carbon is < 0.2%.
Confidence limits of 2.4 to 12 jj,g/goc for
freshwater sediments and 0.44 to 2.2 jj,g/goc for
marine/estuarine sediments are provided as an estimate
of the uncertainty associated with the degree to which
toxicity can be predicted using the Koc and the water-
only effects concentration. Confidence limits do not
incorporate uncertainty associated with water quality
criteria, or unusual, site-specific circumstances. An
understanding of the theoretical basis of the
equilibrium partitioning methodology, uncertainty, and
the partitioning and toxicity of endrin are required in
the use of ESBs and their confidence limits.
The benchmarks presented in this document are
the concentrations of a substance that may be present
in sediment while still protecting benthic organisms
from the effects of that substance. These benchmarks
are applicable to a variety of freshwater and marine
sediments because they are based on the biologically
available concentration of the substance in those
sediments.
These benchmarks do not protect against additive,
synergistic, or antagonistic effects of contaminants or
bioaccumulative effects to aquatic life, wildlife or
human health. Consistent with the recommendations of
EPA's Science Advisory Board, publication of these
documents does not imply the use of ESBs as stand-
alone, pass-fail criteria for all applications; rather,
exceedances of ESBs could trigger collection of
additional assessment data.
6.2 Considerations in the Application and
Interpretation of ESBs (also see
Section 5.5)
6.2.1 Relationship of ESB c to Expected
Effects
The ESBWQC should be interpreted as a chemical
concentration below which adverse effects are not
expected. In comparison, at concentrations above the
ESBWQC effects may occur. In principle, above the
upper confidence limit effects are expected if the
chemical is bioavailable as predicted by EqP theory. In
general terms, the degree of effect expected increases
with increasing endrin concentration in the sediment.
Because the FCV is derived as an estimate of the
concentration causing chronic toxicity to sensitive
organisms, effects of this type may be expected when
sediment concentrations are near the ESB . As
sediment concentrations increase beyond the ESBWQC,
one can expect chronic effects on less sensitive species
and/or acute effects on sensitive species.
6.2.2 Use of EqP to Develop Alternative
Benchmarks
The FCV is used to define a threshold for
unacceptable effects based on its precedence in
establishing unacceptable effects in the development of
WQC. However, the use of EqP to assess sediment
contamination is not limited to the ESB c and the
associated level of protection. By substituting water-
only effect values other than the FCV into the ESB
equation, other benchmarks may be developed that are
useful in evaluating specific types of biological effects,
or that better represent the ecological protection goals
for specific assessments.
6-1
-------
Sediment Benchmark Values: Application and Interpretation
6.2.3 Influence of Unusual Forms of
Sediment Organic Carbon
Partition coefficients used for calculating these
ESBs are based on estimated and measured partitioning
from natural organic carbon in typical field sediments.
Some sediments influenced heavily by anthropogenic
activity may contain sources of organic carbon whose
partitioning properties are not similar, such as rubber,
animal processing wastes (e.g., hair or hide fragments),
or wood processing wastes (bark, wood fiber or chips).
Relatively undegraded woody debris or plant matter
(e.g., roots, leaves) may also contribute organic carbon
that results in partitioning different from that of typical
organic carbon. Sediments with large amounts of these
materials may show higher concentrations of chemicals
in interstitial water than would be predicted using
generic Koc values, making the ESBs underprotective.
Direct analysis of interstitial water can be used to
evaluate this possibility (see U.S. EPA, 2003a,b); if
necessary, derivation of site-specific Koc values may be
warranted.
6.2.4 Relationship to Risks Mediated
through Bioaccumulation and
Trophic Transfer
As indicated above, ESBs are designed to address
direct toxicity to benthic organisms exposed directly to
contaminated sediment. They are not designed to
address risks that may occur through bioaccumulation
and subsequent exposure of pelagic aquatic organisms
(e.g., predatory fish), terrestrial or avian wildlife, or
humans. No inference can be drawn between
attainment of the ESBWQC and the potential for risk via
bioaccumulation; the potential for those risks must be
addressed by separate means.
6.2.5 Exposures to Chemical Mixtures
The methodology described in this document can be
used to derive ESBWQCs that protect against the specific
toxic effects of endrin; it does not account for potential
antagonistic, additive, or synergistic effects that may
occur in sediments containing a mixture of endrin and
other chemicals. Consideration of this potential must
be on a site-specific basis. In general terms, it might
be expected that chemicals with toxicological modes of
action similar to endrin may show additive toxicity
with endrin
6.2.6 Interpreting ESBs in Combination
with Toxicity Tests
Sediment toxicity tests provide an important
complement to ESBs in interpreting overall risk from
contaminated sediments. Toxicity tests have different
strengths and weaknesses compared to chemical-
specific guidelines, and the most powerful inferences
can be drawn when both are used together.
Unlike chemical-specific guidelines, toxicity tests
are capable of detecting any toxic chemical, if it is
present in toxic amounts; one does not need to know
what the chemicals of concern are to monitor the
sediment. Toxicity tests are also useful for detecting
the combined effect of chemical mixtures, if those
effects are not considered in the formulation of the
applicable chemical-specific guideline.
On the other hand, toxicity tests have weaknesses
also; they provide information only for the species
tested, and also only for the endpoints measured. This
is particularly critical given that most sediment
toxicity tests conducted at the time of this writing
primarily measure short-term lethality; chronic test
procedures have been developed and published for some
species, but these procedures are more resource-
intensive and have not yet seen widespread use. In
contrast, the ESBWQC is intended to protect most
species against both acute and chronic effects.
Many assessments may involve comparison of
sediment chemistry (relative to ESBs or other sediment
quality guidelines) and toxicity test results. In cases
where results using these two methods agree (either
both positive or both negative), the interpretation is
clear. In cases where the two disagree, the
interpretation is more complex and required further
evaluation.
Individual ESBs address only the effects of the
chemical or group of chemicals for which they are
derived. For this reason, if a sediment shows toxicity
but does not exceed the ESB for a chemical of
interest, it is likely that the cause of toxicity is a
different chemical or chemicals. This result might
6-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
also occur if the partitioning of the chemical in a
sediment is different from that assumed by the ^Toc
value used (see "6.2.3 Influence of Unusual Forms of
Sediment Organic Carbon" above).
In other instances, it may be that an ESBWQC is
exceeded but the sediment is not toxic. As explained
above, these findings are not mutually exclusive,
because the inherent sensitivity of the two measures is
different. The ESB c is intended to protect relatively
sensitive species against both acute and chronic effects,
whereas toxicity tests are run with specific species that
may or may not be sensitive to chemicals of concern,
and often do not encompass the most sensitive endpoints
(e.g., growth or reproduction). As such, one would not
expect an endrin concentration near the ESBWQC to
cause lethality in a short-term test. It is also possible
for a sediment above the ESBWQC to be non-toxic if
there are site-specific conditions that run counter to the
equilibrium partitioning model and its assumptions as
outlined in this document.
A good method for evaluating the results of toxicity
tests is to calculate effect concentrations in sediment
that are species and endpoint specific. For species
contained in the water-only toxicity data for the endrin
ESB cs (Section 3), effect concentrations in sediment
can be calculated that are specific for that organism
using procedures in Section 5. These values could then
be used to directly judge whether the absence of
toxicity in the toxicity test would be expected from the
concentration of endrin present.
If the exceedance of an ESB is sufficient that one
would expect effects in a toxicity test but they are not
observed, it is prudent to evaluate the partitioning
behavior of the chemical in the sediment. This is
performed by isolating interstitial water from the
sediment and analyzing it for endrin. Predicted
concentrations of endrin in the interstitial water can be
calculated from the measured concentrations in the
solid phase (normalized to organic carbon) as follows
ug chemical/L = (u.g chemical/goc) x 103goc/Kgoc -f- Koc
For chemicals with log Kow greater than 5.5,
corrections for DOC binding in the interstitial water
will be necessary (see Gschwend and Wu 1985;
Burkhard 2000). If the measured chemical in the
interstitial water is substantially less (e.g., 2-3 fold
lower or more), it suggests that the organic carbon in
that sediment may not partition similarly to more
typical organic carbon, and derivation of site-specific
ESBs based on interstitial water may be warranted
(U.S.EPA2003b).
6.3 Summary
Based on the level of protection provided by WQC,
the procedures described in this document indicate that
benthic organisms should be comparably protected from
adverse effects of endrin where endrin concentrations
in sediment are below the ESB values of 5.4 |j,g
endrin/goc for freshwater sediments and 0.99 (j^g
endrin/goc for marine/estuarine sediments, except
possibly where a locally important species is very
sensitive or sediment organic carbon is < 0.2%.
The ESBs do not consider the antagonistic,
additive or synergistic effects of other sediment
contaminants in combination with endrin or the
potential forbioaccumulation and trophic transfer of
endrin to aquatic life, wildlife or humans. Consistent
with the recommendations of EPA's Science Advisory
Board, publication of these documents does not imply
the use of ESBs as stand-alone, pass-fail criteria for all
applications; rather, exceedances of ESBs could trigger
collection of additional assessment data.
6-3
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Section 7
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17:95-101.
7-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Keilty TJ, White DS, Landrum PF 1988b. Sublethal
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Korn S, Earnest RD. 1974. Acute toxicity of 20
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Landrum P. 1991. Memorandum to W Berry, U.S.
Environmental Protection Agency, Narragansett, RI,
May 13, 1991. 1pp.
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Mabey WR, Smith JH, Podoll RT, Johnson HL, Mill T,
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Macek KJ, Hutchinson C, Cope OB. 1969. Effects of
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MackayD, Powers B. 1987. Sorptionofhydrophobic
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Mayer FL, Ellersieck MR. 1986. Manual of acute
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Acute toxicities of selected herbicides to fingerling
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McLeese DW, Metcalfe CD. 1980. Toxicities of eight
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McLeese DW, Burridge LE, Dinter DJ. 1982.
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Nebeker AV, Schuytema GS, Griffis WL, Barbitta JA,
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PoirierS, CoxD. 1991. Memorandum to R. Spehar,
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11, 1991.7pp.
Rapaport RA, EisenreichSJ. 1984. Chromatographic
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
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7-5
-------
Appendix A
Summary of Acute Values for Endrin used to
Calculate the WQC FCV for
Freshwater and Saltwater Species
-------
Appendix A
LC50/EC506 Og/L)
T if- HMAV rh-nii
Common Name, Llfe; f g h
Scientific Name stage Habitat Method Concentration Test Species Genus GMAV
FRESHWATER SPECIES
Oligochaete A I FT M >165.1
worm,
Lumbriculus
variegatus
Oligochaete A I FT M >165.0 >165.0 >165.0 >165.0
worm,
Lumbriculus
variegatus
Cladoceran, X W,E S U 26
Simocephalus
serrulatus
Cladoceran, X W,E S U 45 34.20 34.20 34.20
Simocephalus
serrulatus
Cladoceran, L W S U 4.2
Daphnia
magna
Cladoceran, LWS U 74
Daphnia
magna
Cladoceran, LWS U 41 _ _ _
Daphnia
magna
Cladoceran, L W FT M 230
Daphnia
magna
Cladoceran, L W FT M 88 142.3
Daphnia
magna
Cladoceran, LWS U 20 20 53.35 53.35
Daphnia
pulex
Ostracod, A I,E S U 1.8 1.8 1.8 1.8
Cypridopsis
sp.
Sowbug, A E S U 1.5 1.5 1.5 1.5
Asellus
brevicaudus
Reference
Poirier and
Cox, 1991
Brooke,
1993
Sanders
and Cope,
1966;
Mayer and
Ellersieck,
1986
Sanders
and Cope,
1966;
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Thurston et
al, 1985
Thurston et
al., 1985
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Sanders,
1972;
Mayer and
Ellersieck,
1986
A-2
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Common Name,
Scientific Name
Scud,
Gammarus
fasciatus
Scud,
Gammarus
fasciatus
Scud,
Gammarus
fasciatus
Scud,
Gammarus
lacustris
Glass shrimp,
Palaemonetes
kadiakensis
Glass shrimp,
Palaemonetes
kadiakensis
Crayfish,
Orconectes
immunis
Crayfish,
Orconectes
nais
Crayfish,
Orconectes
nais
Mayfly,
Baetis sp.
Mayfly,
Hexagenia
bilineata
Mayfly,
Hexagenia
bilineata
LC50/EC506 Oig/L)
Lif- HMAV 0-crall
stage Habitat Method Concentration Test Species Genus GMAV Reference
A E S U 4.3 Sanders,
1972;
Mayer and
Ellersieck,
1986
X E S U 1.3 Sanders,
1972;
Mayer and
Ellersieck,
1986
X E FT U 5.5 3.133 Sanders,
1972
A E S U 3.0 3.0 3.066 3.066 Sanders,
1972;
Mayer and
Ellersieck,
1986
A E S U 3.2 Sanders,
1972;
Mayer and
Ellersieck,
1986
X E FT U 0.5 1.265 1.265 1.265 Sanders,
1972;
Mayer and
Ellersieck,
1986
J E FT M >89 >89 Thurston et
al., 1985
X E S U 320 Sanders,
1972;
Mayer and
Ellersieck,
1986
J E S U 3.2 3.2 3.2 16.88 Sanders,
1972;
Mayer and
Ellersieck,
1986
J I S U 0.90 0.90 0.90 0.90 Mayer and
Ellersieck,
1986
J I S U 62 62.99 62.99 62.99 Mayer and
Ellersieck,
1986
X I S U 64 Sanders,
1972
A-3
-------
Appendix A
Common Name,
Scientific Name
Stonefly,
Acroneuria sp.
Stonefly,
Pteronarcella
badia
Stonefly,
Pteronarcys
California*
Stonefly,
Claassenia
sabulosa
Stonefly,
Claassenia
sabulosa
Caddis fly,
Brachycentrus
americanus
Damesfly,
Ischnura
verticalus
Damesfly,
Ischnura
verticalus
Damesfly,
Ischnura
verticalus
Midge,
Tanytarsus
dissimilis
Diptera,
Tipula sp.
Diptera,
Atherix
variegata
Coho salmon,
Oncorhynchus
kisutch
LC50/EC506 Og/L)
T if- HMAV rh-nii
stage Habitat Method Concentration Test Species Genus GMAV Reference
L W,E S U >0.18 >0.18 >0.18 >0.18 Mayer and
Ellersieck,
1986
L I,E S U 0.54 0.54 0.54 0.54 Sanders
and Cope,
1968;
Mayer and
Ellersieck,
1986
A I,E S U 0.25 0.25 0.25 0.25 Sanders
and Cope,
1968;
Mayer and
Ellersieck,
1986
J W,E S U 0.76 Sanders
and Cope,
1968
J W,E S U 0.76 0.2403 0.2403 0.2403 Mayer and
Ellersieck,
1986
X E FT M 0.34 0.34 0.34 0.34 Anderson
and DeFoe,
1980
X W,E S U 1.8 Sanders,
1972
J W,E S U 2.1 Mayer and
Ellersieck,
1986
J W,E S U 2.4 2.086 2.086 2.086 Mayer and
Ellersieck,
1986
L I FT M 0.83 0.83 0.83 0.83 Thurston et
al, 1985
J I,E S U 12 12 12 12 Mayer and
Ellersieck,
1986
J I,E S U 4.6 4.6 4.6 4.6 Mayer and
Ellersieck,
1986
J W S U 0.51 Katz, 1961
A-4
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
LC50/EC506 Oig/L)
HMAV
Common Name, be d fa
Scientific Name stage Habitat Method Concentration Test Species Genus
Coho salmon, J W S U 0.089
Oncorhynchus
kisutch
Coho salmon, J W S U 0.27 0.2306
Oncorhynchus
kisutch
Cutthroat trout, J W S U >1.0 >1.0
Oncorhynchus
clarki
Rainbow trout, J W S U 0.74
Oncorhynchus
mykiss
Rainbow trout, J W S U 0.75
Oncorhynchus
mykiss
Rainbow trout, J W S U 0.75
Oncorhynchus
mykiss
Rainbow trout, J W S U 2.4
Oncorhynchus
mykiss
Rainbow trout, J W S U 1.4
Oncorhynchus
mykiss
Rainbow trout, J W S U 1.11
Oncorhynchus
mykiss
Rainbow trout, J W S U 1.1
Oncorhynchus
mykiss
Rainbow trout, J W S U 0.58
Oncorhynchus
mykiss
Rainbow trout, J W S U 0.90
Oncorhynchus
mykiss
Rainbow trout, J W FT M 0.33 0.33
Oncorhynchus
mykiss
Chinook J W S U 1.2
salmon,
Oncorhynchus
tshawytscha
Overall
GMAV Reference
Mayer and
Ellersieck,
1986
Katz and
Chad wick,
1961
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Macek et
al, 1969
Katz, 1961
Katz and
Chad wick,
1961
Thurston et
al., 1985
Katz, 1961
A-5
-------
Appendix A
Common Name,
Scientific Name
Chinook
salmon,
Oncorhynchus
tshawytscha
Goldfish,
Carassius
auratus
Goldfish,
Carassius
auratus
Goldfish,
Carassius
auratus
Carp,
Cyprinus
carpio
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
A-6
LC50/EC506 Og/L)
T if- HMAV rh-nii
stage Habitat Method Concentration Test Species Genus GMAV Reference
J W S U 0.92 1.051 >0.5318 >0.5318 Katz and
Chadwick,
1961
JW S U 2.1 Henderson
etal., 1959
J W FT U 0.44 Mayer and
Ellersieck,
1986
J W FT M 0.95 0.95 0.95 0.95 Thurston et
al., 1985
J W FT U 0.32 0.32 0.32 0.32 Mayer and
Ellersieck,
1986
J W S U 1.1 Henderson
etal., 1959
JW S U 1.4 __ _ Henderson
etal., 1959
L W S U 0.7 Jarvinenet
al., 1988
J W S U 1.8 Mayer and
Ellersieck,
1986
J W FT U 0.24 Mayer and
Ellersieck,
1986
J W FT M 0.50 Brungsand
Bailey,
1966
U FT M 0.49 Brungsand
Bailey,
1966
J W FT M 0.40 Brungsand
Bailey,
1966
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Common Name,
Scientific Name
Fathead
minnow,
Pimephales
promelas
Fathead
minnow,
Pimephales
promelas
Black
bullhead,
Ictalurus
melas
Black
bullhead,
Ictalurus
melas
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Flag fish,
Jordanella
floridae
Mosquitofish,
Gambusia
affinis
Mosquitofish,
Gambusia
affinis
LC50/EC506 Oig/L)
Life HMAV 0-crall
stage Habitat Method Concentration Test Species Genus GMAV Reference
J W FT M 0.45 Brungsand
Bailey, 1966
J W FT M 0.64 0.4899 0.4899 0.4899 Thurston et
al, 1985
J W,E S U 1.13 Mayer and
Ellersieck,
1986
J W,E FT M 0.45 0.45 Anderson
and DeFoe,
1980
J W,E S U 0.32 Mayer and
Ellersieck,
1986
J W,E S U 1.9 Mayerand
Ellersieck,
1986
J W,E S U 0.8 McCorkleet
al., 1977
J W,E FT M 0.43 Thurston et
al., 1985
J W,E FT M 0.41 0.4199 0.4347 0.4347 Thurston et
al., 1985
J W FT M 0.85 0.85 0.85 0.85 Hermanutz,
1978;
Hermanutz
etal., 1985
J W S U 1.1 Mayerand
Ellersieck,
1986
X W S U 0.75 Katzand
Chadwick,
1961
A-7
-------
Appendix A
Common Name,
Scientific Name
Mosquitofish,
Gambusia
af finis
Guppy,
Poecilia
reticulata
Guppy,
Poecilia
reticulata
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
Bluegill,
Lepomis
macrochirus
LC50/EC506 Og/L)
T if- HMAV rh-nii
stage Habitat Method Concentration Test Species Genus GMAV Reference
J W FT M 0.69 0.69 0.69 0.69 Thurston et
al, 1985
X W S U 0.90 Katzand
Chadwick,
1961
X W S U 1.6 1.200 1.200 1.200 Henderson
etal., 1959
J W S U 0.60 Katzand
Chadwick,
1961
J W S U 8.25 Katzand
Chadwick,
1961
J W S U 5.5 Katzand
Chadwick,
1961
J W S U 2.4 Katzand
Chadwick,
1961
J W S U 1.65 Katzand
Chadwick,
1961
J W S U 0.86 Katzand
Chadwick,
1961
J W S U 0.33 Katzand
Chadwick,
1961
J W S U 0.61 Maceket
al., 1969;
Mayer and
Ellersieck,
1986
J W S U 0.41 Maceket
al., 1969;
Mayer and
Ellersieck,
1986
J W S U 0.37 Maceket
al., 1969;
Mayer and
Ellersieck,
1986
A-8
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
LC50/EC506 Oig/L)
Life HMAV
Common Name, a b c d f
Scientific Name stage Habitat Method Concentration Test Species Genus
Bluegill, J W S U 0.53
Lepomis
macrochirus
Bluegill, J W S U 0.73
Lepomis
macrochirus
Bluegill, J W S U 0.68
Lepomis
macrochirus
Bluegill, J W S U 0.19
Lepomis
macrochirus
Bluegill, J W S U 0.66
Lepomis
macrochirus
Bluegill, U S U 0.61
Lepomis
macrochirus
Bluegill, J W FT M 0.19
Lepomis
macrochirus
Bluegill, J W FT M 0.23
Lepomis
macrochirus
Largemouth J W S U 0.31 0.31 0.31
bass,
Micropterus
dolomieu
Yellowperch, J W FT U 0.15 0.15 0.15
Perca
flavescens
Tilapia, J W S U <5.6 <5.6 <5.6
Tilapia
mossambica
Bullfrog, L E FT M 2.5 2.5
Rana
catesbiana
Southern E W FT M 25 25 2.5(E)
leopard frog, 25(W)
Rana
sphenocephala
Fowler's toad, L E S U 120 120 120
Bufofirwleri
Overall
GMAV Reference
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Henderson
etal, 1959
Sanders,
1972
Thurston et
al, 1985
Thurston et
al., 1985
0.31 Mayer and
Ellersieck,
1986
0.15 Mayer and
Ellersieck,
1986
<5.6 Mayer and
Ellersieck,
1986
Thurston et
al., 1985
7.906 Hall and
Swineford,
1980
120 Mayer and
Ellersieck,
1986
A-9
-------
Appendix A
Common Name, ' e" , .
Scientific Name stage Habitat Method Concentration
Western L E S U
chorus frog,
Psuedocris
triseriata
SALTWATER SPECIES
Eastern oyster, E,L W S U
Crassostrea
virginica
Sand shrimp, A E S U
Crangon
septemspinosa
Hermit crab, A E S U
Pagurus
longicarpus
Korean A W,E S U
shrimp,
Palaemon
macrodactylus
Korean A W,E FT U
shrimp,
Palaemon
macrodactylus
Grass shrimp, L W FT M
Palaemonetes
pugio
Grass shrimp, J W FT M
Palaemonetes
pugio
Grass shrimp, A W,E FT M
Palaemonetes
pugio
Grass shrimp, A W,E FT M
Palaemonetes
pugio
Grass shrimp, A W,E S U
Palaemonetes
vulgar is
Pink shrimp, A I,E FT M
Penaeus
duorarum
American eel, J E S U
Anguilla
rostrata
LC50/EC506 Og/L)
HMAV ,,
f g h
Test Species Genus GMAV Reference
180 180 180 180 Mayer and
Ellersieck,
1986
7901 790 790 790 Davis and
Hidu, 1969
1.7 1.7 1.7 1.7 Eisler,
1969
12 12 12 12 Eisler,
1969
4.7 Schoettger,
1970
0.3 1.187 1.187 1.187 Schoettger,
1970
1.2 Tyler-
Schroeder,
1979
0.35 Tyler-
Schroeder,
1979
0.69 Tyler-
Schroeder,
1979
0.63 0.6536 Schimmel
et al., 1975
1.8 1.8 1.085 1.085 Eisler,
1969
0.037 0.037 0.037 0.037 Schimmel
et al., 1975
0.6 0.6 0.6 0.6 Eisler,
1969
A-10
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Common Name,
Scientific Name
Chinook
salmon,
Oncorhynchus
tshawytscha
Sheepshead
minnow,
Cyprinodon
variegatus
Sheepshead
minnow,
Cyprinodon
variegatus
Sheepshead
minnow,
Cyprinodon
variegatus
Sheepshead
minnow,
Cyprinodon
variegatus
Mummichog,
Fundulus
heteroclitus
Mummichog,
Fundulus
heteroclitus
Striped
killifish,
Fundulus
majalis
Sailfm molly,
Poecilia
latipinna
Atlantic
silverside,
Menidia
menidia
Threespine
stickleback,
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
LC50/EC506 Oig/L)
Life HMAV 0-crall
stage Habitat Method Concentration Test Species Genus GMAV Reference
J W FT U 0.048 0.048 0.048 0.048 Schoettger,
1970
J W,E FT M 0.37 Hansen et
al., 1977
J W,E FT M 0.34 Hansen et
al., 1977
A W,E FT M 0.36 Hansen et
al., 1977
J W,E FT M 0.38 0.3622 0.3622 0.3622 Schimmel
etal., 1975
A W,E S U 0.6 Eisler,
1970b
A W,E S U 1.5 0.9487 Eisler,
1970b
J W,E S U 0.3 0.3 0.5334 0.5334 Eisler,
1970b
A W FT M 0.63 0.63 0.63 0.63 Schimmel
etal., 1975
J W S U 0.05 0.05 0.05 0.05 Eisler,
1970b
J W,E S U 1.65 Katzand
Chadwick,
1961
J W,E S U 1.50 Katzand
Chadwick,
1961
A-ll
-------
Appendix A
Common Name,
Scientific Name
Threespine
stickleback,
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
Striped bass,
Morone
saxatilis
Shiner perch,
Cymatogaster
aggregata
Shiner perch,
Cymatogaster
aggregata
Dwarf perch,
Micrometrus
minimus
Dwarf perch,
Micrometrus
minimus
Bluehead,
Thalassoma
bifasciatum
LC50/EC506 Og/L)
T if- HMAV rh-nii
stage Habitat Method Concentration Test Species Genus GMAV Reference
J W,E S U 1.20 Katzand
Chad wick,
1961
J W,E S U 1.57 Katzand
Chad wick,
1961
J W,E S U 1.57 Katzand
Chad wick,
1961
J W,E S U 0.44 Katz, 1961
J W,E S U 0.50 1.070 1.070 1.070 Katz, 1961
J E FT U 0.094 0.094 0.094 0.094 Korn and
Earnest,
1974
J W S U 0.8 Earnest
and
Benville,
1972
J W FT U 0.12 0.3098 0.3098 0.3098 Earnest
and
Benville,
1972
A W S U 0.6 Earnest
and
Benville,
1972
A W FT U 0.13 0.2793 0.2793 0.2793 Earnest
and
Benville,
1972
A W S U 0.1 0.1 0.1 0.1 Eisler,
1970b
A-12
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Endrin
Common Name,
Scientific Name
Life-
LC50/EC50 (,ug/L)
HMAV
Overall
stage Habitat Method Concentration Test Species Genus GMAV Reference
Striped mullet,
Mugil
cephalus
Northern
puffer,
Sphaeroides
maculatus
A
A W
U
U
0.3 0.3
3.1 3.1
0.3
3.2
0.3 Eisler,
1970b
3.1
Eisler,
1970b
A-13
-------
Appendix B
Summary of Data from Sediment-Spiking
Experiments with Endrin (Data from these
experiments were used to calculate Koc values
(Figure 2-2) and to compare mortalities of amphipods
with interstitial water toxic units (Figure 4-1) and
predicted sediment toxic units (Figure 4-2)).
-------
Appendix B
Sediment Concentration 0-ig/g)
Sediment Source, Mortality
Species tested (%)
Soap Creek Pond
No. 7, OR
Hyalella azteca
1 : 1 Mixture Soap
Creek Pond And
Mercer Lake, OR
Hyalella azteca
Mercer Lake, OR
Hyalella azteca
Soap Creek
Pond, OR
Hyalella azteca
Mercer Lake, OR
Hyalella azteca
Mercer Lake, OR
Hyalella azteca
Lake Michigan
Diporeia sp.
20
32
90
100
100
9
44
95
100
100
5
2
52
100
100
1.5
8.5
100
100
100
10
5
25
45
100
100
2.5
12.5
10
100
100
Dry
Weight Organic Carbon
2.2
3.4
8.1
17.9
45.9
1.1
4.9
17.7
31.7
56.4
1.1
1.3
6.7
26.8
73.8
3.0
8.7
19.6
40.4
62.1
2.0
5.3
13.3
13.3
100
267
1.3
1.3
8.0
20.0
66.7
0.012b
0.171b
0.224b
73
113
270
597
1,530
18
80
290
520
924
10
12
60
239
659
100
290
653
1,350
2,070
18
48
121
121
909
2,430
12
12
73
182
606
17b
31b
13b
Interstitial Water
Concentration
1.1
1.5
4.7
9.8
23.8
0.5
1.7
6.8
10.6
24.5
0.3
0.3
2.3
7.2
15.6
1.1
3.1
6.1
13.9
22.2
0.4
1.0
2.4
3.2
20.1
65.0
0.3
0.2
0.8
3.9
10.8
1.07
2.20
0.63
TOC b
(%) Log Koc References
3.0
3.0
3.0
3.0
3.0
6.1
6.1
6.1
6.1
6.1
11.2
11.2
11.2
11.2
11.2
3.0
3.0
3.0
3.0
3.0
11.0
11.0
11.0
11.0
11.0
11.0
11.0
11.0
11.0
11.0
11.0
0.07
0.55
1.75
MEAN
SE
4.82 Nebekeretal.,
4.88 1989
4.76
4.78
4.81
4.56 Nebeker et al . ,
4.67 1989
4.63
4.69
4.58
4.59 Nebekeretal.,
4.60 1989
4.42
4.52
4.63
4.96 Schuytemaet
4.97 al., 1989
5.03
4.99
4.97
4.65 Schuytema et
4.68 al., 1989
4.70
4.58
4.66
4.57 Schuytemaet
4.60 al., 1989
4.60
4.96
4.67
4.75
4.20 Stehly, 1992
4.15
4.31
= 4.67
= 0.04
B-2
-------
Appendix C
Quality Assurance Summary for the ESB Document:
Procedures for the derivation of equilibrium
partitioning sediment benchmarks (ESBs)
for the protection of benthic organisms: Endrin
c-i
-------
All data were obtained either from the WQC document for endrin (USEPA, 1980) or from a
comprehensive literature search completed in 1995.
All data used in the example benchmark calculations were evaluated for acceptability using the
procedures outlined in the Stephan et al. (1985): Guidelines for deriving numerical national
water quality criteria for the protection of aquatic organisms and their uses. Data not
meeting the criteria were rej ected. All calculations were made using the procedures in Stephan
et al. (1985). All calculations were checked by at least one other EPA scientist and then the
document was distributed for public comment. All data and intermediate values are presented in
tables in the document, and all original data were made available as part of the public comment
process. Any errors of omission or calculation discovered during the public comment process
were corrected and included in the revised document and can be found in Comment Response
Document for the Proposed Equilibrium Partitioning Sediment Guidelines for the Protection
ofBenthic Organisms. Office of Water, Office of Science & Technology, (U.S. EPA, 2000).
Hard copies of all literature cited in this document reside at ORD/NHEERL Atlantic Ecology
Division - Narragansett, Rhode Island.
C-2
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