United State*
Environmental Protection Agency
Office of Water &
Office of Research and
Development
Office of Science and Technology
Health and Ecological Criteria Div.
Washington, D.C. 2046O
EPA-822-R-93-O12
September 1993 •
Sediment Quality Criteria
for the Protection of
Benthic Organisms:
FLUORANTHENE
Printed on Recycled Paper
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CONTENTS
Foreword ii
Acknowledgments iv
Tables vi
Figures vii
Introduction 1-1
Partitioning ........ 2-1
Toxicity of Fluoranthene: Water Exposures 3-1
Toxicity of Fluoranthene (Actual and Predicted): Sediment Exposures .... 4-1
Criteria Derivation for Fluoranthene 5-1
Criteria Statement 6-1
References -. 7-1
Appendix A: Summary of Acute Values for Fluoranthene for Freshwater and
Saltwater Species ; . A-l
Appendix B: Summary of Data from Sediment Spiking Experiments with
Fluoranthene . 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 kind and extent of effects
on human health and welfare, including effects on plankton, fish, shellfish, and wildlife, which
may be expected from the presence of pollutants in any body of water, including ground water,
(2) the concentration and dispersal of pollutants, or their byproducts, through biological, physical
and chemical processes, and (3) the effects of pollutants on biological community diversity,
productivity, and stability. Section 304(a)(2) directs the Administrator to develop and publish
information on, among other things, the factors necessary for the protection and propagation of
shellfish, fish, and wildlife for classes and categories of receiving waters.
To meet this objective, U.S. EPA has periodically issued ambient water quality criteria
(WQC) guidance beginning with the publication of "Water 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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Principal Author
David J. Hansen
Coauthors
Walter J. Beny
Dominic M. Di Toro
Paul R. Paquin
Laurie D. De Rosa
Frank E. Stancil, Jr.
Christopher S. Zaiba
Technical and Clerical Support
Heinz P. Kollig
Glen B. Thursby
Maria R. Paruta
BettyAnne Rogers
ACKNOWLEDGEMENTS
U.S. EPA, Environmental Research Laboratory,
Narragansett, RI
Science Applications International Corporation,
Narragansett, RI
Manhattan College, Bronx, NY;
HydroQual, Inc., Mahwah, NJ
HydroQual, Inc.,
Mahwah, NJ
HydroQual, Inc.,
Mahwah, NJ
U.S. EPA, Environmental Research Laboratory, Athens, GA
U.S. EPA Headquarters, Office of Water, Washington, DC
U.S. EPA, Environmental Research Laboratory, Athens, GA
Science Applications International Corporation,
Narragansett, RI
NCSC Senior Environmental Employment Program
Narragansett, RI
Science Applications International Corporation,
Narragansett, RI
IV
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Persons who have made significant contributions to the development of the approach and
supporting science used in the derivation of sediment criteria for nonionic organic contaminants
are as follows:
Herbert E. Allen
Gerald T. Ankley
University of Delaware, Newark, DE
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
Christina E. Cowan Battelle, Richland, WA
Dominic M. Di Tore
David J. Hansen
Paul R. Paquin
Spyros P. Pavlou
Richard C. Swartz
Nelson A. Thomas
HydroQual, Inc., Mahwah, NJ;
Manhattan College, Bronx, NY
U.S. EPA, Environmental Research Laboratory,
Narragansett, RI
HydroQual, Inc., Mahwah, NJ
Ebasco Environmental, Bellevue, WA
U.S. EPA, Environmental Research Laboratory,
Newport,OR
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
Christopher S. Zarba U.S. EPA Headquarters, Office of Water, Washington, DC
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Table 2-1
Table 2-2.
Table 3-1.
Table 3-2.
Table 3-3.
TABLES
i
Fluoranthene measured and estimated Log10KoW values.
Table 4-1.
Table 4-2.
Table 5-1.
Table 5-2.
Table 5-3.
Appendix A.
Appendix B.
Summary of logioKow values for fluoranthene measured by the U.S. EPA
Environmental Research Laboratory, Athens, GA. '
Chronic sensitivity of freshwater and saltwater organisms to fluoranthene
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 fluoranthene.
Results of approximate randomization test for the equality of freshwater and
saltwater FAV distributions for fluoranthene and approximate randomization test
for the equality of bentbic and combined benthic and water column (WOC) FAV
distributions.
Summary of tests with fluoranthene-spiked sediment.
Water-only and sediment LCSOs used to test the applicability of the equilibrium
partitioning theory for fluoranthene.
Sediment quality criteria for fluoranthene.
Analysis of variance for derivation of sediment quality criteria confidence limits
for fluoranthene.
Sediment quality criteria confidence limits for fluoranthene.
- Summary of acute values for fluoranthene for freshwater and saltwater species.
- Summary of data from sediment spiking experiments with fluoranthene. Data
from these experiments were used to calculate K^ 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 fluoranthene.
Figure 2-1. Organic carbon-normalized sorption isotherm for fluoranthene (top) and
probability plot of KQC (bottom) from sediment toxicity tests conducted by Swartz
et al., (1990) and DeWitt et al., (1992). The line in the top panel represents the
relationship predicted with a log KQC of 5.00, that is C^oc = KOC • 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. Effect of UV light on toxicity of fluoranthene to the mysid, Mysidopsis bahia.
Labels refer to light source (Vita-light and UVA-340 are artificial UV light
sources). UV measurements (in fiW/cm2) were made with a Macam
Photometries model UV-103 radiometer which read UV-A at 365 ± 36 nm. and
UV-B at 310 ± 34 nm. The UV-A irradiances were 57.5 for Vita-lights at 16
cm above the water, 252 for UVA-240 at 45 cm, 473 for UVA-340 at 26 cm, and
1,725 for sunlight. The curve was similar whether the data were plotted against
UV-A or UV-B. (From Thursby, 1992)
Figure 3-4. 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, 1991a), endrin (Nebeker et al., 1989; Schuytema et al.,
1989), or fluoranthene (Swartz et al., 1990; De Witt et al., 1992), and midge in
sediments spiked with dieldrin (Hoke, 1992) or kepone (Adams et al., 1985)
relative to pore water toxic units. Pore water toxic units are ratios of
concentrations of chemicals measured in individual treatments divided by the
water-only LC50 value from water-only tests. (See Appendix B in this SQC
document, Appendix B in the endrin, dieldrin, acenaphthene, and phenanthrene
SQC documents, and original references for raw data.)
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Figure 4-2.
Figure 5-1.
Figure 5-2.
Figure 5-3.
Percent mortality of amphipods in sediments spiked with acenaphthene or
phenanthrena (Swartz, 1991a), 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 in sediments
0*g/goc) divided by the predicted LC50 0*g/goc) in sediments (K^ x Water-only
LC50 Otg/L) x 1 Kgoc/^OOOgoc). (See Appendix B in this document and
Appendix B in the dieldrin, endrin, acenaphthene, and phenanthrene SQC
documents for raw data).
Comparison between SQC concentrations and 95% confidence intervals effect
concentrations from benthic organisms exposed to fluoranthene-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 Koc values. PGMCV = (GMAV -f- 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
JLC-jUs.
Comparison between SQC concentrations and 95% confidence intervals effect
concentrations from benthic organisms exposed to fluoranthene-spiked sediments
and sediment concentrations predicted to be chronically safe in salt water
sediments. . Concentrations predicted to be chronically safe (Predicted Genus
^f^^^C ValueS} PGMCy) a*6 derfved from the Genus Mean Acute Values
(GMAV) from water-only 96-hour lethality tests, Acute Chronic Ratios (ACR)
and Koc values. PGMCV = (GMAV + ACR)!^. 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 ranee of
LCSOs.
Probability distribution of concentrations of fluoranthene 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 fluoranthene SQC values
of 62 jtg/g in freshwater sediments having TOC = 10% and 6.2 ug/g in
freshwater sediments having TOC = 1% and compared to SQC values for
saltwater sediments of 30 /*g/g when TOC = 10% and 3.0 /*g/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 %.
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Figure 5-4. Probability distribution of concentrations of fluoranthene 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 300 jtg/goc-
DTSfTTATMHR
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.
AVAILABILITY NOTICE
This document is available to the public through the National Technical Information
Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161. NTIS Accession Number
XXXX-XXXXXX.
IX
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SECTION 1
INTRODUCTION
1.1 GENERAL INFORMATION:
\
Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U.S.
EPA) is responsible for protecting the chemical, physical and biological integrity of the nation's
waters. In keeping with this responsibility, the 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
i
water-column concentrations comply with established WQC. In addition, contaminated
sediments can be a significant pollutant source that may cause water quality degradation to
persist, even when other pollutant sources are stopped. The absence of defensible sediment
quality criteria (SQC) makes it difficult to accurately assess the extent of the ecological risks
of contaminated sediments and to identify, jdoritize and implement appropriate clean up
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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 concentrations in pore
water.
3. The distribution of sensitivities of benthic and water column organisms to
chemicals are similar; thus, the currently established WQC final chronic values
(FCV) can be used to define the acceptable effects concentration of a chemical
freely-dissolved in pore water.
The EqP approach, therefore, assumes that: (1) the partitioning of the chemical between
sediment organic carbon and interstitial water is at equilibrium; (2) the concentration in either
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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 via ingestion, sediment-
integument exchange, or from a mixture of exposure routes; (4) for nonionic chemicals, effect
concentrations in sediments on an organic carbon basis can be predicted using the organic carbon
partition coefficient (Koc) and effects concentrations in water; (5) the FCV concentration is an
appropriate effects concentration for freely-dissolved chemical in interstitial water; and (6) the
SQC 0*g/goc) derived as the product of the KQC and FCV is protective of benthic organisms.
SQC concentrations presented in this document are expressed as peg chemical/g sediment organic
carbon and not on an interstitial water basis because: (1) pore water is difficult to adequately
sample; and (2) significant amounts of the dissolved chemical may be 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 fluoranthene 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.
t
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 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. For
example, spill situations, in which the chemical equilibrium between water and sediments has
not yet been reached, a sediment chemical concentration that is 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 (TOG) in the sediment does not
greatly affect chemical binding (DeWitt et al., 1992). However, the physical form of the
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chemical in the sediment may have an effect. At some sites concentrations in excess of the SQC
;
may not pose risks to benthic organisms, because the compound may be a component of a
particulate, such as coal or soot, or 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 Deriving Site-specific
Sediment Quality Criteria for the Protection of Benthic Organisms" (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 fluoranthene. An understanding of the "Guidelines for Deriving
Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their
Uses" (Stephan et al., 1985), response to public comment (U.S. EPA, 1985) and "Technical
Basis for Deriving Sediment Quality Criteria for Nonionic Organic Contaminants for the
Protection of Benthic Organisms by Using Equilibrium Partitioning " (U.S. EPA, 1993a) is
necessary in order to understand the following text, tables and calculations. Guidance for the
acceptable use of SQC values is contained in "Guide for the Use and Application of Sediment
Quality Criteria for Nonionic Organic Chemicals" (U.S. EPA, 1993c).
1.2 GENERAL INFORMATION: FLUORANTHENE
Fluoranthene is a member of the polycyclic aromatic hydrocarbon (PAH) group of organic
compounds. Some sources of fluoranthene are crude oil, coal tar and motor oil (Verschueren,
1983). Fluoranthene is also produced naturally by plants, algae and bacteria (Suedel, 1989).
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Some PAHs are of environmental concern because they are known to be carcinogens and/or
t
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,
(Eadieetal., 1982).
Fluoranthene has a three ring structure and exists as pale yellow plates or needles
(Figure 1-1). It has a solubility in water of 0.26 mg/L at 25°C and is a solid at room
temperature (melting point of 111°C) (Verschueren, 1983). Two significant processes which
can influence the fate of fluoranthene in the sediment are sorption and biodegradation (U.S.
EPA, 1980). Sorption of fluoranthene onto solids in the water column and subsequent settling,
as well as partitioning onto organics in the sediment, can significantly affect fluoranthene
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 fluoranthene
in the sediment are oxidation, hydrolysis and volatilization (U.S. EPA, 1980).
In the absence of ultraviolet (UV) light the acute toxicity of fluoranthene ranges from 36
to >212 /tg/L for freshwater and 52 to >20,000 /tg/L for saltwater organisms (Appendix A).
Differences between fluoranthene concentrations causing acute lethality and chronic toxicity in
the absence of UV light in invertebrates are small; acute-chronic ratios (ACR) range from 3.385
to 3,404 for two species. The ACR for one species of fish is much larger (14.09). Sunlight
photoactivates fluoranthene to increase its toxicity to up to over an order of magnitude greater
than non-UV activated chemical (Appendix A). Although fluoranthene bioaccumulates hi 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
C16H10
202.26
1.252 g/cc (25°C)
Pale yellow plates or needles
5.0xlO
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1.3 OVERVIEW OF DOCUMENT:
Section 1 provides a brief review of the EqP methodology, and a summary of the
physical-chemical properties and aquatic toxicity of fluoranthene. Section 2 reviews a variety
of methods and data useful in deriving partition coefficients for fluoranthene and includes the
KQC recommended for use in the derivation of the fluoranthene SQC. Section 3 reviews aquatic
toxicity data contained in the fluoranthene WQC document (U.S. EPA, 1980) and new data that
were used to derive the FCV used in this document to derive the SQC concentration. In
addition, the comparative sensitivity of benthic and water column species is examined as the
justification for the use of the FCV for fluoranthene in the derivation of the SQC. Section 4
reviews data on the toxicity of fluoranthene in sediments, the need for organic carbon
normalization of fluoranthene 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 fluoranthene and its
uncertainty. The SQC for fluoranthene is then compared to STORET (U.S. EPA, 1989b) and
National Status and Trends (NOAA, 1991) data on fluoranthene's environmental occurrence in
sediments. Section 6 concludes with the criteria statement for fluoranthene. 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 fig
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
(EqP) methodology for deriving SQC. The methodology is discussed in detail in the "Technical
Basis for Deriving Sediment Quality Criteria for Nonionic Organic Contaminants for the
Protection of Benthic Organisms by Using Equilibrium Partitioning" (U.S. EPA, 1993a),
hereafter referred to as the SQC Technical Basis Document.
Bioavailability of a chemical at a particular sediment concentration often differs from one
sediment type to another. Therefore, a method is necessary for determining a SQC based on the
bioavailable chemical fraction in a sediment. For nonionic organic chemicals, the
concentration-response relationship for the biological effect of concern can most often be
correlated with the interstitial water (i.e., pore water) concentration 0*g chemical/liter pore
water) and not to the sediment chemical concentration (jig chemical/g sediment) (Di Toro et al.,
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1991). From a purely practical point of view, this correlation suggests that if it were possible
s
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 fluoranthene 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 Otg/L) be the acceptable concentration in water
for the chemical of interest; then compute the SQC using the partition coefficient, (Kp;
^fcgicdtmmt). between sediment and water:
SQC = KP FCV (2_D
This is the fundamental equation used to generate the SQC. Its utility depends upon the
existence of a methodology for quantifying the partition coefficient, KP.
Organic carbon appears to be the dominant sorption phase for nonionic organic chemicals
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in naturally occurring sediments and thus controls the 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
fluoranthene is discussed in this section and section 4. The organic carbon binding of a chemical
in sediment is a function of that chemical's organic carbon partition coefficient (Koc) and the
weight fraction of organic carbon in the sediment (foe). The relationship is as follows:
It follows that:
(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 KoWfor fluoranthene.
2.2 DETERMINATION OF KQW FOR FLUORANTHENE:
Several approaches have been used to determine KQW for the derivation of SQC, as
discussed in the SQC Technical Basis Document. At the U.S. EPA, Environmental Research
Laboratory at Athens, GA (ERL,A) three methods were selected for measurement and two for
estimation of KQW. The measurement methods were shake-centrifugation (SC), generator column
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(GCol) and slow-stir-flask (SSF) and the estimation methods were SPARC (SPARC Performs
T
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. FLUORANTHENE MEASURED AND ESTIMATED LOG10KoW VALUES.
METHOD
Measured
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
Estimated
LOG10KoW
5.15
4.90
4.95
4.95
5.21
5.22
5.29
5.33
REFERENCE
DeBruijnetaL, 1989
Mabey et al., 1982
CLOGP1
GEMSb
SPARC0
Yalkowsky et al., 1983
Mackay et al., 1980
Callahan et al., 1979
•CLOGP 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).
"GEMS is the Graphical Exposure Modeling System for the U.S. EPA. It is an interactive
computer system located on the VAX Cluster in the National Computer Center in Research
Triangle Park, North Carolina, under management of EPA's Office of Toxic Substances. PC
GEMS is the version for personal computers.
CSPARC is from SPARC Performs Automated Reasoning in Chemistry, (Karickhoff et al.,
1989).
• In the examination of the literature data for fluoranthene, only one primary reference was
found listing a measured logioKoW. This value was 5.15 (Table 2-1). Several primary
references were found for estimated log10KoW values in the literature ranging from 4.90 to 5.33
(Table 2-1). Although the range of reported values for fluoranthene is significantly lower than
2-4
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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 EKL, 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 fiuoranthene is 5.21.
The CLOGP program estimate of the log10KoW value for fiuoranthene using structure activity
relationships is 4.95.
We had little confidence in the available measured or estimated values for KOW, therefore
the SCj GCol, SSF methods were used to provide additional data from which to define KQW for
fiuoranthene (Table 2-2). The SC method yielded a log10KoW = 5.00 (n=3), the GCol method
yielded a log10KoW = 5.39 (n=4), and the SSF method yielded a log10KoW = 5.09 (n=6).
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,
fiuoranthene 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 5.09 is the value for log10KoW
recommended for SQC derivation. This value agrees with the SPARC estimated value and the
average of the values measured by the three methods under carefully controlled conditions at
2-5
-------
ERL,A. This KQW is the logarithm of the mean of six KQW measurements made by SSF. The
!
logs of the KOW values measured by SSF range from 4.98 to 5.23.
TABLE 2-2. SUMMARY OF LOG10KoW VALUES FOR FLUORANTHENE MEASURED BY
THE U.S. EPA ENVIRONMENTAL RESEARCH LABORATORY, ATHENS, GA.
SHAKE-
CENTR1FUGATION
rso
4.99
5.00
5.01
5.00s
GENERATOR
COLUMN
fGCoD
5.19
5.35
5.47
5.48
5.39*
SLOW-STIR
FLASK
4.98
5.02
5.02
5.10
5.14
5.23
5.09a
"Logic of mean of measured values.
2.3 DERIVATION OF KOC FROM ADSORPTION STUDIES:
Two types of experimental measurement of the K^ 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 fluoranthene, sediment organic carbon (OC) and non-dissolved organic carbon (DOC)
associated fluoranthene dissolved in pore water were used to compute KQC-
2.3.1 KOC FROM PARTICLE SUSPENSION STUDIES:
Laboratory studies to characterize adsorption are generally conducted using particle
suspensions. The high concentrations of solids and turbulent conditions necessary to keep the
2-6
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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 KOC / ux
where m is the particle concentration in the suspension (kg/L), and u* = 1.4, an empirical
constant. In this expression the KOC is given by:
logioKoc = 0.00028 + 0.983 log10KoW (2-5)
Experimental data demonstrating the effect of particle suspensions was not found for
fluoranthene during a comprehensive literature search for partitioning information. Nonetheless
the above discussion highlights the need to consider particle interaction effects when interpreting
partitioning data in particle suspension studies.
In the absence of particle effects, KQC is related to KQW via Equation 2-5. For
log10KoW = 5.09 (See section 2.2), this expression results in an estimate of logtoKoc = 5.00.
2.3.2 KOC FROM SEDIMENT TOXICITY TESTS:
Measurements of KOC are available from sediment toxicity tests using fluoranthene
(Swartz et al., 1990, DeWitt et al., 1992). The tests from Swartz et al. (1990) used three
2-7
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saltwater sediments with a range of. organic carbon contents of 0.2 to 0.5 percent. The tests
}
from DeWitt et al. (1992) used a marine sediment amended with five types of organic matter
yielding sediments with organic caibon contents of 0.3 to 0.43 percent. (Table 4-1; Appendix
B.) Fluoranthene concentrations were measured in the sediment and pore waters providing the
data necessary to calculate the partition coefficient for an undisturbed bedded sediment. Swartz
et al. (1990) measured total pore water concentrations and DeWitt et al. (1992) measured total
and freely dissolved pore water concentrations. The relationship of total to free pore water in
the DeWitt et al. (1992) analysis was applied to the total pore water data in Swartz et al. (1990)
to estimate free pore water in that study. Also an adjustment was made to pore water to account
for 28% loss due to filtration during experimental measurements (Swartz, 1992).
The upper panel of Figure 2-1 is a plot of the organic carbon-normalized sorption
isotherm for fluoranthene, where the sediment fiuoranthene concentration, G*g/gOC) is plotted
versus free (dissolved) pore water concentration, fttg/L). The data used to make this plot are
included in Appendix B. One of the amended sediments for the DeWitt analysis was omitted
because the author believed that elevated dissolved organic carbon may have interfered with the
method used to determine free interstitial water fiuoranthene concentrations. The line of unity
slope corresponding to the log10Koc = 5.00 derived from SSF is compared to the data. A
probability plot of the observed experimental logl^Koc values is shown in the lower panel of
Figure 2-1. The log10KoC values are approximately normally distributed with a mean of log10KoC
- 5.05 and a standard error of the mean of 0.071. This value agrees with the log10BQjC of 5.00,
which was computed from the SSF determined (Section 2.2) fluoranthene log10KoW of 5.09 using
Equation 2-5.
2-8
-------
100000
10000
O
jg
g
o°
ZQ
81
Q
CO
10
= I I I I I III!
I I III! I I I I I lilt I III I LKI
I I Ig
LEGEND
— Swortz et al., 1990
- De Witt et al., 1992
11 mi
i i i MHI i in inn
i i MIII i i i i mi
0.1
10
100
1000
10000
PORE WATER CONCENTRATION (ug/L)
oc = KQC • Cd.
2-9
-------
2.4 SUMMARY OF DERIVATION OF KQC FOR FLUORANTBDENE:
r
The KQC selected to calculate the SQC for fluoranthene is based on the regression of
logioKoc to logioKoW (Equation 2-5), using the fluoranthene logioKoW of 5.09 recently measured
by ERL, A. This approach, rather than the use of the KQC from the toxicity tests, was adopted
because the regression equation is based on the most robust data set available that spans a broad
range of chemicals and particle types, thus encompassing a wide range of KQW and foc. The
regression equation yields a log10Koc of 5.00. This value is in agreement with the log10Koc of
5.05 measured in the sediment toxicity tests.
2-10
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SECTION 3
TOXICITY OF FLUORANTHENE: WATER EXPOSURES
3.1 TOXICITY OF FLUORANTHENE IN WATER: DERIVATION OF FLUORANTHENE
WATER QUALITY CRITERIA:
The equilibrium partitioning (EqP) method for the derivation of sediment quality criteria
(SQC) uses the fluoranthene water quality criterion (WQC) final chronic value (FCV) and
partition coefficients (K<>c) 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 fluoranthene 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. Tests conducted in the dark or under normal laboratory lighting are discussed in
Sections 3.2 and 3.3. The phototoxicity of fluoranthene is discussed in Section 3.4.
3.2 ACUTE TOXICITY - WATER EXPOSURES:
Twenty standard acute toxicity tests with fluoranthene have been conducted under normal
laboratory lighting on 13 freshwater species from 12 genera (Appendix A). Overall genus mean
acute values (GMAVs) range from 36 to > 211.7 /tg/L. The acute values for three species
tested, Lumbriculus variegatus. Physella virgata. and Ophiogomphus sp., were all > 178.5
/tg/L. The acute values for the 4 most sensitive species differed by only a factor of 3; 32 to 95
3-1
-------
/*g/L. Nine tests on benthic life-stages of nine species from eight genera are contained in this
t
database (Figure 3-1; Appendix A). Benthic organisms were similar to water column species
in sensitivity to fluoranthene; GMAVs range from 61.1 to > 178.5 fig/L. One infaunal species,
the annelid L variegatus. had an LC50 of > 178.5 /tg/L. The final acute value (FAV) derived
from the overall GMAVs (Stephan et al., 1985) for freshwater organisms is 33.58 ng/L (Table
3-2).
Fifteen acute tests have been conducted on 8 saltwater species from 8 genera (Appendix
A). Overall GMAVs range from 51.54 to > 20,000 jtg/L, with crustaceans the most sensitive.
There are results from 6 tests on benthic life-stages of 6 species from 6 genera (Figure 3-2;
Appendix A). Benthic organisms were among both the most sensitive and most resistant
saltwater genera to fluoranthene. The most sensitive benthic species is the mysid, Mvsidopsis
bahia, with an average flow-through 96 hour LC50 of 51.54 /tg/L based on two tests with
measured concentrations. Other benthic species for which there are data are only slightly less
sensitive while others are resistant to fluoranthene and had GMAVs ranging from 66.93 to >
20,000 (ig/L. The FAV derived from the overall GMAVs (Stephan et al., 1985) for saltwater
organisms is 16.13 jig/L (Table 3-2).
3.3 CHRONIC TOXECITY - WATER EXPOSURES:
Chronic toxicity tests have been conducted with fluoranthene using a freshwater
cladoceran, (Daphnia magna) and fathead minnows (Pimepheles promelas) and a saltwater mysid
Mi. frahia (Table 3-1). I), magna and M- bahia were tested in life-cycle exposures. P. promelas
were exposed in an early life-stage toxicity test.
3-2
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O Rshes
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3W
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D Other Invertebrates
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Sea Urchin
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Neanthes (J)
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Palaemonetes (J)
Grass Shrimp
Menidia (J)
Silverside
Ampelisca (J)
Amphipod
_L
20 40 60 80
PERCENTAGE RANK OF SALTWATER GENERA
100
Figure 3-2. Genus mean acute values from water only acute toxicity tests using saltwater
species vs. percentage rank of their sensitivity. Symbols representing bentbic
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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n, magna exposed to fluoranthene for 21 days were not affected in 10.6 /tg/L or less.
t
There was a reduction, relative to controls, of 23.1% in growth at 2L2 /tg/L and 36.9% in
survival at 43.5 /tg/L. P. r>romelas exposed to fluoranthene for 28 days in an early life-stage
toxicity test were not affected in 10.4 /tg/L or less. There was a reduction of 67% in survival
and a 50.2% reduction in growth relative to controls in 21.7 /tg/L. In the 96 hour acute lethality
test, 25% of the P. promelas died in 211.7 /tg/1, saturation for fluoranthene (Brooke, 1991).
The concentration of 211.7 /tg/1 was used as the acute value for P. promelas because this value
approximated an LC50 and greater concentrations could not be tested in this measured flow-
through test.
Saltwater M. bjhja were tested in two life-cycle toxicity tests. In the first, they were
exposed to fluoranthene for 28 days (EG&G, 1978). There was no effect on survival or
reproduction (growth was not measured) after 28 days of exposure to fluoranthene at
concentrations <.12 /tg/L. At a fluoranthene concentration of 21 /tg/L survival and reproduction
were reduced by 26.7 and 91.7%, respectively, relative to the controls. At the highest
concentration of fluoranthene, 43 /tg/L, all M. bjhia died. In the second test, M- bahia were
exposed to fluoranthene for 31 days (Champlin and Poucher, 1991b). Effect concentrations were
similar to the first test. M- bahia were not affected at fluoranthene concentrations <. 11.1 /tg/L.
Survival was reduced 30%, growth 12% and reproduction 100% relative to controls, in 18.8
/tg/L, the highest concentration tested.
The difference between acute and chronic sensitivity to fluoranthene, in tests where UV
activation did not occur, varied minimally between species (Table 3-2). Three species mean
acute-chronic ratios (ACR) are available; 3.385 for D. magna. 3.404 for M. bahia and 14.09
3-6
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3-7
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for E promelas. The final acute chronic ratio final (ACR) is the geometric mean of these three
values (5.455).
The FCVs (Table 3-2) are used as the effect concentrations for calculating the SQC for
protection of freshwater and saltwater benthic species. The FCV for freshwater organisms of
6.16 ftg/L is the quotient of the FAV of 33.6 pg/L and the final ACR of 5.45. Similarly, the
FCV for saltwater organisms of 2.96 jtg/L is the quotient of the FAV of 16.1 /tg/L and the final
ACR.
3.4 PHOTOTOXrCTTY OF FLUORANTHENE:
Under laboratory conditions many PAHs are predicted to be not acutely toxic at or below
their solubility in water (Veith et aL, 1983). Under ultra-violet light (UV), however, the
toxicity of some PAHs can increase by several orders of magnitude. This effect has been shown
to be a result of photoactivation rather than photodegradation of the parent compound to more
toxic metabolites. With some PAHs, toxicity occurs by activation by UV light of chemical
present on or within an organism. Bluegills (Lepomis macrochims^ exposed to anthracene in
sunlight in outdoor artificial streams died; L. macrochirus in the same stream, but downstream
in the shade survived. L. macrochirus exposed in the shade died within 24 hours when placed
into clean water and brought into the sunlight. Likewise, D. magna were much more sensitive
to anthracene in the presence of sunlight than when exposed under laboratory light, with toxicity
proportional to UV intensity (Allred and Giesy, 1985). UV-A wavelengths (UVA) were
implicated as responsible for most of the photoinduced toxicity.
The mechanism by which UV light activates PAHs is the same as that for electron
3-8
-------
excitation of plant pigments during photosynthesis by visible light. This process of excitation
;
of PAH electrons and the probable consequence of that excitation are reviewed by Newsted and
Giesy (1987). Briefly, if a compound absorbs light, then electrons can be elevated to higher
energy states to form the excited singlet state. If the excited electrons return immediately to
their ground state then the extra energy is lost harmlessly through fluorescence. However, if
the electrons pass through a triplet state, then the energy can be transferred to other molecules
(thought to be oxygen in the case of PAHs). Singlet oxygen formed in this process is capable
of denaturing biomolecules. Singlet oxygen is very reactive with water and unless organisms,
PAH, and sunlight are present simultaneously, photoactivation does not enhance toxicity.
Therefore, benthic organisms which remain buried or organisms in the shade can survive PAH
concentrations which would be lethal if they emerged from the sediment or shade into sunlight.
PAHs are concentrated in the non-polar environments of cells, such as the phospholipids of
membranes. Singlet oxygen in tissues is longer lived, thus greatly increasing the likelihood that
it would denature biomolecules. This also explains why membrane damage is one of the
probable mechanisms for this type of toxicity (Kagan et al., 1987) and why organisms exposed
to PAHs out of direct sunlight die when placed in the sun in PAH-free water.
Fluoranthene has exhibited photoinduced toxicity during standardized toxicity tests with
a variety of organisms (Appendix A). Although, the toxicity of fluoranthene appears to increase
with increases in intensity of UVA at low UV intensities (Figure 3-3), the acute toxicity of
fluoranthene to saltwater organisms is similar under commercially available UV lights and
sunlight (Figure 3-3; Appendix A). This is important since conducting acute and particularly
chronic toxicity tests outside hi sunlight would be extremely difficult and expensive. The
3-9
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3-10
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magnitude of increase in fluoranthene's toxicity following UV activation can be great. The ratio
j
of LCSOs from acute tests conducted in the dark or under cool-white fluorescent light {"dark")
to LCSOs for the same species exposed in the same laboratory using either UV lights or sunlight
ranges from 2 to 5,000 (Appendix A). This enhanced toxicity also can occur with relatively
short exposures to UV light. Kagan et aL, (1985) observed that 1,000 /tg/L of fluoranthene,
pyrene or anthracene was not toxic to five aquatic species in 30 minute exposures. Exposures
to these PAHs for 30 minutes followed by 30 to 60 minutes exposures to UV resulted in LCSOs
from 4 to 360 /tg/L. Data from chronic tests conducted in both the "dark" and under UV light
are available for two freshwater species, D. magna and P. promelas. and one saltwater species,
M. bahia. The chronic values decreased by factors of 33 for D. magna. 5.8 for P. promelas.
and 2.1 for M. bahia in the presence of UV light (Table 3-2). The magnitude of increase in
acute toxicity under UV light was greater than the magnitude of increase in chronic toxicity.
under UV light, and so the acute-chronic ratios under UV light are somewhat lower (Table 3-2).
There are not enough acute toxicity data from tests using UV light to calculate final acute
and chronic values for freshwater and saltwater aquatic life. However, if existing freshwater
and saltwater UV data are combined, the magnitude of the possible decrease in the FCV for
organisms in photic zones can be approximated. The FAVuv derived using overall GMAVs
from all UV and sunlight tests (Appendix A) is 0.5067 /&g/L. This value is approximately 66
times lower than the "dark" freshwater FAV, and approximately 32 times lower than the "dark"
saltwater FAV.
The difference between acute and chronic sensitivity to fluoranthene, in tests under UV
light, varied minimally between species (Table 3-2). Three species mean ACR are available;
3-11
-------
1.05 for I>. daphnia, 1.492 for M- bahia and 4.72 for P. promelas. The final ACR under UV
j
light is the geometric mean of these three values (1.948). The estimated FCVuv for freshwater
and saltwater organisms combined of 0.2601 is the quotient of the FAVuv of 0.5067 and the
final ACR.
At first glance it might seem that photoinduced increases in toxicity are not relevant to
benthic organisms and that SQC should not be derived using data from UV toxicity tests. This
may not be true and for this reason SQC in this document may be under protective in some
instances. There are many examples of specific benthic organisms where exposure to
fluoranthene (and other PAHs) and sunlight can co-occur. For example, fiddler crabs which
typically occupy burrows within the sediment, could accumulate fluoranthene from that sediment,
and when they come out onto the surface of the sediment at low tide during daylight hours could
be affected by PAHs in their tissues. Most freshwater insects that inhabit sediment during early
developmental stages could also be affected by photo-induced PAHs in their tissues when they
mature and emerge from water during daylight as adults. The importance of PAH's transferred
from benthic species in aquatic food chains to aquatic predators which may be exposed to
sunlight is unknown. Rooted aquatic plants also could be directly affected by fluoranthene
contaminated sediment if they were to accumulate fluoranthene and translocate it to their leaves.
Plants may be a source of photoactivated PAHs to herbivores.
At this time, U.S. EPA does not recommend a SQC value that considers fluoranthene
toxicity data from UV tests. This is partly because data are insufficient to calculate a FCV.
More importantly, there is an absence of data demonstrating a causal linkage between exposure
of sediment-associated fluoranthene and increased risks of UV enhanced effects on benthic
3-12
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organisms or organisms coupled to benthic organisms via food chains. EPA. encourages research
efforts on these topics.
3.5 APPLICABILITY OF THE WATER QUALITY CRITERION AS THE EFFECTS
CONCENTRATION FOR DERIVATION OF THE FLUORANTHENE 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 fluoranthene-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 fluoranthene.
For fluoranthene, benthic species account for 8 out of 12 genera tested in freshwater,
and 6 out of 8 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
3-13
-------
to fluoranthene was performed using the Approximate Randomization method (Noreen, 1989).
i
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 3-4). 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 fluoranthene, the test-statistic falls at the 78 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
3-14
-------
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Probability distribution of FAV difference statistics to compare water-only data
from freshwater vs. saltwater (upper panel) and benthic vs. WQC (lower panel)
data.
3-15
-------
TABLE 3-3. RESULTS OF APPROXIMATE RANDOMIZATION TEST FOR
THE EQUALITY OF THE FRESHWATER AND SALTWATER LC50
DISTRIBUTIONS FOR FLUORANTHENE AND APPROXIMATE
RANDOMIZATION TEST FOR THE EQUALITY OF BENTfflC AND
COMBINED BENTfflC AND WATER COLUMN (WQC) LC50
DISTRIBUTIONS.
Compar-
ison Habitat or Water Type* ARStatisticb Probability0
Fresh Fresh (12) Salt (8) 17~4 78~
vsSalt
Benthic Benthic (14) WQC (20) 3.64 74
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.
"Probability that the theoretical AR statistic .<_ than the observed AR statistic
given that the samples came from the same population.
Since freshwater and saltwater species showed similar sensitivity, a test of difference in
sensitivity for benthic and all (benthic and water column species combined, hereafter referred
to as "WQC") organisms combining freshwater and saltwater species using the Approximate
Randomization method was performed. The test statistic in this case is the difference between
the WQC FAV, computed from the WQC LC50 values, and the benthic FAV, computed from
the benthic organism LCio 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
3-16
-------
the number of benthic organisms are selected from the WQC data set. A "benthic" FAV is
computed. The original WQC FAV and the "benthic" FAV are then used to compute the
difference statistic. This is done many times and the distribution that results is representative
of the population of FAV difference statistics. The test statistic is compared to this distribution
to determine its level of significance. The probability distribution of the computed FAV
differences are shown in the bottom panel of Figure 3-4. The test statistic for this analysis falls
at the 74 percentile and the hypothesis of no difference.in sensitivity is accepted (Table 3-3).
This analysis suggests that the FCV for fluoranthene based on data from all tested species is an
appropriate effects concentration for benthic organisms.
3-17
-------
-------
SECTION 4
TOxicrrY OF FLUORANTHENE (ACTUAL AND PREDICTED):
SEDIMENT EXPOSURES
4.1 TOXKTTY OF FLUORANTHENE IN
The toxicity of fluoranthene spiked into sediments has been tested with three saltwater
amphipod species, and one .amphipod, one midge, one cladoceran, and two fish species from
freshwater. Data from all species tested have been included in Table 4-1, but only data from
tests with benthic species have been included in Table 4-2 and Figures 4-1 and 4-2. All
concentrations of fluoranthene 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 fluoranthene 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
fluoranthene. Data from these studies have not been included here because it is not possible to
determine the contribution of fluoranthene to the observed toxicity.
Gendusa (1990) exposed fathead minnows, Pimephales promelas. and channel catfish,
Ictalurus punctatus. to fluoranthene-spiked sediments with a total organic carbon content (TOC)
4-1
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of 0.7%. Sediments were spiked with fluoranthene dissolved in acetone which was added
directly to the sediment. After the acetone was allowed to evaporate for 15 minutes, crystalline
fluoranthene was stirred into the sediment, and overlying water was added. Animals were added
24 hours later. I. punctatus were exposed for 96 hours, P. promelas were exposed for 96 hours
and 30 days. Fluoranthene concentrations at the end of the sediment tests were much lower than
at the beginning of the tests, presumably because the sediments did not have time to equilibrate.
The 4-day LCSOs for I. punctatus and P. promelas and the 30-day LC50 for P. promelas were
3.683, 0.97 and 0.437 jtg/g dry wt. respectively, and 526.2, 131.0, and 62.46 jtg/g organic
carbon respectively. These LCSOs are means of the LCSOs calculated on the basis of the initial
and final measurements of fluoranthene in the tests.
Suedel et al., (1993) exposed the amphipod, Hyalella azteca. the midge, Chironomus
tentans, and the cladoceran, Daphnia magna. to three fluoranthene-spiked sediments of similar
TOC content, 0.44 to 0.50%. Sediments were spiked with fluoranthene dissolved in acetone.
Microliter quantities were added to the sediment surface and allowed to dry for two minutes,
then the sediments were mixed by hand for 60 seconds before the addition of test water.
Animals were added to the test beakers 18-24 hours later. LCSOs normalized to dry wt. differed
by a factor of 3.2 (2.3 to 7.4 jig/g) for H. azteca. a factor of 2.9 (3.0 to 8.7 /zg/g) for C.
tentans. and a factor of 3.6 (4.2 to 15.0 /tg/g) for D. magna. The organic carbon normalized
LCSOs for H. azteca differed by a factor of 3.0 (500 to 1,480 /ig/goc), a factor of 2.6 (682 to
1,740 g/goc) for C. tentans. and a factor of 3.4 (955 to 3,261 A*g/goc) f°r D- magna. Organic
carbon normalization had little effect because the TOC contents of the three sediments were so
similar.
4-5
-------
Swartz et al. (1990) exposed the amphipods Corophium spinicorne and Rhepoxvnius
t
abronjus to three fluoranthene-spiked sediments with TOC contents of 0.18%, 0.31% and
0.48 %. Sediments were prepared using the methods of Swartz et al. (1985) by mixing varying
amounts of organically-rich fine sediment into sand with a low organic content. Fluoranthene,
dissolved in acetone, was added to sediment aliquots in rolling mill jars and rolled. The
sediments were allowed to equilibrate for approximately two weeks before sediment was added
to test chambers. The 10-day LCSO's for R. abronius increased with increasing organic carbon
concentration when the fluoranthene concentration was expressed on a dry weight basis, but were
not different when concentration was expressed on an organic carbon basis. LCSO's normalized
to dry weight differed by a factor of 3.1 (3.4 to 10.7 jtg/g) for R. abronius over a 2.7-fold range
of TOC. The organic carbon normalized LCSO's for R. abronius differed by a factor of 1.2
(1,890 to 2,230 jig/goc). Because less than 50% mortality of C. spinicorne resulted in the
highest fluoranthene treatments in two of the three sediments used in this experiment, it was not
possible to make similar comparisons with this species.
De Witt et al. (1989) exposed the saltwater amphipod E. estuarius to fluoranthene-spiked
sediments at five different salinities and R.. abronius and the freshwater amphipod IL azteca to
fluoranthene-spiked sediments at single salinities (Table 4-1). Sediments were spiked with
fluoranthene dissolved in acetone and mixed on a rolling mill intermittently over a 24 hour
period. Overlying water was then added to the test chambers and allowed to equilibrate for 24
hours before the addition of test animals. Fluoranthene toxicity to E. estuarius was not affected
by interstitial water salinity. Nominal LC50 values (fluoranthene was not measured at all
salinities) varied by a factor of 1.3 (range 13.8 to 17.5 pg/g) on a dry weight basis. TOC was
4-6
-------
not measured in these sediments. The 10-day LC50 for R. abronius (5.1 /tg/g, dry wt) was
;
similar to those reported by Swartz et al. (1990).
De Witt et al. (1992) exposed R. abronius to five fluoranthene-spiked sediments of
similar organic carbon content amended with organic carbon from five sources: Zostera marina
(eelgrass); fine grained material which had settled from the water column of Yaquina Bay, OR;
organic-rich sediment from a small slough in Alsea Bay, OR; feces of a suspension-feeding
oyster (Crassostrea gigas) and feces of a deposit-feeding shrimp fCallianassa californiensis').
Sediments were spiked by shell coating fiuoranthene onto glass jars and rolling for 24 hours.
The sediments were then allowed to equilibrate for 5 weeks before use in the experiments. The
TOC content of the sediment varied from 3.1 to 4.0%. The authors concluded that the source
of organic carbon had little effect on the 10-day LC50 values which varied by a factor of 2.2
(range 8.65 to 19.1jtg/g) on a dry weight basis and by a factor of 2.0 (range 2,790 to 5,620
/tg/g on an organic carbon basis.
Combining the results of Swartz et al., (1990) and De Witt et al. (1992) for R. abronius.
10-day LC50 values for fluoranthene varied by a factor of 5.6 (3.4 to 19.1 /tg/L) on a dry
weight basis and by a factor of 3.0 (1,890 to 5,620) on an organic carbon normalized basis
(Table 4-1).
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 fluoranthene is also 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
4-7
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sediments (>315 /tg/L). This result runs counter to previous observations that the pore water
!
chemical concentration shows strong correlation with toxicity. De Witt et al. (1992) note that
this eelgrass amended sediment was the only sediment tested where the organic carbon originated
from fresh plant material. Similarity of LCSOs on an organic carbon basis and dry weight basis
suggest that the pore water fluoranthene may not have been entirely bioavailable. Free (not
associated with dissolved organic carbon, DOC) pore water fluoranthene concentration was
measured using a modification of the Landrum et al. (1984) reverse-phase separation method.
Free pore water fluoranthene concentrations were generally 60 to 90% of the total pore water
concentrations. For the five sediments tested, LC50 values based on free flouranthene in pore
water (LC50 = 8.09 to > 179 /tg/L) were as variable as those based on total pore water
fluoranthene (9.38 to > 315 /*g/L). The fact that the organic-carbon normalized LC50s may be
better predictors of toxicity than pore water concentrations was also observed with dieldrin
(Hoke and Ankley, 1991). Partitioning to dissolved organic carbon was proposed to explain the
lack of similarity of LC50 values based on total pore water dieldrin concentrations. This
explanation is not applicable to results with fluoranthene because the total and free pore water
fluoranthene concentrations and LC50 values were simitar and uniform across all sediment
types.
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 acenanphthene and
phenanthrene (Swartz, 1991a), endrin (Nebeker et al., 1989; Schuytema et al., 1989), dieldrin
(Hoke 1992), fluoranthene (Swartz et al., 1990, DeWitt et al. 1992), or kepone (Adams et al.,
4-11
-------
1985) (Figure 4-1; Appendix B). The data included in this analysis come from tests conducted
r ' i
at EPA laboratories or from tests which utilized designs at least as rigorous as those conducted
at the EPA laboratories. Tests with acenaphthene and phenanthrene used two saltwater
amphipods QU. plumulosus and E.. estuarius') and marine sediments. Tests with fluoranthene used
a saltwater amphipod QjL. abroniusl and marine sediments. Freshwater sediments spiked with
endrin were tested using the amphipod H. azteca: while kepone-spiked sediments were tested
using the midge, £. tentans. Figure 4-1 presents the percentage mortalities of the benthic
species tested in individual treatments for each chemical versus "pore water toxic units" (PWTU)
for all sediments tested. PWTUs are the concentration of the chemical in pore water fyig/L)
divided by the water only LC50 fttg/L). Theoretically, 50% mortality should occur at one
interstitial water toxic unit. At concentrations below one PWTU there should be less than 50%
mortality, and at concentrations above one PWTU there should be greater than 50% mortality.
Figure 4-1 shows that at concentrations below one PWTU mortality was generally low, and
increased sharply at approximately one PWTU. Therefore, this comparison supports the concept
that interstitial water concentrations can be used to predict the response of an organism to a
chemical that is not sediment specific. This pore water normalization was not used to derive
SQC in this document because of the complexation of nonionic organic chemicals with pore
water DOC (Section 2) and the difficulties of adequately sampling pore waters.
4.3 TESTS OF THE EQUILIBRIUM PARTITIONING PREDICTION OF SEDIMENT
TOXICITY:
SQC derived using the EqP approach utilize partition coefficients and FCVs from WQC
documents to derive the SQC concentration for protection of benthic organisms. The partition
4-12
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4-13
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coefficient (Koc) is used to normalize sediment concentrations and predict biologically available
r
concentrations across sediment types. The data required to test the organic carbon normalization
for fluoranthene 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
/tg/L, (2) an identical sediment effect concentration on an organic carbon basis, such as a 10-day
LC50 value in /tg/goc, and (3) a partition coefficient for the chemical, KQC in L/kgoc. This
section presents evidence that the observed effect concentration in sediments (2) can be predicted
utilizing the water only effect concentration (1) and the partition coefficient (3).
Predicted ten-day LC50 values from fluoranthene-spiked sediments tests with H. azteca
and £. tentans were calculated (Table 4-2) using the LogjoKoc value of 5.00 from Section 2 of
this document and the sediment LC50's in Suedel (1993). Ratios of actual to predicted LCSOs
for fluoranthene averaged 0.217 (range 0.111 to 0.330) for H. azteca and 0.387 (range 0.214
to 0.545) for C. tentans. The ratio of actual to predicted LC50 for I. punctatus (Gendusa, 1990)
was 0.146. Data on P.. promelas (Gendusa, 1990) and D. maena were not used for prediction
because they are not benthic organisms.
Predicted ten-day LC50 values on a ftg/goc basis from fluoranthene-spiked sediment tests
with IL. abronius (Swartz et al., 1990; De Witt et al., 1992) were calculated (Table 4-2) using
the value of KQC (105-00) from Section 2 of this document and the 10-day water-only EC50 values
in Swartz (1991a). Ratios of actual to predicted LCSOs for fluoranthene averaged 2.13 (range
1.36 to 4.04) for IL. abronius. The data from De Witt et al (1989) can not be used for
prediction because the TOC of the sediments was not measured.
4-14
-------
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 fiuoranthene.
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 for kepone obtained
using the slow stir flask method is not available. Swartz (1991a) exposed the saltwater
amphipods R estuarius and L.. plumulosus to acenaphthene in three marine sediments having
organic carbon contents ranging from 0.82 to 4.2% and to phenanthrene in three marine
sediments having organic carbon contents ranging from 0.82 to 3.6%. Swartz et al. (1990)
exposed the saltwater amphipod IL abronius to fiuoranthene 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 C. tentans to two freshwater dieldren-spiked sediments
having 2.0 and 1.5% organic carbon. Nebeker et al. (1989) and Schuytema et al. (1989)
exposed H. azteca to three endrin-spiked sediments having 3.0, 6.1 and 11.2% organic carbon.
Figure 4-2 presents the 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
Gig/goc) in 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
4-15
-------
a
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4-16
-------
these tests calculated on a PSTU basis were 1.90 for acenaphthene, 1.16 for dieldrin, 0 .44 for
j
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.
4-17
-------
-------
SECTIONS
CRITERIA DERIVATION FOR FLUORANTHENE
5.1 CRITERIA DERIVATION:
The water quality criteria (WQC) Final Chronic Value (FCV), without an averaging
period or return frequency (See section 3), is used to calculate the sediment quality criteria
(SQC) because it is probable that the concentration of contaminants in sediments are relatively
stable over tune, 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 areal
extent and vertical stratification of contamination of a sediment at a site in much the same way
that averaging periods or mixing zones are used in WQC.
The FCV is the value that should protect 95%of the tested species included in the
calculation of the WQC from chronic effects of the substance. The FCV is the quotient of the
Final Acute Value (FAV), and the final Acute Chronic Ratio (ACR) for the substance. The
FAV is an estimate of the acute LC50 or EC50 concentration of the substance corresponding to
a cumulative probability of 0.05 for the genera from eight or more families for which acceptable
acute tests have been conducted on the substance. The ACR is the mean ratio of acute to
chronic toxicity for three or more species exposed to the substance that meets minimum database
5-1
-------
requirements. For more information on the calculation of ACRs, FAVs, and FCVs see the
t
National Water Quality Criteria Guidelines (Stephan et al., 1985). The FCV used in this
document differs from the FCV in the fluoranthene WQC document (U.S. EPA, 1980) because
it incorporates recent data not included in that document, and omits some data which does not
meet the data requirements established in the WQC Guidelines (Stephan et al., 1985).
The equilibrium partitioning (EqP) method for calculating SQC is based on the following
procedure. If FCV Otg/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 nonionic organic chemicals
in naturally occurring sediments, (salinity, grainsize and other sediment parameters have
inconsequential roles in sorption, see sections 2.1 and 4.3) the organic carbon partition
coefficient, (Koc) can be substituted for KP. Therefore, on a sediment organic carbon basis, the
SQCoc G*g/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 fluoranthene SQC.
The organic carbon normalized SQC is applicable to sediments with an organic carbon
fraction of foe ^ 0.2%. For sediments with f^ < 0.2%, organic carbon normalization and
SQC may not apply.
5-2
-------
TABLE 5-1. SEDIMENT QUALITY CRITERIA FOR FLUORANTHENE
Type of
Water Body
Fresh Water
Salt Water
Log10Kow
(L/kg)
5.09
5.09
Logi^
(L/kg)
5.00
5.00
FCV
6.16
2.96
(Mg/goc)
620"
300b
= (105'00 L/kgoc)«(10-3 kgoc/gocW6.16 fig fluoranthene/L) = 620 fig
fluoranthene/goc
= (105-00 L/kgocXlO'3 kgoc/goc)»(2.96 fig fluoranthene/L) = 300 fig
fluoranthene/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:
Hg Chemical/goc = fig Chemical/gDRY WT + (% TOC 4- 100)
= fig Chemical/gDRYWT • 100 * % TOC
For example, a freshwater sediment with a concentration of 6.00 fig chemical/gDRYWT
and 0.5 % TOC has an organic carbon-normalized concentration of 1,200 fig/goc (6.00 ftg/gDRrwT
5-3
-------
• 100 H- 0.5 = 1,200 /tg/goc) which exceeds the SQC of 620 /ig/goc- Another freshwater
i
sediment with the same concentration of fiuoranthene (6.00 Mg/gnRvwr) but a TOC concentration
of 5.0% would have an organic carbon normalized concentration of 120 /ng/gbc (6.00 /tg/gDRY
vr • 100 -s- 5.0 = 120 Atg/goc), which is below the SQC for fluoranthene.
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 (SQCDRYWr.). This "conversion" must be done for each level of TOC of interest:
SQCDRYWT = SQCoc Otg/goc) • (% TOC -s- 100)
where SQCDRYVrr is the dry weight normalized SQC value. For example, the SQC value for
freshwater sediments with 1 % organic carbon is 6.2 /tg/g:
SQCDRYWT. = 620 jig/a*, • 1% TOC * 100 = 6.2 /*g/gDRYWT
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 fluoranthene 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 (^g/gocD can be estimated from the product of the effects concentration
from water only exposures G-ig/L) and the partition coefficient KQC (L/kg). The uncertainty
associated with the SQC can be obtained from a quantitative estimate of the degree to which the
available data support these assertions.
5-4
-------
The data used in the uncertainty analysis are from the water-only and sediment toxicity tests
;
that have been conducted to fulfill the minimum database requirements for the development of
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
fluoranthene are summarized in Appendix B. LC50s 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; /tig/L) are related to the
organic carbon-normalized LCSOs from sediment exposures (LC50s>0c; /*g/goc) via the
partitioning equation:
T f^n — TT T /"^n /c o\
i-A^JVg QC — X^Q^J-A^JU^r 1j—j 1
The EqP model asserts that the toxicity of sediments expressed on an organic carbon basis equals
the toxicity in water tests multiplied by the KQC. Therefore, both LC50S(OC and KoC*LC50w
are estimates of the true LC50OC for each chemical-organism pair. In this analysis, the
5-5
-------
uncertainty of KQC is not treated separately. Any error associated with KQC will be reflected in
: - i
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 or 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
o; represents the uncertainty due to the approximations inherent in the model and G represents
the experimental error. Let (erj2 and (o-e)2 be the variances of these random variables. Let i
index a specific chemical-organism pair. Let j index the exposure media, water-only, or the
individual sediments. Let k index the replication of the experiment. Then the equation that
describes this relationship:
ln(LC50i>i)fc) = ft + ay + €ij)k (5-4)
where ln(LC50)iJifc, are either InCLCSOw) or ln(LC50s,oc) corresponding to a water-only or
sediment exposure, and 0, are the population of ln(LC50) for the chemical-organism pair i. The
error structure is assumed to be lognormal which corresponds to assuming that the errors are
proportional to the means, e.g. 20%, rather than absolute quantities, e.g. 1 /ig/goc. The
statistical problem is to estimate n-a (oj2, and (
-------
TABLE 5-2: ANALYSIS OF VARIANCE FOR DERIVATION OF
SEDIMENT QUALITY CRITERIA CONFIDENCE LIMITS FOR
FLUORANTHENE.
Source of Uncertainty
Exposure media
Replication
Sediment Quality Criteria
Parameter Value
0*g/goc)
aa 0.39
ae 0.21
aSQc a 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 the
95% confidence interval limits, the significance level is 1.96 for normally distributed errors.
Hence:
= ln(SQCoc) + 1.96
-------
TABLE 5-3. SEDIMENT QUALITY CRITERIA
CONFIDENCE LIMITS FOR FLUORANTHENE
Sediment Quality Criteria
95% Confidence Limits
Type of SQCOC
Water Body A*g/goc Lower Upper
Freshwater 620 290 1300
Saltwater 300 140 640
5.3 COMPARISON OF FLUORANTHENE SQC CONCENTRATIONS TO SEDIMENT
CONCENTRATIONS THAT ARE TOXIC OR PREDICTED TO BE CHRONICALLY
ACCEPTABLE:
Insight into the magnitude of protection afforded to benthic species by SQC
concentrations and 95% confidence intervals can be inferred using effect concentrations from
toxicity tests with benthic species exposed to sediments spiked with fluoranthene and sediment
concentrations predicted to be chronically safe to organisms tested in water-only exposures
(Figures 5-1 and 5-2). This is because effect concentrations in sediments can be predicted from
water-only toxicity data and KQC values (See Section 4). Chronically acceptable concentrations
are extrapolated from genus mean acute values (GMAV) from water-only, 96-hour lethality tests
using acute-chronic ratios (ACR). Therefore, it may be reasonable to combine these two
predictive procedures to estimate, for fluoranthene, chronically acceptable sediment
concentrations (Predicted Genus Mean Chronic Value, PGMCV) from GMAVs (Appendix A),
ACRs (Table 3-2) and the KQC (Table 5-1):
PGMCV = (GMAV + ACR)«Koc. (5-7)
In Figures 5-1 and 5-2 each PGMCV for fishes, arthropods or other invertebrates tested
5-8
-------
O
s
1
N^
UJ
O
O
<
UJ
to
O
LU
Q
UJ
106
105
104
0 103
102
Water-only tests: (96HR LC5Q -*• ACR) KQC
A Arthropods
D Other Invertebrates
O Fishes
ACR = 5.45
Sediment Tests: 10 d LC50
if Q. tentgns = 1235 ng/goc
range 3 tests = 682 to 1 740
® H.gztecg = 974(ig/goc
range 3 tests = 500 to 1 480
53 I. punctatus = 526 ng/goc (4d)
T T
O O
T t
t
A
T
O
• • upper:
-SQC:
lower: 290(ig/goc
20
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 fluoianthene-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)!^. 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.
5-9
-------
o
I
UJ
O
O
G:
31
1
CO
LLJ
CD
Q
o
Q
UJ
105
,04
102|
Water-only tests: (96HR LC50•*- ACR) KQC
A Arthropods
D Other Invertebrates
O Rshes
log, nKoc^ 5.00 •
ACR = 5.45
Sediment Tests: 10 d LC50
•^ C.. splnlcorne « 2830 ng/gog
® R. qbronlus = 2960 ng/g^ .
range 8 tests = 1890 to 5620
upper: 640^^00
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 fluoranthene-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 -J- 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 ranee of
LC50s.
5-10
-------
in water is plotted against the percentage rank of its sensitivity. Results from toxicity tests with
;
benthic organisms exposed to sediments spiked with fluoranthene (Table 4-1) are placed in the
PGMCV rank appropriate to the test-specific effect concentration. For example, the 10-day
JLC50 for H. azteca (974 /ig/goc) is placed between the PGMCV of 660 jtg/goc for the catfish
(Ictalurus) and the PGMCV of 1,121 /Jg/goc for the amphipod (Gammarus). Therefore,
sediment test 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 fluoranthene 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 fluoranthene.
The freshwater SQC for fluoranthene (620 /*g/g0c) is less than any of the PGMCVs and
all but one of the LC50 values from spiked sediment toxicity tests. The PGMCVs for 8 of 12
freshwater genera are greater than the upper 95 % confidence interval of the SQC (1300 ftg/goc)-
The PGMCV for the catfish Ictalurus (660 pg/goc), the amphipod Gammarus (1,121 /tg/goc),
the cladoceran Daphnia (1,247 /tg/goc) and the hydrpid Hydra (1,285 /ig/goc) are below the SQC
upper 95% confidence interval. This illustrates why the slope of the species sensitivity
distribution is important. It also suggests that if the extrapolation from water only acute lethality
tests to chronically acceptable sediment concentrations is accurate, these or similarly sensitive
genera may be chronically impacted by sediment concentrations marginally above the SQC and
5-11
-------
possibly less than the 95% upper confidence interval. For fluoranthene, the PGMCVs of
freshwater organisms range over less than one order of magnitude from the most sensitive to the
most tolerant genus. However, many of the LC50 values for the less sensitive species are
"greater than" values, presumably because actual LC50 values would exceed saturation. Chronic
effect concentrations may, however, occur at concentrations below saturation. A sediment
concentration five times the SQC would include the PGMCVs of five of the eight benthic genera
tested including stoneflies, amphipods, hydroids, catfish, and snails. Tolerant benthic genera
such as the annelid Lumbriculus. and the dragonfly Ophiogomphus might be expected to not be
chronically impacted in sediments with fluoranthene concentrations five times the SQC. We
speculate that sediment concentrations far in excess of this may be unlikely to chronically impact
benthic genera that tolerate chronic water-only exposures up to fluoranthene's water solubility.
The saltwater SQC for flouranthene (300 ftg/goc) and the upper confidence limits are less
than any of the PGMCVs for saltwater genera. For fluoranthene, PGMCVs from the most
sensitive to the most tolerant saltwater genus range over two orders of magnitude. As with the
freshwater data, many of the values for less sensitive species are "greater than" values dictated
by fluoranthene's water solubility. A sediment concentration five times the SQC would include
the PGMCVs of all three benthic arthropod genera tested. Less sensitive benthic genera include
molluscs, polychaetes and fish, some of which might not be expected to be chronically impacted
in sediments with fluoranthene concentrations _>. 1,OOOX the SQC.
The above extrapolation using the PGMCV approach for fluoranthene may be reasonable
given (1) the accuracy of the equilibrium partioning prediction using KQC and water-only LC50
values as demonstrated by the ratio of the actual and predicted LC50 value (Predicted Sediment
5-12
-------
Toxic Units, PSTU; Table 4-2) and (2) the measured toxicity of fluoranthene-spiked sediments
i
to benthic taxa when compared to PGMCV predictions of their relative sensitivities. If water
only LC50 and KQC values perfectly predict actual sediment toxicity then the ratio of the actual
to the predicted LC50 values (PSTU) would be 1.0. PSTU values for fluoranthene averaged
0.217 for Hvalella azteca. 0.387 for C. tentans. 0.146 for I. punctatus and 2.13 for Rhepoxynius
abronius; overall mean 0.80. Ten-day LC50 values for H. azteca and C. tentans were greater
than the SQC (620 MS/goc) but less than the upper 95%- confidence limit (1,300 jig/goc)- The
measured LC50 for I. punctatus (526 /Kg/goc) and the PGMCV (660 /*g/g0c) for catfish are the
lowest values on Figure 5-1 as might be expected, because this catfish was the most acutely
sensitive species tested in water-only exposures (Figure 3-1). However, the fact that the
PGMCV, which estimates the safe concentration is less than the 96 hour LC50 value for this fish
suggests that one or both of these values may be suspect.
5.4 COMPARISON OF FLUORANTHENE SQC TO STORET AND NATIONAL STATUS
AND TRENDS DATA FOR SEDIMENT FLUORANTHENE:
A STORET (U.S. EPA, 1989b) data retrieval was performed to obtain a preliminary
assessment of the concentrations of fluoranthene in the sediments of the nation's water bodies.
Log probability plots of fluoranthene concentrations on a dry weight basis in sediments are
shown in Figure 5-3. Fluoranthene is found at varying concentrations in sediments from rivers,
lakes and near coastal water bodies in the United States. Median concentrations are between 0.1
/tg/g to 0.3 /Kg/g in the three water bodies. There is significant variability with fluoranthene
concentrations in sediments ranging over seven orders of magnitude within the country.
The SQC for fluoranthene can be compared to existing concentrations of fluoranthene in
5-13
-------
w
,
o
Figure 5-3.
•10 3 = STREAM ,
i"
« 2L
10 1 ^_
10 °-r '
10 ~1i-
10 "Z~r
10 "a5
- r *u
10 ~*l. <«««**
mr =
1Q "«1_i , •.....!
ai 1
10 4 __,_,„,,„„, ,
10 ^S-L^
!
10 2 ^
10 1 k-
» °j.
« -';.
10 ^r
i
10 -«r
10 ^F
0.1 1
10 4g . .,mll| , ,
w 3 = ESTUARY
10 '••„ ..
10 !•• -
« °r
10 "1i-
10 *-T
10 ^r < «
10 ^J.
K) "C~ I 1 Illllll 1 L
0.1 1
TOTAL SAMPLES: 786
MEASURED SAMPLES: 254
^<«««Jic<^<«^
.-/
10 20 60
TOTAL SAMPLES: 57*
MEASURED SAMPLES: 28
*<«*^
^gp
jl^*
c < ««< «
I ' I ml 1 .il..
10 20 60
TOTAL SAMPLES: 88 '
MEASURED SAMPLES: 32
^^t«««
^^^^^^
^^
«« «««
10 20 60
•* — r Turn i r T- -^|iTn 1 1 i i •
e
® 1
-~^*-1
" ^**^~~ -
0^1 1
i
i
i
1 I..,,,,, , I,,,,,, , . =
80 90 99 99.9
1
V -• m ~*
m++i* ^
-
1
1
'BO 90 99 99.9
I TH 1 1 1 i i — i — limn i i |
jj
T
_-.•
:««<^*^" 1
T
r
i ||,, , | , , , i,,.,.. , , =
80 90 99 99.9
PROBABILITY
Probability distribution of concentrations of fluoranthene 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 fluoranthene SQC values
of 62 jig/g in freshwater sediments having TOC = 10% and 6.2 /tg/g in
freshwater sediments having TOC = 1% and compared to SQC values for
saltwater sediments of 30 /*g/g when TOC = 10 % and 3.0 /*g/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%.
5-14
-------
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% are compared to fluoranthene's distribution in sediments as examples
only. For fresh water sediments, SQC values are 6.2 /*g/g dry weight in sediments having 1 %
organic carbon and 62 pg/g dry weight in sediments having 10% organic carbon; for marine
sediments SQC are 3.0 /*g/g dry weight and 30 /ig/g dry weight, respectively. Figure 5-3
presents the comparisons of these SQC to probability distributions of observed sediment
fluoranthene levels for streams and lakes (fresh water systems, shown on the upper panels) and
estuaries (marine systems, lower panel). For streams (n = 786) the SQC of 6.2 jcg/g dry weight
for 1 % organic carbon fresh water sediments is exceeded for 2% of the data and the SQC of 62
jttg/g dry weight, for fresh water sediments having 10% TOG is exceeded by less than 1% of
the data. For lakes (n = 57) the SQC for 1 % organic carbon fresh water sediments is exceeded
by about 5% of the data, but the SQC for 10% organic carbon fresh water sediments is not
exceeded by any of the sample data. In estuaries, the data (n = 88) indicate that the criteria of
3.0 jtg/g dry weight for salt water sediments having 1 % organic carbon is exceeded by less than
2% of the data and the criteria of 30 pg/g dry weight for salt water sediments having 10%
organic carbon is not exceeded by the post 1986 samples.
The fluoranthene distribution in Figure 5-3 includes data from some samples in which
the fluoranthene 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
5-15
-------
SQC values are actually exceeded may be less than the percentage reported.
i
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 fluoranthene (300 /tg/goc) on Figure 5-4. Data presented are from sediments
with 0.2 to 16.2 percent organic carbon. The median organic carbon normalized fluoranthene
concentration (about 7.0 jig/goc) is two orders of magnitude below the SQC of 300 Atg/goc- Less
than 1% of the measured values (n=797) are greater than the SQC for fluoranthene. Hence,
these results are consistent with the preceding comparison of the marine SQC to STORET data.
Regional differences in fluoranthene 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 fluoranthene sediment
concentrations and the extent to which they may exceed the SQC.
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
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
5-16
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all of the contaminants in a particular sediment, there might be additive or synergistic effects
t ;
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 fluoranthene 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 jig/goc, salt: 300 jig/goc), acenaphthene (fresh: 130 /ng/goc, salt: 230 A*g/goc) and
phenanthrene (fresh: 180 fig/goo salt: 240 ftg/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, antagonism has rarely been demonstrated. What should be much more common are
instances where toxic effects occur at concentrations below the SQC because of the additivity
of toxicity of many common contaminants (Alabaster and Lloyd, 1982), e.g; heavy metals and
PAHs, and instances where other toxic compounds for which no SQC exist occur along with
5-18
-------
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
disequilibrium discussed above, site-specific methodologies may be applied (U.S. EPA, 1993b).
Finally, it should be remembered that in some situations the phototoxicty of fluoranthene
5-19
-------
may be important. The current SQC for fluoranthene does not take phototoxic effects into
account.
5-20
-------
SECTION 6
CRITERIA
The procedures described in the "Technical Basis for Deriving Sediment Quality Criteria
for Nonionic Organic Contaminants for the Protection of Benthic Organisms by Using
Equilibrium Partitioning" (U.S. EPA, 1993a) indicate that benthic organisms should be
acceptably protected in freshwater sediments containing <_ 620 /tg fluoranthene/g organic carbon
and saltwater sediments containing <_ 300 /tg fluoranthene/g organic carbon, except possibly
where a locally important species is very sensitive or sediment organic carbon is < 0.2%.
Confidence limits of 290 to 1300 jig/goc for freshwater sediments and 140 to 640 ftg/goc
for saltwater sediments are provided as an estimate of the uncertainty associated with the degree
to which the observed concentration in sediment (/tg/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 fluoranthene, 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 fluoranthene in sediments that would be protective of benthic species. It is the philosophy of
the Agency and the EPA Science Advisory Board that the use of sediment quality criteria (SQCs)
as stand-alone, pass-fail criteria is not recommended for all applications and should frequently
6-1
-------
trigger additional studies at sites under investigation. The upper confidence limit should be
.*
interpreted as a concentration above which impacts on benthic species should be expected.
Conversely, the lower confidence limit should be interpreted as a concentration below which
impacts on benthic species should be unlikely.
6-2
-------
SECTION 7
REFERENCES
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Alabaster, J.S. and R. Lloyd. 1982. Water Quality Criteria for freshwater fish. Chapter 11.
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.AUred, P.M and J.P. Giesy. 1985. Solar radiation-induced toxicity of anthracene to Daphnia
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Birge, W.J., J.A. Black, S.T. Ballard and W.E. McDonnell. 1982. Acute toxicity testing with
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Bowling, J.W., G.J. Leversee, P.P. Landrum and J.P. Giesy. 1983. Acute mortality of
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Brooke, L. 1992. Letter to Walter J. Berry documenting result of Daphnia magna chronic test
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7-1
-------
1979. Water-related environmental fate of 129 priority pollutants. Volume n:
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(UV lights). Memomandum to Brian Melzian. September 11. 9pp.
Champlin, D.M. and S. L. Poucher. 1992b. Flow-through chronic toxicity of fluoranthene
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7-2
-------
Environmental Toxiciology and Chemistry. 10:(12)1541-1583.
i '
Eadie, B.J., P.P. Landrum, W. Faust. 1982. Polycyclic aromatic hydrocarbons hi sediments,
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EG&G, Bionomics. 1978. Preliminary Research Report. Acute and chronic toxicity of
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Gendusa, A.C. 1990. Toxicity of chromium and fluoranthene from aqueous and sediment
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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
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and D. Di Toro. June 18, 1991. 9 pp.
Home, J,D. and B.R. Oblad. 1983. Aquatic toxicity studies of six priority pollutants. Final
Report Task H. U.S. EPA. Contract No. 68-01-6201.
Kagan, J., E.D. Kagan, I.S. Kagan, P.A. Kagan, and S. Quigley. 1985. The phototoxicity
of Non-carcinogenic polycyclic aromatic hydrocarbons in aquatic organisms."
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Kagan, J., A. Stokes, H. Gong, and R.W. Tuveson. 1987. Light-dependent cytotoxicity of
fluoranthene: Oxygen-dependent membrane damage. Chemosphere 16(10-12):2417-2422.
Karickhoff, S.W., L.A. Carreira, C. Melton, V.K. McDaniel, A.N. Vellino, and D.E. Nute.
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Landrum, P.P., S.R. Nihart, B.J. Eadie and W.S. Gardner. 1984. Reverse-phase separation
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carbon in natural waters. Environ. Sci. Technol. 18:187-192.
7-3
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Landrum, P.P., S.R. Nihart, B.J. Eadie, and L.R. Herche. 1987. Reduction in bioavailability
of organic contaminants to the amphipod Pontoporeia hoyi by dissolved organic matter
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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,
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.
Mackay, D. and B. Powers. 1987. Sorption of Hydrophobic Chemicals From Water: A
Hypothesis for the Mechanism of the Particle Concentration Effect. Chemosphere
16(4):745-747.
National Academy of Sciences/National Academy of Engineering. 1973. Water Quality Criteria
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Nebeker, A.V., G.S. Schuytema, W.L. Griffis, J.A. Barbitta, and L.A. Carey. 1989. Effect
of sediment organic carbon on survival of Hyalella azteca exposed to DDT and endrin.
Environ. Toxicol. Chem. 8(8):705-718.
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Rossi, S.S. and J.M. Neff. 1978. Toxicity of polynuclear aromatic hydrocarbons to the
polycheate Neanthes arenaceodentata. Mar. Pollut. bull. 9:220-223.
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Schuytema, G.A., A.V. Nebeker, W.L. Griffis, and C.E. MiUer. 1989. Effects of freezing
on toxicity of sediments contaminated with DDT and endrin. Environ. Toxicol. and
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Service, Springfield, VA. 98 pp.
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to a marine infaunal amphipod: Cadmium and its interaction with sewage sludge. Mar.
Envir. Res. 18:133-153.
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1080.
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Poucher. December 12, 1991. 25pp.
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86-001. May 1, 1987. U.S. Government Printing Office No. 955-002-000008. 406pp.
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Yalkowsky, S.H., S.C. Valvani, and D. Mackay. 1983. Estimation of the aqueous solubility
of some aromatic compounds. Residue Rev. 85:43-55.
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