United Statas Office of Science and Technology
Environmental Protection Agency Health and Ecologies/ Criteria OiV.
Office of Water & Washington, O.C. 20460
Office of Research and
Development
EPA-822-R-93-013
September 1993
Sediment Quality Criteria
for the Protection of
Benthic Organisms:
ACENAPHTHENE
Recycled/Recyclable
Printed with Soy/Canoia Ink on paper that
contains at least 50% recycled fiber
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CONTENTS
PAGE
Foreword '. ii
Acknowledgements iv
Tables vi
Figures vii
Introduction 1-1
Partitioning 2-1
Toxicity of Acenaphthene: Water Exposures '. . . . 3-1
Toxicity of Acenaphthene (Actual and Predicted): Sediment Exposures . 4-1
Criteria Derivation for Acenaphthene 5-1
Criteria Statement 6-1
References 7-1
Appendix A: Summary of Acute Values for Acenaphthene for
Freshwater and Saltwater species A-l
Appendix B: Summary of Data from Sediment Spiking Experiments with
Acenaphthene B-l
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FOREWORD
Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U.S.
EPA) 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 kmd 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 Quality 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
estuaries, 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 contaminant levels 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 and
regulations on the basis of the specific facts.
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ACKNOWLEDGEMENTS
Principal Author
David J. Hansen
Coauthors
Walter J. Berry
U.S. EPA, Environmental Research Laboratory,
Narragansett, RI
Science Applications International Corporation,
Narragansett, RI
Dominic M. Di Toro Manhattan College, Bronx, NY;
HydroQual, Inc., Mahwah, NJ
Paul R. Paquin
Laurie De Rosa
HydroQual, Inc.,
Mahwah, NJ
HydroQual, Inc.,
Mahwah, NJ
Frank E. Stancil, Jr. U.S. Environmental Research Laboratory, Athens, GA
Christopher S. Zarba U.S. EPA Headquarters, Office of Water, Washington, DC
Technical and Clerical Support
Heinz P. Kollig U.S. Environmental Research Laboratory, Athens, GA
Glen B. Thursby
Maria R. Paruta
Dinalyn Spears
BettyAnne Rogers
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 nave 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. Anldey
University of Delaware, Newark, DE
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
Christina E. Cowan Battelle, Richland, WA
Dominic M. Di Toro HydroQual, Inc., Mahwah, NJ;
Manhattan College, Bronx, NY
David J. Hansen
Paul R. Paquin
Spyros P. Pavlou
Richard C. Swartz
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
Nelson A. Thomas
Christopher S. Zarba 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.
TABLES
*
Acenaphthene measured and estimated log10KoW values.
Summary of log10KoW values for acenaphthene measured by the U.S. EPA,
Environmental Research Laboratory, Athens, GA.
Table 4-1.
Table 4-2.
Table 5-1.
Table 5-2.
Table 5-3.
Appendix A.
Appendix B.
Summary of K^. values for acenapthene derived from literature sorption isotherm
data.
Chronic sensitivity of freshwater and saltwater organisms to acenaphthene. Test
specific data.
Summary of Freshwater and Saltwater acute and chronic values, acute-chronic
ratios, and derivation of final acute values, final acute-chronic ratios, and final
chronic values for acenaphthene.
Results of approximate randomization test for the equality of freshwater and
saltwater FAV distributions for acenaphthene and approximate randomization test
for the equality of benthic and combined benthic and water column (WQC) FAV
distributions.
Summary of tests with acenaphthene-spiked sediment.
Water-only and sediment LCSOs used to test the applicability of the equilibrium
partitioning theory for acenaphthene.
Sediment quality criteria for acenaphthene.
Analysis of variance for derivation of sediment quality criteria confidence limits
for acenaphthene.
Sediment quality criteria confidence limits for acenaphthene.
- Summary of acute values for acenaphthene for freshwater and saltwater species.
- Summary of data from sediment spiking experiments with acenaphthene. Data
from these experiments were used to calculate KQC values (Figure 2-1) and to
compare mortalities of amphipods with pore water toxic units (Figure 4-1) and
predicted sediment toxic units (Figure 4-2).
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FIGURES
Figure 1-1. Chemical structure and physical-chemical properties of acenaphthene. v
Figure 2-1. Organic carbon-normalized sorption isotherm for acenaphthene (top) and
probability plot of Koc (bottom) from sediment toxicity tests conducted by Swartz
(1991). The line in the top panel represents the relationship predicted with a log
of 3.76, that is Cij08=KOB • 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, N = nymph, 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. E = embryo, 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., 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 hi this SQC document, Appendix B hi the endrin, dieldrin
fluoranthene and phenanthrene SQC documents, and original references for raw
data.)
Figure 4-2. Percent mortality of amphipods hi 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; De
Witt 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 hi sediments 0*g/goc)
divided by the predicted LC50 Otg/g0c) to sediments (KoC x Water-only LC50,
/Kg/L). (See Appendix B hi this document and Appendix B hi the dieldrin,
endrin, fluoranthene, and phenanthrene SQC documents for raw data).
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Figure 5-1. Comparison between SQC concentrations and 95% confidence intervals, effect
concentrations from benthic organisms exposed to acenaphthene-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 water-only 96-hour lethality tests, Acute Chronic Ratios (ACR)
and KQC values. PGMCV = (GMAV -s- 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
LCSOs.
Figure 5-2. Comparison between SQC concentrations and 95% confidence intervals, effect
concentrations from benthic organisms exposed to acenaphthene-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 KQC 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
LCSOs.
Figure 5-3. Probability distribution of concentrations of acenaphthene in sediments from
streams, lakes and estuaries in the United States from 1986 to 1990 from the
STORET (U.S. EPA, 1989b) database, compared to the acenaphthene SQC values
of 13 pg/g in freshwater sediments having TOC = 10% and 1.3 pg/g in
freshwater sediments having TOC = 1% and compared to SQC values for
saltwater sediments of 23 pg/g when TOC =10% and 2.3 pg/g when TOC=1 %.
The upper dashed line on each figure represents the SQC value when TOC =
10%, the lower dashed line represents the SQC when TOC = 1 %.
Figure 5-4. Probability distribution of concentrations of acenaphthene in sediments from
coastal and estuarine sites from 1984 to 1989 as measured by the National Status
and Trends Program (NOAA, 1991). The horizontal line is the saltwater SQC
value of 230
vui
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DISCLAIMER
This report has been reviewed by the Health and Ecological Criteria Division, Office of
Science and Technology, U.S. Environmental Protection Agency, and approved for publication.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
AVATT.ABTTJrTY
This document is available to the public through the National Technical Information
Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161. NITS 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
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contaminated sediments and to identify, prioritize and implement appropriate clean up activities
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 concentration 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 value
(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 stable at equilibrium; (2) the concentration in
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either phase can be predicted using appropriate partition coefficients and the measured
concentration in the other phase; (3) organisms receive equivalent exposure from water-only
exposures or from any equilibrated phase: either from pore water via respiration, sediment
integument exchange, sediment via ingestion or from a mixture of exposure routes; (4) for
nonionic chemicals, effect concentrations in sediments on an organic carbon basis can be
predicted using the organic carbon partition coefficient (Koc) and effects concentrations in water;
and (5) the FCV concentration is an appropriate effects concentration for freely-dissolved
chemical in interstitial water; and (6) the SQC (/tg/goc) derived as the product of the K^ and
FCV is protective of benthic organisms. SQC concentrations presented in this document are
expressed as jtg chemical/g sediment organic carbon and not on an interstitial water basis
because: (a) pore water is difficult to adequately sample; and (b) significant amounts of the
dissolved chemical may be 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 acenaphthene 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
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5. protective of benthic organisms.
As is the case with WQC, the SQC reflect the use of available scientific data to: 1) assess the
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 periodically
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 or site specific considerations.
SQC values may also need to be adjusted because of site specific considerations. In spill
situations, where chemical equilibrium between water and sediments has not yet been reached,
a sediment chemical concentration less than the 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 total organic carbon (TOC) in the sediment does not greatly affect chemical
binding (DeWitt et al., 1992). However, the physical form of the chemical in the sediment may
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have an effect. At some sites concentrations in excess of the SQC may not pose risks to benthic
organisms, because the compound may be a component of a particulate, such as coal or soot,
exceed solubility such as undissolved oil. 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 the Derivation of Site-Specific Sediment Quality Criteria" (U.S.
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 acenaphthene. 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 By Using
Equilibrium Partitioning for the Protection of Benthic Organisms" (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 Chemicals" (U.S. EPA, 1993c).
1.2 GENERAL INFORMATION: ACENAPHTHENE:
Acenaphthene is a member of the polycyclic aromatic hydrocarbon (PAH) group of
organic compounds. It occurs both naturally in coal tar, and as a by product of manufacturing
processes such as petroleum refining, shale oil processing and coal tar distilling (Verschueren,
1983). Other man made sources of acenaphthene include its generation as a by product of the
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combustion of tobacco, and its presence in asphalt and in soots generated by the combustion of
aromatic fuels amt;ii=d with pyridine (Verschueren, 1983). Acenaphthene is used in
manufacturing processes to produce dyes, plastics, insecticides and fungicides (Windholtz et al.,
1983). Some PAHs are of environmental concern because they are known to be carcinogens
and/or mutagens (Brookes, 1977). With an increase in fossil fuel consumption in the United
States an increase in emissions of PAHS to the environment can be expected over the next
several decades (Eadie et al., 1982).
Acenaphthene has a two ring bridged structure (Figure 1-1). It has a solubility in water
at 25°C of 3.94 mg/1 (Miller et al., 1985), and is a solid at room temperature (melting point of
116°C). Two significant processes which can influence the fate of acenaphthene in sediment
are sorption and biodegradation (U.S. EPA, 1980). Sorption of acenaphthene onto solids in the
water column and subsequent settling, as well as partitioning onto organics in the sediment, can
significantly affect acenaphthene transport. Bioaccumulation is a short-term process in which
PAHs with 4 rings or less are metabolized and long-term partitioning into biota is not considered
a significant fate process (U.S. EPA, 1980). Other processes found to have little or no effect
on the fate of acenaphthene in the sediment are oxidation, hydrolysis and volatilization (U.S.
EPA, 1980).
The acute toxicity of acenaphthene from individual toxicity tests ranges from 120.0 to
2,045 /tg/L for freshwater and 160 to 16,440 jcg/L for saltwater organisms (Appendix A).
Differences between concentrations of acenaphthene causing acute lethality and chronic toxicity
are small; acute-chronic ratios range from 1.5 to 6.7 (Table 3-3). Although acenaphthene
bioaccumulates in aquatic biota, the associated health or ecological risks are unknown.
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MOLECULAR FORMULA
MOLECULAR WEIGHT
DENSITY
MELTING POINT
PHYSICAL FORM
VAPOR PRESSURE
154.21
1.069 g/cc @ 20°C
90-95°C
Orthorhpmbic
bipyramidal needles
0.0026 mPa (25°C)
CAS NUMBER:
CHEMICAL NAME:
83-32-9
1,2-Dihydroacenaphthylene;
periethylenenaphthalene;
1,8-ethylenenaphthalene
Figure 1-1. Chemical structure and physical-chemical properties of acenaphthene.
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1.3 OVERVIEW OF DOCUMENT:
Section 1 provides a brief review of EqP methodology, and a summary of the physical-
chemical properties and aquatic toxicity of acenaphthene. Section 2 reviews a variety of methods
and data useful in deriving partition coefficients for acenaphthene and includes the KQC
recommended for use in the derivation of the acenaphthene SQC. Section 3 reviews aquatic
toxicity data contained in the acenaphthene WQC document (U.S. EPA, 1980) and new data that
were used to derive the Final Chronic Value (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 acenaphthene in the derivation of the
SQC. Section 4 reviews data on the toxicity of acenaphthene in sediments, the need for organic
carbon normalization of acenaphthene 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 for the derivation of the SQC for
acenaphthene and its uncertainty. The SQC for acenaphthene is then compared to STORET
(U.S. EPA, 1989b) and National Status and Trends (NOAA, 1991) data on acenaphthene's
environmental occurrence in sediments. Section 6 concludes with the criteria statement for
acenaphthene. 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 non-ionic organic chemicals, SQC are expressed as
Hg chemical/g organic carbon and apply to sediments having ^ 0.2% organic carbon by dry
weight. A brief overview follows of the concepts which underlie the equilibrium partitioning
methodology for deriving SQC. The methodology is discussed in detail in the "Technical Basis
for Deriving Numerical National Sediment Quality Criteria for Nonionic Organic Contaminants
by Using Equilibrium Partitioning for the Protection of Benthic Organisms" (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
correlated with the interstitial water (i.e., pore water) concentration 0*g chemical/liter pore
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water) and not to the sediment chemical concentration (jj.g chemical/g sediment) (Di Toro et al.,
1991). From a purely practical point of view, this correlation suggests that if it were possible
to measure the 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 species for a wide range of
chemicals. (The data showing this for acenaphthene 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 (fig/L) be the acceptable concentration in water
for the chemical of interest; then compute the SQC using the partition coefficient, (Kp,
L/Kg»cdiment)> between sediment and water:
SQC = KpFCV (2-1)
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.
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Organic carbon appears to be the dominant sorption phase for nonionic organic chemicals
in naturally occurring sediments and thus controls the bioavailability of these compounds in
sediments. Evidence for this can be found in numerous toxicity tests, bioaccumulation studies
and chemical analyses of pore water and sediments (Di Toro et al., 1991). The evidence for
acenaphthene 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 and the
weight fraction of organic carbon in the sediment (foe)- The relationship is as follows:
(2-2)
It follows that:
SQCoc = KocFCV (2-3)
where SQCOC is the sediment quality criterion on a sediment organic carbon basis.
KQC is not usually measured directly (although it can be done, see section 2.3).
Fortunately, KQC is closely related to the octanol-water partition coefficient (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 KQW for acenaphthene.
2.2 DETERMINATION OF KQW FOR ACENAPHTHENE:
Several approaches have been used to determine KQW for 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
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of KOW'S. The measurement methods were shake-centrifugation (SC) and generator column
(GCol), and slow stir-flask (SSF) and the estimation methods were SPARC (SPARC Performs
Automated Reasoning in Chemistry; Karickhoff et al., 1989) and CLOGP (Chou and Jurs,
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.
TABLE 2-1. ACENAPHTHENE MEASURED AND ESTIMATED LOGioKoW VALUES
METHOD
Measured
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
LOG10KoW
3.92
3.70
3.88
3.92
3.98
4.03
4.07
4.15
4.22
4.33
4.43
REFERENCE
Banerjee etal., 1980
Yalkowsky etal., 1983
SPARC4
Miller etal., 1985
Mabey etal., 1982
Yalkowsky et al., 1979
CLOGP"
Mackay etal., 1980
Kamletetal., 1988
Callahan et al., 1979
Arbuckle 1983
•SPARC is from SPARC Performs Automated Reasoning in Chemistry, (Karickhoff et
al., 1989).
bCLOGP is an algorithm that is included in the database QSAR located at the U.S. EPA,
Environmental Research Lab., Duluth, MN (Chou and Jurs, 1979).
In an examination of the literature data for acenaphthene, only one primary reference was
found, with a measured log10KoW value of 3.92 (Banerjee et al., 1980) (Table 2-1). Several
primary references were found for estimated log10KoW values in the literature ranging from 3.70
to 4.43 (Table 2-1). Although the range of reported values for acenaphthene is significantly
2-4
-------
lower than the range of values for some other compounds, it is relatively large, and we were not
* ' *•
able to determine from studying the primary articles that any value was more likely to be
accurate than any other.
KQW values for SPARC and CLOGP are also included in Table 2-1. SPARC is a
computer expert system under development at ERL,A, and the University of Georgia, at Athens.
The CLOGP algorithm is included in the database QSAR located at EPA's Environmental
Research Laboratory (ERL,D) at Duluth, Minnesota. For more information on SPARC and
CLOGP see U.S. EPA (1993a). The SPARC estimated log10KoW value for acenaphthene is
3.88. The CLOGP program estimate of the logioKoW value for acenaphthene using structure
activity relationships is 4.07.
We had littie confidence in the available measured or estimated values for KQW, therefore
the SC, GCol, SSF methods were used to provide additional data from which to define KQW
acenaphthene (Table 2-2). The SC method yielded a log10KoW = 3.84 (n=4), the GCol method
yielded a log10KoW = 4.17 (n=4), and the SSF method yielded a log10KoW = 3.83 (n=3).
Comparison of the results from the SC, GCol, SSF and SPARC KQW determination methods for
the five chemicals for which SQC are currently being developed (acenaphthene, dieldrin, endrin,
fluoranthene and phenanthrene) indicate that the SSF method provides the best estimate of KQW
(U.S. EPA, 1993a). The SSF method had less variability, less experimental bias (Bias is defined
as the mean difference between the best-fit estimate of KQW using all four methods and the
estimates from each method.) and was generally in the range of the SC, GCol, and SPARC
methods (U.S. EPA, 1993a). Therefore, the SSF value of 3.83 is the value for log10KoW
recommended for SQC derivation. This value agrees with the SPARC estimated value and the
2-5
-------
average of the values measured by the three methods under carefully controlled conditions at
*
ERL, A. This KQW is the logarithm of the mean of three KQW measurements made by SSF. The
logs of the KQW values measured by SSF range from 3.81 to 3.84.
TABLE 2-2. SUMMARY OF LOG10KoW VALUES FOR ACENAPHTHENE MEASURED
BY THE U.S. EPA, ENVIRONMENTAL RESEARCH LABORATORY,
ATHENS , GA.
SHAKE-
CENTRIFUGATION
3.82
3.84
3.88
3.84
3.84
GENERATOR
COLUMN
4.18
4.17
4.16
4.17
4.17
SLOW STIR
FLASK
3.81
3.84
3.84
3.83
•Logic of mean measured values.
2.3 DERIVATION OF KOC FROM ADSORPTION STUDIES:
Two types of experimental measurements of the KOC are available. The first type
involves experiments which were designed to measure the partition coefficient in particle
suspensions. The second type of measurement is from sediment toxicity tests in which
measurements of sediment acenaphthene, sediment organic carbon (OC) and non-dissolved
organic carbon (DOC) associated acenaphthene in pore water were used to compute KOC.
2.3.1 KOC FROM PARTICLE SUSPENSION STUDIES:
Laboratory studies to characterize adsorption are generally conducted using particle
2-6
-------
suspensions. The high concentrations of solids and turbulent conditions necessary to keep the
•*
mixture in suspension make data interpretation difficult as a result 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).
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:
oe
(2-4)
1 + mf
oc
where m is the particle concentration in the suspension (kg/L), and % = 1.4, an empirical
constant.
In this expression the KQC is given by:
log10Koc = 0.00028 + 0.983 log10KoW (2-5)
A sorption isotherm experiment that demonstrates the effect of particle suspensions was
found in a comprehensive literature search for partitioning information for acenaphthene (Table
2-2) (Mihelcic and Luthy, 1988). The experiment with four different concentrations of particles
in suspension showed an observed Kp of 52 L/kg for an acenaphthene solution and soil (2.9 +.
0.28% organic carbon). Calculated Kp using KQC (Equation 2-5) and foc is 175 L/kg. The
difference between the observed and calculated Kp can be explained by particle interaction
effects. Particle interaction results in a lower observed partition coefficient. The particle
interaction model (Equation 2-4) predicts Kp of 36.8 L/kg to 6.7 L/kg for respective solids
2-7
-------
concentrations of 0.03 kg/L to 0.20 kg/L which is in order with the observed Kp. Log10KoC
computed from observed Kp and f^ is 3.25. This value is lower than K<,c from laboratory
measurements as a result of particle effects. This data is presented as an example of particle
effects only, as 100 percent reversibility is assumed in the absence of a desorption study and an
actual KQC can not be computed.
TABLE 2-3. SUMMARY OF KQC VALUES FOR ACENAPHTHENE DERIVED
FROM LITERATURE SORPTION ISOTHERM DATA.
Observed Solids
LogjoKoc n (kg/L) Reference
3.25 1 0.03-0.20 Milhelcic and
Luthy, 1988
In the absence of particle effects, KOC is related to KQW via Equation 2-5, shown above.
For log10Kow = 3.83 (ERL,A mean measured value), this expression results hi an estimate of
log10Koc = 3.76
2.3.2 KQC FROM SEDIMENT TOXKTTY TESTS:
Measurements of KQC are available from sediment toxicity tests using acenaphthene
(Swartz, 1991). These tests are from three marine saltwater sediments having a range of organic
carbon contents of 1.02 to 4.37 percent (Table 4-1; Appendix B). Acenaphthene concentrations
were measured in the sediments and pore waters providing the data necessary to calculate the
partition coefficient for an undisturbed bedded sediment. The pore water measurements did not
2-8
-------
increase at values greater than 1,000 /ttg/L suggesting that the limit of aqueous solubility was
being approached. These tests were run at 15°C, but a literature search for acenaphthene
revealed no solubility data at this temperature. Solubility for acenaphthene is reported as 3.94
mg/L at 25°C (Miller et al., 1985), supporting the idea of saturation limitation. As a result,
computations for the partition coefficient did not include treatments where pore water
concentrations were greater than 1,000/tg/L.
The upper panel of Figure 2-1 is a plot of the organic carbon-normalized sorption
isotherm for acenaphthene where the sediment acenaphthene concentration Og/goc) is plotted
versus pore water concentration 0*g/L). The data used to make this plot are included in
Appendix B. Data from treatments where pore water concentrations were greater than 1,000
Atg/L were not included on the plot. The line of unity slope corresponding to the log10KoC =
3.76 is compared to the data. A probability plot of the observed experimental log10KoC values
is shown in lower panel of Figure 2-1. The logi^oc values are approximately normally
distributed with a mean of log10KoC = 3.58 and a standard error of the mean of 0.012. This
value is statistically indistinguishable from logxoKoc = 3.76, which was computed from the
experimentally determined acenaphthene log10KoW of 3.83 Equation 2-5.
2.4 SUMMARY OF DERIVATION OF KQC FOR ACENAPHTHENE:
The Koc selected to calculate the sediment quality criteria for acenaphthene is based on
the regression of log10KoC from log10KoW (Equation 2-5), using the acenaphthene log10ECoW of
3.83 recently measured by ERL,A. This approach rather than the use of the KQC from the
toxicity test was adopted because the regression equation is based on the most robust data set
2-9
-------
ACENAPHTHENE
100000
10000
>£> 1000
100
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1 1 IIIIIII 1 1 1 1 1 (Ml
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10 20 60 80 80
PROBABILITY
Figure 2-1. Organic carbon-normalkdd sorption isotherm for acenaphthene (top) and
probability plot of Koc (bottom) from sediment toxicity tests conducted by Swartz
(1991). The line in the top panel represents the relationship predicted with a log
of 3.76, that is C^-Ko. • Cd.
2-10
-------
available that spans a broad range of chemicals and particle types, thus encompassing a wide
range of KQW and f^. The regression equation yields a log10KoC = 3.76. This value is in very
good agreement with the log10Koc of 3.58 measured in the sediment toxicity tests.
2-11
-------
-------
SECTION 3
TOXICITY OF ACENAPHTHENE: WATER EXPOSURES
3.1 TOXICITY OF ACENAPHTHENE IN WATER: DERIVATION OF ACENAPHTHENE
WATER QUALITY CRITERIA:
The equilibrium partitioning (EqP) method for derivation of sediment quality criteria (SQC)
uses the acenaphthene water quality criteria (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
live 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
acenaphthene WQC document (U.S. EPA, 1980) 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:
Twenty standard acute toxicity tests with acenaphthene have been conducted on 10
freshwater species from 10 genera (Appendix A). Overall genus mean acute values (GMAVs)
range from 120 to 2,045 jtg/L. Three invertebrates and two fishes were among the most
sensitive species; overall GMAVs for these taxa range from 120 to 670 /*g/L. Tests on the
benthic life-stages of 5 species from 5 genera are contained in this database (Figure 3-1;
Appendix A). Benthic organisms were among both the most sensitive, and most resistant,
freshwater genera to acenaphthene; GMAVs range from 240 and > 2,040 jtg/L. Three
3-1
-------
10000 r
A Arthropods
D Other Invertebrates
O Fishes
Lepomis (J)
Bluegill
Snail
a
LU
3
Ul
<
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Ictalurus (J) Mid9e
Catfish
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Minnow
Oncorhynchus (J)
Trout
Gammarus (X)
Amphipod
Pettopeita (X)
Stonefly
100
JL
Daphnia (X)
Cladoceran
—I
_L
J.
_L
20 40 60 80
PERCENTAGE RANK OF FRESHWATER GENERA
100
Figure 3-1. Genus mean acute values from water only acute toxicity tests using freshwater
species 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, N = nymph, X =
unspecified life stage.
3-2
-------
epibenthic species, stoneflies, a snail and channel catfish, were tested. Two infaunal species
were tested, including the amphipod, Gammarus minus (LC50 = 460 pg/lL), and the midge,
Paratanvtarsus sp. (LC50 = 2,045 /tg/L). The FCV derived from the overall GMAVs (Stephan
et al., 1985) for freshwater organisms is 80.01 (Table 3-2).
Twenty-one acute toxicity tests have been conducted on 10 saltwater species from 10 genera
(Appendix A). Overall GMAVs range from 245.0 to 8,163 jtg/L., similar to the range for
freshwater genera. Crustaceans were most sensitive; GMAVs range from 245.0 to 1,125 /tg/L.
Benthic life-stages from 6 species from 6 genera have been tested (Figure 3-2; Appendix A).
They are among both the most sensitive, and most resistant, saltwater genera to acenaphthene.
The most sensitive benthic species is the sand shrimp, Crangon septemspinosus. with a 96-hour
LC50 of 245.0 jig/L based on unmeasured concentrations. The mysid, Mysidopsis bahia. has
a similar sensitivity with an average, flow-through 96-hour LC50 of 317.7 jig/L based on
measured concentrations. Other benthic species for which there are data appear less sensitive;
GMAVs range from 589.4 to 7,693 /*g/L. The FAV derived from the overall GMAVs (Stephan
et al., 1985) for saltwater organisms is 140.8 jtg/L (Table 3-2).
3.3 CHRONIC TOXICITY - WATER EXPOSURES:
Chronic life-cycle toxicity tests have been conducted with the freshwater midge
(Paratanytarsus sp.) and the saltwater mysid (M. bahia) (Table 3-1) and early life stage tests
have been conducted with the fathead minnow (Pimephales promelas) and sheepshead minnow
(Cyprinodon variegatus) (Table 3-1). For each of these species, except for P. promelas. one or
more benthic life stages were exposed to acenaphthene. Other chronic toxicity tests have been
3-3
-------
10000 A Arthropods
D Other Invertebrates
O Fishes
Neanthes (X)
Annelid
Arbacia (E)
Sea urchin
Cyprinodon
Sheepshead minnow
Crepidula (L)
Slipper limpet
'Menidia (J,X)
Silverside
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Amphipod
Ampelisca (X)
Amphipod
Mysidopsis (J)
Mysid
Crangon (X)
Sand shrimp
100
20
40
60
80
100
PERCENTAGE RANK OF SALTWATER GENERA
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. E = embryo, J = juvenile.
3-4
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conducted with these two freshwater species (Lemke, 1984; Lemke et al.,1983; Leinke and
Anderson, 1984) but insufficient documentation is available to permit use of these results
(Thursby, 1991a).
Two acceptable life cycle toxicity tests have been conducted with Paratanytarsus sp.
(Northwestern Aquatic Sciences, 1981). In the first test there was a 59% reduction in growth
and an 85 % reduction in reproduction in 575 /tg/L relative to control animals (Table 3-2). Eggs
produced by animals in this first test failed to hatch at 575 /*g/L. There was no significant effect
on parents or egg hatchability in acenaphthene concentrations from 32 to 295 /tg/L. In the
second test with Paratanytarsus sp. there was a 21 % reduction in survival in 315 /tg/L relative
to control animals; egg hatchability was not affected at the highest concentration tested (676
jKg/L); although survival of hatched Paratanytarsus sp. larvae was reduced 64% in this
concentration.
A total of six early life-stage toxicity tests have been conducted with the P. promelas part
of a round-robin test series; two each from three laboratories (Table 3-2). The effect
concentrations across laboratories and tests ranged from 98 to 509 /*g/L, a factor of 5.2.
Growth (dry weight), survival or both growth and survival were reduced. Only one of these test
pairs had a suitable measured acute value, allowing calculation of an acute-chronic ratio (Cairns
and Nebeker, 1982). The concentration-response relationships were similar for these two tests.
Parental fish were unaffected in the first test at acenaphthene concentrations ranging from 67 to
332 jfg/L, while fish exposed to 495 /xg/L had a 54% reduction in growth relative to control
fish. In the second test, Cairns and Nebeker (1982) observed a 30% reduction in growth in
parental fish in 509 jtg/L while there was no effect on fish exposed to 197 to 345 /*g/L.
3-7
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Data from three saltwater chronic toxicity tests are available, two with M- bahia and one
«
with (£. variegatus. M. bahia reproduction was affected by acenaphthene in two tests from two
different laboratories. In the first test (Home et al., 1983) there was a 93% decrease in
reproduction in 340 /tg/L relative to control M- bahia: all M- bahia in 510 /tg/L died. No
effects were observed at 100 or 240 /tg/L in the parental generation, and juveniles released in
>. 340 /ig/L were not affected. In the second test (Thursby et al., 1989b) there was a 34%
decrease in growth in 168 /tg/L and 96% increase in mortality at 354 /tg/L. There was a 91 %
decrease in reproduction in M. bahia exposed to 91.8 /tg/L and M- bahia exposed to 168 and
354 /tg/L did not reproduce. M- bahia exposed to _<. 44.6 /tg/L were not affected.
£_.. variegatus exposed to acenaphthene in an early life stage test (Ward et al., 1981) were
affected at acenaphthene concentrations of _>.970 /tg/L (Table 3-2). There was a 70% reduction
in survival of fish hatched in 970 /tg/L. Fewer than 10% of the embryos at >_ 2,000 /tg/L
hatched and all fish that hatched died. There was no effect on either survival or growth in fish
exposed to 240 or 520 /tg/L.
The difference between acute and chronic toxicity of acenaphthene is small (Table 3-2).
Species mean acute-chronic ratios are 1.475 for P. promelas. 3.424 for M. bahia. 4.365 for C.
variegatus and 6.683 for Paratanytarsus sp. The final ACR, the geometric mean of these four
values, is 3.484.
The FCV (Table 3-2) are used as the effect concentrations for calculating the SQC for
protection of bentbic species. The FCV for freshwater organisms of 22.96 /tg/L is the quotient
of the FAV of 80.01 /tg/L and the final ACR of 3.484. Similarly, the FCV for saltwater
organisms of 40.41 /tg/L is the quotient of the FAV of 140.8 /tg/L and the final ACR of 3.484.
3-10
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3.4 APPLICABILITY OF THE WATER QUALITY CRITERION AS THE EFFECTS
CONCENTRATION FOR DERIVATION OF THE ACENAPHTHENE 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 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 acenaphthene-specific comparisons in sensitivity of benthic and all (benthic and water-
column) species conducted. The following paragraphs examine the data on the similarity of
sensitivity of benthic and all species for acenaphthene.
For acenaphthene, benthic species account for 4 out of 10 genera tested in freshwater,
and 6 out of 10 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 acenaphthene was performed using the Approximate Randomization method (Noreen, 1989).
3-11
-------
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, then 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 two 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). The 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 acenaphthene, the test-statistic falls at the 33 percentile of the generated FAV
differences. Since the probability is less 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
3-12
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600
800
ACENAPHTHENE
I I I I I—T
Bil
- FRESHWATER VS SALTWATER
99.9
LU
O
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600
600
400
300
200
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iCEOO O~
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0.1
10 20
60
80 90
99
99.9
PROBABILITY
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. '
3-13
-------
TABLE 3-3. RESULTS OF APPROXIMATE RANDOMIZATION TEST FOR
THE EQUALITY OF THE FRESHWATER AND SALTWATER-LC50
DISTRIBUTIONS FOR ACENAPHTHENE AND APPROXIMATE
RANDOMIZATION TEST FOR THE EQUALITY OF BENTHIC AND
COMBINED BENTHIC AND WATER COLUMN (WQC) LC50
DISTRIBUTIONS.
Compar-
ison Habitat or Water Type' AR Statistic" Probability"
Fresh Fresh (10) Salt (10) -59.04 33
vsSalt
Benthic Benthic (10) WQC (20) -41.34 31
vs Water
Column +
Benthic (WQC)
•Values in parentheses are the number of LC50 values used in the comparison.
bAR statistic = FAV difference between original compared groups.
Tr(AR statistic theoretical ^ AR statistic observed) given that the samples
came from the same population.
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 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 tunes and the distribution that results i representative
3-14
-------
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 31 percentile and the hypothesis of no difference in sensitivity is
accepted (Table 3-3). This analysis suggests that the FCV for acenaphthene based on data from
all tested species is an appropriate effects concentration for benthic organisms.
3-15
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-------
SECTION 4
TOXICITY OF ACENAPHTHENE (ACTUAL AND PREDICTED):
SEDIMENT EXPOSURES
4.1 TOXICITY OF ACENAPHTHENE IN SEDIMENTS:
The toxicity of acenaphthene spiked into sediments has been tested with two saltwater
amphipod species. Freshwater benthic species have not been tested in acenaphthene-spiked
sediments. All concentrations of acenaphthene in sediments or interstitial water where effects
were observed in benthic species (Table 4-1) are greater than SQC or FCV concentrations
reported in this document. Details about exposure methodology are provided because, unlike
aquatic toxicity tests, sediment testing methodologies have not been standardized.
Generalizations across species or sediments are limited because of the limited number of
experiments. Therefore, insights into relative sensitivities of aquatic species to acenaphthene can
only be obtained from results of water-only tests (Section 3). Data are available from many
experiments using both field and laboratory sediments contaminated with mixtures of PAHs and
other compounds which include acenaphthene. Data from these studies have not been included
here because it is not possible to determine the contribution of acenaphthene to the observed
toxicity.
Swartz (1991) exposed the amphipods Eohaustorius estuarius and Leptocheirus
plumulosus to three acenaphthene-spiked sediments with total organic carbon content (TOC)
ranging from 0.82 to 4.21 %. Sediments were rolled (1) for four hours in acenaphthene-coated
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(experiments with L.. plumulosus). (3) rolled for an additional four hours, and (4) then stored
for 7 days at 4°C. The 10-day LCSO's for both species increased slightly with increasing
organic carbon concentration when the acenaphthene concentration was expressed on a dry
weight basis, but decreased for E. estuarius and no pattern was apparent for L. plumulosus when
concentration was expressed on an organic carbon basis. LCSO's normalized to dry weight
differed by less than a factor of 1.5 (44.4 to 68.4 /*g/g) for R estuarius and less than a factor
of 1.9 for L.. plumulosus over a 5.3-fold range of TOC. The organic carbon normalized LCSO's
for R estuarius differed by a factor of 2.7 (1,630 to 4,330 /tg/goc) wnile for L- plumulosus they
differed by a factor of > 3.0 (7730 to > 23,500 Aig/goc).
Overall, the need for organic carbon normalization of the concentration of nonionic
organic chemicals in sediments is presented in the SQC Technical Basis Document (U.S. EPA,
1993a). The need for organic carbon normalization for acenaphthene is somewhat supported by
»
the results of spiked-sediment toxicity tests described above. Although it is important to
demonstrate that organic carbon normalization is necessary if SQC are to be derived using the
EqP approach, it is fundamentally more important to demonstrate that KQC and water only effects
concentrations can be used to predict effects concentrations for acenaphthene and other nonionic
organic chemicals on an organic carbon basis for a range of sediments. Evidence supporting this
prediction for acenaphthene and other nonionic organic chemicals follows in Section 4.3.
4.2 CORRELATION BETWEEN ORGANISM RESPONSE AND PORE WATER
CONCENTRATION:
One corollary of the EqP theory is that pore-water LCSO's for a given organism should
4-3
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be constant across sediments of varying organic carbon content (U.S. EPA, 1993a). Appropriate
pore-water LC50 values are available for two bentMc species (Table 4-2). Swartz (1991) found
10-day LC50 values based on pore-water concentrations varied by a factor of 1.5 (542 to 800
pg/L) for IL estuarius and by a factor of > 1.2 (1,410 to > 1,720 /*g/L) for L. plumulosus.
This variability is somewhat less than that shown when either dry weight (factors of 1.5 and 1.9)
or organic carbon (factors of 2.7 and 3.0) normalization are used to determine LCSOs based on
acenaphthene concentration in sediments.
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), endrin (Nebeker et al., 1989; Schuytema et al., 1989), dieldrin
(Hoke, 1992), fluoranthene (Swartz et al., 1990, De Witt 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 utilize designs at least as rigorous as those conducted
at the EPA laboratories. Tests with acenaphthene and phenanthrene used two saltwater
amphipods (L. plumulosus and JL estuarius) and marine sediments. Tests with fluoranthene used
the saltwater amphipod (Rhepoxynius abronius) and marine sediments. Freshwater sediments
spiked with endrin were tested using the amphipod Hyalella azteca: while the midge,
Chironomus tentans. was tested using kepone-spiked sediments. Figure 4-1 presents the
percentage mortalities of the benthic species tested in individual treatments for each chemical
versus "pore water toxic units" (PWTUs) for all sediments tested. PWTUs are the concentration
of the chemical in pore water (/tg/L) divided by the water only LC50 (/tg/L). Theoretically,
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50% mortality should occur at one interstitial water toxic unit. At concentrations below one
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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. Data from the dieldrin experiments (Hoke and Ankley, 1991) are not
included because more knowledge of the pore water DOC will be required because dieldrin has
a high KQC value.
4.3 TESTS OF THE EQUILIBRIUM PARTITIONING PREDICTION OF SEDIMENT
TOXICITY:
SQC derived using the EqP approach utilize partition coefficients (Koc) and FCV from
WQC documents to derive the SQC concentration for protection of benthic organisms. The KQC
is used to normalize sediment concentrations and predict biologically available concentrations
across sediment types. The data required to test the organic carbon normalization for
acenaphthene in sediments are available for four benthic species. Data from tests with water
column species were not included in this analysis. Testing of this component of SQC derivation
requires three elements: (1) a water-only effect concentration, such as a 10-day LC50 value in
ftg/L, (2) an identical sediment effect concentration on an organic carbon basis, such as a 10-day
LC50 value in /wg/goc, and (3) a partition coefficient for the chemical, KQC in L/Kgoc. This
4-7
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section presents evidence that the observed effect concentration in sediments (2) can be predicted
utilizing the water effect concentration (1) and the partition coefficient (3).
Predicted 10-day LC50 values from acenaphthene-spiked sediment tests on a Mg/goc basis
with IL estuarius and L.. plumulosus were calculated (Table 4-2) using the log10KoC value of 3.78
from Section 2 of this document and the sediment LCSOs in Swartz (1991). Ratios of actual to
predicted LC50s for acenaphthene averaged 1.11 (range 0.76 to 2.01) for R estuarius and 3.25
(range 1.98 to >6.02) for ^ plumulosus. The overall mean for both species was 190.
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 KQW obtained using the slow-
stir flask method is not available. Swartz (1991) exposed the saltwater amphipods R estuarius
and LJ. plumulosus to acenaphthene in three marine sediments having organic carbon contents
ranging from 0.82 to 4.2% and to phenanthrene in three marine sediments having organic
carbon contents ranging from 0.82 to 3.6%.. Swartz et al. (1990) exposed the saltwater
amphipod JR.. abronius to fluoranthene in three marine sediments having 0.18, 0.31 and 0.48%
organic carbon. Hoke and Ankley (1991) exposed the amphipod H. azteca to three dieldrin-
spiked freshwater sediments having 1.7, 3.0 and 8.5% organic carbon and Hoke (1992) exposed
the midge £L. tentans to 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-
4-8
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spiked sediments having 3.0, 6.1 and 11.2% organic carbon. Figure 4-2 presents the percentage
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 0*g/goc) divided by the predicted LC50 0*g/goc)m sediments (the product of KQC
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 LCSOs 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.
4-9
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4-10
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SECTION 5
CRITERIA DERIVATION FOR ACENAPHTHENE
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-the calculate the 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 are mixing zones used 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
5-1
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acute tests have been conducted on the substance. The ACR is the mean ratio of acute to
chronic toxicity for three or more species exposed to the substance that meets minimum database
requirements. For more information on the calculation of ACRs, FAVs, and FCVs see the
National Water Quality Criteria Guidelines (Stephan et al., 1985). The FCV used in this
document differs from the FCV in the acenaphthene WQC document (U.S. FJ*A, 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 fyig/L) is the chronic concentration from the WQC for the chemical of
interest, then the SQC Otg/g sediment), is computed using the partition coefficient, KP (L/g
sediment), between sediment and pore water:
SQC = KP FCV (5-1)
Since organic carbon is the predominant sorption phase for conionic 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, (K^ can be substituted for KP. Therefore, on a sediment organic carbon basis,
SQCOC Og/goc), is:
SQCoc = KOC FCV (5-2)
Since (Koc) is presumably independent of sediment type for non-ionic organic chemicals, so also
is SQCoc. Table 5-1 contains the calculation of the acenaphthene SQC.
The organic carbon normalized SQC is applicable to sediments with an organic carbon
fraction of foe ^ 0.2%. For sediments with foe < 0.2%, organic carbon normalization and
5-2
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SQC may not apply.
TABLE 5-1. SEDIMENT QUALITY CRITERIA FOR ACENAPHTHENE.
Type of
Water Body
Fresh Water
Salt Water
(L/kg) W
3.83
3.83
LogioKoc
(L/kg)
3.76
3.76
FCV
23.0
40.4
Otg/goc)
130
230
aSQCoc = (103'76 L/kgoc)»(10-3 kgoc/goc)»(23.0 /*g acenaphthene/L) = 130 /tg
acenaphthene/goc
bSQCoc = (103-76 L/kgoc)«(10-3 kgoc/goc)«(40.4Mg acenaphthene/L) = 230 pg
acenaphthene/goc
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 fig 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:
jug Chemical/goc = A*g Chemical/gDRy ^ + (% TOC -s- 100)
= jug Chemical/gDRYWT • 100 •*• % TOC
For example, a freshwater sediment with a concentration of 6.00 jig chemical/gDRY m
5-3
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and 0.5 % TOG has an organic carbon-normalized concentration of 1,200 /*g/goc (6.00 /*g/gDRYWT
• 100 -^ 0.5 = 1,200 Atg/goc) which exceeds the SQC of 130 jig/goc- Another freshwater
sediment with the same concentration of acenaphthene (6.00 /tg/gDRY wr) but a TOC
concentration of 5.0% would have an organic carbon normalized concentration of 120 pcg/goc
(6.00 /tg/gDRY wr * 100 •* 5.0 = 120 /tg/goc), which is below the SQC for acenaphthene.
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 vr.). This "conversion" must be done for each level of TOC of interest:
SQCDRYWT = SQCocOtg/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 1.3 jtg/g:
SQCDRYWT. = 130 /tg/goc • 1% TOC H- 100 = 1.3 ftg/goRYwr
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 acenaphthene 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 (/tg/goc) can be estimated from the product of the effects concentration
from water-only exposures 0*g/L) and the partition coefficient KQC (L/kg). The uncertainty
5-4
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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 fulfill the minimum database requirements for the development of 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
acenaphthene are summarized in Appendix B. LCSOs for sediment and water-only tests were
computed from these data. The EqP model can be used to normalize the data in order to put
it on a common basis. The LCSOs from water-only exposures (LC50W; ftg/L) are related to the
organic carbon-normalized LCSOs from sediment exposures (LC50S(OC; /tg/goc) via the
partitioning equation:
LC50S,OC = KocLC50w (5-3)
The EqP model asserts that the toxicity of sediments expressed on an organic carbon basis equals
5-5
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the toxicity in water tests multiplied by the KQC. Therefore, both LC50SjOC and Koc*LC50w
are estimates of the true LC50oc £°r each chemical-organism pair. In this analysis, the
uncertainty of KQC is not treated separately. Any error associated with KQC will be reflected in
the uncertainty attributed to varying the exposure media.
In order to perform an analysis of variance, a model of the random variations is required.
As discussed above, experiments that seek to validate equation 5-3 are subject to various sources
of random variations. A number of chemicals and organisms have been tested. Each chemical-
organism pair was tested in water-only exposures and in different sediments. Let a represent
the random variation due to this source. Also, each experiment is replicated. Let G 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 e represents
the experimental error. Let (
-------
makie these estimates (U.S. EPA, 1993a). The results are shown in Table 5-2.
Table 5-2: ANALYSIS OF VARIANCE FOR DERIVATION OF
SEDIMENT QUALITY CRITERIA CONFIDENCE LIMITS
FOR ACENAPHTHENE.
Source of Uncertainty
Exposure media
Replication
Sediment Quality Criteria
Parameter
0*g/goc)
°«
«e
a
°SQC
Value
0.39
0.21
0.39
The last line of Table 5-2 is the uncertainty associated with the SQC; i.e., the variance
associated with the exposure media variability.
The confidence limits for the SQC are computed using this estimate of uncertainty for SQC.
For the 95% confidence interval limits, the significance level is 1.96 for normally distributed
errors. Hence:
= ln(SQCoc) + 1.96aSQC (5-5)
= ln(SQCoc) - 1.96_ 0.2 % For sediments with foe < 0.2 % , organic carbon normalization does not
5-7
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apply and the SQC do not apply.
TABLE "3. SEDIMENT QUALITY CRITERIA
CGKrIDENCE LIMITS FOR ACENAPHTHENE
Sediment Quality Criteria
95% Confidence Limits
Type of
Water Body MS/Soc Lower Upper
Freshwater 130 62 280
Saltwater 240 110 500
5.3 COMPARISON OF ACENAPHTHENE 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 determined from effect concentrations from
toxicity tests with benthic species exposed to sediments spiked with acenaphthene and sediment
concentrations predicted to be chronically safe to organisms tested in water-only exposures
(Figure 5-1; 5-2). Effect concentrations in sediments can be predicted from water-only toxicity
data and KOC values (See Section IV). Acute-chronic ratios (ACRs) are used to extrapolate from
Genus Mean Acute Values (GMAV) from water-only 96-hour lethality tests to chronically
acceptable concentrations. Therefore, it may be reasonable to predict for acenaphthene
chronically acceptable sediment concentrations (Predicted Genus Mean Chronic Value
(PGMCV)) from GMAVs (Appendix A), ACRs (Table 3-3) and the KQC (Table 5-1):
PGMCV = (GMAV -^ ACR)«KoC
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Water-only tests: (96HR LC50+ ACR) KQC
A Arthropods
D Other Invertebrates
O Rshes
Log10Koc»3-76
ACR »3.48
upper: 280 ng/goe
k3wer:62Mg/goc
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 acenaphthene-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 water-only 96-hour lethality tests, Acute Chronic Ratios (ACR)
and KQC 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.
5-9
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Water-only tests: (96HR LC50 •«• ACR) KQC
A Arthropods
D Other Invertebrates
O Rshes
ACR = 3.48
Sediment Tests: IQdLCSO
*£, estuarlus » 2384 fig/Qoc
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range 3 tests -7730 - 723500
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upper:
SQC:
lower:
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 acenaphthene-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 KOC 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
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Each predicted PGMCV for tested fishes » , arthropods A or other invertebrates • is
*
plotted against the percentage rank of its sensitivity. Results from toxicity tests with benthic
organisms exposed to sediments spiked with acenaphthene (Table 4-1) are placed in the
PGMCV rank appropriate to the test-specific effect concentration. (For example, the 10-day
LC50 for E. estuarius (2,384 /*g/goc) is placed between the PGMCV of 1,858 /Kg/goc for the
amphipod, Leptocheirus. and the PGMCV of 5,120 jtg/goc for the minnow, Cyrmnodon.)
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 acenaphthene 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 acenaphthene.
The freshwater SQC for acenaphthene (130 /tg/goc) is less than any of the PGMCVs.
PGMCVs for 9 of 10 freshwater genera are greater than the upper 95% confidence interval of
the SQC (280 Mg/goc)- The PGMCVs for the cladoceran Daphnia (198 /*g/goc) is below the
SQC upper 95% confidence interval. This suggests that if the extrapolation from water only
acute lethality tests to chronically acceptable sediment concentrations is accurate, this or
similarly sensitive genera may be chronically impacted by sediment concentrations marginally
above the SQC and possibly less than the 95 % upper confidence interval. For acenaphthene,
the PGMCVs range over an order of magnitude from the most sensitive to the most tolerant
5-11
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genus. Chronic effect concentrations may, however, occur at concentrations below saturation.
<•«
A sediment concentration four times the SQC would include the PGMCVs of two of the three
benthic genera tested including stoneflies, and amphipods. The tolerant benthic midge
Paratanytarsus might be expected to not be chronically impacted in sediments with acenaphthene
concentrations 25X the SQC.
The saltwater SQC for acenaphthene (230 /ig/goc) is less than any of the PGMCVs or
LC50 values from spiked sediment toxicity tests. The PGMCV for the sand shrimp Crangon
septemspinosus (404 /tg/goc) is lower than the upper 95 % confidence interval for the SQC. For
acenaphthene, PGMCVs from the most sensitive to the most tolerant saltwater genus range over
an order of magnitude. A sediment concentration five times the SQC would include the
PGMCVs of one-half of the six benthic genera tested including three arthropod genera. Other
genera of benthic arthropods, polychaetes, and fishes are less sensitive. Data from lethality tests
with two saltwater amphipods, Eohaustorius and Leptocheirus. exposed to acenaphthene spiked
into sediments substantiates this projection; the 10 day LCSOs from these tests range from 10
to 40 times the SQC of 230 /tg/goc-
5.4 COMPARISON OF ACENAPHTHENE SQC TO STORET AND STATUS AND
TRENDS DATA FOR SEDIMENT ACENAPHTHENE:
A STORET (U.S. EPA, 1989a) data retrieval was performed to obtain a preliminary
assessment of the concentrations of acenaphthene in the sediments of the nation's water bodies.
Log probability plots of acenaphthene concentrations on a dry weight basis in sediments since
1986 are shown in Figure 5-3. Acenaphthene is found at detectable concentrations in sediments
from rivers, lakes and near coastal water bodies in the United States. Median concentrations are
5-12
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Figure 5-3.
10 •
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Probability distribution of concentrations of acenaphthene in sediments from
streams, lakes and estuaries in the United States from 1986 to 1990 from the
STORET (U.S. EPA, 1989b) database, compared to the acenaphthene SQC values
of 13 pg/g in freshwater sediments having TOC = 10% and 1.3 pg/g in
freshwater sediments having TOC = 1% and compared to SQC values for
saltwater sediments of 23 pg/g when TOC =10% and 2.3 pg/g when TOC=1 %.
The upper dashed line on each figure represents the SQC value when TOC =
10%, the lower dashed line represents the SQC when TOC = 1%.
5-13
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generally at about 0.1 /ag/g in each of these three types of water bodies. Acenaphthene
concentrations in sediments range over seven orders of magnitude throughout the country.
The SQC for acenaphthene can be compared to existing concentrations of acenaphthene
in sediments of natural water systems in the United States as contained in the STORET database
(U.S. EPA, 1989b). These data are generally 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 10 percent are compared to acenaphthene's distribution in sediments as
examples only. For fresh water sediments, SQC values are 1.3 /*g/g in sediments having 1 %
organic carbon and 13 /ag/g dry weight in sediments having 10% organic carbon; for marine
sediments SQC are 2.3 /tg/g and 23 j*g/g, respectively. Figure 5-3 presents the comparisons
of these SQC to probability distributions of observed sediment acenaphthene levels for streams
and lakes (fresh water systems, shown on the upper panels) and estuaries (marine systems, lower
panel). For streams (n = 681) the SQC of 1.3 /*g/g for 1 % organic carbon fresh water
sediments is exceeded by less than 4 % of the data; the 13 /tg/g criteria for 10% organic carbon
freshwater sediments is exceeded in about 2 % of the samples. For lakes (n = 56), the SQC for
1 % organic carbon fresh water sediments is exceeded by about 2% of the samples, while the
SQC for 10 % organic carbon fresh water sediments is not exceeded by any of the lake samples.
In estuaries, the data (n = 74) indicate that neither of the criteria of 2.3 /ig/g dry weight for salt
water sediments having 1 % organic carbon or 23 jig/g dry weight for salt water sediments
having 10 % organic carbon are exceeded, although the STORET database for marine sediments
is not as extensive as the database for freshwater sediments.
The acenaphthene distribution in Figure 5-3 includes data from some samples in which
5-14
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the acenaphthene 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 contaminant 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 acenaphthene (230 /tg/goc) on Figure 5-4. Data presented are from sediments
with 0.20 to 31.9% organic carbon. None of these samples (n=288) exceeded the criteria.
Hence, these results are consistent with the preceding comparison of the marine SQC to
STORET data.
Regional differences in acenaphthene 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
areas. It is presented as an aid in assessing the range of reported acenaphthene sediment
concentrations and the extent to which they may exceed the sediment quality criteria.
5.5 LIMITATIONS 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
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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.
The concerns about mixtures of contaminants are particularly important with the PAHs,
which almost invariably occur as complex mixtures. Some guidance on interpretations of PAH
concentrations is possible given the presence of SQC for acenaphthene and other individual
PAHs. This is because much is known about the toxicity and structure-activity relationships of
the so-called narcosis chemicals, a group of nonionic organic chemicals to which the PAHs
belong. The toxicity of the narcosis chemicals is additive (Broderius and Kahl, 1985). The
toxicity of these chemicals increases with increasing KQW (Veith et al., 1983) and their
bioavailability in sediments decreases as a function of its KQW. Therefore, the toxicities of many
PAHs in sediments are likely to be similar. This explains why SQC values for fluoranthene
(fresh: 620 Atg/goc, salt: 300 /*g/goc)> acenaphthene (fresh: 130 ^tg/goc, salt: 230 Mg/goc^ and
phenanthrene (fresh: 180 /ig/goc, salt: 240 jtg/goc) differ little and why it is theoretically
possible to develop an SQC for total PAHs. EPA is currently conducting research aimed at
development of SQC for combined PAHs.
It is theoretically possible that antagonistic reactions between chemicals could reduce the
toxicity of a given chemical such that it might not cause unacceptable effects on benthic
organisms at concentrations above the SQC when it occurs with the antagonistic chemical.
However, antagonisms have rarely been demonstrated. What should be much more common are
instances where toxic effects occur at concentrations below the SQC because of the additivity
5-17
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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
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 like cinder,
soot, or oil droplets contain PAHs, 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 KQW 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 paniculate 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
5-18
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disequilibrium discussed above, site-specific methodologies may be applied (U.S. EPA, 1993b).
5-19
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-------
SECTION 6
CRITERIA STATEMENT
The procedures described in the "Technical Basis for Deriving Sediment Quality Criteria
for Nonionic Organic Contaminants by Using Equilibrium Partitioning" (U.S. EPA, 1993a)
indicate that benthic organisms should be acceptably protected in freshwater sediments containing
_<. 130 fig acenaphthene/g organic carbon and saltwater sediments containing _<. 230 fig
acenaphthene/g organic carbon, except possibly where a locally important species is very
sensitive or sediment organic carbon is < 0.2%.
Confidence limits of 62 to 280 /ttg/goC for freshwater sediments and 110 to 500 A*g/goc
for saltwater sediments are provided as an estimate of the uncertainty associated with the degree
to which the observed concentration in sediment 0*g/goc)> which may be toxic, can be predicted
using the organic carbon partition coefficient (KoC) 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 acenaphthene, 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 acenaphthene 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
6-1
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frequently trigger additional studies at sites under investigation. The upper confidence limit
'i
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
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SECTION 7
REFERENCES
Academy of Natural Sciences, 1981. Early life stage studies using the fathead minnow
(Pimephales promelas^ to assess the effects of isophorone and acenaphthene. Final report
to U.S. EPA, Cinn., OH. Academy of Natural Sciences, Philadelphia, PA. 26 pp.
Adams, W.J., R.A. Kimerle and R.C. Mosher. 1985. Aquatic safety assessment of chemicals
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Materials, Philadelphia, PA. STP 854. pp. 429-453.
Alabaster, J.S. and R. Lloyd. 1982. Water Quality Criteria for freshwater fish. Chapter 11.
Mixtures of Toxicants. London, Butterworth Scientific.
Arbuckle, W.B. 1983. Estimating activity coefficients for use in calculating environmental
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Banerjee, S.; S.H. Yalkowsky, and S.C. Valvani, 1980. Water solubility and octanol/water
partition coefficients of organics: Limitations of the solubility-partition coefficient
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Broderius, S. and M. Kahl. 1985. Acute toxicity of organic chemical mixtures to the fathead
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Brookes, P. 1977. Mutagenicity of polycyclic aromatic hydrocarbons. Mutation Res. 39:257-
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Buccafusco, R.J., S.J. Ells and G.A. LeBlanc. 1981. Acute toxicity of priority pollutants to
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Cairns, M. A. and A. V. Nebeker. 1982. Toxicity of acenaphthene and isophorone to early life
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1979. Water-related environmental fate of 129 priority pollutants. Volume H:
Halogenated aliphatic hydrocarbons, halogenated ethers, monocyclic aromatics, phthalate
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Chapman, G. A. 1987. Establishing sediment criteria for chemicals-regulatory perspective. In:
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Dickson, A.W. MaM and W.A. Brungs. Pergamon Press, New York. pp. 355-376.
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De Witt, T.H., R.J. Ozretich, R.C. Swartz, J.O. Lamberson, D.W. Shults, G.R. Ditsworth,
J.K.P. Jones, L. Hoselton, and L.M. Smith. 1992. The effect of organic matter quality
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amphipod, Rhepoxynius abronius. Environmental Toxicology and Chemistry 11:197-208.
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Chemosphere. 14(10): 1503-1538.
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Eadie, B.J., P.P. Landrum, W. Faust. 1982. Polycyclic aromatic hydrocarbons in sediments,
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EG&G Bionomics. 1982. Acute toxicity of selected chemicals to fathead minnow, water flea
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ERGO, 1981. Toxicity testing inter-laboratory comparison early life stage test with fathead
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PP-
Heitmuller, P.T., T.A. Hollister and P.R. Parrish. 1981. Acute toxicity of 54 industrial
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Hoke, R. 1992. Results of the third dieldrin sediment-spiking experiment. Memorandum to D.
Hansen, D. Di Toro and G. Ankley. December 2. 5 pp. -
Hoke, R., and G.T. Ankley. 1991. Results of dieldrin sediment spiking study conducted in
support of USEPA development of sediment quality criteria. Memorandum to D. Hansen
and D. Di Toro. June 18, 1991. 9 pp.
Holcombe, G.W., G.L. Phipps and J.T. Fiandt. 1983. Toxicity of selected priority pollutants
to various aquatic organisms. Ecotoxicol. Environ. Safety 7:400-409.
Home, J.D., M.A. Swirsky, T.A. Hollister, B.R. Oblad and J.H. Kennedy. 1983. Aquatic
toxicity studies of five priority pollutants. Report No. 4398. EPA Contract No. 68-01-
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Kamlet. M.J., R.M. Doherty, P.W. Carr, D. Mackay, M.H. Abraham, and R.W. Taft. 1988.
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prediction of octanol/water partition coefficients and other solubility and toxic properties
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Karickhoff, S.W., L.A. Carreira, C. Melton, V.K. McDaniel, A.N. Vellino, and D.E. Nute.
1989. Computer prediction of chemical reactivity - The ultimate SAR. U.S. EPA,
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EPA/600/M-89/017.
LeBlanc, G.A. 1980. Acute toxicity of priority pollutants to water flea (Daphnia magma"). Bull.
Environ. Contam. Toxicol. 24:684-691.
Lemke, A.E. 1984. Inter-laboratory comparison of continuous flow, early life stage testing with
fathead minnows. EPA-600/3-84-005 or PB84-129493. National Technical Information
Service, Springfield, VA. 26 pp.
Lemke, A.E. and R.L. Anderson. 1984. Insect interlaboratory toxicity test comparison study
for the chironomid (Paratanvtarsus sp.) procedure. EPA-600/3-84-054 or PB84-180025.
National Technical Information Service. Springfield, VA. 15 pp.
Lemke, A.E., E. Durban and T. Felhaber. 1983. Evaluation of a fathead minnow Pimephales
promelas embryo-larval test guideline using acenaphthene and isophorone. EPA-600/3-
83-062 or PB83-243436. National Technical Information Service, Springfield, VA. 26
pp.
Mabey, W.R., J.H. Smith, R.T. Podoll, H.L. Johnson, T. Mill, T.W., Chou, J. Gates, I.W.
Partridge, H. Jaber, and D. Vandenberg. 1982. Aquatic fate process data for organic
priority pollutants. U.S. EPA, Office of Water Regulations and Standards, Washington,
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DC, Final Report, EPA-440/4-81-041.
•*•
Mackay, D., A. Bobra, and W.Y. Shui. 1980. Relationships between aqueous solubility and
octanol-water partition coefficients. Chemosphere 9:701-711.
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minnows (Pimephales promelas). Final report to U.S. EPA, Duluth, MN. Marine
Bioassay Laboratories, 1234 Highway One, Watsonville, CA. 71 pp.
Milhelcic, J.R., and R.G. Luthy. 1988. Microbial degradation of polycyclic aromatic
hydrocarbons under denitrification conditions in soil-water suspensions. Final Report.
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Between Octanol-Water Partition Coefficient and Aqueous Solubility. Env. Sci. Technol.
19(6):522-528.
National Academy of Sciences/National Academy of Engineering. 1973. Water Quality Criteria
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of sediment organic carbon on survival of Hyalella azteca exposed to DDT and endrin.
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Technical Memorandum NOS OMA 59. NOAA Office of Oceanography and Marine
Assessment, Rockville, MD. 29 pp + appendices.
Noreen, E.W. 1989. Computer intensive methods for testing hypotheses: An introduction.
John Wiley and Sons Inc., New York, N.Y.
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Acute and chronic toxicity tests of 2,4,6-trichlorophenol and acenaphthene. (Contract No.
68-03-3081) Report to U.S. EPA, ERL-Duluth, MN. Northwestern Aquatic Sciences,
Inc., Newport, OR. 66 pp.
Randall, T.L. and P.V. Knopp. 1980. Detoxification of specific organic substances by wet
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oxidation. J. Water Pollut. Control Fed. 52:2117-2130.
«
Schuytema, G.A., A.V. Nebeker, W.L. Griffis, and C.E. Miller. 1989. Effects of freezing
on toxicity of sediments contaminated with DDT and endrin. Environ. Toxicol. and
Chem. 8(10):883-891.
Stephan, C.E., D.I. Mount, DJ. Hansen, J.H. Gentile, G.A. Chapman, and W.A. Brungs.
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of aquatic organisms and their uses. PB85-227049. National Technical Information
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Swartz, R.C. 1991. Acenaphthene and penanthrene files. Memorandum to David J. Hansen,
June 26, 1991. 160 pp.
Swartz, R.C., G.R. Ditsworth, D.W. Schults, and J.O. Lamberson. 1985. Sediment toxicity
to a marine infaunal amphipod: Cadmium and its interaction with sewage sludge. Mar.
Envir. Res. 18:133-153.
Swartz, R.C., D.W. Schults, T.H. DeWitt, G.R. Ditsworth, and J.O. Lamberson. 1990.
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7-5
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7-7
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