United States Office of Science and Technology
Environmental Protection Agency Health and Ecological Criteria OiV. September 1993
Office of Water & Washington. O.C. 2046O
Office of Research and
Development . '
Sediment Quality Criteria
for the Protection of
Benthic Organisms:
DIELDR1N
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CONTENTS
PAGE
Foreword
Acknowledgments ' " .»
Tables . ".".'.'.'.".'.".*.'.'.".". m
Figures V-
Introduction ™
Partitioning .'*.......!!..... 21
Toxicity of Dieldrin: Water Exposures . . . . . 31
Toxicity of Dieldrin (Actual and Predicted): Sediment Exposures 4-\
Criteria Derivation for Dieldrin " " 5 J
Criteria Statement gli
References
Appendix A: Summary of Acute ValueYfor Dieldrin for Freshwater and Saltwater"
Species A _i
Appendix B: Summary of Data from Sediment Spiking Experiments with
Dieldrin" " ' "
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FOREWORD
Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U S
and the States develop programs for protecting the chemical, physical, and biological
integrity of the nation's waters. Section 304(a)(l) 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 which
may be expected from the presence of pollutants in any body of water, including ground water
(2) the concentration and dispersal of pollutants, or their byproducts, through biological, physical
and chemical processes, and (3) the effects of pollutants on biological community diversity
productivity, and stability. Section 304(a)(2) directs the Administrator to develop and publish
information on, among other things, the factors necessary for the protection and propagation of
shellfish, fish, and wildlife for classes and categories of receiving waters.
To meet this objective, U.S. EPA has periodically issued ambient water quality criteria
(WQC) guidance beginning with the publication of "Water QHiality Criteria 1972" (NAS/NAE
1973). All criteria guidance through late 1986 was summarized in an U.S. EPA document
entitled "Quality Criteria for Water, 1986" (U.S. EPA, 1987). Additional WQC documents that
update criteria for selected chemicals and provide new criteria for other pollutants have also been
published. In addition to the development of WQC and to continue to comply with the mandate
of the CWA, U.S. EPA has conducted efforts to develop and publish sediment quality criteria
(SQC) for some of the 65 toxic pollutants or toxic pollutant categories. Section 104 of the CWA
authorizes the administrator to conduct and promote research into the causes, effects extent
prevention, reduction and elimination of pollution, and to publish relevant information.' Section
104(n)(l) in particular provides for study of the effects of pollution, including sedimentation in
estuanes, on aquatic life, wildlife, and recreation. U.S. EPA's efforts with respect to sediment
criteria are also authorized under CWA Section 304(a).
Toxic contaminants in bottom sediments of the nations's lakes, rivers, wetlands and
coastal waters create the potential for continued environmental degradation even where water
column contaminantlevels meet established WQC. In addition, contaminated sediments can lead
to water quality impacts, even when direct discharges to the receiving water have ceased EPA
intends SQC be used to assess the extent of sediment contamination, to aid in implementing
measures to limit or prevent additional contamination, and to identify and implement appropriate
remediation activities when needed.
The criteria presented in this document are the U.S. 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 criteria are applicable to a variety of
freshwater and marine sediments because they are based on the biologically available
concentration of the substance in sediments. These criteria do not protect against additive
synergistic or antagonistic effects of contaminants or bioaccumulative effects to aquatic life'
wildlife or human health. '
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The criteria derivation methods outlined in this document are proposed to provide
protection of benthic organisms from-biological impacts from chemicals present in sediments.
Guidelines and guidance are being developed by U.S. EPA to assist in the application of criteria
presented in this document, in the development of sediment quality standards, and in other
water-related programs of this Agency.
These criteria are being issued in support of U.S. EPA'S regulations and policy
initiatives. This document is Agency guidance only. It does not establish or affect legal rights
or obligations. It does not establish a binding norm and is not finally determinative of the issues
addressed. Agency decisions in any particular case will be made by applying the law ana
regulations on the basis of the specific facts.
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Principal Author
David J. Hansen
Coauthors
Walter J. Berry
Dominic M. Di Toro
PaulR. Paquin
Laurie D. De Rosa
Frank E. Stahcil, Jr.
Robert A. Hoke
Christopher S. Zarba
Technical and Clerical Support
Heinz P. Kollig
Glen B. Thursby
Maria R. Paruta
Dinalyn Spears
Deborah Robson
ACKNOWLEDGEMENTS
U.S. EPA, Environmental Research Laboratory,
Narragansett, RT
Science Applications International Corporation,
Narragansett, RI
Manhattan College, Bronx, NY
HydroQual, Inc., Mahwah, NJ
HydroQual, Inc.,
Mahwah, NJ
HydroQual, Inc.,
Mahwah, NJ
•,
U.S. EPA, Environmental Research Laboratory, Athens, GA
Science Applications International Corporation,
Hackensack, NJ
U.S. EPA Headquarters, Office of Water, Washington, DC
U.S. EPA, Environmental Research Laboratory, Athens, GA
Science Applications International Corporation,
Narragansett, RI
NCSC Senior Environmental Employment Program,
Narragansett, RI
Computer Science Corporation, Narragansett, RI
Science Applications International Corporation
Narragansett, RI
IV
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Persons who have made significant contributions to the development of the approach and
supporting-science used.in the derivation of sediment criteria for nonionic organic contaminants
are as follows:
Herbert E. Allen
Gerald T. Ankley
Christina E. Cowan
Dominic M. Di Tore
David J. Hansen
PaulR. Paquin
Spyros P. Pavlou
Richard C. Swartz
Nelson A. Thomas
Christopher S. Zarba
University of Delaware, Newark, DE
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
Battelle, Richland, WA
HydroQual, Inc., Mahwah, NJ;
Manhattan College, Bronx, NY
U.S. EPA, Environmental Research Laboratory,
Narragansett, RI
HydroQual, Inc., Mahwah, NJ
Ebasco Environmental, Bellevue, WA
U.S. EPA, Environmental Research Laboratory,
Newport, OR
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
U.S. EPA Headquarters, Office of Water, Washington, DC
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Table 2-1.
Table 2-2.
Table 2-3.
Table 3-1.
Table 3-2.
Table 3-3.
Table 4-1.
Table 4-2.
V
Table 5-1.
Table 5-2.
Dieldrin measured and estimated log10KoW values..
Summary of log10KoW values for dieldrin measured by the U.S EPA
Environmental Research Laboratory, Athens, GA.
Summary of KQC values for dieldrin derived from literature sorption isotherm
data.
Chronic sensitivity of freshwater and saltwater organisms to dieldrin
Test specific data.
Summary of freshwater and saltwater acute land chronic values, acute-chronic
ratios, and derivation of final acute values, final acute-chronic ratios, and final
chronic values for dieldrin.
Results of approximate randomization test for the equality of freshwater and
saltwater FAV distributions for dieldrin and approximate randomization test for
the equality of benthic and combined benthic and water column (WQC) FAV
distributions.
Summary of tests with dieldrin-spiked sediment.
Water-only and sediment LC50s used to test the applicability of the equilibrium
partitioning theory for dieldrin.
Sediment quality criteria for dieldrin.
Analysis of variance for derivation of sediment quality criteria confidence
limits for dieldrin.
Table 5-3. Sediment quality criteria confidence limits for dieldrin.
Appendix A. - Summary of acute values for dieldrin for freshwater and saltwater species.
Appendix B. -Summary of data from sediment spiking experiments with dieldrin. Data
from these experiments were used to calculate KOC values (Figure 2-2) and to
compare mortalities of test organisms with pore water toxic units (Figure 4-1)
and predicted sediment toxic units (Figure 4-2).
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FIGURES
• J
, t
Figure 1-1. Chemical structure and physical-chemical properties of dieldrin.
Figure 2-1 Observed versus predicted (equation 2-4) partition coefficients for nonionic
organic chemicals (dieldrin datum is highlighted).
Figure 2-2. Organic carbon-normalized sorption isotherm for dieldrin (top) and probability
plot of KOC (bottom) from sediment toxicity tests conducted by Hoke and Ankley
(1991). The line in the top panel represents the relationship predicted with a log
of 5.25, that is 0.^=1^ • Cd
Figure 3-1. Genus mean acute values from water-only acute toxicity tests using freshwater
species vs. 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, X = unspecified life
stage.
Figure 3-2. Genus mean acute values from water-only acute toxicity tests using saltwater
species vs. 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.
Figure 3-3. Probability distribution of FAV difference statistics to compare water-only data
from freshwater vs. saltwater (upper panel) and benthic vs. WQC (lower panel)
data.
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.s 1990), and midge in sediments spiked with
dieldrin (Hoke, 1992) or kepone (Adams et al. , 1985) relative to pore water toxic
units. Pore water toxic units are ratios of concentrations of chemicals measured
in individual treatments divided by the water-only LC50 value from water-only
tests. (See Appendix B in this SQC document, Appendix B in the endrin,
acenaphthene, fluoranthene and phenanthrene SQC documents, and original
references for raw data.)
Figure 4-2. Percent mortality of amphipods in sediments spiked with acenaphthene or
phenanthrene (Swartz, 1991), dieldrin (Hoke and Ankley, 1991), endrin (Nebeker
et al. , 1989; Schuytema et al. , 1989) or fluoranthene (Swartz et al. , 1990; DeWitt
- et al., 1992) and midge, in dieldrin spiked sediments (Hoke, 1992) relative to
"predicted sediment toxic units. " Predicted sediment toxic units are the ratios of
measured treatment concentrations for each chemical in sediments Otg/goc)
divided by the predicted LC50 Otg/goc) in sediments (Koc x Water-Only LC50
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G*g/L) • IKg^/l.OOO&J. (See Appendix B in this document and Appendix B in
• theace,naphthene, endrin, fluoranthene, andphenanthrene SQC documents for raw
data).
Figure 5-1. Comparison between SQC concentrations and 95% confidence intervals effect
concentrations from benthic organisms expoajd to dieldrin-spiked sediments and
sediment concentrations predicted to be chronically safe in fresh water sediments
Concentrations predicted to be chronically safe (Predicted Genus Mean Chronic
Values, PGMCV) are derived from the Genus Mean Acute Values (GMAV) from
*™™?y 9,™°U/ lethaHty ***• Acute Chronic ""
VUl
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This report has been reviewed by the Health and Ecological Criteria Division, Office oi
Science and Technology, U.S. Environmental Protection Agency, and approved for publication.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
AVAILABILITY NOTTCE
This document is available to the public through the National Technical Information
Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161. NTIS Accession Number
XXXX-XXXXXX.
IX
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SECTION 1
INTRODUCTION
1.1 GENERAL INFORMATION:
Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U.S.
EPA) is responsible for protecting the chemical, physical and biological integrity of the nation's
waters. In keeping with this responsibility, U.S. 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 the U.S. EPA's best estimate of concentrations protective of human
health and the presence and uses of aquatic life. While these WQC play an important role in
assuring 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
quality criteria (SQC) makes it difficult to accurately assess the extent of the ecological risks of
contaminated sediments and to identify, prioritize and implement appropriate clean up activities
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and source controls. As a result of the need for a procedure to assist regulatory agencies in
***••*• • ,
' • '
making decisions concerning contaminated sediment problems, a U.S. EPA Office of Science
and Technology, Health and Ecological Criteria Division (OST/HEC) research team was
established to review alternative approaches (Chapman, 1987). All of the approaches reviewed
had both strengths and weaknesses and no single approach was found to be applicable for SQC
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 SQC applicable across a broad range of sediment
types. The three principal observations that underlie the EqP method of establishing SQC are:
1. The concentrations of nonionic organic chemicals in sediments, expressed on an
•,
organic carbon basis, and in pore 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 chemical concentrations
in pore water.
3. The distribution of sensitivities of benthic and water column organisms to
chemicals are similar, thus, the currently established WQC final chronic values
(FCV) can be used to define the acceptable effects concentration of a chemical
freely-dissolved in pore water.
The EqP approach, therefore, assumes that: (1) the partitioning of the chemical between
sediment organic carbon and interstitial water is at equilibrium; (2) the concentration in either
. phase can be predicted using appropriate partition coefficients and the measured concentration
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in the other phase; (3) organisms receive equivalent exposure from water-only exposures or from
.'.''**
any equilibrated phase: either from pore water via respiration, from sediment via ingestion or
other sediment-integument exchange or from a mixture of both exposure routes; (4) for nonionic
chemicals, effect concentrations in sediments on an organic carbon basis can be predicted using
the organic carbon partition coefficient (Koc) and effects concentrations in water, (5) the FCV
concentration is an appropriate effects concentration for freely-dissolved chemical in interstitial
water; and (6) the SQC (pg/goc) derived as the product of the K^ and FCV is protective of
benthic organisms. SQC concentrations presented in this document are expressed as pg
chemical/g sediment organic carbon and not on an interstitial water basis because: (1) pore water
is difficult to adequately sample; and (2) significant amounts of the dissolved chemical may be
if
associated with dissolved organic carbon; thus, total chemical concentrations in interstitial water
may overestimate exposure.
The data that support the EqP approach for deriving SQC for nonionic organic
chemicals are reviewed by Di Toro et al. (1991) and U.S. EPA, (1993a). Data supporting these
observations for dieldrin are presented in this document.
SQC generated using the EqP method 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; and
5. protective of benthic organisms.
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As is the case with WQC, the SQC reflect the use of available scientific data to: (1) assess the
i
likelihood of significant environmental effects to benthic organisms from chemicals in sediments;
and (2) to derive regulatory requirements which will protect against these effects.
It should be emphasized that these criteria are intended to protect benthic organisms from
the effects of chemicals associated with sediments. SQC are intended to apply to sediments
permanently inundated with water, intertidal sediment and to sediments inundated periodicaUy
for durations sufficient to permit development of benthic assemblages. They do .not apply to
occasionally inundated soils containing terrestrial organisms. These criteria do not address the
question of possible contamination of upper trophic level organisms or the synergistic, additive
or antagonistic effects of multiple chemicals. SQC addressing these issues may result in values
lower or higher than those presented in this document. The SQC presented in this document
represent the U.S. EPA's best recommendation at this time of the concentration of a chemical
in sediment that will not adversely affect most benthic organisms. SQC values may be adjusted
to account for future data.
SQC values may also need to be adjusted because of site specific consideration. In spill
situations, where chemical equiUbrium between water and sediments has not yet been reached,
sediment chemical concentrations less than SQC 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 TOG in the sediment does not effect 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 SQC may not pose risks to benthic organisms, because the
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compound may be a component of a paniculate, such as coal or soot, or exceed solubility such
'',-.-•
as undissolved oil or chemical. In these situations, the national SQC would be overly protective
of benthic organisms and should not be used unless modified using the procedures outlined in
the "Guidelines for Deriving Site-specific Sediment Quality Criteria for the Protection of Benthic
Organisms" (US EPA, 1993b). The SQC may be underprotective where the toxicity of other
chemicals are additive with the SQC chemical or species of unusual sensitivity occur at the site.
This document presents the theoretical basis and the supporting data relevant to the
derivation of the SQC for dieldrin. An understanding of the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses"
(Stephan et aL, 1985), response to public comment (U.S. EPA, 1985) and "Technical Basis for
Deriving Sediment Quality Criteria for Nonionic Organic Contaminants for the Protection of
Benthic Organisms by Using Equilibrium Partitioning" (U.S. EPA 1993a) is necessary in order
to understand the following text, tables and calculations. Guidance for the acceptable use of
SQC values is contained in "Guide for .the Use and Application of Sediment Quality Criteria for
Nonionic Organic Contaminants" (U.S. EPA, 1993c).
1.2 GENERAL INFORMATION: DIELDRIN
Dieldrin is the common name of a persistent, non-systemic organochlorine insecticide used
for control of public health insect pests, termites and locusts. It is formulated for use as an
emulsifiable concentrate, wettable and dustable powder and granular product. Other than direct
usage of dieldrin, another source of dieldrin in the environment stems from the quick
transformation of aldrin, also an organochlorine pesticide, to dieldrin. Both dieldrin and aldrin
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usage peaked in the mid-1960s and declined until the early 1970s. Dieldriri and aldrin have been
restricted from registration and production in the United States since 1974 (U.S. EPA, 1980a).
Dieldrin is a cyclic hydrocarbon having a chlorine substituted methanobridge (Figure 1-1).
It is structurally similar to endrin, its endo-endo stereoisomer, and has similar physico-chemical
chlorine properties, except that it is more difficult to degrade in the environment (Wang, 1988).
Dieldrin is a colorless crystalline solid at room temperature, having a melting point of about
176°C and specific gravity of 1.75 at 20°C. It also has a vapor pressure of 0.4 mEa and a
solubility of 0.19 mg/L at 20°C (Hartley and Kidd, 1987).
Dieldrin is considered to be toxic to aquatic organisms, bees and mammals (Hartley and
Kidd, 1987). The acute toxicity of dieldrin ranges from 0.5 to 740 ug/L for freshwater and 0.7
to > 100 /tg/L for saltwater organisms (Appendix A). Differences between dieldrin
concentrations causing acute lethality and chronic toxicity in specie.? acutely sensitive to this
insecticide are small; acute-chronic ratios range from 2.417 to 12.82 for three species (Table 3-
3). Dieldrin bioconcentrates in aquatic animals from 400 to 68,000 times the concentration in
water (U.S. EPA, 1980a). The WQC for dieldrin (U.S. EPA, 1980a) is derived using a Final
Residue Value calculated using bioconcentration data and the FDA action level to protect
marketability of fish and shellfish; therefore, the WQC is not "effects based". The SQC for
dieldrin is effects based. It is calculated from the Final Chronic Value (FCV) derived in section
3.
1.3 OVERVIEW OF DOCUMENT:
Section 1 provides a brief review of the EqP methodology, and a summary of the
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JA, V
ci
MOLECULAR FORMULA C12H,C160
MOLECULAR WEIGHT 380.93
DENSITY ,, 1.75 g/cc (20-0
MELTING POINT 176°C
PHYSICAL FORM Colorless crystal
VAPOR PRESSURE 0.40 mPa (20eC)
CAS NUMBER: 60-57-1
TSL NUMBER: IO 15750
COMMON NAME: Dieldrin (also dieldrine and ndieldrin)
^TRADE NAME: Endrex (Shell); Hexadrin
CHEMICAL NAME: l,2,3,4,10,10shexachloro-lR,4S,4aS,5R,6R,7S,8SR,8aR-
octahydro-6 J-epoxy-l,4:5,8-dimethanoaphthalene (IUPAC)
FIGURE 1-1. Chemical structure and physical-chemical properties of dieldrin.
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physical-chemical properties and aquatic toxicity of dieldrin. Section 2 reviews a variety of
methods and'data useful in deriving partition coefficients for dieldrin and includes the KQC
recommended for use in the derivation of the dieldrin SQC. Section 3 reviews aquatic toxicity
data contained in the dieldrin WQC document (U.S. EPA, 1980a) and new data that were used
to derive the FCV used in this document to derive the SQC concentration. In addition, the
comparative sensitivity of benthic and water column species is examined as the justification for
the use of the FCV for dieldrin in the derivation of the SQC. Section 4 reviews data on the
toxicity of dieldrin in sediments, the need for organic carbon normalization of dieldrin 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 are used in Section 5 as the basis
»f
for the derivation of the SQC for dieldrin and its uncertainty. The SQC for dieldrin is then
compared to STORET (U.S. EPA, 1989b) and National Status and Trends (NOAA, 1991) data
on dieldrin's environmental occurrence in sediments. Section 6 concludes with the criteria
statement for dieldrin. The references used in this document are listed in Section 7.
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SECTION 2
PARTITIONING
2.1 DESCRIPTION OF THE EQUILIBRIUM PARTITIONING METHODOLOGY:
Sediment quality criteria (SQC) are the numerical concentrations of individual chemicals
which 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 consequence, they can be used in much the same way as water
quality criteria (WQC); ie., the concentration of,a chemical which is protective of the intended
use such as aquatic life protection. For nonionic organic chemicals, SQC are expressed as ;tg
chemical/g organic carbon and apply to sediments having Ss 0.2% organic carbon by dry
weight. A brief overview follows of the concepts which underlie the equilibrium partitioning
(EqP) methodology for deriving SQC. The methodology is discussed in detail in the "Technical
Basis for Deriving Sediment Quality Criteria for Nonionic Organic Contaminants for the
Protection of Benthic Organisms by Using Equilibrium Partitioning" (U.S. EPA, 1993a),
hereafter referred to as the SQC 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 a SQC 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
lj :
correlated with the interstitial water (i.e., pore water) concentration 0»g chemical/liter pore
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1991). From a purely practical point of view, this correlation suggests that if it were possible
to measure the pore 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 SQC. It is for this reason that the methodology described below is called the
equilibrium partitioning (EqP) method.
It is shown in the SQC Technical Basis Document (U.S. EPA, 1993a) that the final acute
values (FAVs) in the WQC documents are appropriate for benthic organisms for a wide range
of chemicals. (The data showing this for dieldrin are presented in Section 3). Thus, a SQC can
be established using the final chronic value (FCV) derived using the WQC Guidelines (Stephan
et aL, 1985) as the acceptable effect concentration in pore or overlying water (see Section 5),
and the partition coefficient can be used to relate the pore water concentration to the sediment
concentration via the partitioning equation. This acceptable concentration in sediment is the
SQC.
The calculation is as follows: Let FCV fog/L) be the acceptable concentration in water
for the chemical of interest; then compute the SQC using the partition coefficient, (Kp)
O-^giediment)} between sediment and water:
SQC = Kp FCV (2-l)
This is the fundamental equation used to generate the SQC. Its utility depends upon the
existence of a methodology for quantifying the partition coefficient, Kp.
Organic carbon appears to be the dominant sorption phase for nonionic organic chemicals
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in naturally occurring sediments and thus controls the bioavailabffity of these compounds in
sediments. Evidence for this can be found in numerous toxicity tests, bioaccumulation studies
and chemical analyses of pore water and sediments (Di Toro et aL, 1991). The evidence for
dieldrin is discussed in this section and section 4. The organic carbon binding of a chemical in
sediment is a function of that chemical's organic carbon partition coefficient (Koc) and the
weight fraction of organic carbon in the sediment (f^. The relationship is as follows:
KP = foe KOC (2-2)
It follows that:
SQCOC = KOCFCV p_3)
where SQCoc is the sediment quality criterion on a sediment organic carbon basis.
Koc is not usually measured directly (although it can be done, see section 2.3).
Fortunately, KOC is closely related to the octanol-water partition coefficient (KoW) (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 KoWfor dieldrin.
2.2 DETERMINATION OF KOW FOR DIELDRIN:
Several approaches have been used to determine KOW for the derivation of SQC, as discussed
in the SQC Technical Basis Document. At the U.S. EPA, Environmental Research Laboratory
at Athens, GA (ERL,A) three methods were selected for measurement and two for estimation
of KOW. The measurement methods were shake-centrifugation (SC), generator column (GCol),
and slow-stir-flask (SSF), and the estimation methods were SPARC (SPARC Performs
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Automated Reasoning in Chemistry; Karickhoff et al., 1989) and CLOGP (Chou and Jure,
1979). Data were also extracted from the literature. The SC method is a standard procedure in
the Organization for Economic Cooperation and Development (OECD) guidelines for testing
chemicals, therefore, it has regulatory precedence.
In the examination of the literature data primary references were found listing measured
KQWS for dieldrin ranging from 4.09 to 6.2 (Table 2-1). Primary references were found in the
literature for estimated log10KoW ranging from 3.54 to 5.40 (Table 2-1). The range of reported
values for dieldrin is significantly greater than the range of values for some other compounds,
and we were not able to determine from studying the primary articles that any value was more
likely to be accurate than any other.
•.
TABLE 2-1. DIELDRIN MEASURED AND ESTIMATED LOG,0KoW VALUES .
METHOD
Measured
Measured
Measured
Measured
Measured
Estimated
Estimated
LOG10KoW
4.09
4.54
4.65
5.40
6.2
3.54
5.40
REFERENCE
Ellington and Stancil, 1988
Brooke, et al., 1986
De Kock and Lord, 1987
De Bruijn et al., 1989
Briggs, 1981
Mabey etal., 1982
SPARC'
•SPARC is from SPARC Performs Automated Reasoning in Chemistry, (Karickhoff et al.,
1989).
A KOW value for SPARC is also included in Table 2-1. SPARC is a computer expert system
under development at ERL,A, and the University of Georgia, at Athens. For more information
on SPARC see U.S. EPA (1993a). The SPARC estimated log10KoW value for dieldrin is 5.40.
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We had little confidence in the available measured or estimated values for K
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Laboratory studies to characterize adsorption are generally conducted using particle
» • ' "' ' i '•'
suspensions. The high concentrations of solids and turbulent conditions necessary to keep the
mixture in suspension make data interpretation difficult as a inesult of a particle interaction effect.
This effect suppresses the partition coefficient relative to that observed for undisturbed sediments
(Di Toro, 1985; Mackay and Powers, 1987).
TABLE 2-2. SUMMARY OF LOG10Kow VALUES FOR DIELDRIN MEASURED BY THE
U.S. EPA, ENVIRONMENTAL RESEARCH LABORATORY, ATHENS, GA.
SHAKE
CENTRIFUGATION
(SC)
5.04
5.00
5.04
5.03
5.04
4.88
4.99
5.01"
GENERATOR
COLUMN
(GCol)
4.89
4.88
5.1S
5.15
5.26
5.38
5.67'
5.04
5.16b
SLOW STIR
FLASK
(SSF)
5.33
5.43
5.38
5.33
5.43
5.08
5.28
5.34b
"Value considered outlier and omitted from mean Computation.
bLog10 of mean measured values.
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 KP:
(2-4)
1 +
KQC / V x
2-6
-------
where m is the particle concentration in the suspension (kg/L), and u* = 1.4, an empirical
constant. The K^ is given by:
= 0.00028 + 0.983 log10KoW (2-5)
Figure 2-1 compares observed partition coefficient data for the reversible component with
calculated values estimated with the particle interaction model (Equation 2-4 and Equation 2-5)
for a wide range of compounds (Di Toro, 1985). The observed partition coefficient for dieldrin
using adsorption data (Sharom et aL, 1980) is highlighted on this plot. The observed log10Kp
of 1.68 reflects significant particle interaction effects. The observed partition coefficient is more
than an order of magnitude lower than the value expected in the absence of particle effects ( i.e. ,
log10Kp = 3.32 from the fooKoc = 2100 L/kg).' KQC was computed from equation 2-5.
Several sorption isotherm experiments with particle suspensions that provide an additional
way to compute KQC were found in a comprehensive literature search for partitioning information
for dieldrin (Table 2-3). TheKoc values derived from these data are lower than KQC values from
laboratory measurements of KOW. The lower KQC can be explained from the particle interaction
effects. Partitioning in a quiescent setting would result in less desorption and higher KOC. These
data are presented as examples of particle interaction effects only as 100 percent reversibility
is assumed in the absence of desorption studies and actual K^ can not be computed.
In the absence of particle effects, KQC is related to KQW via Equation 2-5. For logioKow
= 5.34 (ERL,A, mean measured value), this expression results in an estimate of log10Koc =
5.25.
2-7
-------
O)
\
a
o
TJ
d>
a>
€0
JO
O
Partition Coefficient
Reversible Component
Predicted log 10 Kp (L/kg)
Figure 2-1
Ob^ived versus predicted (equation 2-4) partition coefficients for nonionic
organic chemicals (dieldrin datum is highlighted).
2-8
-------
2.3.2. KQC FROM SEDIMENT TOXTCITY TESTS:
Measurements of K^ are available from sediment toxicity tests using dieldrin (Hoke and
Ankley, 1992). These tests were with a sediment having an average organic carbon content of
1.75 percent (Table 4-1; Appendix B). Dieldrin concentrations were measured in sediments and
unfiltered pore waters providing the data necessary to calculate the partition coefficient for an
undisturbed bedded sediment. Since it is likely that organic carbon complexing in pore water
is significant for dieldrin, organic carbon concentrations were also measured in pore water.
Figure 2-2 is a plot of the organic carbon-normalized sorption isotherm for dieldrin, where the
sediment dieldrin concentration frg/g^ is plotted versus calculated free (dissolved) pore water
concentration (pg/L). Using pore water organic carbon concentrations (DOC), and assuming
KBOC equal to K^, the calculated tree pore water dieldrin concentration CD 0*g/L) is presented
in Figure 2-2 is given by:
l+mpoeKpoc (2'6>
where Croro is the measured total pore water concentration and m^ is the measured DOC
concentration (U.S. EPA, 1993a).
The data used to make this plot are included in Appendix B. The line of unity slope
corresponding to the log^ = 5.25 derived from SSF is compared to the data. The data from
the sediment toxicity test fall on the line of unity slope for logl^oc = 5.25.
A probability plot of the observed experimental log10Koc values is shown in the lower
,panel of Figure 2-2. The lo&oKoc values are approximately normally distributed with a mean
of Log10 KOC = 5.32 and a standard error of the mean of 0.109. This value is in agreement
2-9
-------
i-Hiiii—i i i IIIHI—r-rn
LEGEND
- Hoke ond Ankley, 1992
0.1 — » i i HUH — '
001
0.1
i t i mi
i tun § i i
100
1000
CALCULATED PORE WATER CONCENTRATION
-------
with log10Koc = 5.25, which was computed from the SSF determined (Section 2.2) dieldrin
5.34 (Equation 2-5).
TABLE 2-3. SUMMARY OF KOC VALUES FOR DIELDRIN
DERIVED FROM LITERATURE SORPTION ISOTHERM DATA.
Observed
LogioKoc (SD)
4.20 (0.14)
4.14 (0.15)
4.10
n
4
3
1
Solids (SD)
(g/L)
5.0
16.4 (4.6)
100.0
References
Eye, 1968
Betsffl, 1990
Briggs, 1981
2.4 SUMMARY OF DERIVATION OF KOC FOR DIELDRIN:
The KOC selected to calculate the sediment quality criteria for dieldrin is based on the
regression of log^ to log10KoW (Equation 2-5), using the dieldrin log10KoW of 5.34 recently
measured by ERL,A. This approach, rather than the use of the K^ from 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 K<,w and f^.
The regression equation yields a log^ = 5.25. This value is in agreement with the log
of 5.32 measured in the sediment toxicity tests.
2-11
-------
-------
SECTIONS
TOXICITY OF DIELDRIN: WATER EXPOSURES
3.1 TOXICTrY OF DffiLDRIN IN WATER: DERIVATION OF DffiLDRIN WATER
QUALITY CRITERIA:
The equilibrium partitioning (EqP) method for derivation of sediment quality criteria
(SQC) uses the dieldrin water quality criterion (WQC) Final Chronic Value (FCV) and partition
coefficients (Koc) to estimate the maximum concentrations of nonionic organic chemicals in
sediments, expressed on an organic carbon basis, that will not cause adverse effects to benthic
organisms. For this document, life stages of species classed as benthic are either species that
five in the sediment (infauna) or on the sediment surface (epibenthic) and obtain their food from
either the sediment or water column (U.S. EPA, 1989c). In this section (1) the FCV from the
dieldrin WQC document (U.S. EPA, 1980a) is revised using new aquatic toxicity test data; and
(2) the use of this FCV is justified as the effects concentration for SQC derivation.
3.2 ACUTE TOXICITY - WATER EXPOSURES:
One hundred and forty five standard toxicity tests with dieldrin have been conducted on
25 freshwater species from 19 genera (Appendix A). Eighty six of these tests are from one
study with the guppy, Poecilla reticulata (Chadwick and Kiigemagi, 1968). Overall genus mean
acute values (GMAVs) range from 0.5 to 740 /tg/L. Fishes, damselflys, isopods, glass shrimp,
stoneflies, and mayflies were most sensitive; GMAVs for these taxa range from 0.5 to 24 pg/L.
Seventeen tests on thirteen benthic species from twelve genera are contained in this database
(Figure 3-1; Appendix A). Benthic organisms were among both the most sensitive, and most
3-1
-------
1000
100
i
ui
I
HI
1
IU
S
CO
i
o
10
0.1
A Arthropods
D Other Invertebrates
O.Fishes
Orconectes (A)'
Gammarus (A,X)
Simocephalus (J,X)
Daphnia (A,J,X)
Ephemerella (X)
Micropterus
_'Asellus (X)
fQnoprhynchus (J.X)
Pteronarcys (J, naiads)
[ Acroneuria (naiads)
*alaemonetes (X)
*-
Ischnura (J)
gpomis(J)
arassius (J,X)
Claassenfa (J)
Pleronarcella (J)
_L
J_
J_
20 40 (50 80
PERCENTAGE RANK OF FRESHWATER GENERA
100
Figure 3-1. Genus mean acute values from water-only acute toxicity tests using freshwater
species vs. 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, X = unspecified life
stage.
3-2
-------
resistant, freshwater species to dieldrin; GMAVs range from 0.5 to 740 pg/L. Of the epibenthic
species tested, channel catfish, stoneflies, mayflies, damselflies, and isopods were most
sensitive; GMAVs range from 0.5 to 12 /tg/L. infaunal species tested include only the
stoneflies, Pteronarcvs califomica (LC50 = 4.416 ftgfL) and Pteronarcella badia (LC50 = 0.5
jtg/L). The final acute value (FAV) derived from the overall GMAVs (Stephan et al. 1985) for
freshwater organisms is 0.3595 /xg/L (Table 3-2).
Thirty two acute tests have been conducted on 23 saltwater species from 21 genera
(Appendix A). Overall GMAVs range from 0.70 to > 100 jtg/L. Sensitivities of saltwater
organisms were similar to those of freshwater organisms. Fishes and crustaceans were the most
sensitive. Within this database there are results from 23 tests on benthic life-stages of 16 species
from 14 genera (Figure 3-2; Appendix A). Benthic organisms were among both the most
sensitive, and most resistant, saltwater genera to dieldrin. The most sensitive benthic species
is the pink shrimp, Peneaus duorarum. with a flow-through 96 hour LC50 of 0.70 pg/L based
on measured concentrations. The American eel, Anquilla rostrata. has a similar sensitivity to
dieldrin with a 96 hr LC50 of 0.9 jtg/L. other benthic species for which there are data appear
less sensitive; GMAVs range from 4.5 to > 100 jtg/L. The FAV derived from the overall
GMAVs (Stephan et al., 1985) for saltwater organisms is 0.6594 jtg/L (Table 3-2), less than the
acute value for the economically important P. duorarum.
3.3 CHRONIC TOXZCITY - WATER EXPOSURES:
Chronic toxicity tests have been conducted with dieldrin using two freshwater fish:
rainbow trout, Oncorhynchus myldss. and the guppy, P. reticulata. and a saltwater mysid,
3-3
-------
1000
100
i
(II
I
I
CO
til
o
10
A Arthropods
D Other Invertebrates
O Fishes
Ophyryotrocha *(A)
Poecilia (J)
Cyprinodon (A)
Sphaeroides (A)
Fundulus (J)
~ Crangon (A)
Thalassoma (A)
. 'MenidiafJ)
Mysidopsis (fl)
'Micrometrus(A) '
Morone (J) s^ Crassostrea (A)
^BMugit(A)
Palaemonetes (A}
Pagurus(A)
Gasterosteus (J)
Palaemon (A)
Cymatogaster (J)
'Oncorhynchus (J)
'Anguilla (J)
Penaeus (A)
0.1
20 40 60 80
PERCENTAGE RANK OF SALTWATER GENERA
100
Figure 3-2.
Genus mean acute values ftom water-only acute toxicity tests using saltwater
species vs. percentage rank of their sensitivity. Symbok represen
species are solid, those representing water column species a£ open
indicate greater than values. A - adult, J = juvenile
3-4
-------
Mysidopsis £*> (Table 3-1).. Both Q. mvkiss and the M- bahia have benthic life stages.
(Chronic toxicity tests using Q. myldss and P. reticulata fail to meet the test requirement of
measured concentration for use in deriving WQC. Recently, an early life-stage test was
successfully completed using rainbow trout, Q. mvkiss (Brooke, 1993). The acute-chronic ratio
ACR, from this test (11.39) was almost identical to the value of 12.82 from unmeasured tests
with this fish (Table 3-1; 3-2). This new value will be added to this document following public
comment. Time did not permit its inclusion in this draft.
Dieldrin concentrations were not measured in freshwater tests. However, the nominal
and measured concentrations in the salt water M- bahia chronic test differed by less than 20%
at all concentrations. One life x;ycle test has been conducted with Q. mvkiss (Chadwick and
>^
Shumway 1969). There was a 97% reduction in survival and a 36% reduction in growth of the
survivors in 0.39 jtg/L relative to control fish; all fish died at 1.2 jtg/L. Q. mvkiss were not
significantly affected at concentrations of 0.012 to 0.12 pg/L. No progeny were tested. The
other freshwater chronic test was a three-generation study using the guppy, P. reticulata
(Roelofs, 1971). Because exposure concentrations were increased from the test with the first
generation to the tests with the next two generations, and because there was no effect at any
concentration in the first test, only results from the second two tests are reported here (Table
3-2). There was no effect on P. reticulata survival at dieldrin concentrations from 0.2 to 1.0
/tg/L. Mean brood size was reduced by 32% at 2.5 jtg/L.
Saltwater M- bahia exposed to dieldrin in a life-cycle test were affected at concentrations
similar to those affecting the two freshwater fish mentioned above. M. bahia exposed to 1.1 and
1.6 /tg/L (U.S. EPA, 1980b) had a 35% and 58% reduction in survival, respectively, relative
3-5
-------
3-6
-------
r-
in
A
01
in
A
«n
10
in
A
e
H
I
0
a
«
M
w
e
o
e
o
Mysid
i
ete
Pol
Ooh
3-7
-------
to control_M.bjhja. There were no significant effects at 0.10 to 0.49 jtg/L. No effects were
j ' ' *
observed on reproduction at any concentration tested and progeny response was not recorded.
One life-cycle and one partial life-cycle test were conducted with the polychaete worm,
.Ophryptrogha diadema (Hooftman and Vink, 1980; Tables 3-1 and 3-2). The observed nominal
no effect concentration was of 0.1 ^g/L (below limit of analytical detection) for the life-cycle
test initiated with larvae and 1.2 jtg/L (based on measured concentrations) for the partial life-
cycle test initiated with adults. For the life-cycle test with larvae there were 40, 37 81 and 99 %
decreases in reproductive potential, (combined effect on number of egg masses and embryo
survival), relative to carrier control worms at 0.3,1.5,3.1 and 13 jtg/L, respectively. Embryo
survival was reduced by 35, 16, 61 and 71 % at dieldrin concentrations of 0.3,1.5, 3.1 and 13
pg/L, respectively. At 13 jig/L dieldrin survival was reduced to 34% relative to the controls.
In the Q.diadema partial life-cycle test, reproductive potential was reduced by 57, 92, 97 and
100% relative to the carrier control in concentrations of 2.6, 8, 23 and 72 pgfL. Sixty-three
percent of adults in 72 /tg/L died. Reductions in egg survival were 39, 70, 62 and 100%
relative to controls in concentrations of 2.6, 8, 23 and 72 jigfL, respectively. The chronic
sensitivity of this species appears similar to that of the other species tested chronically but acute
sensitivity is low: 96 hr LC50 > 100 jtg/L for adults and larvae.
The difference between acute and chronic sensitivity to dieldrin for acutely sensitive
species is approximately an order-ofcmagnitude or less (Table 3-2). The acute-chronic ratio
(ACR) for acutely insensitive polychaetes was > 56.63 in one test and > 577.4 in a second.
The available ACRs for acutely sensitive species are 2.417 forP. reticulata, 6.129 for M. bahia
and 12.82 for.O. myMss. The Final Acute-Chronic Ratio (ACR), the geometric mean of these
3-8
-------
three values, is 5.748. - .
Hie FCVs (Table 3-2), are used as the effect concentrations for calculating the SQC for
benthic species. The FCV for freshwater organisms of 0.0625 pg/L is the quotient of the FAV
of 0.3595 ftg/L and the final ACR of 5.748. Similarly, the FCV for saltwater organisms of
0.1147 ftg/L is the quotient of the FAV of 0.6594 jtg/L and the final ACR of 5.748.
3.4 APPLICABILITY OF THE WATER QUALITY CRITERION AS THE EFFECTS
CONCENTRATION FOR DERIVATION OF THE DffiLDRIN SEDIMENT
QUALITY CRITERION:
The use of the FCV (the chronic effects-based WQC concentration) as the effects
concentration for calculation of the EqP-based SQC assumes that benthic (infaunal and
epibenthic) species, taken as a group, have sensitivities similar to all benthic and water column
species tested to derive the WQC concentration. Data supporting the reasonableness of this
assumption over all chemicals for which there are published or draft WQC documents are
presented in Di Toro et al. (1991), and the SQC Technical Basis Document (U.S. EPA, 1993a).
The conclusion of similarity of sensitivity is supported by comparisons between (1) acute values
for the most sensitive benthic species and acute values for the most sensitive water column
species for all chemicals; (2) acute values for all benthic species and acute values for all species
in the WQC documents across all chemicals after standardizing the LC50 values; (3) FAVs
calculated for benthic species alone and FAVs calculated for all species in the WQC documents;
and (4) individual chemical comparisons of benthic species vs. all species. Only in this last
comparison are dieldrin-specific comparisons of the sensitivity of benthic and all (benthic and
water-column) species conducted. The following paragraphs examine the data on the similarity
3-9
-------
of sensitivity of benthic and all species for dieldrin.
For dieldrin,'benthic species account for 12 out of 19 genera tested in fteshwater, and
14 out of 21 genera tested in saltwater (Figures 3-1, 3-2),, An initial test of the difference
between the freshwater and saltwater FAVs for all species (water column and benthic) exposed
to dieldrin was performed using the Approximate Randomization method (Noreen, 1989). The
Approximate Randomization method tests the significance level of a test statistic when compared
to a distribution of statistics generated from many random subsamples. The test statistic in this
case is the difference between the freshwater 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 3-1). In the
Approximate Randomization method, the freshwater LC50 values and the saltwater LC50 values
are combined into one data set. The data set is shuffled, them separated back so that randomly
generated "freshwater" and "saltwater" FAVs can be computed. The LC50 values are separated
back such that the number of LC50 values used to calculate the sample FAVs are the same as
the number used to calculate the original FAVs. These too FAVs are subtracted and the
difference used as the sample statistic. This is done many times so that the sample statistics
make up a distribution that is representative of the population of FAV differences (Figure 3-3).
Hie test statistic is compared to this distribution to determine it's level of significance. The null
hypothesis is that the LC50 values that comprise the saltwater and freshwater data bases are not
different. If this is true, the difference between the actual freshwater and saltwater FAVs should
be common to the majority of randomly generated FAV differences. For dieldrin, the test-
statistic falls at the 31 percentile of the generated FAV differences. Since the probability is less
3-10
-------
than 95%, the hypothesis of no significant difference in sensitivity for freshwater and saltwater
species is accepted (Table 3-3).
Since freshwater and saltwater species showed similar sensitivity, a test of difference in
sensitivity for benthic and all (benthic and water column species combined, hereafter referred
to as "WQC") organisms combining freshwater and saltwater species using the Approximate
Randomization method was performed. The test statistic in this case is the difference between
the WQC FAV, computed from the WQC LC50 values, and the benthic FAV, computed from
the benthic organism LC50 values. This is slightly different then the previous test for saltwater
and freshwater species. The difference is that saltwater and freshwater species in the first test
represent two separate groups. In this test the benthic organisms are a subset of the WQC
organisms set. In the Approximate Randomization method for this test, the number of data
points coinciding with the number of benthic organisms are selected from the WQC data set.
A "benthic" FAV is computed. The original WQC FAV and the "benthic" FAV are then used
to compute the difference statistic. This is done many times and the distribution that results is
representative of the population of FAV difference statistics. The test statistic is compared to
this distribution to determine its level of significance. The probability distribution of the
computed FAV differences are shown in the bottom panel of Figure 3-3. The test statistic for
this analysis falls at the 72 percentile and the hypothesis of no difference in sensitivity is
accepted (Table 3-3). This analysis suggests that the FCV for dieldrin based on data from all
tested species is an appropriate effects concentration for benthic organisms.
3-11
-------
TABtE 3-3. RESULTS OF APPROXIMATE RANDOMIZATION TEST FOR
THE EQUALITY OF THE FRESHWATER AND SALTWATER FAV
DISTRIBUTIONS FOR DIELDRIN AND APPROXIMATE
RANDOMIZATION TEST FOR THE EQUALITY OF BENTHIC AND
COMBINED BENTHIC AND WATER COLUMN (WQC) FAV
DISTRIBUTIONS.
Compar-
ison Habitat or Water Type1 AR Statistic" Probability6
Fresh. Fresh (19) Salt (21) loiioS31
vs Salt
Bentbic Bentbic (26) WQC (40) 0.090 72
vs Water
Column +
Bentbic (WQC)
•Values in parentheses are the number of LC50 values used in the comparison.
AR statistic = FAV difference between original compared groups.
"Probability that the theoretical AR statistic <. tiiiat the observed AR statistic given
that the samples came from the same population.
3-12
-------
6
4
8
2
1
0
-1
-2
-3
DIELDRIN
~i—i—i—i—i—i—r
1111111 i i "lull 1—i—r—i—i—i—i inin i i i
FRESHWATER VS SALTWATER
TT
O
O H
tf*9
0CD
» Himn
0.1
10 20
60
80 90
89
99.9
U
O
III
cc^^
Wsj
4
3
2
1
0
-1
•2
-3
- BENTHIC VS WQC
i i i i i—r
mil I
- o
-o
' "iimi i i u i
0.1
...
10 20
60
80 90
99
99.9
Figure 3-3.
PROBABILITY
Robabffity distribution of FAV difference statistics to compare water-only data
from freshwater vs. saltwater (upper panel) and benthic vs. WQC (lower panel)
3-13
-------
-------
SECTION 4
TOXIOTY OF DIELDRIN (ACTUAL AND PREDICTED): SEDIMENT EXPOSURE
4.1 TOXICITY OF DIELDRIN IN SEDIMENTS:
The toxicity of dieldrin spiked into clean sediments has been tested with two saltwater
species (a polychaete and the sand shrimp) and two freshwater species (an amphipod and a
• *,
midge) (Table 4-1). Therefore, generalizations of dieldrin's toxicity across species or sediments
are limited. The endpoint reported in these studies was mortality. Details about exposure
methodology are provided because, unlike aquatic toxicity tests, sediment testing methodologies
have not been standardized. Data are available from many experiments using both field and
laboratory sediments contaminated with mixtures of dieldrin and other compounds. Data from
these studies have not been included here because it is not possible to determine the contribution
of dieldrin to the observed toxicity.
The only saltwater experiments that tested dieldrin-spiked sediments were conducted by
McLeese et al. (1982) and McLeese and Metcalfe (1980). These began with clean sediments
that were added to dieldrin-coated beakers just prior to the addition of test organisms. This is
in marked contrast to tests with freshwater sediments that were spiked with dieldrin days or
weeks prior to test initiation. As a result, the dieldrin concentrations in the sediment and
4-1
-------
c
I)
a
.s
£
u
»
5 «4 y tn
Bt ™ »5 3.
ot
ot
il
ot
Ok
ot
ot
d
ot
01
i
n
CO
in
CO
u>
in
in
10
oo
§ f
Sf
e
^
E
a
•
n
fa
la
•H «
ArH
II
in
a o
4-2
-------
overlying water varied greatly over the course of these saltwater experiments and exposure
,.•'•*'
conditions are uncertain. In addition, transfer of test organisms to freshly prepared beakers
every 48 hours further complicates interpretation of results of McLeese et al. (1982) because
exposure conditions change several times during the course of the test. McLeese et al. (1982)
tested the effects of dieldrin on the polychaete worm, Nereis virens. in sediment with 2% TOC
(17% sand and 83% silt and clay) in 12 day toxicity tests. No worms died in 13 /*g/g dry wt
sediment, the highest concentration tested. McLeese and Metcalfe (1980) tested the effects of
dieldrin in sand with a TOC content of 0.28 % on the sand shrimp, Crangon septemspinosa. The
4 day LC50 was 0.0041 ug/g dry wt. sediment, and 1.46 ug/g^. Concentrations of dieldrin in
water overlying the sediment were 10 times the LC50 in water. The authors conclude that
sediment-associated dieldrin contributed little towards the toxicity observed.
The effects of dieldrin-spiked sediments from three fresh-water sites in Minnesota on the
fresh-water amphipod, Hvalella azteca have been studied by Hoke and Ankley (1991). The total
organic carbon (TOC) concentrations in the three sediments were 1.7%, 2.9%, and 8.7%. The
sediments were rolled in dieldrin-coated jars at 4°C for 23 days. Mortality of H. azteca in these
flow-through tests was related to sediment exposure because dieldrin concentrations in overlying
water were generally below detection limits. There was no dose-response relationship observed
in the results from the definitive test with one of the sediments (Airport Pond), or in the results
from further testing with this sediment using H- azteca (Hoke and Ankley, 1992; Hoke 1992).
For this reason only the data from the range finder test with this sediment are used in the
analysis of the toxicity data (sections 4.1, 4.2, 4.3), and in Figures 4-1 and 4-2. The ten-day
LCSO's increased with increasing TOC when dieldrin concentration was expressed on a dry
4-3
-------
weight basis, but increased only slightly with increasing organic carbon when dieldrin
'.<•''
concentration was expressed on an organic carbon basis (Table 4-1). LCSO's normalized to dry
weight differed by a factor of 21.2 (18.2 to 386 jtg/g)over a 5..0 fold range of TOC. In contrast,
the organic carbon normalized LCSO's differed by a factor of 3.4 (1,073 to 3,682 Mg/goc)-
The effects of dieldrin-spiked sediments from two freshwater sites in Minnesota on the
fresh water midge, Chironomus tentans. have been studied by Hoke (1992). The TOC contents
in the two sediments were 1.5 and 2.0%. The sediments weie rolled in dieldrin coated jars at
4°C for one month, stored at 4°C for two months, and then rolled at 4°C for an additional
month. LCSOs normalized to dry weight differed by a factor of 2.89 (0.53 to 1.53 ng/g dry wt).
LC50s normalized to organic carbon differed by a factor of 2.22 (35.33 to 78.46). It is not
v
surprising that organic carbon normalization had little effect, given the small range of TOC (1.5
to 2.0%).
Overall, the need for organic normalization of the concentration of nonionic organic
chemicals in sediments is presented in the Technical Basis Document (U.S.EPA, 1993a). The
need for organic carbon normalization for dieldrin is supported by the dieldrin-spiked toxicity
tests described above. Although it is important to demonstrate that organic carbon normalization
is necessary if SQC are to be developed using the EqP approach, it is fundamentally more
important to demonstrate that KQC and water only effects concentrations can be used to predict
the effects concentration for dieldrin and other nonionic organic chemicals on an organic carbon
basis for a range of sediments. Evidence supporting this pirediction for dieldrin and other
nonionic organic chemicals follows in section 4.3.
4-4
-------
4.2 CORRELATION BETWEEN ORGANISM RESPONSE AND PORE WATER
CONCENTRATION:
One corollary of the EqP theory is that freely dissolved pore-water LCSQs for a given
organism should be constant across sediments of varying organic carbon content (U.S.EPA,
1993a). Appropriate pore-water values are available from two studies (Table 4-2). Data from
tests with water column species were not considered in this analysis. Hoke and Ankley (1991)
found 10-day LC50 values for H. azteca based on pore-water concentrations differed by a factor
of 8.0 (57.6 to 458 jig/L) for three sediments containing from 1.7 to 8.7% TOC. Therefore,
pore water normalized LC50 values provide only a slight improvement over LC50s for dieldrin
expressed on a dry weight basis which varied by a factor of 21.2 (18.2 to 386 /tg/L). Hoke
(1992) found 10-day LC50 values for the,.£. tentans based on predicted pore water
concentrations (the sediment concentration multiplied by the KQC) differed by a factor of 2.17
(0.23 to 0.50). This variability is slightly less than that shown when dry wt (factor of 2.89) is
used, but similar to that shown when organic carbon (factor of 2.22) normalization is used.
Partitioning to dissolved organic carbon was proposed to explain the lack of similarity of LC50
values based on total pore water dieldrin concentrations.
A more detailed evaluation of the degree to which the response of benthic organisms can
be predicted from toxic units of substances in pore water can be made utilizing results from
toxicity tests with sediments spiked with other substances, including acenaphthene and
phenanthrene (Swartz, 1991), dieldrin (Hoke 1992), endrin (Nebeker et al., 1989; Schuytema
et al., 1989), fluoranthene (Swartz et al., 1990; DeWitt et al., 1992), or kepone (Adams et al.,
1985) (Figure 4-1; Appendix B). The data included in this analysis come from tests conducted
at EPA laboratories or from tests which utilized designs at least as rigorous as those conducted
4-5
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at the EPA laboratories. Tests with acenaphthene and phenanthrene used two saltwater
,.-•'"
amphipods (Leptocheirus plumulosus and Eohaustorius estuarius't and marine sediments. Tests
with fluoranthene used a saltwater amphipod (Rhepoxvnius abronius^ and marine sediments.
Freshwater sediments spiked with endrin were tested using the amphipod IL azteca: while
kepone-spiked sediments were tested using the midge, £, tentans. Figure 4-1 presents the
percentage mortalities of the benthic species tested in individual treatments for each chemical
versus "pore water toxic units"(PWTU) for all sediments tested. PWTUs are the concentration
of the chemical in pore water fctg/L) divided by the water only LC50 (pg/L). Theoretically,
50% mortality should occur at one interstitial water toxic unit. At concentrations below one
PWTU there should be less than 50% mortality, and at concentrations above one PWTU there
should be greater than 50% mortality. Figure 4-1 shows that at concentrations below one PWTU
mortality was generally low, and increased sharply at approximately one PWTU. Therefore this
comparison supports the concept that interstitial water concentrations can be used to predict the
response of an organism to a chemical that is not sediment specific. This pore water
normalization was not used to derive SQC in this document because of the complexation of
nonionic organic chemicals with pore water DOC (Section 2) and the difficulties of adequately
sampling pore waters.
4.3 TESTS OF THE EQUILIBRIUM PARTITIONING PREDICTION OF SEDIMENT
TOXICITY:
SQC derived using the EqP approach utilize partition coefficients and FCVs from WQC
documents to derive the SQC concentration for protection of benthic organisms. The partition
coefficient (KQC) is used to normalize sediment concentrations and predict biologically available
4-7
-------
concentrations across sediment types. The data required to test the organic carbon normalization
• • ' ' ' '
for dieldrin in sediments are available for 2 benthic species. Data from tests with water column
species were not included in this analysis. Testing of this component of SQG derivation requires
three elements: (1) a water-only effect concentration, such as a 10-day LC50 value in j*g/L; (2)
an identical sediment effect concentration on an organic caitbon basis, such as a 10-day LC50
value in jig/god and (3) a partition coefficient for the chemical, KQC in L/Kgoc. This section
presents evidence that the observed effect concentration iin sediments (2) can be predicted
utilizing the water effect concentration (1) and the partition coefficient (3).
Predicted ten-day LC50 values from dieldrin-spiked sediment tests with IL azteca (Hoke and
Ankley, 1991) wese calculated (Table 4-2) using the logw KOC value of 5.25 from Section 2 of
>f
this document and the water-only LC50 value (7.3 ug/L). Batios of actual to predicted LCSO's
for dieldrin averaged 1.26 (range 0.827 to 2.83) in tests with three sediments (Table 4-2).
Similarly, predicted 10-day LC50 values for dieldrin-spiked isedimenf tests with C. tentans were
calculated using the logw KQC of 5.25 and a 10-day water only LC50 value of 0.29 jig/L.
Ratios of predicted to actual LCSOs for dieldrin averaged 1.02 (range 0.69 to 1.52) in tests with
two sediments (Table 4-2). The overall mean for both species was 1.16.
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 came from tests conducted at EPA laboratories or from tests
which utilized designs at least as rigorous as those conducted at the EPA laboratories. Data
from the kepone experiments are not included because a measured K^ for kepone obtained using
4-8
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the slow stir flask method is not available. Swartz (1991) exposed the saltwater amphipods £,
SSfo^rius and'L. plumulpsus to acenaphthene in three marines sediments having organic carbon
contents ranging from 0.82 to 4.2% and to phenanthrene fin three marine sediments having
organic carbon contents ranging from 0.82 to 3.6%. Swartz et al. (1990) exposed the saltwater
j
amphipod E, abronius to fluoranthene in three marine sediments having 0.18, 0.31 and 0.48%
organic carbon. Hoke and Ankley (1991) exposed the ampMpod g, azteca to three dieldrin-
spiked freshwater sediments having 1.7, 2.9 and 8.7% organic carbon and Hoke (1992) exposed
the midge £. tentans to two freshwater dieldrin-spiked sediments having 2.0 and 1.5 % organic
carbon. Nebeker et al. (1989) and Schuytema et al. (1989) exposed H- azteca to three endrin-
spiked sediments having 3.0, 6.1 and 11.2% organic carbon. Figure 4-2presents thepercentage
mortalities of amphipods in individual treatments of each chemical versus "predicted sediment
toxic units" (PSTU) for each sediment treatment. PSTUs are the concentration of the chemical
in sediments fcg/goc) divided by the predicted LC50 Otg/goc) in sediments (the product of KOC
and the 10-day water-only LC50). In this normalization, 50% mortality should occur at one
PSTU. At concentrations below one PSTU mortality was generally low, and increased sharply
at one PSTU. The means of the LC50s for these tests calculated on a PSTU basis were 1.90,
for acenaphthene, 1.16 for dieldrin, 0.44 for endrin, 0.80 for fluoranthene, and 1.22 for
phenanthrene. The mean value for the five chemicals is 0.99. This illustrates that the EqP
method can account for the effects of different sediment properties and properly predict the
effects concentration in sediments using the effects concentration from water only exposures.
4-10
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','.••••" SECTIONS
CRITERIA DERIVATION FOR DffiLDRIN
5.1 CRITERIA DERIVATION:
The water quality criteria (WQC) Final Chronic Value (FCV), without an averaging
period or return frequency (See section 3), is used to calculate of sediment quality criteria (SQC)
because it is probable that the concentration of contaminants in sediments are relatively stable
over time, thus exposure to sedentary benthic species should be chronic and relatively constant.
This is in contrast to the situation in the water column, where a rapid change in exposure and
exposures of limited durations can occur due to fluctuations in effluent concentrations, dilutions
in receiving waters or the free-swimming or planktonic nature of water column organisms. For
some particular uses of the SQC it may be appropriate to use the area! extent and vertical
stratification of contamination of a sediment at a site in much the same way that averaging
periods or mixing zones are used with WQC.
The FCV is the value that should protect 95% of the tested species included in the
calculation of the WQC from chronic effects of the substance. The FCV is the quotient of the
Final Acute Value (FAV), and the final Acute Chronic Ratio (ACR) 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 for the genera from eight or more families 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
5-1
-------
requirements. For more information on the calculation of ACRs, FAVs, and FCVs see the
National Water Quality Criteria Guidelines (Stephan et al., 1985). The FCV used in this
document differs from the FCV in the dieldrin WQC document (U.S. EPA, 1980) because it
incorporates recent data not included in that document, and omits some data which does not meet
the data requirements established in the WQC Guidelines (Stephan et al., 1985).
The equilibrium partitioning (EqP) method for calculating SQC is based on the following
procedure. If FCV (pg/L) is the chronic concentration from the WQC for the chemical of
interest, then the SQC 0*g/g sediment), is computed using the partition coefficient, KP (L/g
sediment), between sediment and pore water:
SQC = KP FCV (5_!)
Since organic carbon is the predominant sorption phase for nonionic organic chemicals
in naturally occurring sediments, (salinity, grainsize and other sediment parameters have
inconsequential roles in sorption, see sections 2.1 and 4.3) the organic carbon partition
coefficient, (Koc) can be substituted for KP. Therefore, on a siediment organic carbon basis, the
, is:
SQCOC = KOCFCV (5_2)
Since (K^) is presumably independent of sediment type for non-ionic organic chemicals, so also
\
is SQCoc. Table 5-1 contains the calculation of the dieldrin SQC.
The organic carbon normalized SQC is applicable to sediments with an organic carbon
/
fraction of foe ^ 0.2%. For sediments with f^, < 0.2%, organic carbon normalization and
SQC may not apply.
5-2
-------
TABLE 5-1. SEDIMENT QUALITY CRITERIA FOR DffiLDRIN
Type of
Water Body
Fresh Water
Salt Water
(L/kg) W
5.34
5.34
(L/kg)
5.25
5.25
FCV
0*g/L)
0.0625
0.1147
SQCoc
0*g/goc)
11*
20"
= (105-25 L/kgocWlO-* kgoc/gocKO.0625 ug dieldrin/L) = 11 M
"SQCoc = (105-25 L/kgocWIO-* kgoc/goc)»(0.1147 jig dieldrin/L) - 20 pg di
Since organic carbon is the factor controlling the bioavailability of nonionic organic
compounds in sediments, SQC 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
concentration and organic carbon data are available, it is best to convert the sediment
concentration to pg chemical/gram organic carbon. These concentrations can then be directly
compared to the SQC value. This facilitates comparisons between the SQC and field
concentrations relative to identification of hot spots and the degree to which sediment
concentrations do or do not exceed SQC values. The conversion from dry weight to organic
carbon normalized concentration can be done using the following formula:
Ug Chemical/goc = A*g Chemical/gDRywT -~- (% TOC -^ 100)
= ftg Chemical/gDRYWT • 100 -5- % TOC
For example, afreshwater sediment with a concentration of 0.1 (tg chemical/gDRYWT and
0.5% TOC has an organic carbon-normalized concentration of 20 /tg/goc (0.1 ng/gDKrw[ • 100
- 0.5 = 20 Mg/goc) which exceeds the freshwater SQC of 11 ^g/g^. Another freshwater
sediment with the same concentration of dieldrin (0.1 Mg/gDRY w) but a TOC concentration of
5-3
-------
5.0% would have an organic carbon normalized concentration of 2.0 /ig/goc (0.1 /«g/gDRYWT *
i • •
100 -s- 5.0 = 2.0 /*g/goc), which is below the SQC for dieldrin.
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 organic
carbon-normalized SQC values (SQCoc) may be "converted" to dry weight-normalized SQC
values (SQCDRY WT.). This "conversion" must be done for each level of TOC of interest:
SQCDRywT = SQCoc Gtg/goc) • (% TOC * 100)
where SQCDRYWT is the dry weight normalized SQC value. For example, the SQC value for
freshwater sediments with 1% organic carbon is 0.11 /xg/g:
SQCDRYWT. = 11 jtg/goc • 1% TOC -5- 100 == 0.11 /*g/gDRYWT
i
This method is used in the analysis of the STORET data in section 5.4.
5.2 UNCERTAINTY ANALYSIS:
Some of the uncertainty in the calculation of the dieldrirt SQC can be estimated from the
degree to which the EqP model, which is the basis for the criteria, can rationalize the available
sediment toxicity data. The EqP model asserts that (1) the bioavailability of nonionic organic
chemicals from sediments is equal on an organic carbon basis, and (2) that the effects
concentration in sediment Otg/goc) can be estimated from thepiraduct of the effects concentration
from water only exposures 0*g/L) and the partition coefficient KQC (L/kg). The uncertainty
associated with the SQC 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 have been conducted to fulfill the minimum database requirements for the development of
5-4
-------
SQC (See Section 4.3 and Technical Basis Document, U.S. EPA, 1993a). These freshwater and
saltwater tests span a range of chemicals and organisms; they include both water-only and
sediment exposures and they are replicated within each chemical-organism-exposure media
treatment. These data were analyzed using an analysis of variance (ANOVA) to estimate the
uncertainty (i.e. the variance) associated with varying the exposure media and that associated
with experimental error. If the EqP model were perfect, then there would be only experimental
error. 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 dieldrin
are summarized in Appendix B. LCSOs for sediment and water-only tests were computed from
i^
these data. The EqP model can be used to normalize the data in order to put it on a common
basis. The LCSOs from water-only exposures (LC50*; ftg/L) are related to the organic carbon-
normalized LCSOs from sediment exposures (LC50S(OC; ^g/g^ via the partitioning equation:
LC50S>OC = oc (5.3)
The EqP model asserts that the toxicity of sediments expressed on an organic carbon basis equals
the toxicity in water tests multiplied by the KOC. Therefore, both LC50SfOC and KOC»LC50W are
estimates of the true LCSOoc for each chemical-organism pair. In this analysis, the uncertainty
of KOC 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.
5-5
-------
As discussed above, experiments that seek to validate equation 5-3 are subject to various sources
,.'''"' ' '
of random variations. A number of chemicals and organisms! have been tested. Each chemical -
organism pair was tested in water-only exposures and in different sediments. Let a represent
the random variation due to this source. Also, each experiment is replicated. Let € represent
the random variation due to this source. If the model were perfect, there would be no random
variations other than that due to experimental error which is reflected in the replications. Hence
a represents the uncertainty due to the approximations inherent in the model and € represents
the experimental error. Let (trj2 and (rresponding to a water-only or
sediment exposure; ^ are the population of ln(LC50) for chemical-organism pair i. The error
structure is assumed to be lognormal which corresponds to assuming that the errors are
proportional to the means, e.g. 20%, rather than absolute quantities, e.g. 1 jtg/L. The statistical
problem is to estimate /iis (oj2, and (
-------
Table 5-2: ANALYSIS OF VARIANCE FOR DERIVATION OF
SEDIMENT QUALITY CRITERIA CONFIDENCE LIMITS FOR
DIELDRIN.
Source of Uncertainty
Exposure media
Replication
Sediment Quality Criteria
Parameter Value
0.2%. For sediments with r^, < 0.2%, organic carbon normalization and
SQC do not apply.
5-7
-------
TABLE 5-3. SEDIMENT QUALITY CRITERIA
• CONFIDENCE LIMITS FOR DIELDRIN
Sediment Quality Criteria
95% Confidence Limits fW/«r__^
Type of
Water Body
Fresh Water
Salt Water
p j^\^i
*3>^^*Oi"^
A*g/goc
11
20
Lower
5.2
9.5
Upper
24
44
• 5.3 COMPARISON OF DIELDRIN SQC 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 SQC
concentrations and 95% confidence intervals can be inferred using effect concentrations from
toxicity tests with benthic species exposed to sediments spiked with dieldrin and sediment
concentrations predicted to be chronically safe to organisms tested jn water-only exposures
(Figures 5-1 and 5-2). Effect concentrations in sediments can be predicted from water-only
toxicity data and KOC values (See Section 4). Chronically acceptable concentrations are
extrapolated from genus mean acute value (GMAV) from water-only, 96-hour lethality tests
using acute-chronic ratios (ACR). Therefore, it may be reasonable to combine these two
predictive procedures to estimate, for dieldrin, chronically acceptable sediment concentrations
(Predicted Genus Mean Chronic Value, PGMCV)) from GMAVs (Appendix A), ACRs (Table
3-2) and the KOC (Table 5-1):
PGMCV = (GMAV -^ ACR)* K^ (5.7)
5-8
-------
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: A
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Other invertebrate
Fishes A A
Log^Koc-525
ACR.5.75 A A
SEDIMENTTESTS: lOdLCSO T
* C. tentans • 56.9ua/goc 1
range 2 tests -35.3 to 78.5»ig/goc ©
® H.,azteea » IddOfig/ggg -^
range 3 tests- 1070 to 3680 Q A A
A O 0
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- .-----..-„..„«....„„„„„ lower. S^g/goc
— ' -1 — — «- • ' « • ,
20 40 60 80 100
PERCENTAGE RANK OF FRESHWATER GENERA
Figure 5-1.
Comparison between SQC concentrations and 95% confidence intervals, effect
concentrations from benthic organisms exposed to dieldrin-spiked sediments and
sediment concentrations predicted to be chronically safe in fresh water sediments
SZ^^**0!?15° * °hronicaUy «* (Predicted Genus Mean Chronic
Values, PGMCV) are denved from the Genus Mean Acute Values (GMAV) from
S?V - £SfJ8?1??£?• ACUtC Chn)niC ""^ (ACR) ^ ** values.
PGMCy = (GMAV + ACR)Koc. Symbols for PGMCVs are A for arthropods
O for fishes and D for other invertebrates. Solid symbols are benthic genenr
open symbols water column genera. Arrows indicate greater than values Error
bars around sediment LC50 values indicate observed range of LC50s
5-9
-------
10'
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ill
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10
Water-only tests; (96HR LC50 •*• ACR)
A Arthropod
D Other Invertebrate
O Fishes
AC R- 5.75
upper:44fig/goc
• SQC:
lower: Q.
20
40
60
80
100
PERCENTAGE RANK OF SALTWATER GENERA
Figure 5-2. Comparison between SQC concentrations and 95% confidence intervals, effect
concentrations from benthic organisms exposed to dieldrin-spiked sediments and
sediment concentrations predicted to be chronically safe in salt water sediments.
Concentrations predicted to be chronically safe (Predicted Genus Mean Chronic
Values, PGMCV) are derived from the Genus Mean Acute Values (GMAV) from
water-only 96-hour lethality tests, Acute Chronic Ratios (ACR) and K— values
PGMCV = (GMAV •*• ACR)Koc. Symbols for PGMCVs are A for arthropods,
O for fishes and D for other invertebrates. Solid symbols are benthic genera;
open symbols water column genera. Arrows indicate greater than values. Error
bars around sediment LC50 values indicate observed range of LC50s.
5-10
-------
In Figures 5-1 and 5-2 each PGMCV for fishes, arthropods or other invertebrates tested
in water is plotted against the percentage rank of its sensitivity. Results from toxicity tests with
benthic organisms exposed to sediments spiked with dieldrin (Table 4-1) are placed in the
PGMCV rank appropriate to the test-specific effect concentration. (For example, the 10-day
LC50 for H. izteca, (1,640 ^g/goc) is placed between the PGMCV of 742 fig/g^ for the
stonefly, Acroneuria, and the PGMCV of 6,605 pg/g^ for the cladoceran, SJr
Therefore, LC50 or other effect concentrations are intermingled in this figure with
concentrations predicted to be chronically safe. Care should be taken by the reader in
interpreting these data with dissimilar endpoints. The following discussion of SQC, organism
sensitivities and PGMCVs is not intended to provide accurate predictions of the responses of taxa
or communities of benthic organisms relative to specific concentrations of dieldrin in sediments
in the field. It is, however/intended to guide scientists and managers through the complexity
of available data relative to potential risks to benthic taxa posed by sediments contaminated with
dieldrin.
The freshwater SQC for dieldrin (11 ^g/g^ is less than any of the PGMCVs or LC50
values from spiked sediment toxicity tests. The PGMCVs for 17 of 19 freshwater genera are
greater than the upper 95% confidence interval of the SQC (23 pg/g^. The PGMCVs for the
stonefly Bejonarcella (15 ^g/goc) and Claassenia (18 Mg/goc) are below the SQC 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 impacted by sediment concentrations marginally above the SQC and possibly
5-11
-------
less than the 95% upper confidence interval. For dieldrin, PGMGVs range over three orders
» • . - • • •
of magnitude 'from the most sensitive to the most tolerant genus. A sediment concentration 20
times the SQC would include the GMCVs of one-half of the 12 benthic genera tested including
stoneflies, mayflies, isopods and catfish. Tolerant benthiic genera such as the amphipod
•
-------
assessment of the concentrations of dieldrin in the sediments of the nation's water bodies. Log
probability plots of dieldrin concentrations on a dry weight basis in sediments are shown in
Figure 5-3. Dieldrin is found at varying concentrations in sediments from rivers, lakes and near
coastal water bodies in theTJnited States. This is due to its widespread use and quantity applied
during the 1960s and early 1970s. It was restricted from register and production in the
United States in 1974. Median concentrations are generally at or near detection limits in most
water bodies for data after 1986. There is significant variability with dieldrin concentrations in
sediments ranging over nine orders of magnitude within the country.
lie SQC for dieldrin can be compared to existing concentrations of dieMrta in sediments
of natural water systems in the United States as contained in the STORET database (U.S. EPA,
1989b). These data are generaUy reported on a dry weight basis, rather than an organic carbon
normalized basis. Therefore, SQC values corresponding to sediment organic carbon levels of
1 to 10SS are compared to dieldrin's distribution in sediments as examples only. For fresh
water sediments, SQC values are 0.11 Mg/g dry weigh, in sediments having 1 * organic carbon
and 1.1 «fe dry weight in sediments having 10% organic carbon; for marine sediments SQC
are 0.20 nfe dry weigh, and 2.0 Mg/g, dry weight respectively. Figure 5-3 presents the
comparisons of these SQC to probability distributions of observed sediment dieldrin levels for
steams and lakes (fresh water systems, shown on the upper panels) and estuaries (marine
systems, lowerpanel). For both streams (n = 3075) and lakes (n ^ 457), both the SQC of 0.11
n/g dry weight for 1% organic carbon fresh water sediments and the SQC of 1.1 ,,g/g dry
weight for 10% organic carbon fresh water sediments are exceeded by less than 1 % of the data.
m estuaries, the dab (n=160) indicate tha, neither criteria, 0.20 ug/g dry weight for sediments
5-13
-------
Uj5»
Wfc
•"TIT] 1 1—i—i—i—r-
TOTAL SAMPLES: 3075
MEASURED SAMPLES: 590
J_
80 90
Jin 11.
99
99.9
10
10
'TOTAL SAMPLES:' <57' '
MEASURED SAMPLES: 124
p""" ' """"
1 '••[•« • • • •
80 90
99
99.9
10'
10'
| ESTUARY
"""1 1 1 — i — | — i — r — | — inn
TOTAL SAMPLES: 160
MEASURED SAMPLES: 3
' • ' in
10 20
50
EIO 90
99 . 99.9
Figure 5-3.
PROBABILITY
on each figure represents the SQC vlfwhe^roc - iST 27 ? ?"
line represents the SQC when TOC ™% ~ ' ^lower da*ed
5-14
-------
having 1 % organic carbon or 2.0 Mg/g dry weight for sediments having 10% organic carbon are
exceeded by the post '1986 samples. Concentrations of dieldrin in sediments from estuaries are
two order of magnitude below the SQC value for 1 % organic carbon sediments and three orders
. of magnitude below the SQC value for sediments with TOCs of 10 %.
Hie dieldrin distribution in Figure 5-3 includes data from some samples in which the
dieldrin concentration was below the detection limit. These data are indicated on the plot as
"less than" symbols «), and plotted at the reported detection limits. Because these values
represent upper bounds and not measured values the percentage of samples in which the SQC
values are actually exceeded may be less than the percentage reported.
A second database developed as part of the National Status and Trends Program (NOAA,
1991) is also available for assessing contairdnant levels in marine sediments that are
representative of areas away from sources of contamination. The probability distribution for
these data, which can be directly expressed on an organic carbon basis, is compared to the
saltwater SQC for dieldrin (20 ^g/g^ on Figure 5-4. Data presented are from sediments with
0.20 to 31.9 percent organic carbon. The median organic carbon normalized dieldrin
concentration (0.08 Atg/goc) is 2 orders of magnitude below the SQC of 20 Mgoc. Noneofthese
samples (n=408) exceeded the criteria. Hence, these results are consistent with the preceding
comparison of the marine SQC to STORE! data.
Regional differences in dieldrin concentrations may affect the above conclusions
concerning expected criteria 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
5-15
-------
i iimni i mini i nunii
0
i o
uiiii 111 uiiii 11
mm i nun 111 •mm , pi,,,,,
(OO 0/On) 1N3WIQ38
-8 §
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5-16
-------
areas. It is presented as an aid in assessing the range of reported dieldrin sediment
concentrations and the extent to which they may exceed the SQC.
5.5 UMTTATIONS TO THE APPLICABILITY OF SEDIMENT QUALITY CRITERIA:
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 SQC
does not mean that other chemicals, for which there are no SQC available, are not present in
concentrations sufficient to cause harmful effects. Furthermore, even if SQC were available for
all of the contaminants in a particular sediment, there might be additive or synergistic effects
that the criteria do not address. In this sense the SQC represent "best case" criteria.
It is theoretically possible that antagonistic reactions between chemicals could reduce the
toxiciry of a given chemical such that it might not cause unacceptable effects on benthic
organisms at concentrations above the SQC when it occurs with the antagonistic chemical.
However, antagonism has rarely been demonstrated. What should be much more common are
instances where toxic effects occur at concentrations below the SQC because of the additivity
of toxicity of many common contaminants (Alabaster and Lloyd, 1982), e.g. heavy metals and
PAHs, and instances where other toxic compounds for which no SQC exist occur along with
SQC chemicals.
Care must be used in application of EqP-based SQC in disequilibrium conditions. In
some instances site-specific SQC may be required to address this condition. EqP-based SQC
assume that nonionic organic chemicals are in equilibrium with the sediment and IW and are
associated with sediment primarily through adsorption into sediment organic carbon. In order
5-17
-------
for these assumptions to be valid, the chemical must be dissolved in IW and partitioned into
sediment organic carbon. The chemical must, therefore, be associated with the sediment for a
sufficient length of time for equilibrium to be reached. In sediments where particles of
undissolved dieldrin occur, disequilibrium exists and criteria are over protective. In liquid
chemical spill situations disequilibrium concentrations in interstitial and overlying water may be
proportionately higher relative to sediment concentrations. In this case criteria may be
underprotective.
In very dynamic areas, with highly erosional or depositional bedded sediments,
equilibrium may not be attained with contaminants. However, even high KOW nonionic organic
compounds come to equilibrium in clean sediment in a period of days, weeks or months.
Equilibrium times are shorter for mixtures of two sediments, each previously at equilibrium.
This is particularly relevant in tidal situations where large volumes of sediments are eroded and
deposited, yet near equilibrium conditions may predominate over large areas. Except for spills
and particulate chemical, near equilibrium is the rule and disequilibrium is uncommon. In
instances where it is suspected that EqP does not apply for a particular sediment because of
disequilibrium discussed above, site-specific methodologies may be applied (U.S. EPA, 1993b).
5-18
-------
SECTION 6
CRITERIA STATEMENT
The procedures described in the "Technical Basis for Deriving National Sediment Quality
Criteria for Nonionic Organic Contaminants for the Protection of Benthic Organisms by Using
Equilibrium Partitioning" (U.S. EPA, 1993a) indicate that benthic organisms should be
acceptably protected in freshwater sediments containing .<. 11 /tg dieldrin/g organic carbon and
saltwater sediments containing <. 20 /tg dieldrin/g organic carbon, except possibly where a
locally important species is very sensitive or sediment organic carbon is < 0.2%.
•„
Confidence limits of 5.2 to 24 jtg/gbc for freshwater sediments and 9.5 to 44 jtg/goc for
saltwater sediments are provided as an estimate of the uncertainty associated with the degree to
which the observed concentration in sediment fag/goc), which may be toxic, can be predicted
using the organic carbon partition coefficient (K^ 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,
the partitioning and toxicity of dieldrin, and sound judgement are required in the regulatory use
of SQC and their confidence limits.
These concentrations represent the U.S. EPA's best judgement at this time of the levels
of dieldrin in sediments that would be protective of benthic species. It is the philosophy of the
Agency and the EPA Science Advisory Board that the use of sediment quality criteria (SQCs)
as stand-alone, pass-fail criteria is not recommended for all applications and should frequently
6-1
-------
trigger additional studies at sites under investigation. The upper confidence limit should be
interpreted as a concentration above which impacts on benthic species should be expected.
Conversely, the lower confidence limit should be interpreted as a concentration below which
impacts on benthic species should be unlikely.
6-2
-------
SECTION 7
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A Review of Fish and Wildlife Service Investigations During 1961 and 1962. U.S Fish
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7-2
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»•''':"' :-; -
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Hoke, R. and G. Ankley, 1991. Results of dieldrin sediment spiking study conducted in
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Hoke, R. and G. Ankley, 1992. Results of re-test of Airport Pond dieldrin-spiked
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7-3
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Hoqftman, R.N. and G. J. Vink, 1980. The determination of toxic effects of pollutants with the
marine polycheate worm Oohrvotrocha diadema. Ecotoxicol. Environ. Safety 4:252-262.
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7-4
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Nebeker, A.V., G.S. Schuytema, W.L. Griffis, J.A. Barbitta, and L.A. Carey. 1989. Effect
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7-5
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June 26, 1991. 160 pp.
',.>'•"'
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Sediment Quality Criteria for the Protection of Benthic Organisms. (In Review).
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lower lakes at Rocky Mountain Arsenal. Ph.D. Thesis, Colorado State
7-7
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