oEPA
United States Office of Science and Technology and
Environmental Protection Office of Research and Development
Agency Washington, DC 20460
Equilibrium Partitioning
Sediment Guidelines (ESGs)
for the Protection of Benthic
Organisms: Endrin
>.-;-: s
f * * v* * V T.': -*-^T~ ' '^ r
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&EPA
United States
Environmental Protection
Agency
Office of Science and Technology and
Office of Research and Development
Washington, DC 20460
Equilibrium Partitioning
Sediment Guidelines (ESGs)
for the Protection of Benthie
Organisms: Endrin
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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 meet the objectives of the CWA, EPA has periodically issued ambient water
quality criteria (WQC) beginning with the publication of "Water Quality Criteria, 1972" (NAS,
1973). The development of WQC is authorized by Section 304(a)(l) of the CWA, which directs
the Administrator to develop and publish "criteria" reflecting the latest scientific knowledge on
(1) the kind and extent of effects on human health and welfare, including effects on plankton, fish,
shellfish, and wildlife, that may be expected from the presence of pollutants in any body of water,
Including ground water; and (2) the concentration and dispersal of pollutants on biological
community diversity, productivity, and stability. All criteria guidance through late 1986 was
summarized in an EPA document entitled "Quality Criteria for Water, 1986" (U.S. EPA, 1987).
Updates on WQC documents for selected chemicals and new criteria recommendations for other
pollutants have been more recently published as "National Recommended Water Quality Criteria-
Correction" (U.S. EPA, 1999). EPA will continue to update the nationally recommended WQC
as needed in the future.
In addition to the development of WQC and to continue to meet the objectives of the CWA, EPA
has conducted efforts to develop and publish equilibrium partitioning sediment guidelines (ESGs)
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. These guidelines are
authorized under Section 304(a)(2) of the CWA, which directs the Administrator to develop and
publish information on, among other things, the factors necessary to restore and maintain the
chemical, physical, and biological integrity of all navigable waters.
The ESGs and associated methodology presented in this document are EPA's best recommendation
as to the concentrations of a substance that may be present in sediment while still protecting
benthic organisms from the effects of that substance. These guidelines 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 ESGs 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 ESGs do not protect against synergistic or antagonistic effects of
contaminants or bioaccumulative effects to benthos. They are not protective of wildlife or human
health endpoints.
EPA recommends that ESGs be used 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. EPA is developing guidance to assist in
the application of these guidelines in water-related programs of the States and this Agency.
This document provides guidance to EPA Regioas, States, the regulated community, and the
public. It is designed to implement national policy concerning the matters addressed. It does not,
however, 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 guidance where appropriate. EPA may change this guidance in the future.
ill
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This document has been reviewed by EPA's Office of Science and Technology (Health and
Ecological Criteria Division, Washington, DC) and Office of Research and Development (Mid-
Continent Ecology Division, Duluth, MN; Atlantic Ecology Division, Narragansett, RI), and
approved for publication.
i
Mention of trade names or commercial products does not constitute endorsement or
recommendation of use.
Front cover image provided by Wayne R. Davis and Virginia Lee.
IV
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Contents
Acknowledgments ..! be
~ Executive Summary xi
Glossary Xiii
Section 1
Introduction 11
1.1. General Information 1-1
12 General Information: Endrin 1-2
13 Applications of Sediment Guidelines 1-4
1.4 Overview .>. ......: 1-4
Section 2
Partitioning 2-1
2.1 Description of EqP Methodology 2-1
22 Determination of KQvf for Endrin 2-2
23 Derivation of KQC from Adsorption Studies 2-2
23.1 AT0,, from Particle Suspension Studies 2-2
2.3.2 K^ from Sediment Toxicity Tests 2-3
2.4 Summary of Derivation of K^ for Endrin ....2-4
Section 3
Toxicity of Endrin in Water Exposures 3-1
3.1 Derivation of Endrin WQC 3-1
32 Acute Toxicity in Water Exposures 3-1
33 Chronic Toxicity in Water Exposures 3-1
3.4 Applicability of the WQC as the Effects Concentration
for Derivation of the Endrin ESG 3-5
Section 4
Actual and Predicted Toxicity of Endrin
in Sediment Exposures 4-1
4.1 Toxicity of Endrin in Sediments 4-1
42 Correlation Between Organism Response and Interstitial Water Concentration 44
43 Tests of the Equilibrium Partitioning Prediction of Sediment Toxicity 4-6
Section 5
Guidelines Derivation for Endrin 5-1
5.1 Guidelines Derivation 5-1
52 Uncertainty Analysis 5-2
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Tables
Table 2-1. Endrin measured and estimated log,0ATow values 2-2
Table 3-1. Test-specific data for chronic sensitivity of freshwater and saltwater
organisms to endrin 3.4
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
Table4-l. Summary of tests with endrin-spiked sediment 4-2
Table4-2. 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 guidelines (ESGs) for endrin 5-1
Table 5-2. Analysis of variance for derivation of confidence limits of the ESGs for endrin 5-3
Table 5-3. Confidence limits of the ESGs for endrin 5-3
Figures
Figure 1-1. Chemical structure and physical-chemical properties of endrin 1-3
Figure 2-1. Observed versus predicted partition coefficients for nonionic organic chemicals
using Equation 2-4 2-3
Figure2-2. Organic carbon-normalized sorption isotherm for endrin and probability plot
of AQJ, from sediment toxicity tests 2-4
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
Figure4-1. 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
VII
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53 Comparison of Endrin ESG and Uncertainty Concentrations to Sediment
Concentrations that are Toxic or Predicted to be Chronically Acceptable 5-3
5.4 Comparison of Endrin ESG to STORET and Corps of Engineers,
San Francisco Bay Databases for Sediment Endrin .5-6
55 Limitations to the Applicability of ESGs t 5-9
Section 6
Guidelines Statement
Section 7
References
6-1
7-1
Appendix A A-i
Appendix B B-i
VI
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Figure 5-1. Predicted genus mean chronic values calculated from water-only toxicity values
using freshwater species versus percentage rank of their sensitivity S4
Figure 5-2. Predicted genus mean chronic values calculated from water-only toxicity values
using saltwater species versus percentage rank of their sensitivity ป 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 ESG 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 San Francisco Bay ...5-8
VHl
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Acknowledgments
Coauthors
Walter J. Berry*
David J. Hansen
Dominic M. Di Toro
Laurie D, De Rosa
Heidi E. Bell*
Mary C. Re iky
Frank E. Stancil, Jr.
Christopher S. Zarba
Robert L, Spehar
U.S. EPA, NHEERL, Atlantic Ecology Division,
Narragansett, RJ
HydroQual, Inc., Mahwah, NJ; Great Lakes Environmental
Center, Traverse City, MI (formerly with U.S. EPA)
Manhattan College, Riverdale, NY; HydroQual, Inc.,
Mahwah, NJ
HydroQual, Inc., Mahwah, NJ
U.S. EPA, Office of Water, Washington, DC
U.S. EPA, Office of Water, Washington, DC
U.S. EPA, NERL, Ecosystems Research Division, Athens, GA
U.S. EPA, Office of Research and Development, Washington, DC
U.S. EPA, NHEERL, Mid-Continent Ecology Division,
Duluth, MN
Significant Contributors to the Development of the Approach and Supporting Science
Herbert E. Allen
Gerald T. Ankley
Christina E. Cowan
Dominic M. Di Toro
David J. Hansen
Paul R. Paquin
Spyros P. Pavlou
Richard C. Swartz
Nelson A. Thomas
University of Delaware, Newark, DE
U.S. EPA, NHEERL, Mid-Continent Ecology Division,
Duluth, MN
The Procter & Gamble Co., Cincinnati, OH
Manhattan College, Riverdale, NY; HydroQual, Inc.,
Mahwah, NJ
HydroQual, Inc., Mahwah, NJ; Great Lakes Environmental
Center, Traverse City, MI (formerly with U.S. EPA)
HydroQual, Inc., Mahwah, NJ
Ebasco Environmental, Bellevue, WA
Environmental consultant (formerly with U.S. EPA)
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 OAO Corporation, Narragansett, RI
Robert A. Hoke E.I. DuPont deNemours and Company, Newark, DE
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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Executive Summary
This equilibrium partitioning sediment guideline (ESG) document recommends a sediment
concentration for the insecticide endrin that is EPA's best estimate of the concentration protective
of the presence of benthic organisms. The equilibrium partitioning (EqP) approach was chosen
because it accounts for the varying biological availability of chemicals indifferent sediments and
allows for incorporation of the appropriate biological effects concentration. This provides for the
derivation of a guideline that is 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 concentration 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^), which is a constant for each chemical. The ESG Technical Basis
Document (U.S. EPA, 2000a) demonstrates that biological responses of benthic organisms lo
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 ;^g
chemical/g organic carbon basis (//g/g^). 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 ^g/g^ basis can be accurately predicted by multiplying
the effect concentration in water by the chemical's A^,. Because the water quality criteria
(WQC) is the concentration of a chemical in water that is protective of the presence of aquatic
life, and is appropriate for benthic organisms, the product of the final chronic value (FCV) from
the WQC and K^ represents the concentration in sediments that, on an organic carbon basis, is
protective of beathic organisms. For eodrin this concentration is 5.4 /j.g endrin/g^ for freshwater
sediments and 0.99 MS/goc fฐr saltwater sediments. Confidence limits of 2.4 to li^g/g^ for
freshwater sediments and 0.44 to 2.2 /^g/g,^ for saltwater sediments were calculated using the
uncertainty associated with the degree to which toxicity could be predicted by multiplying the K^
and the water-only effects concentration. The ESG should be interpreted as a chemical
concentration below which adverse effects are not expected. In comparison, at concentrations
above the ESG effects are likely, and 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 ESG and associated uncertainty limits.
These guidelines do not protect against additive, synergistic, or antagonistic effects of
contaminants or bioaccumulative effects to aquatic life, wildlife, or human health. The Agency
and the EPA Science Advisory Board do not recommend the use of ESGs as stand-alone, pass-fail
criteria for all applications; rather, exceedances of ESGs could trigger additional studies at sites
under investigation. This ESG applies only to sediments having iO.2% organic carbon.
EPA has developed both Tier 1 and Tier 2 ESGs to reflect the differing degrees of data availability
and uncertainty. Requirements for a Tier 1 ESG include a ATOW, FCV, and sediment toxicity tests to
verify EqP assumptions. In comparison, a Tier 2 ESG requires a Kovi and a FCV or secondary
chronic value (SCV); sediment toxicity tests are recommended but not required. The ESGs derived
for endrin in this document, as well as the ESGs for dieldrin, metal mixtures (Cd, Cu.Pb, Ni, Ag,
Zn), and polycyclic aromatic hydrocarbon (PAH) mixtures represent Tier 1 ESGs (U.S. EPA,
2000d,e,f). Information on how EPA recommends ESGs be applied in specific regulatory programs
is described in the "Implementation Framework for the Use of Equilibrium Partitioning Sediment
Guidelines (ESGs)" (EPA, 2000c),
XI
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Glossary of Abbreviations
ACR
ANOVA
AR
CFR
CWA
DOC
EC50
EPA
EqP
ESG(s)
ESG
drywt
ESG,
oc
i
FACR
FAV
FCV
FDA
/oc
FRY
GMAV
Soc
HECD
HMAV
IUPAC
IWTU
K.
oc
K,
ow
Acute-chronic ratio i
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 guideline(s); for nonionic organics, this term
usually refers to a value that is organic carbon-normalized (more formally
ESGo,-.) unless otherwise specified
Dry weight-normalized equilibrium partitioning sediment guideline
Organic carbon-normalized equilibrium partitioning sediment guideline
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
International Union of Pure and Applied Chemistry
Interstitial water toxic unit
Organic carbon-water partition coefficient
Octanol-water partition coefficient
XIII
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Kf
LC50
LC50
S oc
NAS
NERL
NHEERL
NOEC
NTIS
OC
OEC
OST
PAH
PGMCV
PSTU
SE '
SMACR
STORET
TOC
TU
WQC
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
Standard error
Species mean acute-chronic ratio
EP A ' s computerized database for STOrage and RETrieval of
water-related data
Total organic carbon
Toxic unit
Water quality criteria
XIV
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Section 1
Introduction
1.1 General Information
Under the Clean Water Act (CWA) the U.S.
Environmental Protection Agency (EPA) is responsible
for protecting the chemical, physical, and biological
integrity of the nation's waters. In keeping with this
responsibility, EPA published ambient water quality
criteria (WQC) in 1980 for 64 of the 65 toxic pollutants
or pollutant categories designated as toxic in the
CWA. Additional water quality documents that update
criteria for selected consent decree chemicals and new
criteria have been published since 1980. These WQC
are numerical concentration limits that are EPA's best
estimate of concentrations protective of human health
and the presence and uses of aquatic life. Although
these WQC play an important role in ensuring a
healthy aquatic environment, they alone are not
sufficient to ensure the protection of environmental or
human health.
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
sediment guidelines makes 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) established a research
team to review alternative approaches (Chapman,
1987). All of the approaches reviewed bad both
strengths and weaknesses, and no single approach was
found to be applicable for guidelines derivation in all
situations (U.S. EPA, 1989a). The equilibrium
partitioning (EqP) approach was selected for nonionic
organic chemicals because it presented the greatest
promise for generating defensible, national, numerical
chemical-specific guidelines applicable across a broad
range of sediment types. The three principal
observations that underlie the EqP approach of
establishing sediment guidelines 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 final chronic values (FCV) can be used to
define the acceptable effects concentration of a
chemical freely-dissolved hi 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 (K^) and effects concentrations in water;
(5) the FCV concentration is an appropriate effects
concentration for freely-dissolved chemical hi
interstitial water; and (6) the equilibrium partitioning
1-1
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sediment guideline (ESG) derived as the product of the
A^ and FCV is protective of benthic organisms. ESG
concentrations presented in this document are
expressed as /ig chemical/g sediment organic carbon
O^g/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 guidelines generated using the EqP
approach (i.e., ESGs) are suitable for use in providing
guidance to regulatory agencies because they are:
1. Numerical values
2. Chemical specific
3. Applicable to most sediments
4. Predictive of biological effects
5. Protective of benthic organisms
ESGs 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, ESGs 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. ESGs 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 ESGs under these conditions may result
in values lower or higher than those presented in this
document.
not yet been reached, sediment chemical concentrations
less than the ESG 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 the ESG 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, the national ESG would be overly
protective of benthic organisms and should not be used
unless modified using the procedures outlined In
"Methods for the Derivation of Site-Specific
Equilibrium Partitioning Sediment Guidelines (ESGs)
for the Protection of Benthic Organisms" (U.S. EPA,
2000b). The ESG may be underprotective where the
toxiciry of other chemicals are additive with the ESG
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 the
ESG for eodiin. The data that support the EqP
approach for deriving an ESG for nonionic organic
chemicals are reviewed by Di Toro et al. (1991) and
EPA (U.S. EPA, 2000a). Before proceeding through
the following test, tables, and calculations, the reader
should consider reviewing "Guidelines for Deriving
Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and Their Uses"
(Stephanetal., 1985), "Response to Public Comment"
(U.S. EPA, 1985), and "Technical Basis for the
Derivation of Equilibrium Partitioning Sediment
Guidelines (ESGs) for the Protection of Benthic
Organisms: Nonionic Organics" (U.S. EPA, 2000a).
Guidance for the acceptable use of the ESG values is
contained in "Implementation Framework for Use of
Equilibrium Partitioning Sediment Guidelines (ESGs)"
(U.S.EPA,2000c).
The ESG values presented herein represent EPA's
best recommendation of the concentration of endrin in
sediment that will not adversely affect most benthic
organisms. EPA recognizes that these ESG 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
1.2 General Information: Endrin
Endrin is the common name of a "broad spectrum"
organochlorine iosecticide/rodenticide. It was
formulated for use as an emulsifiable concentrate, as a
wettable or dusiable powder, or as a granular product.
It has been used with a variety of crops including
cotton, tobacco, sugar cane, rice, and ornamentals.
1-2
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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 ฐC and specific gravity of 1.7 g/cc at
20ฐC. It has a vapor pressure of 0.026 mPa (25ฐC)
(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 j/g/L for
freshwater organisms and 0.037 to 790 /^g/L for
saltwater organisms (Appendix A). There is little
difference between the acute and chronic toxicity of
endrin to 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
MOLECULAR FORMULA
MOLECULAR WEIGHT
DENSITY
MELTING POINT
PHYSICAL FORM
VAPOR PRESSURE
380.93
1.70 g/cc (20ฐQ
235 ฐC
Colorless crystal
0.026 mPa (25ฐC)
CAS NUMBER:
TSL NUMBER:
COMMON NAME:
TRADE NAME:
CHEMICAL NAME:
72-20-8
IO 15750
Endrin (also endrine and nendrin)
End rei (Shell); Hexadrin
1,2,3,4,10,10, hexachloro-lR, 4S, 4aS, 5nS, 6,7R, 8R, 8aR-
octahydro-6,7-epoiy-l, 4:5,8-dimethanonaphthalene (IUPAQ
or Hexachlorocpoxy-octahydro-endo-endo-di methanonaphthalene
Figure 1-1. Chemical structure and physical-chemical properties of endrin (from Hartley and Kidd, 1987).
1-3
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(FRY) 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 ESG
for endrin is effects based. It is calculated from the
FCV derived in Section 3.
1.3 Applications of Sediment Guidelines
ESGs are meant to be used with direct toxicity
testing of sediments as a method of evaluation. They
provide a chemical-by-chemical specification of what
sediment concentrations are protective of benthic
aquatic life. TheEqP 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), and polycyclic aromatic hydrocarbon (PAH)
mixtures.
EPA has developed both Tier 1 and Tier 2 ESGs to
reflect the differing degrees of data availability and
uncertainty. The minimum requirements to derive a Tier
1 ESG include (1) an octanol-water partitioning
coefficient (Kow) of the chemical, measured with
current experimental techniques, which appears to
remove the large variation in reported values; (2)
derivation of the FCV, which should also be updated to
include the most recent lexicological information; and
(3) sediment toxicity "check" tests to verify EqP
predictions. Check experiments can be used to verify
the utility of EqP for a particular chemical. As such, the
ESGs derived for nonionic organics, such as dieldrin
and endrin, metal mixtures, and PAH mixtures represent
Tier 1 ESGs (U.S. EPA, 2000d,e,f). In comparison, the
minimum requirements for a Tier 2 ESG include a Kov
for the chemical (as described above) and the use of
either a FCV or secondary chronic value (SC V). The
performance of sediment toxicity tests is recommended,
but not required for the development of Tier 2 ESGs.
Therefore, in comparison to Tier 1 ESGs, the level of
protection provided by the Tier 2 ESGs would be
associated with more uncertainty due to the use of the
SCV and absence of sediment toxicity tests. Examples
of Tier 2 ESGs for nonionics are found in U.S. EPA
(ZOOOg). Information on how EPA recommends ESGs be
applied in specific regulatory programs is described in
the "Implementation Framework for the Use of
Equilibrium Partitioning Sediment Guidelines (ESGs)"
(EPA,2000c).
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 Koc recommended for use in deriving the endrin
ESG. 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 the ESG
concentration. 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 the ESG. 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 KQC 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 ESG for endrin and its uncertainty. The ESG for
endrin is then compared with two databases on endrin's
environmental occurrence in sediments. Section 6
concludes witi the guideline statement for endrin.
The references cited in this document are listed in
Section?.
1-4
-------
__-^ซ.MM'uii
Section 2
Partitioning
2.1 Description of EqP Methodology
ESGs are the numerical 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. As a result, they can be
used in much the same way as WQC", tbat is, the
concentration of a chemical that is protective of the
intended use, such as aquatic life protection. For
nonionic organic chemicals, ESGs are expressed as pg
chemical/goc and apply to sediments having iO.2 %
organic carbon by dry weight. A brief overview
follows of the concepts that underlie the EqP
methodology for deriving ESGs. The methodology is
discussed in detail in "Technical Basis for the
Derivation of Equilibrium Partitioning Sediment
Guidelines (ESGs) for the Protection of Benthic
Organisms: Nonionic Organics" (U.S. EPA, 2000a),
hereafter referred to as the ESG 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 for
determining ESGs 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 0*g chemical/L interstitial water) and
not with the sediment chemical concentration 0/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 ESGs. For this
reason, the methodology described below is called the
EqP method.
The ESG 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 ESG 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 the ESG.
The ESG is calculated as follows. Let FCV
O^g/L) be the acceptable concentration in water for the
chemical of interest, then compute the ESG using the
partition coefficient, Kf (L/kgsedjroent), between sediment
and water
ESG = KP FCV
(2-1)
This is the fundamental equation used to generate the
ESG. Its utility depends on the existence of a
methodology for quantifying Kf.
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 hi 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 K^-. and the weight fraction of organic
carbon (/Jx;) in the sediment. The relationship is as
follows
^p "~ foe ^o
It follows that
(2-2)
2-1
-------
(2~3)
where ESGOC is die ESG on a sediment organic carbon
basis. For nonionic organics, "ESG" usually refers to a
value that is organic carbon-normalized (more formally
ESGoc) unless otherwise specified.
KQC is not usually measured directly (although it
can be done; see Section 2.3). Fortunately, X^ is
closely related to the octanol-water partition
coefficient (^ow) (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 ATOW for endrin.
2.2
Determination of Kovf for Endrin
Several approaches have been used to determine
^ow ^or ^e derivation of an ESG, as discussed in the
ESG Technical Basis Document. In an examination of
the literature, primary references were found listing
measured log10ATow values for endrin ranging from 4.40
to 5.19 and estimated Iog1(fiovi values ranging from
3.54 to 5 .60 (Table 2-1 ). Karickhoff and Long ( 1 995 ,
1996) established a protocol for recommending ATQW
values for uncharged organic chemicals based on the
best available measured, calculated, and estimated
data. The recommended log,,^^ value of 5.06 for
endrin from Karickhoff and' Long (1995) will be used to
derive the ESG for endrin.
2.3 Derivation of Koc from Adsorption
Studies
Two types of experimental measurements of K^
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 K^; endrin associated
with dissolved organic carbon (DOC) was nqt included.
2.3.1 Kocfrom Particle Suspension Studies
Laboratory studies to characterize adsorption 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
Kf =
(2-4)
where m is the particle concentration in the suspension
(kg/L) and ox, an empirical constant, is 1.4. The KQC
is given by
ฐ-983
ow
(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
Method
Measured
Measured
Measured
Measured
Estimated
Estimated
Estimated
and estimated log19Kow values
Log10A:ow
4.40
4.92
5.01
5.19
354
5.40
5.60
Reference
Rapaport and Eisenreich,
1984
Ellington and Stancil, 1988
Eadsforth, 1986
DeBruijnetai.,1989
Mabeyetal., 1982
Karickhoff et al., 1989
Neeley eta). ,1974
2-2
-------
(DiToro, 1985). The observed partition coefficient for
endrin using adsorption data (Sharom et al., 1980) is
highlighted on this plot. The observed IogloKp 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., log,^ = 2.98 fiomf^^ = 958 L/kg).
In the absence of particle effects, KQ^. is related to KOVJ
via Equation 2-5. For log10ATow = 5.06 (see Section
2.2), this expression results in an estimate of log,
= 4.97.
2.3.2 K^from Sediment Toxicity Tests
Measurements of K^ 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 hi 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 (ug/goc) is plotted versus freely-dissolved
interstitial water concentration (^g/L). The data used
to make this plot are included in Appendix B. The line
of unity slope corresponding to the log,^^ = 4.97
derived from the endrin log10Kow of 5.06 from
Karickhoff and Long (1995) is compared with the data.
A probability plot of the observed experimental
logjg/Toc values is shown in Figure 2-2B. The log^^,
values were approximately normally distributed, with a
mean of logj^^ 4.67 and a standard error of the
mean (SE) of 0.04. This value agrees with the
Iog10 XQC = 4.97, which was computed using the
endrin log10Xow of 5.06 from Karickhoff and Long
(1995) using Equation2-5.
&S
-2
024
Predicted log, A (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
-------
2.4 Summary of Derivation of Koc for
Endrin
The XQC seated to calculate the ESG for endrin
was based on the regression of logj^^ to Iog10ฃ"ow
(Equation 2-5) using the endrin logI0^ow of 5.06 from
Karickhoff and Long (1995). This approach, rather than
use of the AT^ 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 ATOW and/^ values. The regression equation
yieldedalog10A"ocof4.97. This value was ta
agreement with the log,^^ of 4.67 measured in the
sediment toxicity tests.
a
o
a
2
u
a
o
U
u
o
a
I
10000
1000
100
10
- A
0.1
Nebeker et aL, 1989
Schuytcma et aL, 1989
Stehly, 1992
T 11 nra
I I I U
I I I 11111
0.91
9.1 1 10 100
Interstitial Water Concentration
1000
6.0
5.5
5.0
4.5 -
4.0
3.0
TTTtTIffl TTT
B
!
U_i_JWULLL
0.1 1 19 29 50 89 99 99
Probability
Figure 2-2. Organic carbon-normalized sorption isotherm for endrin (A) and probability plot of
K^. (B) from sediment toxkity tests (Nebeker ct aL, 1989; Schuytema et aL, 1989;
Stehly, 1992). The solid line represents the relationship predicted with a logJT^,
of 4.97.
2-4
-------
Section 3
Toxicity of Endrin in
Water Exposures
3.1 Derivation of Endrin WQC
The EqP method for derivation of the ESG for
endrin uses the WQC FCV and K^ 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, 2000a). 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
effects concentration for the endrin ESG derivation.
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 ^g/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
Hg/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 among both 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 ,ug/L for the
midge, Tanytarsus, to > 165 /j.g/L for the oligochaete,
iMmbriculus.
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 A*g/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 1 1 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 ^g/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
3.3 Chronic Toxicity in Water Exposures
Life-cycle toxicity tests have been conducted with
the freshwater flagfish (Jordanella floridae) and fathead
minnow (Pimephales promelas) and with the saltwater
sheepshead minnow (Cyprinodon variegatus) and grass
shrimp (Palaemonetespugio). Each of these species,
except for P. promelas, has one or more bendiic life-
stages.
Two life-cycle toxicity tests have been conducted
with /. 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 pg/L endrin relative to
response of control fish; progeny were unaffected
(Table 3 - 1 ) . Neither parental nor progeny (F,)
generation /, floridae were significantly affected when
exposed to endrin concentrations from 0.051 to 0.22 /ig/
L. The chronic value from this test was 0.2569.
Combined with the 96-hour companion acute value of
0. 85 ptg/L (Hermanutz etal., 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 hi
parental fish exposed to 0.29 //g/L endrin, and
reductions of 73% in survival, 18% in (growth) length,
3-1
-------
1000
100
~
I
3
J 10
rฃ
1
u
8
4)
s. ป
M
a
4)
O
0.1
\
A Arthropods
- Other Invertebrates Pwdocrisw
Fish and Amphibians Lumbiiculta* (A) mr>
Bufo(L)/
- Hexagenia (J) J
Daphnia (L) br
Simocephalus (X) A
li Orconectes (J)
_ y* Tipala (J)
: /* Rana (L)
y Atherix (J)
X Gammarus (A)
Jonltmdla (J) . .^'^ Wi^Cv^^(A)
\/'^OCJ*WJ fj\/Q j^ Jf r \ j*
- Tanylarsvs (I^^.^or Palaemonetes (A)
~ Gambusia (J) cr^ \ Carassius (J)
Panepholes (J)-Y.n-.jฃ Baetis (J)
Brackycentna (X) <^Cr\^ Pleronarcella (L)
Microptenu (J)fy^.Jr\, Oncorhyncfna (J)
^ฑ~i(~ \ Ictalurvs (J)
n^S]] \ CyprinusfJ)
Y 1 1 I Pleronarcys (A)
_ 1 1 ' Claassenia (A)
| Lepomis(J)
| Acroneuria * (L)
Perca(J)
i 1 . ! i 1 , ! i
t.
I
0 20 40 60 80 100
Percentage Rank of Freshwater Genera
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.
and 92% in numbers of eggs per female in 0.39 /ug/L.
No significant effects were detected in parental or
progeny generation flagfish in 0.21 ^g/L. The chronic
value from this test was 0.2468. Combined with the
96-hour companion acute value of 0.85 //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
(Jarvinen and Tyo, 1978). Parental fish in 0.25 ^g/L in
water-only exposures exhibited about 60% mortality
relative to controls. Mortality of Fj progeny was 70%
in 0.14 ^g/L, the lowest concentration tested, and 85 %
in0.25/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 fj.g/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.
3-2
-------
1000
100
3
I
a
8
1
o
10
0.1
A Arthropods
Other Invertebrates
Fish and Amphibians
Crassostrea (EJ.)
Psguras (A)
SphaeroiJes (A)
Gasterosteus (J)
MugU(A)
Micrometres (A)
Anguilla (J)
Funduho (A)
Cyprinodon (J,A)
Cymatogaster (J)
/Crangon (A)
Palaemon (A)
Palaemonetes (A)
1 PoeciUa (A)
Morone (J)
Th&lassoma (A)
Msnidia (J)
Oncorhynchus (J)
Penaeus (A)
20
40
60
SO
ISO
Percentage Rank of Saltwater Genera
Figure 3-2. Genus mean acute values from water-only acute toxicity tests using saltwater species versus
percentage rank of their sensitivity. Symbols representing benthic species are solid; those
representing water column species are open. A - adult, E = embryo, J = juvenile, L = larvae.
Onset of spawning was delayed, duration of spawning
was lengthened, and the number of female P. pugio
spawning was less in all exposure concentrations from
0.03 to 0.79 ngfL. These effects on reproduction may
not be important because embryo production and
hatching success were apparently not affected. Larval
mortality and time to metamorphosis increased and
growth of juvenile progeny decreased in endrin
concentrations ^0.11 Mg/L- The chronic value from this
test was 0.07416. Combined with the 96-hour
companion acute value of 0.35 /ug/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 /ug/L and about half of those exposed to
0.31 //g/L died. Females died during spawning, fewer
eggs were fertile, and survival of exposed progeny
decreased in 0.31 /^g/L. No significant effects were
observed on survival, growth, or reproduction in fish
exposed to 0.027 to 0.12 yug/L endrin. The chronic
value from this test was 0.1929. Combined with the 96-
hour companion acute value of 0.3629 Mg/L (Hansen et
al., 1977; Schimmeletal., 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 fotP.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 (Stephanet al., 1985), also
3-3
-------
Table 3-1. Test-specific data for chronic sensitivity of freshwater and saltwater organisms to endrin
Common
Name,
Scientific
Name
Habitat
(life-
Test stage)
Duration
(days)
NOECs0
O^g/L)
OECsฐ
Og/L)
Observed
Effects
(relative to
controls)
Chronic
Value
O^g/L)
*
Reference
Freshwater Species
Flagfish, LC E(E,L) 110 0.051- 0.30
Jordanella W (J,A) 0.22
floridae
Flagfish, LC E(E,L) 140 0.21 0.29,
Jordanella W (J,A) 0.39
floridae
Fathead LC W 300 <0.14 0.14-
minnow, (E,L,J,A) 0.25
Pimephales
promelas
Saltwater Species
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 Hermanutz,
1978
0.2468 Hermanutz
et al., 1985
<0.14 Jarvinen and
Tyo, 1978
Grass shrimp,
Palaemonetes
pugio
Sheepshead
minnow,
Cyprinodon
variegatus
PLC W (L) 145
E,W
(E,J,A)
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;
15% reduction
in growth and in
adult
reproduction;
87% decrease in
progeny
survival
0.07416 Tyler-
Schroeder,
1979
0.1929 Hansenet
al., 1977
"Test: LC = life-cycle, PLC = partial life-cycle, ELS = early life-stage.
Habitat: I = infauna, E = epibenthic, W = water column. Life-stage: E = embryo, L = larvai, J
cNOECs = no observed effect concentrations; OECs = observed effect concentrations.
; juvenile, A = adult.
indicated little difference between acute and chronic
toxicity of endrin. These include tests with the
caddisfly, Brachycentrus americanus; stonefly,
Pteronarcys dorsata (Anderson and DeFoe, 1980);
bhintaose minnow, Pimephales notatus (Mount, 1962);
fathead minnow, P. promelas (Jarvinen et al., 1988);
brown bullhead, Ictalurus meias (Anderson and DeFoe,
1980); largemouth bass, Micropterus salmoides
(Fabacher, 1976); spot, Leiostomm xanlhurus (Lowe,
1966); and mummichog, Funchdus heteroclitus (Eisler,
1970a).
The final acute value (FAV) derived from the
overall GMAVs (Stephan et al., 1985) for freshwater
organisms was 0.1803 Mg/L. The FAV for saltwater
species was 0.03282 /j.g/L (Table 3-2). The FCVs were
used as the effect concentrations for calculating the
ESG for protection of benthic species. The FCV of
3-4
-------
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
0
-------
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 backrso 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 3-3 A). 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 (2000a). 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 die 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
Freshwater vs Saltwater
Freshwater: Benthic vs Water
rViInmซ -L RontKis* f\JJf~\f~'\
Habitat
Fresh (32)
Benthic (21)
or Water T>pe
Salt (19)
WQC (32)
AR StatisticC
0.149
0.042
Probability
99
7
Saltwater: Benthic vs Water
Column + Benthic (WQC)
Benthic (11)
WQC (19)
0.012
68
"Values in parentheses are the number of LC50 values used in the comparison.
bNote thai in both the freshwater vs. saltwater and beothic 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 s the observed AR statistic, given lhat the samples came from the same population.
3-6
-------
ys
1 u
ti
u 9.1
a
;ซ! -03
to
-0.4
~ฎ.$
* f
W.5
0.4
9> 0.2
3
ซ> 0.1
u
5 0.9
hi
U n <
is -*1
Q -93
^ -ซ3
a c
-0.5
0.1
_ 1 1 lllllll MM Hill 1 1 1 1 1 1"
: A
Freshwater vs Saltwater
-
_
QOGDCC**^
lo
~ 1 1 lllllll 1 1 1 I 1 1111 1 1 1 1 I 1
_ 1 1 lllllll 1 1 1 1 1 Illl 1 1 I 1 1 1
: B
: Benthic vs WQC
Freshwater
_ rplJ 1 ft j ', "i,"-1 a '
h QQcpocxrca****111^
_o
~ 1 1 IHIIH 1 1 I 1 Hill | | | 1 I |
_ 1 1 lllllll 1 1 1 1 1 Illl 1 1 1 1 1 1
: c
: Benthic vs WQC
Saltwater
.
I ฐ
~o
* i i mini i i i 1 1 mi i i i i i i
1 10 20 50
Probability
i "" mill 1 1 i nun H i -
-
-
n o-
^^^^-pjpCDCDO CT0
-
~"
-
1 HIM) 1 1 1 lllllll 1 1 ~
i inn 1 1 i i mini i i
-
-
-
iimTtTimiim i ll'OOOOOO O~
-
_
ซ
_
~
-
i nun i i i IHIIH i i "
i mil 1 1 i i mini i i
-
-
_
_
! l||lll 1 1 1 lllllll I 1 ~
8fl 90 99 993
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
-------
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 ESG 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 >ig endrin/g dry weight sediment and
ranged from 1,050 to 5,400 ng endrin/g dry weight
sediment. Four-day LC50 values for three tests with
Umnodrilus hoffmeisteri averaged 3,390 jig/g dry
weight sediment and ranged from 2,050 to 5,600 /ug/g
dry weight sediment. Four-day LC50 values from
these tests averaged 194,000 ^g/grx- for L. hoffmeisteri
and 121,000 ^glg^ for 5. 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
211.5 /ig/g dry weight sediment and possibly at ^0.54
jjg/g dry weight sediment. EC50s, based on sediment
avoidance, were 59.0 fig/g dry weight (3,371 ^g/goc) for
L. hoffmeisteri and 15.3 and 19.0 ngJg dry weight (874
and 1,086 ^glg^ sediment for two tests using Sr
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 5. heringianus in
11.5 Mg/g dry weight sediment after a 54-day exposure
and 26% mortality in 42.0 /^g/g dry weight sediment.
The sediment reworking rate was reported to be
significantly different from the control in sediments
containing ;:0.54 uglg dry weight sediment. Dry
weights of worms in 2:2.33 uglg dry weight sediment
were reduced after 54 days. Keilty and Stehly (1989)
observed no effect of a single, nominal concentration of
50 Mg/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 atfeca to two endrin-
spiked sediments, one with a TOC of 11 % and the other
a 3 % TOC. 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 uglg) 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/gf)C, a factor of 2.7 (Table 4-1).
4-1
-------
Table 4-1. Summar
y 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
hoffmeisteri
Tubificid worm,
Limnodrilus
hoffmeisteri
Tubificid worm,
Limnodrilus
hoffmeisteri
Tubificid worm,
Limnodrilus
hoffmeisteri
TOC
Sediment Source (%)
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75
0.25mm sieved
Lake Michigan; i.75b
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan ; 1.75
0.25mm sieved
Lake Michigan; 1.75
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Lake Michigan; 1.75b
0.25mm sieved
Method,2
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
Drywt
Response Cซg/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 5,600ฐ
EC50 59.0
sediment
avoidance
Water
OC LC50 ซ
(pg/g) O^g/L) Reference
80,000 Keilty et a!.,
1988a
60,000 Keilty et a!.,
1988a
143,000 Keilty et al.,
1988a
309,000 - Keilty etal.,
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 et al.,
1988b
2,860 Keilty and Stehly,
1989
117,000 Keilty etal.,
1988a
194,000 Keilty etal.,
1988a
320,000ฐ Keilty et al.,
1988a
3,371 Keilty etal.,
1988a
4-2
-------
Table 4-1 . Summary of tests with endrin-spiked
sediment
(continued)
Sediment Endrin LCSO Interstitial
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
TOC
Sediment Source (%)
Cake Michigan; 0.07
depth 29m
Lake Michigan; 0.55
depth 45m
Lake Michigan; 1.75
depth 100m
Soap Creek 3.0
Pond No. 7, OR
1 : 1 mixture of 6. 1
Soap Creek and
Mercer Lake, OR
Mercer Lake, OR 11.2
Soap Creek Pond 3
No. 7, OR;
refrigerated
Soap Creek Pond 3
No. 7, OR; frozen
Mercer Lake, OR; 11
refrigerated
Mercer Lake, OR; 1 1
frozen
Mercer Lake, OR; 1 1
refrigerated
Mercer Lake, OR; 1 1
frozen
17% sand, 83% 2
silt and clay
Sand, wet- 0.28
sieved
between l-2mm
d
sieves
Method,3
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
Dry wt
Response (wg/g)
LCSO 0.012
LCSO 0.172
LCSO 0.224
LCSO 4.4
LCSO 4.8
LCSO 6.0
LCSO 5.1
LCSO 7.7
LCSO 19.6
LCSO 21.7
LCSO 10.3
LCSO 9.8
2 of 5 28
worms
died
LCSO 0.047
Water
OC LCSO 5
C^g/g) (^g/L) Reference
17.0 1.07 Stehly, 1992
31.3 2.2 Stehly, 1992
12.8 0.63 Stehly, 1992
147 2.1 Nebekeretal.,
1989
78.7 1.9 Nebekeretal.,
1989
53.6 1.8 Nebekeretal.,
1989
170 Schuytema et al.,
1989
257 Schuytema et al.,
1989
178 Schuytema et al.,
1989
197 Schuytema et al.,
1989
93.6 Schuytema et al.,
1989
89.1 Schuytema et al.,
1989
1,400 McLeese et al.,
1982
16.8 McLeese and
Metcalfe, 1980
aS = static, R = renewal, M = measured, N = nominal.
"Value from Landrum (1991).
CL. hoffmeisteri and S. heringianus tested together.
Clean sediment placed in endrin-coated beakers at beginning of exposure.
4-3
-------
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 4ฐC 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 ^g/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/ig/g
dry weight), which had a TOC of 11 %, than LC50
values from two tests using Soap Creek sediments (5.1
and 7.7 ^g/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 ng 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 ^g/Eoc- 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 ESG Technical Basis
Document. For endrin, mis 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 A^g/g^ (5 tests) for
sediments from Mercer Lake having a TOC of 1 1 % and
186 /^g/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). On a dry weight
basis, 4-day LC50 values forDiporeia sp. ranged from
0.012 to 0,224 ,ug/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 .
Although it is important to demonstrate that
organic carbon normalization is necessary if ESGs are
to be derived using the EqP approach, it is
fundamentally more important to demonstrate that K^.
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, 2000a).
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
4-4
-------
endrin, based on interstitial water concentrations,
ranged from 1.8 to 2.1 ng/L for H. azteca exposed to
three sediments. Overlying water LC50 values from
these static tests (Nebeker et al., 1989) and those
conducted using the same sediments by Schuytema et
al. (1989) were similar; 1.1 to 3.9 pg/L. Stehly (1992)
found that 4-day interstitial water LC50 values for
Diporeia sp. rangedfrom 0.63 to 2.2 /zg/L (a factor of
3.5); this is considerably less than the range in LC50
values expressed as dry weight, 0.012 to 0.224 //g/g (a
factor of 18.7), for three sediments from Lake
Michigan having 0.07 % to 1.75 % organic carbon.
A more detailed evaluation of the degree to which
the response of benthic organisms can be predicted
from toxic units (TUs) of substances in interstitial
water can be made utilizing results from toxicity tests
with sediments spiked with a variety of nonionic
compounds, including acenaphthene and phenanthrene
(Swartz, 1991), endrin (Nebeker etal., 1989;
Schuytema et al., 1989), fluoranthene (Swartz et al.,
1990; DeWitt et al., 1992), and kepone (Adams et al.,
1985) (Figure 4-1). The endrin data included in this
analysis were from tests conducted at laboratories or
from tests that utilized designs at least as rigorous as
those conducted at EPA laboratories. Note that
dieldrin data from Hoke et al. (1995) were aot used in
the interstitial water TU plot either because interstitial
water was not measured or because of inconsistencies
in the mortality results that have been attributed to
DOC complexing in the interstitial water. This is
discussed in Hoke et al. (1995) and in the EPA dieldrin
ESG document (U.S. EPA, 2000d). Tests with
acenaphthene and phenanthrene used two saltwater
amphipods (Leptocheints plwnulosus and Eohaustorius
estuarius) and saltwater sediments. Tests with
fluoranthene used a saltwater amphipod (Rhepoxynius
abronius) and saltwater sediments. Freshwater
sediments spiked with endrin were tested using the
amphipod H. azteca, and kepone-spiked sediments were
tested using the midge, C. tertians.
Figure 4-1 presents the percent mortalities of the
benthic species tested in individual treatments for each
i
1W
80
60
20
+ Endrin
D
A Fluoranthciac
\7 Acenaphthene
O Kepone
LJ_J_
_t__l l_J_ULlJ_
8.81
0.1
18
100
Interstitial Water Toxic Units
Figure 4-1. Percent mortality of amphipods in sediments spiked with acenaphthene or phenanthrene (Swartz,
1991), 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 kepone (Adams et al., 1985) relative to
interstitial water toxic units.
4-5
-------
chemical versus interstitial water TUs (IWTUs) for all
sediments. IWTUs are the concentration of the
chemical in interstitial water Gug/L) divided by the
water-only LC50 (ngfL). 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% mortalityr 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. This interstitial water
normalization was not used to derive the ESG in this
document because of the complexation of nonionic
organic chemicals with interstitial water DOC (Section
2) and the difficulties of adequately sampling
interstitial waters.
4.3 Tests of the Equilibrium Partitioning
Prediction of Sediment Toxicity
Sediment guidelines derived using the EqP
approach utilize partition coefficients and FCVs from
updated or final WQC documents to derive the ESG
concentration that is protective of benthic organisms.
The partition coefficient A^ 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 the ESG derivation required three
elements: (1) a water-only effects concentration, such
as a 10-day LC50 value, in^g/L; (2) an identical
sediment effect concentration on an organic carbon
basis, in /^g/g^; and (3) a partition coefficient for the
chemical, K^, in L/kg^. 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 ah, 1989) were calculated
(Table 4-2) using the logj^^ value of 4,97 from
Section 2 of this document and the geometric mean of
the water-only LC50 value (4.1 ngfL). 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 ATQW for kepone obtained from Karickhoff
and Long (1995) was evaluated using only one
laboratory measured value, whereas the remaining
chemical KOVf 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%, and8.7% organic carbon, and also
exposed the midge C. tertians 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 G^g/goc) divided by the
predicted sediment LC50 (i.e., the product of K^ and
the 10-day water-only LC50 expressed in ^glg^. 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
4-6
-------
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
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 ฑ2. The error bars shown in Figure 4-2
are computed as ฑ 1.96 x (ESG uncertainty). The
value of 1.96 is the t statistic, which provides a 95 %
confidence interval around the ESG.
Table 4-2. Water-only and sediment LC50 values used to test the applicabUity of the EqP theory for endrin
Common Water-
Name, Method, Only
Scientific Duration LC50
Name (days) (A
-------
100
so
I" ซ
i -
20
0
O.fi
' ' ' i ' i '
EBdrin
O DfeWrin
D PhtDanthrtne
A Flซoroanthene
V Acenaphthene
1. ^
o o *
& .y
+ A*. ^ฎ ^fa_fl
^ jj^ j*iP>- vj^y^^i^s
. i . 1 1 i
i
*lf%
1 o'
J^JfiLSL
47 ^
f ^
.....1 .
o
v
3
1
1 0.1 1 10 100
Predicted Sediment Toxic Units with Uncertainty Bars
Figure 4-2. Percent mortality 0f amphipods in sediments spiked with acenapbthene or phenanthrene (Swartz,
1991), dieldrin (Hoke et al., 1995), endrin (Nebeker et a!., 1989; Schuytema et al., 1989), or
fluoranthenc (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
Guidelines Derivation for Endrin
5.1 Guidelines Derivation
The WQC FCV (see Section 3), without an
averaging period or return frequency, is used to
calculate the ESG 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 hi 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 the
ESG, it may be appropriate to use the areal extent and
vertical stratification of contamination at a sediment
site in much me 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 die 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 WQC Guidelines.
The EqP method for calculating ESGs is based on
the following procedure (also described in Section
2-1). If the FCV Cug/L) is the chronic concentration
from die WQC for the chemical of interest, then the
ESG (Mg/g sediment) is computed using the partition
coefficient, Kp (L/g sediment), between sediment and
interstitial water
ESG = Kp FCV
(5-D
can
The organic carbon partition coefficient,
be substituted for Kf, because organic carbon is die
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 ESG (ESG^ in
is
(5-2)
And because KQ^. is presumably independent of
sediment type for nonionic organic chemicals, so too is
ESGoc. Table 5-1 contains the calculation of the endrin
ESG.
The ESGoc is applicable to sediments
iO.2%. For sediments with/oc <0.2%, organic
carbon normalization and the ESGs do not apply.
Because organic carbon is the factor controlling the
bioavailability of nonionic organic compounds in
Table 5-1. Equilibrium partitioning sediment guidelines (ESGs) for endrin
Type of Water Body
Freshwater
Saltwater
Log K0w
(L/kg)
5.06
5.06
(L/kg)
4.97
4.97
FCV
0.05805
0.01057
ESGoc
(Pg/goc)
5.4a
0.99b
ESGOC = (10* 97 L/kgo,-) x (1CT3 kgw./gor) x (0.05805 Mg endrin/L) = 5.4 /^g endrm/goc.
bESGoc = (lO^'L/kgoc) x (10ฐ kgo.Jgoc-) x (0.01057 ,/g endrin/L) = 0.99 ^g endrin/g(K,
5-1
-------
sediments, ESGs have been developed on an organic
carbon basis, not on a dry weight basis . When the
chemical concentrations in sediments are reported as
dry weight concentrations and organic carbon data are
available, it is best to convert the sediment
concentrations to /^g chemical/go^.. These
concentrations can then be directly compared with the
ESG value. This facilitates comparisons between the
ESG and field concentrations relative to identification
of hot spots and the degree to which sediment
concentrations do or do not exceed the ESG values.
The conversion from dry weight to organic carbon-
normalized concentration can be done using the
following formula
* (% TOC + 100)
= jug chemical/g^ ^ x 100 * % TOC
For example, a freshwater sediment with a
concentration of 0. 1 ng endrin/g . M and 0.5 % TOC has
an organic carbon-normalized concentration of 20 /ug/
goc (= 0.1 ^g/g.^ X 100 -r 0.5), which exceeds the
freshwater endrin ESG of 5 .4 fj.g/goc. Another
freshwater sediment with the same concentration of
endrin (0. 1 /ig/g^ OT) but a TOC concentration of 5 . 0 %
would have an organic carbon-normalized concentration
of2.0Mg/goc(=0.lMg/gdrywt x 100 * 5,0), which is
below the freshwater ESG 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, the ESG^ values may be "converted" to
dry weight-normalized ESG values (ESG
"conversion" for each level of TOC is
drv wt-
,). This
ESG
w,
X (% TOC + 100)
For example, the ESG,^ M value for freshwater
sediments with 1 % organic carbon is 0.054 jug/g
1%TOC ^ 100 = 0.
This method is used in the analysis of the STORET
data in Section 5. 4.
5.2 Uncertainty Analysis
Some of the uncertainty of the endrin ESG 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 guidelines. 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 in sediment G^g/gor) can be
estimated from the product of the effects concentrations
from water-only exposures, FCV (^g/L), and the
partition coefficient, K^ (L/kg). The uncertainty
associated with the ESG can be obtained from a
quantitative estimate of the degree to which the
available data support these assertions.
The data used in 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 ESG (see Section 4.3 and the
ESG 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 in the uncertainty analysis are
illustrated in Figure 4-2. The data for endrin are
summarized in 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 ^ ^g/goc) via the
partitioning equation
LC50SOC
(5-3)
As mentioned above, one of the assertions of the EqP
model is that the toxicity of sediments expressed on an
organic carbon basis equals the toxicity in water tests
multiplied by the Xoc. Therefore, both LC50S and
KQC x LC50W are estimates of the true LC50Q,, for
each chemical-organism pair. In this analysis, the
uncertainty of K^ is not treated separately. Any error
associated with K^ 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
5-2
-------
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
(o^ and (a^ 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
(54)
where ln(LC50y k) is either ln(LC50w) or ln(LC50s <><,),
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 Mg/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, 2000a). The results are shown in
Table 5-2. The last line of Table 5-2 is the uncertainty
associated with the ESG; i.e. , the variance associated
with the exposure media variability.
The confidence limits for the ESG are computed
using this estimate of uncertainty for the ESG. For the
95% confidence interval limits, the significance level
is 1.96 for normally distributed errors.
Hence,
+ 1 .960
^
(5-5)
- 1 .960
^
The confidence limits are given in Table 5-3.
is applicable to sediments with/^,
2 0 . 2 % . For sediments with/^ < 0. 2 % , organic
carbon normalization and ESGs do not apply. ~
5.3 Comparison of Endrin ESG 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 ESG 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 hi water-only exposures (Figures 5-1
Table 5-2. Analysis of variance for derivation of confidence limits of the ESGs for endrin
Source of Uncertainty
Exposure media
Replication
ESG sediment guidelines
Parameter
Value (Aig/goc)
0.41
0.29
0.41
5-3. Confidence limits of the ESGs for endrin
Type of Water Body
95% Confidence Limits lug/goc)
Lower
Upper
Freshwater
Saltwater
5.4
0.99
2.4
0.44
12
2.2
5-3
-------
greater than the upper 95% confidence interval of the
ESG (12 ^g/goc). The PGMCVs for eight genera,
including four water column fish and four benthic
arthropod genera, are below the ESG 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 ESG 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 ESG
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 times the ESG. 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 tunes the
The saltwater ESG for endrin (0.99 Mg/g^) is less
than any of the PGMCVs for saltwater genera (Figure
5-2). The PGMCVs for the penaeid shrimp Penaeus
(1 . 1 ^g/goc) and the fishes Oncorhynchus (1 .44 ^g/goc)
and Menidia (1 .50 jug/g^) are lower than the upper
95 % confidence interval for the ESG (2 . 2 ngig^ 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
ESG would exceed the PGMCVs of 6 of the 1 1 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 ESG.
5.4 Comparison of Endrin ESG to
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 in databases of sediment
contamination. This means that it is possible that many
of the sediments from the nation's waterways might
exceed the endrin guidelines. ID order to investigate
this possibility, the endrin guidelines 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 in which
the ESG values were actually exceeded may be less
than the reported percentage. Very few of the measured
values from either of the databases exceeded the ESGs.
A STORET (U.S. EPA, 1989b) data retrieval was
performed to obtain a preliminary assessment of the
concentrations of endrin in the sediments of the nation's
water bodies. Log probability plots of endrin
concentrations on a dry weight basis in sediments are
shown in Figure 5-3. Endrin was found at significant
concentrations in 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 in most water bodies. There is
significant variability in endrin concentrations in
sediments throughout the country. Lake samples in EPA
Region 9 appear to have had relatively high endrin
levels (median = 0.030 /ig/g) prior to 1986. The upper
10% of the concentrations were disproportionally found
in streams, rivers, and 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 ESG for endrin can be compared to existing
concentrations of endrin in sediments of natural water
systems in the United States as contained in the
STORET database (U.S. EPA, 1989b). These data
were generally reported on a dry weight basis rather
than an organic carbon-normalized basis. Therefore,
ESG values corresponding to sediment organic carbon
levels of 1 % to 10% were compared with endrin's
distribution in sediments as examples only. For
freshwater sediments, ESG values were 0.054 ng/g dry
weight in sediments having 1 % organic carbon and 0.54
Aig/g dry weight in sediments having 10% organic
carbon; for marine sediments, the ESGs were 0.0099
^g/g dry weight and 0.099 Mg/g dry weight,
respectively. Figure 5-3 presents the comparisons of
these ESGs with probability distributions of observed
sediment endrin levels for streams and lakes
5-6
-------
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(freshwater systems, A and B) and estuaries (marine
systems, C).
For streams (n = 2,677), the ESGs of 0.054 ptg/g
dry weight for 1 % organic carbon sediments and 0.54
Aig/g dry weight for 10% organic carbon freshwater
sediments were exceeded in less than 1 % of the
samples. For lakes (n = 478), the ESG of 0.054 ^g/g
dry weight for 1 % organic carbon sediment was
exceeded in about 2 % of the samples, and the ESG of
0.54 /ig/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 ESG of 0.0099 ^g/g dry weight sediment for
1 % organic carbon sediments was exceeded in about
8% of the samples, and the ESG of 0.099 vg/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
(2000a). These data were collected to examine the
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 ESG developed
using the EqP methodology. A major portion (93 %) of
the samples analyzed had/^ > 0.2%, for which the
ESG 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 hi 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 ESG. 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
*>
s
10 20 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
-------
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 in relation to sediment
guidelines.
Regional-specific differences in endrin
concentrations may affect the above conclusions
concerning expected guidelines 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 ESG.
5.5 Limitations to the Applicability of
ESGs
Rarely, if ever, are contaminants found alone in
naturally occurring sediments. Obviously, the fact that
the concentration of a particular contaminant does not
exceed the ESG does not mean that other chemicals,
for which there are no ESGs available, are not present
in concentrations sufficient to cause harmful effects.
Furthermore, even if ESGs were available for all of
the contaminants in a particular sediment, there might
be additive or synergistic effects that the guidelines do
not address. In this sense the ESG represents a "best
case" guideline.
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 ESG 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 ESG 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 ESGs exist occur along widi
ESG chemicals.
Care must be used in the application of EqP-
derived guidelines in disequilibrium conditions. In
some instances, site-specific ESGs may be required to
address disequilibrium. The ESGs assume that
nonionic organic chemicals are in 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 in 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
guidelines are overprotective. In liquid chemical spill
situations, disequilibrium concentrations in interstitial
and overlying water may be proportionately higher
relative to sediment concentrations. In this case the
guidelines may be underprotective.
Note that the K^ values used in the EqP
calculations described in this document assume that the
organic carbon in sediments is similar in partitioning
properties to "natural" organic carbon found inmost
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
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 ESG underprotective. If such a
situation is encountered, the applicability of literature
AOC 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 AT^ values recommended
herein (after accounting for DOC binding in the
interstitial water), then the national ESGs would be
underprotective and calculation of a site-specific ESG
should be considered (see U.S. EPA, 2000b).
The presence of organic carbon in 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
5-9
-------
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 ESGs 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 ESGs.
In very dynamic areas, with highly erosional or
depositional bedded sediments, equilibrium may not be
attained with contaminants. However, even high ^Tow
nonionic organic compounds come to equilibrium in
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 paniculate 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, 2000b).
5-10
-------
Section 6
Guidelines Statement
The procedures described in the ESG Technical
Basis Document indicate that benthic organisms should
be acceptably protected in freshwater sediments
containing <,5.4 /zg endrin/goc and saltwater sediments
containing <.0.99 /^g endrin/g^, except possibly where
a locally important species is very sensitive or sediment
organic carbon is <0.2%.
Confidence limits of 2.4 to 12/jg/g^^ for freshwater
sediments and 0.44 to 2.2 Pg/goc for saltwater
sediments are provided as an estimate of the
uncertainty associated with the degree to which the
observed concentration in sediment (wg/g,^), which
may be toxic, can be predicted using the T^, and the
water-only effects concentration. Confidence limits do
not incorporate uncertainty associated with water
quality criteria. An understanding of the theoretical
basis of the equilibrium partitioning methodology,
uncertainty, and the partitioning and toxicity of endrin
are required in the regulatory use of ESGs and their
confidence limits.
The guidelines presented in this document are
EPA's best recommendation of the concentrations of a
substance that may be present in sediment while still
protecting benthic organisms from the effects of that
substance. These guidelines 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 guidelines do not
protect against additive, synergistic, or antagonistic
effects of contaminants or bioaccumulative effects to
aquatic life, wildlife or human health. The Agency and
the U.S. EPA Science Advisory Board do not
recommend the use of ESGs as stand-alone, pass-fail
criteria for all applications; rather, exceedances of ESGs
could trigger additional studies at sites under
investigation. The ESG should be interpreted as a
chemical concentration below which adverse effects are
not expected. In comparison, at concentrations above
the ESG effects are likely, and 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 ESG and
associated uncertainty limits.
6-1
-------
Section 7
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7-5
-------
Appendix A
Summary of Acute Values for Endrin
for Freshwater and Saltwater Species
-------
Common Name, 3 (, c
Scientific Name stage Habitat Method Cone
FRESHWATER SPECIES
Oligochaete A I FT
worm,
Lumbriculus
variegatus
Oligochaete A I FT
worm,
Lumbriculus
variegatus
Cladoceran, X W,E S
Simocephalus
serrulalus
Cladoceran, X W,E S
Simocephalus
serrulalus
Cladoceran, L W S
Daphnia
magna
Cladoceran, L W S
Daphnia
magna
Cladoceran, L W S
Daphnia
magna
Cladoceran, L W FT
Daphnia
magna
LC50/EC506 (ME/L)
HMAV
- ' wVCJtUl
entration Test Species Genus GMAV Reference
*
M > 165.1 Poirierand
Cox, 1991
M >165.0 >165.0 >165.0 >165.0 Brooke,
1993
U 26 Sanders
and Cope,
1966;
Mayer and
Ellersieck,
1986
U 45 34.20 34.20 34.20 Sanders
and Cope,
1966;
Mayer and
Ellersieck,
1986
U 4.2 Mayer and
Ellersieck,
1986
U 74 Mayer and
Ellersieck,
1986
U 41 Mayer and
Ellersieck,
1986
M 230 Thurston et
al., 1985
Cladoceran, L W FT M 88 142.3 Thurston et
Daphnia
magna
Cladoceran, L W S
Daphnia
pulex
Ostracod, A I,E S
Cypridopsis
sp.
al., 1985
U 20 20 53.35 53.35 Mayer and
Ellersieck,
1986
U 1.8 1.8 1.8 1.8 Mayer and
Ellersieck,
1986
Sowbug, A E S U 1.5 1.5 1.5 1.5 Sanders,
Asellus
brevicaudus
1972;
Mayer and
Ellersieck,
1986
A-l
-------
Common Name,
Scientific Name
life-
stage Habitat Method Concentration
LC50/ECSO Qjg/L)
HMAV
Overall
Test Species Genus8 GMAv Reference
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
bilineala
Mayfly,
Hexagenia
bilineata
X
E FT
FT
FT
U
U
U
U
U
U
M
U
U
U
4.3
1.3
5.5 3.133
3.0
3.2
3.2
3.0 3.066 3.066
0.5 1.265 1.265 1.265
>S9 >S9
320
3.2 3.2
16.88
0.90 0.90 0.90 0.90
64
62.99 62.99 62.99
Sanders,
1972;
Mayer and
Ellersieck,
1986
Sanders,
1972;
Mayer and
Ellersieck,
1986
Sanders,
1972
Sanders,
1972;
Mayer and
Ellersieck,
1986
Sanders,
1972;
Mayer and
Ellersieck,
1986
Sanders,
1972;
Mayer and
Ellersieck,
1986
Thurston et
al., 1985
Sanders,
1972;
Mayer and
Ellersieck,
1986
Sanders,
1972;
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Sanders,
1972
A-2
-------
Common Name, &\ b c d
Scientific Name stage Habitat Method Concentration
Stonefly, L W.E S U
Acroneuria sp.
Stonefly, L I,E S U
Pleronarcella
badia
Stonefly, A I,E S U
Pteronarcys
califomica
Stonefly, J W,E S U
Claassenia
sabulosa
Stonefly, J W,E S U
Claassenia
sabulosa
Caddis fly, X E FT M
Brachycentrus
americanus
Damesfly, X W.E S U
Ischnura
verticalus
Damesfly, J W,E S U
Ischnura
verticalus
Damesfly, J W.E S U
Ischnura
verticalus
Midge, L I FT M
Tanytarsus
dissimilis
Diptera, J I,E S U
Tipula sp.
Diptera, J I,E S U
Atherix
variegata
Coho salmon, J W S U
Oncorhynchus
kisulch
LC50/EC50e (MK/L)
liMAV Overall
-" WVCI all
Test Species Genus8 GMAv" Reference
>0.18 >0.18 >0.18 >0.18 Mayerand
Ellersieck,
1986
0.54 0.54 0.54 0.54 Sanders
and Cope,
1968;
Mayer and
Ellersieck,
1986
0.25 0.25 0.25 0.25 Sanders
and Cope,
1968i
Mayer and
Ellersieck,
1986
0.76 Sanders
and Cope,
1968
0.76 0.2403 0.2403 0.2403 Mayerand
Ellersieck,
1986
0.34 0.34 0.34 0.34 Anderson
and DeFoc,
1980
1.8 Sanders,
1972
2. 1 Mayer and
Ellersieck,
1986
2.4 2.086 2.086 2.086 Mayerand
Ellersieck,
1986
0.83 0.83 0.83 0.83 Thurston et
a!., 1985
12 12 12 12 Mayerand
Ellersieck,
1986
4.6 4.6 4.6 4.6 Mayer and
Ellersieck,
1986
0.51 Katz, 1961
A-3
-------
Common Name,
Scientific Name
Coho salmon,
Oncorhynchus
kisutch
Coho salmon,
Oncorhynchus
kisutch
Cutthroat trout,
Oncorhynchus
clarki
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Rainbow trout,
Oncorhynchus
mykiss
Chinook
salmon,
Oncorhynchus
ishawylscha
LC50/EC506 (Mg/L)
Life HMAV Overall
stage Habitat Method0 Concentration Test Species Genus8 GMAv Reference
J W S U 0.089 Mayer and
Ellersieck,
1986
J W S U 0.27 0.2306 Katzand
Chadwick,
1961
J W S U >1.0 >1.0 Mayerand
Ellersieck,
1986
J W S U 0.74 Mayerand
Ellersieck,
1986
J W S U 0.75 Mayerand
Ellersieck,
1986
J W S U 0.75 Mayerand
Ellersieck,
1986
J W S U 2.4 Mayerand
Ellersieck,
1986
JWS U 1.4 Mayerand
Ellersieck,
1986
JWS U 1.11 Mayerand
Ellersieck,
1986
JWS U 1.1 Maceket
al., 1969
JWS U 0.58 Katz, 1961
JWS U 0,90 Katzand
Chadwick,
1961
J W FT M 0.33 0.33 Thurstonet
al., 1985
JWS U 1.2 Katz, 1961
A-4
-------
^^M^^^f^^^^l^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^S
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LC50/EC50e (Mg/L)
Common Name, a be d
Scientific Name stage Habitat Method Concentration Test
Chinook J W S U 0.92
salmon,
Oncorhynchus -
ishawytscha
Goldfish, J W S U 2.1
Carassius
auratus
Goldfish, J W FT U 0.44
Carassius
auratus
Goldfish, J W FT M 0.95
Carassius
auratus
Carp, J W FT U 0.32
Cyprinus
carpio
Fathead J W S U 1.1
minnow,
Pimephales
promelas
Fathead J W S U 1.4
minnow,
Pimephales
promelas
Fathead L W S U 0.7
minnow,
Pimephales
promelas
Fathead J W S U 1.8
minnow,
Pimephales
promelas
Fathead J W FF U 0.24
minnow,
Pimephales
promelas
Fathead J W FT M 0.50
minnow,
Pimephales
promelas
Fathead U FT M 0.49
minnow,
Pimephales
promelas
Fathead J W FF M 0.40
minnow,
Pimephales
promelas
HMAV 0-crall
'"- WYCJaJI
Species Genus GMAV Reference
ป
1.051 XX5318 X>.5318 Katzand.
Chadwick,
1961
Henderson
etal., 1959
Mayer and
Ellersieck,
1986
0.95 0.95 0.95 Thurston et
al., t985
0.32 0.32 0.32 Mayer and
Ellersieck,
1986
Henderson
etal., 1959
Henderson
etal., 1959
Jarvinen et
al., 1988
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Brungs and
Bailey,
1966
Brungs and
Bailey,
1966
Brungs and
Bailey,
1966
A-5
-------
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,
Icialunts
punctatus
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Channel
catfish,
Ictalurus
punctatus
Flagfish,
Jordanella
floridae
Mosquitofish,
Gambusia
affinis
Mosquitofish,
Gambusia
affinis
LC50/EC50e Oig/L)
Life HMAV Ova 11
j^iic- ___^____ ^^venm
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 Mayerand
Ellersieck,
1986
J W.E FT M 0.45 0.45 _ _ Anderson
and DeFoe,
1980
J W,E S U 0.32 Mayerand
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
et al., 1985
J W S U 1.1 Mayerand
Ellersieck,
1986
X W S U 0.75 Katzand
Chad wick,
1961
A-6
-------
Common Name, a b c
Scientific Name stage Habitat Method Concentration
Mosquitofish, J W FT M
Gambusia
affinis
Guppy, X W S U
Poecilia
reticulata
Guppy, X W S U
Poecilia
reticulata
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
Bluegill, J W S U
Lepomis
macrochirus
>
LC50/EC50e OiR/L)
HMAV _ ,
Test Species Genus8 GMAv Reference
0.69 0.69 0.69 0.69 Thurston et
al., 1985
0.90 Katzand
Chadwick,
1961
1.6 1.200 1.200 1.200 Henderson
etal., 1959
0.60 Katzand
Chadwick,
196T
8.25 - Katzand
Chadwick,
1961
5.5 Katzand
Chadwick,
1961
2.4 Katzand
Chadwick,
1961
1.65 Katzand
Chadwick,
1961
0.86 Katzand
Chadwick,
1961
0.33 Katzand
Chadwick,
1961
0.61 Maceket
al., 1969;
Mayer and
Ellersieck,
1986
0.41 Maceket
al., 1969;
Mayer and
Ellersieck,
1986
0.37 Maceket
al., 1969;
Mayer and
EUcrsicck,
1986
A-7
-------
LC50/EC50e C*g/L)
HMAV
l_Jfc iuปi^v Overall
Common Name, a b c d f e h
Scientific Name stage Habitat Method Concentration Test Species Genus GMAv
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
BluegilJ, 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 0.31
bass,
Micropterus
dolomieu
Yellowperch, J W FT V 0.15 0.15 0.15 0.15
Perca
flavescens
Tilapia, J W S U <5.6 <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) 7.906
leopard frog, 25(W)
Rana
sphenocephala
Fowler's toad, L E S V 120 120 120 120
Bufofawleri
, _
Reference
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
Henderson
et al., 1959
Sanders,
1972
Thurston et
al., 1985
Thurston et
al., 1985
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Mayer and
Ellersieck,
1986
Thurston et
al., 1985
Hall and
Swineford,
1980
Mayer and
Ellersieck,
1986
A-8
-------
LC50/EC50C CUS/L)
Life HMAV (
Common Name, a b c d f g
Scientific Name stage Habitat Method Concentration Test Species Genus C
Western L E S U 180 180 180
chorus frog,
Psuedocris
triseriata
SALTWATER SPECffiS
Eastern oyster, E.L W S U 790* 790 790
Crassostrea
virginica
Sand shrimp, A E S U 1.7 1.7 1.7
Crangon
septemspinosa
Hermit crab, A E S U 12 12 12
Pagurus
longicarpus
Korean A W,E S U 4.7
shrimp,
Palaemon
macrodactylus
Korean A W,E FT U 0.3 1.187 1.187
shrimp,
Palaemon
macrodactylus
Grass shrimp, L W FT M 1.2
pugio
Grass shrimp, J W FT M 0.35
Palaemonetes
pugio
Grass shrimp, A W,E FT M 0.69
Palaemonetes
pugio
Grass shrimp, A W.E FT M 0.63 0.6536
Palaemonetes
pugio
Grass shrimp, A W,E S U 1.8 1.8 1.085
Palaemonetes
vulgaris
Pink shrimp, A I,E FT M 0.037 0.037 0.037
Penaeus
duorarum
American eel, J E S U 0.6 0.6 0.6
Anguilla
roslrata
>verall
MAY Reference
180 Mayer and
Ellersieck,
1986
790 Davis and
Hidu, 1969
1.7 Eisler,
1969
" 12 Eisler,
1969
Schoettger,
1970
1.187 Schoettger,
1970
Tyler-
Scnroeder,
1979
Tyler-
Schroeder,
1979
Tyler-
Schroeder,
1979
Schimmel
etal., 1975
1.085 Eisler,
1969
0.037 Schimmel
et al., 1975
0.6 Eisler,
1969
A-9
-------
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
Sailfin molly,
Poecilia
latipinna
Atlantic
silverside,
Menidia
menidia
Threespine
stickleback.
Gasterosteus
aculeatus
Threespine
stickleback,
Gasterosteus
aculeatus
LC50/EC506 fcg/L)
Life HMAV Ov -rill
i^nc- _ซ ___~_~__ uverajl
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 Hansenet
al., 1977
J W,E FT M 0.34 Hansenet
al., 1977
__
A W,E FT M 0.36 Hansenet
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-10
-------
LC50/EC50C (pg/L)
Common Name, a j, c j
Scientific Name stage Habitat Method Concentration Test
Threespine J W,E S U 120
stickleback,
Gasterosteus
aculeatus
Threespine J W,E S U 1.57
stickleback,
Gaslerosteus
aculeatus
Threespine J W,E S U 1.57
stickleback,
Gasterosteus
aculeatus
Threespine J W,E S U 0.44
stickleback,
Gasterosteus
aculeatus
Threespine J W,E S U 050
stickleback,
Gasterosleus
aculeatus
Striped bass, J E FT U 0.094
Morone
saxatilis
Shiner perch, J W S U 0.8
Cymatogaster
aggregala
Shiner perch, J W FT U 0.12
Cymatogaster
aggregata
Dwarf perch, A W S U 0.6
Micrometrus
minimus
Dwarf perch, A W FF U 0.13
Micrometrus
minimus
Bluehead, AW S U 0.1
Thalassoma
bifasciatum
HMAV 0
\_/ VC1 till
Species Genus GMA1/1 Reference
$
Katz and
Chad wick.
1961
Katz and
Chad wick,
1961
Katz and
Chadwick,
1961
_ _ _ Katz, 1961
1.070 1.070 1.070 Katz, 1961
0.094 0.094 0.094 Kom and
Earnest,
1974
Earnest
and
Benville,
1972
0.3098 0.3098 0.3098 Earnest
and
Benville,
1972
Earnest
and
Benville,
1972
0.2793 0.2793 0.2793 Earnest
and
Benville,
1972
0.1 0.1 0.1 Eisler,
1970b
A-ll
-------
LC50/EC50e Og/L)
Common Name, a be d
Scientific Name stage Habitat Method Concentration Test
Striped mullet, A E S U 0.3
Mugil
cephalus
Northern A W S U 3.1
puffer,
Sphaeroides
maculatus
HMAV
Species' Genus8 GMAV* Reference
0-3 0.3 0.3 Eisler,
1970b
3.1 3.2 3.1 Eisler,
1970b
aLife-stage: A = adult, J = juvenile, L = larvae, E = embryo, U = life-stage and habitat unknown, X = life-stage unknown but habitat
known.
Habitat: I = infauna, E = epibenthic, W = water column.
cMethod: S = static, R = renewal, FT = flow-through.
Concentration: U = unmeasured (nominal), M = chemical measured.
eAcute value: 96-hour LC50 or EC50, except for 48-hour EC50 for cladocera, barnacles, and bivalve molluscs (Stephan et al., 1985).
HMAV species: Habitat Mean Acute Value - Species is the geometric mean of acute values by species by habitat (epibenthic", infaunal,
and water column).
gHMAV genus: Geometric mean of HMAV for species within a genus.
_ Overall GMAV: Geometric mean of acute values across species, habitats, and life-stages within the genus.
'Abnormal development of oyster larvae.
A-12
-------
Appendix B
Summary of Data from Sediment-Spiking Experiments with Endrin. Data from
these experiments were used to calculate K values (Figure 2-2) and to compare
mortalities of amphipods with interstitialwater toxic units (Figure 4-1) and
predicted sediment toxic units (Figure 4-2).
-------
Sediment Concentration Cng/g)
Sediment Source,
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.
Mortality
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
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
b
0.224
Organic Carbon
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
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
LogKoc"
4.82
4.88
4.76
4.78
4.81
4.56
4.67
4.63
4.69
4.58
4.59
4.60
4.42
4.52
4.63
4.96
4.97
5.03
4.99
4.97
4.65
4.68
4.70
4.58
4.66
4.57
4.60
4.60
4.96
4.67
4.75
4.20
4.15
4.31
References
riebeker et al.,
1989
Nebekeret al.,
1989
Nebeker et al.,
1989
.- - -
Schuytema et
al., 1989
Schuytema et
al., 1989
Schuytema et
al., 1989
Stehly, 1992
MEAN = 4.67
SE = 0.04
Interstitial water concentrations from Schuytema et a]. (1989) are concentrations of "soluble" endrin in water overlying sediments.
Sediments were refrigerated prior to testing.
= sediment concentration
"" calculated free interstitial water concentration C^g/L) X 10^ g/kg.
B-l
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