£EPA
United States Office of Science and Technology EPA xxx/x-xx-xxx
Environmental Protection Agency Health and Ecological Criteria Oiv. November 1991
Office of Water & Washington. DC 20460
Office or Research and Development
WATER
Proposed Sediment Quality
*
Criteria for the Protection
of Benthic Organisms:
FLUORANTHENE
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CONTENTS
EASE
Foreword ii
Acknowledgments iii
Tables v
Figures vi
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: Evaluation of Octanol-Water Partition Coefficient for Flupranthene. . B-l
Appendix C: Summary of Data from Sediment Spiking Experiments with
Fluoranthene . C-l
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FOREWORD
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. 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. Section 304(a)(l) directs the
Administrator to develop-and publish "criteria" reflecting the latest scientific knowledge on the
kind and extent of effects on plankton, fish, shellfish, and wildlife which may be expected from
the presence of pollutants in any body of water, including ground water, the concentration and
dispersal of pollutants, or their byproducts, through biological, physical and chemical processes;
and the effects of pollutants on biological community diversity, productivity, and stability.
Section 304(a)(2) directs the administrator to develop and publish information on the factors
necessary for the protection and propagation of shellfish, fish, and wildlife for classes and
categories of receiving waters.
To meet this objective, in 1980 EPA published ambient water quality criteria
(WQQ for 64 of the 65 toxic pollutants or pollutant categories designated as toxic under
Section 307(a)(l) of the CWA. Additional water quality documents that update criteria for
selected consent decree chemicals and new criteria have also been published since 1980. In
addition to the development of water quality criteria and to continue to comply with the mandate
of the CWA, EPA has conducted efforts to develop and publish sediment quality criteria for
some of the 65 toxic pollutants or toxic pollutant categories.
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 comply with established water quality criteria. In addition,
contaminated sediments can lead to water quality degradation, even when pollutant sources are
stopped. It is intended that sediment quality criteria 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. Environmental Protection
Agency's best recommendation of the concentrations of a substance in sediment that will not
unacceptably affect benthic organisms. 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 or human health.
The criteria and derivation methods outlined in this document are proposed to
provide protection of benthic organisms from biological impacts from chemicals associated with
sediments. Recommendations on the use of these criteria will follow completion of the public
response process. In the interim, until final Sediment Quality Criteria are promulgated, these
criteria should only be used to support site specific assessments for sediments that are consistent
with assumptions of equilibrium partitioning theory on which these Sediment Quality Criteria
are developed.
Guidelines and guidance have been developed by 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.
11
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Principal Author
David J. Hansen
Coauthors
Walter J. Berry
Dominic M. Di Toro
Paul Paquin
Laurie Davanzo
Frank E. Stancil, Jr.
Heinz P. Kollig
Technical and Clerical Support
Glen Thursby
Charito Paruta
Betty Anne Calise
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. Environmental Research Laboratory,
Athens, GA
U.S. Environmental Research Laboratory,
Athens, GA
Persons who have made significant contributions to the development of the approach and
supporting science used in the derivation of sediment criteria for non-ionic organic contaminants
are as follows:
Herbert E. Allen
Gerald Ankley
Christina E. Cowan
Dominic M. Di Toro
David J. Hansen
PaulR. Paquin
Spyros P. Pavlou
University of Delaware, Newark, DE
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
Battelle, Richland, WA
HydroQual, Inc., Mahwah, NJ;
Manhattan College, BlDDX, NY
U.S. EPA, Environmental Research Laboratory,
Narragansett, RI
HydroQual, Inc., Mahwah, NJ
Ebasco Environmental, Bellevue, WA
iii
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Richard C. Swaitz
Nelson A. Thomas
Christopher S. Zarba
U.S. EPA, Environmental Reserach Laboratory,
Newport.OR
U.S. EPA, Environmental Research Laboratory,
Duluth, MN
U.S. EPA Headquarters, Office of Water, Washington, DC
IV
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Table 2-1. Summary of measured and estimated KOV values for fluoranthene by the U.S.
EPA, Environmental Research Laboratory, Athens, GA.
Table 3-1. Acute sensitivity of freshwater and saltwater benthic species to fluoranthene.
Table 3-2. Chronic sensitivity of freshwater and saltwater organisms to fluoranthene.
Test specific data.
Table 3-3. Summary of acute and chronic values, acute-chronic ratios and freshwater and
saltwater final acute values, final acute-chronic ratios, and final chronic values
for fluoranthene.
Table 3-4. Kolmogorov-Smirnov test for the equality of freshwater and saltwater LC50
distributions for fluoranthene. Kplmogorov-Smirnov test for the equality of
benthic and water column LC50 distributions.
Table 4-1. Summary of tests with fluoranthene-spiked sediment.
Table 4-2. Water-only and sediment LCSOs used to test the applicability of the equilibrium
partitioning theory for fluoranthene.
Table 5-1. Sediment quality criteria for fluoranthene.
Table 5-2. Analysis of variance for derivation of sediment quality criteria confidence limits
for fluoranthene.
Table 5-3. Sediment quality criteria confidence limits for fluoranthene.
APPENDIX
Appendix A. - Fluoranthene: Summary of acute values for freshwater and saltwater species.
Appendix B. - The octanol-water partition coefficient, KW for fluoranthene.
Appendix C. - Summary of data from sediment spiking experiments with fluoranthene that
were used to calculate KQC values (Figure 2-2) and to compare mortalities of
amphipods with interstitial water toxic units (Figure 4-1) and predicted
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 KOC (bottom) from sediment toxicity tests conducted by Swartz
(1991).
Figure 3-1. Acute toxicity of fluoranthene to mysids (Mysidopsis b_abja) as a function of
intensity of UVA exposure.
Figure 3-2. Comparison of fluoranthene water only LCSO probability distributions for
freshwater (0) and saltwater (*) species (top panel). Cumulative distribution
functions for calculating the K-S statistic (bottom panel).
Figure 3-3. Comparison of fluoranthene water only LCSO probability distributions for water
column (0) and benthic (*) freshwater and saltwater species (top two panels).
Cumulative distribution functions for calculating the K-S statistic (bottom two
panels).
Figure 4-1. Percent mortality of amphipods in sediments spiked with acenaphthene or
phenanthrene (Swartz, 1991), cadmium (Swartz et al., 1985), endrin (Nebeker
et al., 1989; Schuytema et al., 1989), or fluoranthene (Swartz et al., 1990), and
midge in kepone-spiked sediments (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 LCSO value from
water-only tests. (See Appendix C in this SQC document, Appendix C in the
endrin, dieldrin, acenaphthene, and phenanthrene SQC documents, and original
references for raw data.)
Figure 4-2. Percent mortality of amphipods in sediments spiked with acenaphthene or
phenanthrene (Swartz, 1991), dieldrin (Hoke and Ankley, 1991), endrin (Nebeker
et al., 1989; Schuytema et al., 1989) or fluoranthene (Swartz et al., 1990)
relative to "predicted sediment toxic units." Predicted sediment toxic units are
' the ratios of measured treatment concentrations for each chemical in sediments
Otg/goc) divided by the predicted LCSO 0*g/goc) in sediments (Koc * Water-only
LCSO, pg/L). (See Appendix C in this document and Appendix C in the dieldrin,
endrin, acenaphthene, and phenanthrene SQC documents for raw data).
Figure 5-1. Probability distribution of concentrations of fluoranthene in sediments from
streams (n=786), lakes (n=57) and estuaries (n=88) in the United States from
1986 to 1990, from the STORET (U.S. EPA, 1989c) database, compared to the
fluoranthene SQC values of 102 pg/g in freshwater sediments having TOC =
10% and 10.2 /xg/g in freshwater sediments having TOC = 1%; SQC values for
saltwater sediments are 134 pg/g when TOC =10% and 13.4 pg/g when
TOC=1%. The upper dashed line on each figure represents the SQC value when
TOC - 10%, the lower dashed line represents the SQC when TOC = 1%.
Figure 5-2. 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 SQC value of
1340 Mg/goc-
vi
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DISCLAIMER
This report has been reviewed by the Health and Ecological Criteria Division, Office
of Science and Technology, U.S. Environmental Protection Agency, and approved for
publication. Mention of trade names or commercial products does not constitute endorsement
or recommendation for use.
AyAyr.ARTT.rrY NOTICE
This document is available to the public through the National Technical Information
Service (NTTS), 5285 Port Royal Road, Springfield, VA 22161. NTIS Accession Number
xxxx-xxxxxx.
vu
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SECTION 1
INTRODUCTION
1.1 GENERAL INFORMATION:
Under the Clean Water Act (CWA) the U.S. Environmental Protection Agency (U.S.
EPA) is responsible for protecting the chemical, physical and biological integrity of the nation's
waters. In keeping with this responsibility, U.S. EPA published ambient water quality criteria
(WQC) in 1980 for 64 of the 65 toxic pollutants or pollutant categories designated as toxic in
the CWA, Additional water quality documents that update criteria for selected consent decree
chemicals and new criteria have been published since 1980. These water quality criteria 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 water quality criteria
play an important role in assuring a healthy aquatic environment, they alone are not sufficient
to ensure the protection of environmental or human health.
Toxic pollutants in bottom sediments of the nation's lakes, rivers, wetlands, estuaries
and marine coastal waters create the potential for continued environmental degradation even
where water-column concentrations comply with established water quality criteria. 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 (SQQ makes it difficult to accurately assess the extent of
the ecological risks of contaminated sediments and to identify, prioritize and implement
appropriate clean up activities 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).
1-1
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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 non-ionic organic chemicals because it presented
the greatest promise for generating defensible national numerical chemical-specific sediment
quality criteria applicable across a broad range of sediment types. The three principal
observations that underlie the EqP method of establishing sediment quality criteria are:
1. The concentrations of non-ionic 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 non-ionic organic
chemicals on an organic carbon basis to pore water concentrations.
3. The distribution of sensitivities of benthic and water column organisms to
chemicals are similar, thus, the currently established water quality criteria final
chronic values (FCV) can be used to define the acceptable effects concentration
of a chemical freely-dissolved in pore water.
The EqP approach, therefore, assumes that: (1) the partitioning of the chemical between
sediment organic carbon and interstitial water is at equilibrium; (2) the concentration in either
phase can be predicted using appropriate partition coefficients and the measured concentration
in the other phase; (3) organisms receive equivalent exposure from water-only exposures or
from any equilibrated phase: either from pore water via respiration, from sediment via ingestion.
or from a mixture of both exposure routes; (4) for non-ionic chemicals, effect concentrations;
in sediments on an organic carbon basis can be predicted using the organic carbon partition
coefficient (Koe) and effects concentrations in water, (5) the FCV concentration is art
appropriate effects concentration for freely-dissolved chemical in interstitial water, and (6) the
SQC (/ig/goc) derived as the product of the KOC and FCV is protective of benthic organisms,
Sediment quality criteria concentrations presented in this document are expressed as n\\
1-2
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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, interstitial concentrations may
overestimate exposure.
The data that support the equilibrium partitioning approach for deriving SQC for non-
ionic organic chemicals are reviewed by Di Toro et al. (1991) and in the sediment quality
criteria guidelines (U.S. EPA, 1992a). Data supporting these observations for fluoranthene are
presented in this document.
Sediment quality criteria generated using the equilibrium partitioning method are suitable
for use in providing guidance to regulatory agencies because they are:
1. numerical values;
2. chemical specific;
3. applicable to most sediments;
4. predictive of biological effects; and
5. protective of benthic organisms.
As is the case with water quality criteria, the sediment quality criteria reflect the use of
available scientific data to: (1) assess the likelihood of significant environmental effects to
benthic organisms from chemicals in sediments; and (2) to derive regulatory requirements which
will protect against these effects.
It should be emphasized that these criteria are intended to protect benthic organisms from
the effects of chemicals associated with sediments. SQC are intended to apply to sediments
permanently inundated with water, intertidal sediment and to sediments inundated periodically
for durations sufficient to permit development of benthic assemblages. They do mot apply to
occassionally inundated soils containing terrestrial organisms. In spills, where chemical
equilibrium between water and sediments has not yet been reached, sediment chemical
concentrations in excess of SQC indicate benthic organisms may be at risk. This is because for
1-3
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spills, disequilibrium concentrations in interstitial and overlying water may be proportionally
higher relative to sediment concentrations. In spills, sediments having concentrations less than
SQC may also pose risks to benthic 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. Environmental Protection Agency's best recommendation at this time of the
concentration of a chemical in sediment that will not unacceptably affect benthic organisms.
SQC values may be adjusted to account for future data or site specific considerations.
This document presents the theoretical basis and the supporting data relevant to the
derivation of the sediment quality criterion 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 "Guidelines for Deriving Numerical National Sediment Quality Criteria for
Non-ionic Organic Chemicals for the Protection of Benthic Organisms" (U.S. EPA, 1992a) is
necessary in order to understand the following text, tables and calculations. Guidance into the
acceptable use of SQC values is contained in U.S. EPA, 1992b.
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). Some PAHs are of environmental concern because they are known to be
carcinogens and/or mutagens (Brookes, 1977). With an increase in fossil fuel consumption in
the United States, an increase in emissions of PAHs to the environment can be expected over
the next several decades (Eadie et al., 1982).
1-4
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Fluoranthene has a three ring structure and exists as pale yellow plates or needles
(Figure 1-1). It has a solubility in water at 25°C of 0.26 mg/L 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).
The acute toxicity of fluoranthene ranges from 32 to > 4,000 /tg/L for freshwater and 52
to > 560,000 Mg/L for saltwater organisms (Appendix A). Differences between fluoranthene
concentrations causing acute lethality and chronic toxicity in invertebrates are small; acute-
chronic ratios range from 1.745 to 6,325 for two species. The only available acute-chronic
ratio for rainbow trout is 59 (Table 3-3). Although fluoranthene bioaccumulates in aquatic
biota, the associated health or ecological risks are unknown.
1.3 OVERVIEW OF DOCUMENT:
The goal of this document is to provide a brief review of the equilibrium partitioning
methodology, a summary of the physical-chemical properties and aquatic toxicity of
fluoranthene, and the tarhnirai basis for setting the SQC for fluoranthene. Section 2 reviews
a variety of methods and data useful in deriving partition coefficients for fluoranthene and
includes the K^ recommended for use in the derivation of the fluoranthene SQC. Section
reviews aquatic toxicity data contained in the fluoranthene WQC document (U.S. EPA, 1980)
and new data that were used to derive the Final Chronic Value (FCV) used in this document
to derive the SQC concentration. In addition, the comparative sensitivity of benthic and water
1-5
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MOLECULAR FORMULA
MOLECULAR WEIGHT
DENSITY
MELTING POINT
PHYSICAL FORM
VAPOR PRESSURE
202.26
1.252g/cc(25°C)
lll'C
Pale yellow plates or needles
CAS NUMBER:
TSL NUMBER
CHEMICAL NAME:
206-44-0
LL 40250
Fluoranthene (Idryl,
Benzo{j,k]fluorene)
FIGURE 1-1. Chemical structure and physical-chemical properties of fluoranthene.
1-6
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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 KOC
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SECTION 2
PARTITIONING
2.1 DESCRIPTION OF THE EQUILIBRIUM PARTITIONING METHODOLOGY:
Sediment quality criteria 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; 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 jig
chemical/g organic carbon and apply to sediments having £ 0.2% organic carbon by dry
weight. A brief overview follows of the concepts which underlie the equilibrium partitioning
methodology for deriving sediment quality criteria. The methodology is discussed in detail in
the "Guidelines for Deriving Numerical National Sediment Quality Criteria for Non-ionic
Organic Chemicals for Protection of Benthic Organisms" (U.S. EPA, 1992a), hereafter referred
to as the SQC Guidelines.
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 sediment quality
criterion based on the bioavailable chemical fraction in a sediment. For non-ionic 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 (pg chemical/liter
pore water) and not to the sediment chemical concentration 0*g chemical/g sediment) (Di Toro
et al., 1991). From a purely practical point of view, this correlation suggests that if it were
possible to measure the 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
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chemicals between the solid and liquid phases in a sediment is a necessary component for
establishing sediment quality criteria. It is for this reason that the methodology described
below is called the equilibrium partitioning (EqP) method.
It is shown in the SQC Guidelines (U.S. EPA, 1992a), that benthic and water column
species from freshwater and saltwater environments exhibit similar sensitivities to a wide range;
of chemicals. The data for fluoranthene are presented in Section 3. Thus, a sediment quality
criteria can be established using the final chronic value (FCV) derived using the Water Quality
Criteria Guidelines (Stephan et al., 1985) as the effect concentration, and the partition!
coefficient can be used to relate the pore water concentration (FCV) to the sediment quality'
criteria via the partitioning equation.
The calculation is as follows: Let FCV Gig/L) be the no effect concentration in water
for the chemical of interest; then the sediment quality criteria, SQC G*g/kg sediment), in
computed using the partition coefficient, K, (L/kg sediment), between sediment and water
SQC - K, FCV (2-1)
This is the fundamental equation used to generate the sediment quality criterion. Its utility
depends upon the existence of a methodology for quantifying the partition coefficient, K,.
For fluoranthene, and other hydrophobic non-ionic organic chemicals, the chemical
property of importance is the octanol-water partition coefficient, KQV. It is empirically related
to the partition coefficient via KOC (Equation 2-5), the organic carbon partition coefficient, and
foe, the weight fraction of organic carbon in the sediment (gbc/g sediment). Organic carbon
appears to be the predominant sorption phase for non-ionic organic chemicals in naturally
occurring sediments. The relationship is as follows:
K, = ^ KOC (2-2)
It follows that:
FCV (2-3)
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where SQCoc is the sediment quality criterion on a sediment organic carbon basis.
The next section reviews the available information for
2.2 DETERMINATION OF KOW FOR FLUORANTHENE:
Several approaches have been used to determine KOW for the derivation of a SQC, as
discussed in the SQC Guidelines. At the U.S. EPA, Environmental Research Laboratory at
Athens, GA (ERL,A) three methods were selected for measurement and two for estimation of
KOW The measurement methods were shake-centrifugation (SC), generator column (GCol) and
slow-stir-flask (SSF) and the estimation methods were SPARC and CLOGP (Appendix B). Data
were also extracted from the literature. The shake-centrifugation method is a standard
procedure in the Organization for Economic Cooperation and Development (OECD) guidelines
for testing chemicals, therefore it has regulatory precedence.
Examination of the literature data reveals that the logtoKow values for fluoranthene range
from 4.90 to 5.33. Only one primary reference was found for a measured value in the
literature. Although the range of reported values for fluoranthene is significantly lower than
the range of values for some other compounds, it is relatively large, and we were not able to
determine from studying the primary articles that any value was more likely to be accurate than
any other.
Preliminary experience with the SPARC program suggests that the program can compute
values for partition coefficients for high logP chemicals which may be more reliable and
accurate than laboratory measurements which may suffer from errors inherent in many
methodologies. However, the program needs more validation by comparison with laboratory-
measured data for a wide range of chemical structures before it is recommended as a basis for
computing partition coefficients for regulatory purposes. The SPARC estimated logtoKow value
for fluoranthene is 5.21. The CLOGP program estimate of the logtoKow value for fluoranthene
using structure activity relationships is 4.95.
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Three measurement methods provide additional data from which to define KQV for
fluoranthene ( Table 2-1; Appendix B). The shake-centrifugation method yielded logloKoV =
5.00, the generator column method yielded log10Kow = 5.39, and the slow-stir-flask method
yielded log10KoW = 5.09. There is no clear-cut best value from the data that has been
developed. Considering the agreement among the SPARC estimated value and the average of
the values measured by the three methods under carefully controlled conditions in the Athens
Laboratory, the recommended value for logtoKow is S. 19. This is the logarithm of the average
of three averages: the average of three KQW measurements made by shake-centrifugation, the
average of four KQW measurements made by generator column, and the average of six KOW
measurements made by slow-stir-flask. The logs of the Kow values measured by shake-
centrifugation measurements range from 4.99 to S.01. The logs of the KOW values measured
by the generator column range from 5.20 to 5.48. The logs of the KOW values measured by the
slow-stir-flask range from 4.98 to 5.23.
TABLE 2-1. SUMMARY OF MEASURED AND ESTIMATED K^ VALUES FOR
FLUORANTHENE BY THE U.S. EPA, ENVIRONMENTAL RESEARCH LABORATORY,
ATHENS, GA.
Measurement
Technique
Shake-
Centrifugation
Generator Column
Slow-stir-flask
SPARC
CLOGP
Number of
Analyses
3
4
6
-
-
Lo
Mean
5.00
5.39
5.09
5.21
4.95
e^K-.
CV
0.019
0.242
0.206
-
-
2.3 DERIVATION OF KOC FROM ADSORPTION STUDIES:
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Several types of experimental measurement of the organic carbon partition coefficient
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 sediment fluoranthene and pore water were used to compute
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
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 K,:
*oc **oc
(2-4)
1 + mfocKoc/
where:
m = particle concentration in the suspension (kg/L)
~l^x 1.4, an empirical constant (unitless).
The other quantities are defined previously. In this expression, the organic carbon partition
coefficient is given by:
logwKoc - 0.00028 + 0.983 log,^ (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.
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In the absence of particle effects, KOC is related to KOV via Equation 2-5, shown above.
For logtoKoW = 5.19 (ERL.A, mean measured value), this expression results in an estimate of
log^ = 5.10.
2.3.2 KOC FROM SEDIMENT TOXICITY TESTS:
Measurements of K^ are available from sediment toxicity tests using fluoranthenc:
(Swaitz et al., 1990). These tests represent freshwater sediments having a range of organic
carbon contents of 0.2 to 0.5 percent (Table 4-1; Appendix C). Fluoranthene concentration!;
were measured in the sediment and pore waters providing the data necessary to calculate the
partition coefficient for an undisturbed sediment.
Figure 2-1 is a plot of the organic carbon-normalized sorption isostherm for
fluoranthene, where the sediment fluoranthene concentration Gtg/goc) is plotted versus pore
water concentration 0*g/L). The data used to make this plot are included in Appendix C. The
line of unity slope corresponding to the logloKoc - 5.10 is compared to the data. The intercept
at a pore water concentration of 1 pg/L is equivalent to logloKoc-
A probability plot of the observed experimental logloKoc values is shown in Figure 2 -
1. The logtoKoc values are approximately normally distributed with a mean of log,,^ = 4.96
and a standard error of the mean of 0.041. This value is in agreement with logtoKoc = 5.10,
which was computed from the experimentally determined fluoranthene logtoKow of 5.19
(Equation 2-5). Complexation with pore water DOC has not been accounted for in the
experimentally based «^Hm^a of logtoKoc = 4.96. Though it is not expected to be a major
factor, consideration of DOC effect would increase the estimate of logtoKoc relative to the value
based on total pore water concentrations. If this uncorrected value was used to set SQC, the
SQC concentration would tend to be environmentally conservative.
2.4 SUMMARY OF DERIVATION OF K^ FOR FLUORANTHENB:
The KOC selected to calculate the sediment quality criteria for fluoranthene is based on
the regression of logloKoc to logtoKow (Equation 2-5). using the fluoranthene logloKow of 5.1.9
26
-------
~. 100000
O)
NS
0>
a:
UJ
o
u
H-
I
H
a
LU
en
looooL
1000
= (Log KOC 5-10)
10! 1 i i i HIM
too 1000 10000
PORE WATER CONCENTRATION (ug/LJ
8 3.8
O»
^
^J 9.0
8
* 4.0
CO
o
n <-°
»
oc
S 3.0
S
%^
3.0
A
~T'i rrimi rrrrmn
M
M
m
v
i i IIIIIH i i i linn
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i o o e
t
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0 9 9
1 1 1
inn n i i innrrrr-
~
0
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.
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IIIHI i i i inn n i i
99 80 90
PROBABILITY
M 99.9
Figure 2-1. Organic caxbon-oonnalized sorption isotfaenn fin* fluniaiiUniB (top) and
probability plot of KK (bottom) from «*<«»»*" tenacity tests conducted by Swam
(1991).
2-7
-------
recently measured by ERL.A. This approach 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 f^. The regression equation yields log10Koc
= 5.10. This value is in agreement with the log10Koc of 4.96 measured in the sediment toxicity
tests.
2-8
-------
SECTION 3
TOXICITY OF FLUORANTHENE: WATER EXPOSURES
3.1 TOXXCTTY OF FLUORANTHENE IN WATER: DERIVATION OF FLUORANTHENE
WATER QUALITY CRITERIA:
The equilibrium partitioning method for derivation of sediment quality criteria uses the
fluoranthene water quality criterion Final Chronic Value (FCV) and partition coefficients (Koc)
to estimate the maximum concentrations of non-ionic 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.
3.2 ACUTE TOXICITY - WATER EXPOSURES:
Eighteen standard acute toxicity tests with fluoranthene have been conducted on 12
freshwater species from 11 genera (Appendix A). Overall genus mean acute values (GMAVs)
range from 32 to > 4,000 /tg/L. The acute values for three species tested, Lumbriculus
variegatus. Physella virgata. and Ophiogomphus sp., were all > 178.5 n£fL. The acute values
for the 4 most sensitive species differed by only a factor of 3; 32 to 95 pg/L. Eight tests on
benthic life-stages of eight species from seven genera are contained in this database (Table 3-
1; Appendix A). Benthic organisms were similar to water column species in sensitivity to
fluoranthene; GMAVs range from 32 to > 178.5 /xg/L. One infaunal species, the annelid
Lumbriculus variegatus. was tested, (LC50 > 178.5 /tg/L). The Final Acute Value derived
from the overall GMAVs (Stephan et al., 1985) for freshwater organisms is 27.18 pg/L (Table
3-3).
3-1
-------
TABLE 3-1. - ACOTB SENSITIVITY OP FRESHWATER AND SALTWATER BENTHIC SPECIES TO
PLUORANTH5NB - DARK DATA ONLY.
RANK*
HMAV
GENUS
1
1
3
5
6
8
8
a
i
2
4
6
COMMON/SCI. NAME
Amphipod,
Gammarua minus
Amphipod,
G.ffffpflrufl paaudol imnaaua
Hydra,
Hydra amerieana
Stonefly,
Peltooerla maria
Snail,
Phvaa heteroatropha
Annelid,
Lumbriculua variaqatus
Dragonfly,
Qphi OQomphua sp.
Snail,
Phvaella viraata
Myaid,
MvaidooBia bahia
Amphipod,
flTOfllipca abdita
Grass shrimp.
palaoBtoaataa ouaio
Annelid,
LIFE1* HAB-C
STAGE ITAT
FRESHWATER SPECIES
A E
A B
J B
X B
X B
A I
N B
A .8
SALTWATER SPBCIBS
J B
J I
J B
J I
HMAV
SPECIES0 GENUS*
M9/L MST/L
32.0 32.0
>116.6 32.0
70.06 70.06
135.0 135.0
137.0 137.0
>178.5 >178.5
>178.5 >178.5
>173.5 >178.5
51.54 51.54
66.93 66.93
142.45 142.45
500 500
WflffinthflB arenaceodantata
a
a
Coot clam.
Mulinia lateralia
Sheepshead minnow,
B B
J E,W
10,710 10,710
>560,000 >560,000
Cyprinodon variacratuB
3-2
-------
aRank of HMAVs by genus are from Appendix A which included benthic and water column
species.
Life stage: A » adult, J » juvenile, L, » larvae, B » embryo
cHabitat: I * infauna, B » epibenthic, w a water column.
-------
Fifteen acute tests have been conducted on 10 saltwater species from 10 genera
(Appendix A). Overall genus mean acute values (GMAVs) range from 51.54 to > 560,000
/ig/L, with crustaceans the most sensitive. Within this database there are results from 6 tests
on benthic life-stages of 6 species from 6 genera (Table 3-1; 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 bjhja, with an average flow-through
96 hour LCSO of 51.54 pg/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; GMAVs range from 66.93 to > 560,000 /tg/L- The Final Acute Value derived
from the overall GMAVs (Stephan et al., 1985) for saltwater organisms is 35.63 /ig/L (Table
3-3).
3.3 CHRONIC TOXIOTY - WATER EXPOSURES:
Chronic toxicity tests have been conducted with fluoranthene using a freshwater
cladoceran, fDaphnia magnat and fathead minnows fPifllCpbeles prflffldafl) &nd a saltwater mysid
(Mysidopsis bjhJa; Table 3-2). The cladoceran and the mysid were tested in life-cycle
exposures. Fathead minnows were exposed in an early life-stage toxicity test.
Brooke (1991) conducted tests on both freshwater species. Paphnia magoa were exposed
21 days to mean fluoranthene concentrations ranging from 1.4 to 45.6 pg/L. No effects out
either reproduction or survival relative to controls were detected at any concentration tested,
The 96 hour LC50 of 102.8 pg/L is used as the observed effect concentration to obtain the
chronic value of 58.9 /ig/L for this species. Fathead minnows 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
Saltwater mysids were tested in two life-cycle toxicity tests. In the first, mysids were
3-4
-------
TABLE 3-2. - CHRONIC SENSITIVITY OP FRESHWATER AND SALTWATER ORGANISMS TO FLUORANTHENE.
TEST SPECIFIC DATA.
Common
Name.
Scientific
Name
Habitat0 Parental
Test* (Life stage) NOBC(*'° LOECC
Parental Progeny
Effect** LOBC
Progeny
Effect*1
References
Cladoceran,
Daphnia maona
LC W(J,A)
FRESHWATER SPECIES
1.4 - 45.6 102.8
50% M
*Test: LC - lifecycle, PLC - partial lifecycle, ELS » early life stage
°Habitat: I « infauna, B - epibenthic, W = water column
Lifestage: A « adult, J - juvenile. L » larvae, B « embryo
CNOBC: no observed effect concentration.
LOBC: lowest observed effect concentration.
^EFFECT: percentage decrease relative to controls. M = mortality G « growth R = reproduction
Brooke. 1991
CN>
^
Fathead minnow, ELS W(E,L,J)
Pimeohales promelas
Mysid, LC E(J,A)
Mvaidopsis bahia
Mysid, LC B(J,A)
MvaidopsiB bahia
3.7 - 10.4 21.7
SALTWATER SPECIES
<5 - 12 21
43
0.11 - 11.1 18.8
67% M
50.2% G
26.7% M
91.7% R
100% M
30% M
12% G
100% R
Brooke, 1991
EG&G, 1978
Champlin and
Poucher, 1991
-------
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
concentrations less than or equal to 12 ng/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 pg/L, all mysids died. In the second
test, mysids were exposed to fluoranthene for 31 days (Champlin and Poucher, 1991). Effect
concentrations were similar to the first test. Mysids were not affected at fluoranthene
concentrations _<. 11.1 /ig/L. Survival was reduced 30%, growth 12% and reproduction 100%
relative to controls in 18.8 pg/L, the highest concentration tested.
The difference between acute and chronic sensitivity to fluoranthene, in tests where UV
activation did not occur, is less than an order-of-magnitude (Table 3-3). Three acute-chronic
ratios (ACR) are available; 1.74S for D. magaa, 3.397 for M- bjhja. and 6.325 for £. promelas.
The Final Acute Chronic Ratio is the geometric mean of these three values (3.347).
The Final Chronic Values (Table 3.3) are used as the effect concentrations for
calculating the sediment quality criteria for protection of freshwater and saltwater benthic
species. The Final Chronic Value (FCV) for freshwater organisms of 8.121 pg/L is the
quotient of the Final Acute Value (FAQ of 27.18 pg/L and the Final Acute-Chrome Ratio of
3.347. Similarly, the FCV for saltwater organisms of 10.65 pg/L is the quotient of the Final!
Acute Value of 35.63 fig/L and the Final Acute-Chronic Ratio.
3.4 PHOTOTOXICrrY OF FLUORANTHENE:
Under laboratory conditions most PAHs are predicted to be not acutely toxic at or below
their solubility in water (Newsted and Giesy, 1987). Under ultra-violet light, however, the
toxicity of some PAHs can increase by several orders of magnfriiH* The effect has been shown
to be a result of photoactivation rather than photodegradation of the parent compound to mow
toxic metabolites. With some PAHs, toxicity occurs by activation by UV light of chemical
3-6
-------
TABLE 3-3. -
SUMMARY OP ACUTE AND CHRONIC VALUES, ACUTE-CHRONIC RATIOS AND DERIVATION OP THE FRESHWATER AND SALTWATER
FINAL ACUTE VALUES, FINAL ACUTE-CHRONIC RATIOS AND FINAL CHRONIC VALUES.
Common Name,
Scientific Name
Acute Value(pg/L)
Chronic Value
-------
present on or within an organism. Bluegills (Lepomis macrochirust exposed to anthracene in
sunlight in outdoor artificial streams died (Bowling et al., 1983). Bluegills in the same stream,
but downstream in the shade survived. Bluegills exposed in the shade died within 24 hours
when placed into clean water and brought into the sunlight. Likewise, Bafihaia magna. are
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 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
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.
Benthic organisms by remaining buried or organisms in the shade therefore, 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 standardirad toxicity tests with
a variety of organisms (Appendix A). Although, the toxicity of fluoranthene appears to increase
3-8
-------
with increases in intensity of UVA at low UV intensities (Figure 3-1), the acute toxicity of
fluoranthene to saltwater organisms is similar under commercially available UV lights and
sunlight (Figure 3-1; Appendix A). This is important since conducting acute and particularly
chronic toxicity tests outside in sunlight would be extremely difficult and expensive. The
magnitude of increase in fluoranthene's toxicity following UV activation can be great. The
ratio of LCSOs from acute tests conducted in the dark or under cool-white fluorescent light to
LCSOs for the same species exposed in 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 /ig/L. Chronic tests conducted in both the "dark" and under UV light are
available for two freshwater species, J^ghjoia magna and Pimephales promelas. The chrome
values decreased by factors of 64 for D. magna and 5.8 for P.. proroejag in the presence of UV
light (Table 3-2). Magnitude of increase in acute and chronic toxicity under UV light are
almost identical. Therefore, acute-chronic ratios differ little between non-UV and UV
exposures (Table 3-3).
There are not enough 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 is 0.4333 /tg/L. This value is approximately 70 times lower than
the "dark" FA Vs. If the final acute-chronic ratio from UV tests is similar for tests without UV
activation, and can be pooled with other chronic tests, then the Final Chronic Value^ would
be estimated to be 0.2037 pg/L, a factor of 44 less than FCV derived from non-activated
fluoranthene toxicity data.
3-9
-------
100
Coot-While
Vita-Lite
UVA-340.45 cm
UVA-340.26 cm
Sunlight
300 600 900 1200 1500
UV-A(nWcm2)
Figure 3-1. Acute toxicity of fluoranthene to mysids (Mysidopsis bahia) as a function of
intensity of UVA exposure.
-------
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 SQC in this document may be under protective. There are many examples
of specific benthic organisms where exposure to fluoranthene (and other PAHs) and sunlight can
co-occur. For example, fiddler crabs 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 transfered 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.
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 final chronic value. More
importantly, there is an absence of data demonstrating a causal linkage between exposure of
sediment-associated fluoranthene and increased risks of UV inhanrMi effects on benthic
organisms or organisms coupled to benthic organisms via food chains. EPA encourages
research efforts on these topics.
3.5 APPUCABUJTY OF THE WATER QUALITY CRITERION AS THE UPPHC1S
CONCENTRATION FOR DERIVATION OF THE FLUORANTHENE SEDIMENT
QUALITY CRITERION:
The use of the Final Chrome Value (the chronic effects-based water quality criteria
concentration) as the effects concentration for calculation of the equilibrium pardtioning-based
sediment quality criterion assumes similar sensitivities of benthic (infauna and epibenthic)
3-11
-------
species and species tested to derive the water quality criteria concentration. Data supporting
the reasonableness of this assumption over all chemicals for which there are published or draft
water quality criteria documents are presented in Di Toro et al. (1991) and U.S. EPA (1989c,
1992a). The conclusion of similarity of sensitivity is supported by comparisons between acute
values: (1) for the most sensitive benthic and water column species for all chemicals; (2) all
species across all chemicals after standardizing the LC50 values; and (3) individual chemical
comparisons for benthic and water column species. Only in this last comparison are
fluoranthene-specific comparisons in sensitivity of benthic and water-column species conducted.
The following paragraphs examine tile data for fluoranthene.
An initial test of the difference between the probability distributions of freshwater and
saltwater fluoranthene LCSOs for all species (water column and benthic) is presented in Figure
3-2. The top panel is a log probability plot of the two LC50 distributions on a log scale versus
the rank order on a probability scale. The natural way to judge the equality of these
distributions is to compare the LCSOs at a particular probability, for example a comparison of
the medians at 50% probability. The Kolmogorov-Smimov test compares another difference
(Conover, 1980). This is illustrated on the bottom plot which presents the same data but in a
slightly different way. The rank order, as a percent, is plotted versus the LCSOs. The points
are connected with straight lines to form the empirical cumulative distribution functions from
the two data sets. The Kolmogorov-Smirnov test is based on the maximum difference in
probability between these two distributions, as indicated on the figure. Note that this difference
is the horizontal distance on the top plot in Figure 3-2 (if the probability scale were linear).
Table 3-4 presents the number of LCSOs in each distribution, the maximum difference (0.600),
and the probability (0.980) that a value of this magnitude or less cannot occur given that these
two samples came from the same distribution. The method of computation for this probability
value is given in Massey (1951). Since the probability is greater than 0.9S, the hypothesis that
freshwater and saltwater LC50 values came from the same distribution is rejected at a 95%
3-12
-------
CT
3
O
in
u
inonnn
1COOO
1000
100
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1
= i i Minn i i i iriiu
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in 1 1 1 1 i i linn 1 1 i
00
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0.1
10 20 90 80 90
PROBABILITY
99 99.9
90
80
70
60
' 30
40
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20
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0
i i i linn i^
-
MAXIMUM
DIFFERENCE
-
1
ki i nun
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i iii urn i i i 1 1
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r
i i i mi
-
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i mi i i i 1 1 mi i i i i mi
10
100
1000
10000 100000
LC50 (ug/L)
Figure 3-2.
Comparison of fluoranmene water only LC50 probability distributions for
freshwater (0) and saltwater (*) species (top panel). Cumulative distribution
functions for ^gniating the K-S «*afistic (bottom panel)*
3-13
-------
confidence level. Therefore, for fluoranthene comparisons of LCSOs for benthic and water
column species are conducted separately for freshwater and saltwater LCSO values.
The probability distributions of freshwater and saltwater fluoranthene LCSOs for the
water column and benthic species are presented in Figure 3-3. Comparisons of freshwater
benthic and water column species are presented in the left panels, saltwater species in the right
panels. Table 3-4 presents the number of LCSOs in each distribution, the maximum difference,
and the probability that a value of this magnitude or less cannot occur given that these two
samples came from the same distribution. For freshwater species the maximum difference and
the probability that a value of this magnitude or less cannot occur given that these two samples
came from the same distribution were 0.37S and 0.487 respectively. For saltwater the two
values were 0.333 and 0.30S. Separate analyses of the relative sensitivities of freshwater
benthic and water column organisms and saltwater benthic and water column organisms that
have been tested indicates that they are from the same probability distributions of LCSO's.
Therefore, benthic and water column organisms have similar acute sensitivities. This suggests
that the final chronic value (FCV) for fluoranthene is an appropriate effects concentration for
both benthic and water column organisms.
3-14
-------
TABLE 3-4. KOLMOGOROV-SMIRNOV TEST FOR THE
EQUALITY OF THE FRESHWATER AND SALTWATER LC50
DISTRIBUTIONS FOR FLUORANTHENE. KOLMOGOROV-SMIRNOV
TEST FOR THE EQUALITY OF BENTHIC AND WATER
COLUMN LC50 DISTRIBUTIONS.
Compar-
ison Pat?it?t 9T Water Type* K-S Statistic* Probability0
Fresh
vsSaJt
Benthic
Fresh (12)
Benthic (8)
Salt (12)
Water (4)
0.600
0.375
0.980
0.487
vs Water Column
Column
(Freshwater)
Benthic Benthic (6) Water (4) 0.333 0.305
vs Water Column
Column
(Saltwater)
"Values in parentheses are the number of LC50 values used in the comparison.
"K-S statistic = mayimum difference between the cumulative
distribution functions for benthic and water column species.
°Pr(K-S theoretical £ K-S observed) given that the samples
came from the same population.
3-15
-------
o
in too
o
10
0
I
1
( mum i i ii mil
1 1 10
_l 1 1
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£ PROBABILITY
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(ug/L)
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i
i i iiiim i i i mill 11 i i i i i ii in mill mini i i
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PROBABILITY
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I i ii mil i i i inn
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1
1 10 tOO 1000 10000 |(
i rnn dm/1 i
FRESHWATER PROBABILITY OF LC90 FOR FLUORANTHENE
SALTWATER PROBABILITY OF LC50 FOR FLUORANTHENE
Figure 3-3. Comparison of fluoranthene water only LCSO probability distributions for water
column (0) and benthic (*) freshwater and saltwater species (top panel).
Cumulative distribution^^ptions for calculating the K-S statistic fbottnm
-------
SECTION 4
TOXICITY OF FLUORANTHENE (ACTUAL AND PREDICTED):
SEDIMENT EXPOSURES
4.1 TOXICITY OF FLUORANTHENE IN SEDIMENTS:
The toxicity of fluoranthene spiked into sediments has been tested with three saltwater
amphipod species, and one freshwater amphipod species. 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 a number of 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 toxicity observed.
Swaitz et al. (1990) exposed the amphipods Corophium spinicorne and Rhepoxynius
abronius to three fluoranthene-spiked sediments with total organic carbon content (TOQ of
0.18%, 0.31 % and 0.48%. Sediments were prepared using the methods of Swaitz et al. (198S)
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 10-day LC50'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
4-1
-------
TABLE 4-1: SUMMARY OP TESTS WITH PLUORANTHBNB-SPIKED SEDIMENT.
Common/Sci . Name
Amphipod,
Corophium
BDJnicome
Amphipod,
Corophium
spinicorne
Amphipod,
Corophium
spinicorne
£
Amphipod.
RhepoxyniuB
abroniua
Amphipod,
Rhepoxvn i us
ftkroniuH
Amphipod,
Rhepoxvn i us
abronius
Amphipod,
Hvalella
asteca.
Sediment
Source ; TOC
Description (%)
Yaquina Bay, OR, 0.18
Sand with Organic
Matter Added
Yaquina Bay, OR, 0.31
Sand with Organic
Matter Added
Yaquina Bay, OR, 0.48
Sand with Organic
Matter Added
Yaquina Bay, OR, 0.18
Sand with Organic
Matter Added
Yaquina Bay, OR, 0.31
Sand with Organic
Matter Added
Yaquina Bay, OR, 0.48
Sand with Organic
Matter Added
Pine Sand, 0.5mm
Sieved (2ppt Salinity)
Pore
Method*/ Sediment Pluoranthene Water
Duration LC50.ua /a LC50.
(Days) Response Dry wt. Org. Car. /*g/L
S.M/10 LC50 5.1 2,830 33.4
S.M/10 LC50 >13.6 >4,390 >52.0
S.M/10 LC50 >13.6 >2,830 >27.5
S.M/10 LC50 3.4 1,890 22.7
S.M/10 LC50 6.5 2,100 29.4
S.M/10 LC50 10.7 2.230 24.2
S.M/10 LC50 15.4
(21.4)b
References
Swartz et al., 1990
Swartz et al., 1990
Swartz et al., 1990
Swartz et al., 1990
Swartz et al., 1990
Swartz et al.. 1990
De Witt et al. .
1989
-------
TABLE 4-1: (continued)
Common/Sci. Name
Amphipod,
flphauBtoriuB
estuariufl
Amphipod,
BohauBtorius
eatuariua
Amphipod,
Bohauatoriufl
eatuariuB
f>
u> Amphipod,
pohauBtoi iutt
aatuaiiua
Amphipod,
Bohaustorius
aatuariuB
Amphipod,
RhepoxyniuB
aJoronius
Amphipod,
Rhepoxvniua
abroniua
Sediment
Source ; TOC
Description (%)
Pine Sand, 0.5mm
Sieved (2ppt Salinity)
Pine Sand, 0.5mm
Sieved (Sppt Salinity)
Pine Sand, 0.5mm
Sieved (lOppt Salinity)
Pine Sand, 0.5mm
Sieved (ISppt Salinity)
Pine Sand, 0.5mm
Sieved (28ppt Salinity)
Pine Sand, 0.5mm
Sieved (28ppt Salinity)
Pine Sand, 0.5mm 0.34
Sieved, Zostera added
Pore
Method*/ Sediment Pluoranthene Water
Duration LC50.ua /a LC50.
(Days) Response Dry wt. Org. Car. pg/L
S.M/10 LC50 9.3
(13.8)b
S.U/10 LC50 14. Ob
S.M/10 LC50 10.7
(15.1)
S.0/10 LC50 13. 9b
S.M/10 LC50 11.8
(17.5)b
S.M/10 LC50 5.1
(6.6)b
S.0/10 LC50 19. lb 5,620 >31Sd
(>14.8)° (>179)*
References
De Witt et al . ,
1989
De Witt et al.,
1989
De Witt et al. ,
1989
De Witt et al . ,
1989
De Witt et al.,
1989
De Witt et al..
1989
De Witt et al. ,
in press
-------
TABLE 4-1: (continued)
Common/Sci . Name
Amphipod,
RhepoxynJ,n0
abronius
Amphipod,
Rhepoxyp^uB
abroniua
Amphipod,
RhepoxyniuB
abroniua
Amphipod.
Rhapoxyniui
afarnniua
Sediment
Source ;
Description
Pine Sand, 0.5mm
Sieved, * Suspended
Solids* Added
Pine Sand, 0.5mm
Sieved, Mud Added
Pine Sand, 0.5mm
Sieved, Oyster Peces
Added
Pine Sand, 0.5mm
Sieved, Shrimp Peces
Added
Pore
Method*/ Sediment Pluoranthene Water
TOC Duration LC5Q,fjq/q LCSO,
(%) (Days) Response Dry wt. Org. Car. /*g/L
0.34 S.M/10 LCSO 15.0 4,410 14. ld
(12.5)*
0.40 S.M/10 LCSO 12.6 3,150 19. 2d
(17.4)*
0.31 S.M/10 LCSO 9.56 3,080 26. 6d
(18.6)*
0.31 S.M/10 LCSO 8.65 2,790 9.38d
(8.09)*
References
De Witt et al . ,
in press
De Witt et al . ,
in press
De Witt et al . ,
in press
De Witt et al . ,
in press
*S . Static, 11 m Measured concentration
bNominal LCSO value
cMeasured LCSO value
^Total interstitial water
*Pree interstitial water
-------
normalized to dry weight differed by a factor of 3.1 (3.4 to 10.7 /tg/g) for IL 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 MS/goc)- Because less than 50% mortality of £. spinicome
resulted in the highest fluoranthene treatments in two of the three sediments used in this
experiment, it is not possible to make similar comparisons with this species.
DeWitt et al. (1989) exposed the saltwater ampmpod £. estuarius to fluoranthene-spiked
sediments at five different salinities and fi. abronius and the freshwater amphipod Hyalella
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. Fluoranthene toxicity to £. 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 not measured in these sediments.
The 10-day LC50 for £. abronius (5.1 /ig/g, dry wt) was similar to those reported by Swartz
et al. (1990).
Dewitt et al. (in press) exposed £. 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 (Crassostiea gigast and feces of a deposit-feeding shrimp
-------
Table 4-2: Hater-only and sediment LCSOa used to test the applicability of the equilibrium partioning theory
for tluoranthene.
Common/Sci .
Amphipod,
Rheooxynius
Amphipod,
Rhepoxvnius
Amphipod,
RhepoxvniuB
Amphipod,
RhepoxvniuB
Amphipod,
Rhepoxvnius
Amphipod,
Rhepoxvnius
Amphipod,
RhepoxvniuB
Amphipod.
RhepoxvniuB
Hater Only
Name Method* LC50
Duration (days)
S.M/10
gbrgniiiB
S.M/10
flbroniufi
S.M/10
flb.rpni"H
S.M/10
abroniuB
S.M/10
abronius
S.M/10
abroniup
S.M/10
ttbroniya
S.M/10
flteoniuft
/ig/L
27, 2b
27. 2b
27. 2b
27. 2b
27. 2b
27. 2b
27. 2b
27. 2b
Sediment
Pluoranthene Predicted0 Ratio:
Overlvina Pore LC50 LC50 Actual
Hater Hater TOC ng/g pg/g /ig/g LC50
LCSO LC50 (%) Dry Ht. OC OC Predicted Reference
/ig/L pg/L
22.7 0.18 3.4 1,890 3.420 0.553 Swartz
29.4 0.31 6.5 2,100 3,420 0.614 Swartz
24.2 0.48 10.7 2,230 3,420 0.652 Swartz
>315d 0.34 19.1* 5,620f 3.420 1.64 DeHitt
press
14. ld 0.34 15.0 4,410 3,420 1.29 DeHitt
press
19. 2d 0.40 12.6 3,150 3.420 0.921 DeHitt
press
26. 6d 0.31 9.56 3,080 3.420 0.900 DeHitt
press
9.38d 0.31 8.65 2.790 3.420 0.816 DeHitt
press
et al., 1990
et al. , 1990
et al., 1990
et al . , in
et al . , in
et al., in
et al., in
et al. , in
*S - static, M - measured concentration
b96 hour BC50 (reburial) (Swart*, 1991).
Predicted LC50 (|ig/goa> Hater-only LC50 (pg/L) x
dTotal interstitial water
*Dnmeasured, measured value - 14.81
Unmeasured, measured value 4,360
-------
Overall, the need for organic carbon normalization of the concentration of non-ionic
organic chemicals in sediments is presented in the SQC Guidelines (U.S. EPA, 1992a). 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 EqP approach,
it is fundamentally more important to demonstrate that KOC and water only effects concentrations
can be used to predict effects concentrations for fluoranthene and other non-ionic organic
chemicals on an organic carbon basis for a range of sediments. Evidence supporting this
prediction for fluoranthene and all SQC chemicals follows in Section 4.3.
4.2 CORRELATION BETWEEN ORGANISM RESPONSE AND PORE WATER
CONCENTRATION:
One corollary of the EqP theory is that pore-water LCSO's for a given organism should
be constant across sediments of varying organic carbon content (U.S. EPA, 1989a). Pore-
water LC50 values are available for two species (Tables 4-1; 4-2). Swartz et al. (1990) found
10-day LC50 values based on pore-water concentrations varied by a factor of 1.3 (22.7 to 29.4
/ig/L) for IL abronius. This variability is less than that shown when dry weight (factor of 3.1)
normalization is used to determine LCSOs based on fluoranthene concentration in sediments, but
similar to that shown when organic carbon (factor of 1.2) normalization is used.
DeWitt et al. (in press) found that 10-day LC50 values based on pore water
concentrations varied by a factor of 2.8 (range 9.38 to 26.6 pg/L) for four of the five sediments
tested, but that the pore water LCSO for the firm sediment was much higher than that of the
other four sediments (>315/ig/L). This result runs counter to previous observations that the
pore water concentration shows strong correlation with toxicity (see below). DeWitt et al. (in
press) 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
4-7
-------
bioavailable. Free (not associated with organic carbon) pore water fluoranthene concentration
was measured using a modification of the Landrum et al. (1984) reverse-phase seperation
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 (LCSO = 8.09 to > 179 pg/L) were as variable as those based on
total pore water fluoranthene (9.38 to >31S ng/L). The fact that the organic-carbon
normalized LCSOs 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 lack of similarity of LCSO 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 LCSO values were similar and not 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, 1991), cadmium (Swartz et al., 1985), endrin (Nebeker et al., 1989;
Schuytema et al., 1989), fluoranthene (Swartz et al., 1990), or kepone (Adams et al., 198S)
(Figure 4-1; Appendix Q. Tests with acenaphthene and phenanthrene used two saltwater
amphipods (L. plumulosus and IL tStMarJUiO and marine sediments. Tests with cadmium and
fluoranthene used the saltwater amphipod fRhepoxyniua abrom'ua) and marine sediments.
Freshwater sediments spiked with endrin were tested using the amphipod Hyalella aztfisi; while
the midge, Chiirmomous tftnt3HSi was tested using kepone-spiked y*ti«pMrt«. Figure 4-1
presents the percentage mortalities of the benthic species tested in individual treatments for each
chemical versus "pore water toxic units" for all sediments tested. Pore water toxic units are
the concentration of the chemical in pore water (>tg/L) divided by the water only LCSO G*g/L).
In this normalization, 50% mortality should occur at one interstitial water toxic unit. In
4-8
-------
100
80
>-
n 60
_l
<
g 40
2:
20
a - ENDRIN
A - ACENAPHTHENE c
v - PHENANTHRENE °
+ - FLUORANTHENE
o - KEPONE
+ - CADMIUM
o
a i
° °a«
o v^
° 9 v &** v
.n JDnOOAv
nrr* rt O u jfc J\A^VA
1 1 1 1 1 1 III 1 1 1 1 1 1 II
1 1 1 1 1 1 1 II 1 1 1 1 1 1 1 1
Dt O 4H flDODUHB O Q
0 A
VV
V
* V V
I V V
0 A
'
kVA*
A
VA
. I.I 1 1 Mill 1 1 I I || ||
0.01
0.1 1 10
PORE WATER TOXIC UNITS
100
Figure 4-1. Percent mortality of amphipods in sediments spiked with acenaphlhene or
phenanthrene (Swartz, 1991), cadmium (Swartz et al., 1985), endrin (Nebeker
el al., 1989; Schuytema et al., 1989), or Huoranthene (Swartz et al., 1990), and
midge in kepone-spiked sediments (Adams et al., 198S) 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 C in this SQC document, Appendix C in the
endrin, dieldrin, acenaphthene, and phenanthrene SQC documents, and original
references for raw data.)
-------
general, 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 sediment quality criteria in this document because
of the complexation of non-ionic organic chemicals with pore water DOC (Section 2) and the
difficulties of adequately sampling pore waters. Data from the dieldrin experiments (Hoke and
Ankley, 1991) are not included because more knowledge of the pore water DOC will be
required because dieldrin has a high KQC value.
4.3 TESTS OF THE EQUILIBRIUM PARTITIONING PREDICTION OF SEDIMENT
TOXTCITY:
Sediment Quality Criteria derived using the equilibrium partitioning approach utilize
partition coefficients and final chronic values from water quality criteria documents to derive
the sediment quality criteria concentration for protection of benthic organisms. The partition
coefficient (Koc) is used to normalize sediment concentrations and predict biologically available
concentrations across sediment types. Data are available to test the normalization for
fluoranthene in sediments. Testing of this component of SQC derivation requires three
elements: (1) a water-only effect concentration, such as a 10-day LC50 value in pg/L, (2) an
identical sediment effect concentration on an organic carbon basis, such as a 10-day LC50 value
in /ig/goc, and (3) a partition coefficient for the chemical, K^ in L/kgo,.. This section presents
evidence that the observed effect concentration in sediments (2) can be predicted utilizing the
water effect concentration (1) and the partition coefficient (3).
Predicted ten-day LC50 values from fluoranthene-spiked sediment tests (Swartz et al.,
1990; DeWitt et al., in press) on a pg.goc basis with IL abmaius were calculated (Table 4-2)
using the value of KOC (10s '*) from Section 2 of this document and the 4-day water-only EC50
values in Swartz (1991). Ratios of predicted to actual LCSOs for fluoranthene averaged 0.682
(range 0.45 to 1.0) for & abronius. The data from DeWitt et al (1989) can not be used for
prediction because the TOC of the sediments was not measured.
4-10
-------
A more detailed evaluation of the accuracy and precision of the EqP prediction of the
response of benthic organisms can be made using the results of toxicity tests with amphipods
exposed to sediments spiked with acenaphthene, phenanthrene, dieldrin, endrin, or fluoranthene.
Data from the kepone experiments are not included because we do not have a KQW for kepone
from ERL, Athens. Swartz (1991) exposed the saltwater amphipods JL estuarius and L.
piumulosus to acenaphthene and phenanthrene in three marine sediments having 1.02, 2.61 and
4.37% organic carbon. Swartz et al. (1990) exposed the saltwater amphipod JL abronius to
fluoranthene in three marine sediments having 0.18, 0.31 and 0.48% organic carbon. Hoke and
Ankley (1991) exposed the amphipod Hyalella azteca to three dieldrin-spiked freshwater
sediments having 1.7, 3.0 and 8.5% organic carbon. Nebeker et al. (1989) and Schuytema et
al. (1989) exposed iL 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" for each sediment treatment
Predicted sediment toxic units are the concentration of the chemical in sediments (/tg/goc)
divided by the predicted LCSO Cig/goc) m sediments (the product of KQC and the 10-day water-
only LCSO). In this normalization, 50% mortality should occur at one predicted sediment toxic
unit. Endrin and fluoranthene data indicate a slight under prediction, and dieldrin, acenaphthene
and phenanthrene a slight over prediction, of the observed mortality. In general, however, this
comparison 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. These are two fundamental propositions that underlie
the EqP method for deriving sediment quality criteria.
4-11
-------
£
N>
oc
o
100
BO
60
40
20
0
I i i i 1 II II i I I I I I 1 1
a - ENDRIN
A - ACENAPHTHENE a c
v - PHENANTHRENE °
o - DIELDRIN £
* - FLUORANTHENE
_
*
A
Q
°3» °
* *
1
ov X A*w
A8v °*V /X
: **'******
i i i i i | M> l | I I I I 1 1
i i i i i 1 1 1 1 i i i i i 1 1 r
>A tt^m a a IB a a
a A w
*
w
V
o
O V V
3 V V °
A O -
'
V
A
V
A
1 I | | H 1 1 1 i i I'llll
0.01
0.1 1 10
PREDICTED SEDIMENT TOXIC UNIT
100
Figure 4-2.
Percent mortality of amphipods in sediments spiked with acenaphihene or
pbenanthrene (Swartz, 1991), dieldrin (Hokeand Ankley, 1991), endrin (Nebeker
et al., 1989; Schuytema et al., 1989) or fluoranthene (Swartz et al 1990)
relative to 'predicted sediment toxic units." Predicted sediment toxic units are
the ratios of measured treatment concentrations for each chemical in sediments
Ojg/foc) divided by the predicted LC50 (^g/g^) in sediments (K^ x Water-only
LC50, jig/L). (See Appendix C in this document and Appendix C in the dieldrin
endnn, acenaphthene, and phenanthrene SQC documents for raw data).
-------
SECTION 5
CRITERIA DERIVATION FOR FLUORANTHENE
5.1 CRITERIA DERIVATION:
The equilibrium partitioning method for calculating sediment quality criteria is based
on the following procedure. If FCV 0*g/L) is the Final Chronic Value which is the chronic
effects concentration from the water quality criteria for the chemical of interest, then the
sediment quality criteria, SQC (/ig/g sediment), is computed using the partition coefficient, KP
*
(L/g sediment), between sediment and pore water
SQC = K, FCV (5-1)
On a sediment organic carbon basis, the sediment quality criteria, SQCoc G*g/gbc)> &
SQCoc = Koc FCV (5-2)
where KOC is the organic carbon partition coefficient for the chemical. Since this quantity is
presumably independent of sediment type for non-ionic organic chemicals, so also is
Table 5-1 contains the calculation of the fluoranthene sediment quality criteria.
TABLE 5-1. SEDIMENT QUALITY CRITERIA FOR FLUORANTHENE
Type of
Water Body
Freshwater
Saltwater
(L/kg)
5.19
5.19
(L/kg)
5.10
5.10
FCV
Gig/L)
8.12
10.65
G*g/goc)
1022'
1341b
= (10*" L/kgoc)«(10-J kgoc/gocWS.tt /tg fluoranthene/L) = 1022 Mg
fluoranthene/goc
"SQCoc = (10110 L/kgoc)»(10-J kgoc/gocWlO.65 A*g fluoranthene/L)= 1341 Mg
fluoranthene/goc
5.2 UNCERTAINTY ANALYSIS:
Some of the uncertainty in the calculation of the fluoranthene sediment quality criteria can
5-1
-------
be estimated from the degree to which the equilibrium pardoning model, which is the basis for
the criteria, can rationalize the available sediment toxicity data. The EqP model asserts that (1)
the bioavailability of non-ionic organic chemicals from sediments is equal on an organic carbon
basis, and (2) that the effects concentration in sediment G*g/goc) can be estimated from the
product of the effects concentration from water only exposures (/xg/L) and the partition
coefficient KOC (L/kg). The uncertainty associated with the sediment quality criteria can be
obtained from a quantitative estimate of the degree to which the available data support these
assertions.
The data used in the uncertainty analysis are the water-only and sediment toxicity tests that
have been conducted in support of the sediment criteria development effort. 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 the LCSOs from the water-only and sediment
exposures illustrated in Figure 4.2 in Section 4. The EqP model can be used to normalize the
data in order to put it on a common basis. The LC50 for sediment on an organic carbon basis,
LC50^oc, is related to the LC50 obtained from a water-only exposure, LCSO* via the
partitioning equation:
LC503.0C = KoeLCSO,, (5-3)
Therefore, K^ can be used to define the equivalent sediment toxicity based on free
concentration in pore water
5-2
-------
LC5CU
LC50PW = (5-4)
Koc
The EqP model asserts that toxicity of sediments expressed as the free pore water concentration
equals toxicity in water only tests.
LCSOpw = LC50W (5-5)
Therefore, either LCSO^ or LC50W are estimates of the true LC50 for this chemical - organism
pair. In this analysis, the uncertainty of K^ is not treated separately. Any error associated
with KOC WU1 be reflected in the uncertainty attributed to varying the exposure media.
In order to perform an analysis of variance, a model of the random variations is required.
As discussed above, experiments that seek to validate equation 5-5 are subject to various sources
of random variations. A number of chemicals and organisms have been tested. Each chemical
- organism pair was tested in water-only exposures and in different sediments. Let a represent
the random variation due to this source. Also, each experiment is replicated. Let represent
the random variation due to this source. If the model were perfect, there would be no random
variations other than that due to experimental error which is reflected in the replications. Hence
a represents the uncertainty due to the approximations inherent in the model and represents
the experimental error. Let (oj2 and (
-------
measurement error, (ae)2. The maximum likelihood method is used to make these estimates
(U.S. EPA, 1992a). The results are shown in Table 5-2.
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)
ffa 0.39
-------
TABLE 5-3. SEDIMENT QUALITY CRITERIA
CONFIDENCE LIMITS FOR FLUORANTHENE
Sediment Quality Criteria
95% Cpnfidence Limits
Type of
Water Body jtg/goc Lower Upper
Freshwater 1020 470 2190
Saltwater 1340 620 2880
The organic carbon normalized sediment quality criteria is applicable to sediments with
an organic carbon fraction of f^ S 0.2%. For sediments with f^. < 0.2%, organic carbon
normalization and sediment quality criteria do not apply.
5.3 COMPARISON OF FLUORANTHENE SQC TO STORET DATA FOR SEDIMENT
FLUORANTHENE:
A STORET (U.S. EPA, 1989a) 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 5-1. 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 pg/g to 0.2 /*g/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 sediments of natural water systems in the United States as contained in the STORET database
(U.S. EPA, 1989a). 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
5-5
-------
at
01
z
LU
3
LU '
C/3
LU
CE
0
01
LU
I
s
o
r
10'
10
10'
to°L
STREAM
10
10'
10'
10"
10"
-«r
0000
Ol
\
O)
z
X
II
a
LU
LU
cc
a
0.1 1 1030 90 8090 9999.9
X LESS THAN OR EQUAL TO
ESTUARY
«
JQ "*Jjj^LillML_LjJ^AiJ_^HIIliil
0.1 1 1090 90 0090 9999.9
0.1 1 1020 90 8090 9999.9
X LESS THAN OR EQUAL TO
X LESS THAN OR EQUAL TO
Figure 5-1. Probability distribution of concentrations of fluoianthene in «*tim*nt«
streams (n=»786), lakes (n=57) and estuaries (n=»88) in the United States from
1986 to 1990, from the STORET (U.S. EPA, 1989c) database, compared to the
fluoianthene SQC values of 102 /tg/g in freshwater wtimmts having TOC =
10% and 10.2 pg/g in freshwater sediments having TOC - 1%; SQC values for
saltwater sediments are 134 /jg/g when TOC =10% and 13.4 pg/g when
. TOC=1%. The upper dashed line on each figure represents the SQC value when
TOC = 10%, the lower dashed line represents the SQC when TOC » 1%.
5-6
-------
SECTION 6
CRITERIA STATEMENT
The procedures described in the "Guidelines for Deriving Numerical National Sediment
Quality Criteria for Nonionic Organic Chemicals for the Protection of Benthic Organisms" (U.S.
EPA, 1992a) indicate that, except possibly where a locally important species is very sensitive
or sediment organic carbon is < 0.2%, benthic organisms should be acceptably protected in
freshwater sediments containing _<. 1020 ng fluoranthene/g organic carbon and saltwater
sediments containing _<. 1340 \i% fluoranthene/g organic carbon.
These concentrations are the U.S. EPA's best scientific judgement at this time of the
acceptable concentration of fluoranthene in sediments. Confidence limits of 470 to 2190 /ig/Sbc
for freshwater sediments and 620 to 2880 Mg/goc for saltwater sediments are provided as an
estimate of the uncertainty associated with the degree to which the observed concentration in
sediment Gtg/goc). which may be toxic can be predicted using the 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 sediment quality criteria and their confidence limits. The upper confidence
limit might be interpreted as a concentration above which impacts on benthic species would be
highly likely. The lower confidence limit might be interpreted as .a concentration below which
impacts on benthic species would be unlikely.
6-1
-------
concentrations and the extent to which they may exceed the sediment quality criteria.
5-9
-------
only. For fresh water sediments, SQC values are 10.2 /ig/g in sediments having 1 % organic
carbon and 102 pglg dry wt. in sediments having 10% organic carbon; for marine sediments
SQC are 13.4 /xg/g and 134 uglg, respectively. Figure 5-1 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 10.2 /*g/g for 1 % organic carbon sediments is exceeded for
2% of the data and the SQC of 102/*g/g for sediments having 10% TOC is exceeded by less
than 1 % of the data. For lakes (n = 57) the SQC for 1 % and 10% organic carbon sediments
are not exceeded by any of the sample data. In estuaries, the data (n = 88) indicate that the
criteria of 13.4 ug/g dry weight for sediments having 1 % organic carbon and the criteria of 134
/tg/g dry weight for sediments having 10% organic carbon are also not exceeded by the post
1986 samples.
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 phenanthrene (1340 Mg/goc) on Figure 5-2. Data presented are from
sediments with 0.20 to 31.9 percent organic carbon. The median organic carbon normalized
fluoranthene concentration (about 7.0 Mg/goc) is three orders of magnitude below the SQC of
1340 Mg/goc- About 0.1 % of these samples (n = 982) exceed the criteria. 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 intejisities of sampling in different
study areas. It is presented as an aid in assessing the range of reported fluoranthene sediment
5-7
-------
Ut
oo
o
O
O)
UJ
a
UJ
en
10000
1000
100
10
i iinnr r r i iimr i i
i i
in 1111 i i nuM i
sac
oc
0.1
0 ft1l I I
M HIIII
j
illllll
0.1
10 20 50 60 90
PROBABILITY
99 99.9
Figure 5-2. 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 SQC value of
1340Mg/goc.
-------
APPENDIX C. - SUMMARY OF DATA SEDIMENT EXPERIMENTS WITH FLOORANTHBNB.
SEDIMENT
SEDIMENT MORTALITY
SOURCE (%)
Yaquina Bay, OR, 2.5
RhepQxyniua 0 . 0
abroniua 42.5
100.0
100.0
Yaquina Bay, OR, 2.5
RhaDoxvniua 37.5
O abronius 100.0
H
Yaquina Bay, OR, 0.0
RJie.DCucYnj.ifp 25 . 0
ftbrgn4uB 60.0
CONCENTRATION. fia/Cf
DRY NT. ORO. CAR.
0.9
1.5
3.0
6.3
13.2
3.0
6.3
13.2
3.0
6.3
13.2
450
750
1500
3150
6600
1000
2100
4400
600
1260
2640
PORE WATER
CONCENTRATION
(ug/L)
5.7
9.7
22.5
38.3
56.0
11.3
30.6
58.0
3.0
13.3
27.3
TOC
(%)
0.20
0.20
0.20
0.20
0.20
0.30
0.30
0.30
0.50
0.50
0.50
Log
KOC REFERENCES
4.89 Swartz et al . ,
4.88 1990
4.81
4.92
5.07
4.96 Swartz et al.,
4.83 1990
4.88
5.30 Swartz et al . ,
4.99 1990
4.99
-------
-------
SECTION 7
REFERENCES
Adams, W.J., R.A. Kimerle and R.C. Mosher. 1985. Aquatic safety assessment of chemicals
sorted to sediments. In: Aquatic Toxicology and Hazard Assessment: Seventh
Symposium. Eds: R.D. Cardwell, R. Purdy and R.C. Banner. Amer. Soc. Testing and
Materials, Philadelphia, PA. STP 854. pp. 429-453.
Allied, P.M and J.P. Giesy. 1985. Solar radiation-induced toxicity of anthracene to Daphnia
pulex. Environ. Toxicol. Chem. 4:219-226.
Birge, W.J., J.A. Black, S.T. Ballard and W.E. McDonnell. 1982. Acute toxicity testing with
freshwater fish. Final Report Task n and m. U.S. EPA. Contract No. 68-01-6201.
Bowling, J.W., G.J. Leversee, P.P. Landrum and J.P. Giesy. 1983. Acute mortality of
anthracene-contaminated fish exposed to sundlight. Aquatic Toxicol. 3:79-90.
Brooke, L. 1991. Memorandum to Walter Berry. Summary of results of acute and chronic
exposures of fluoranthene without and with ultraviolet (UV) b'ght to various freshwater
organisms. December 3. 5 pp.
Brookes, P. 1977. Mutagenicity of polycyclic aromatic hydrocarbons. Mutation Res. 39:257-
284.
Buccafusco, R.J., S.J. Elis and G.A. LeBlanc. 1981. Acute toxicity of priority pollutants to
bluegill (Lepomis macrochirus). Bull. Environ. Contain. Toxicol. 26:446-452.
Champlin, D.M. and S. Poucher. 199 la. Acute toxicity of fluoranthene to various marine
organisms. Memorandum to D.J. Hansen. U.S. EPA. Narragansett, RI.
Champlin, D.M. and S. Poucher. 1991b. Chronic toxicity of fluoranthene to the mysid,
Mvsidopsis bahia. Memorandum to D.J. Hansen. U.S. EPA. Narragansett, RI.
Chapman, G.A. 1987. Establishing sediment criteria for chemicals-regulatory perspective. In:
Fate and Effects of Sediment-Bound Chemicals in Aquatic Systems. Editors: K.L.
Dickson, A.W. Maki and W.A. Brungs. Pergamon Press, New York. pp. 355-376.
Conover, W.J., 1980. Practical Nonparametric Statistics, Second Edition, John Wiley and
Sons, New York. 493 pp.
DeWitt, T.H, R.C. Swartz, and J.O. Lamberson. 1989. Measuring the acute toxicity of
estuarine sediments. Environ. Toxicol. Chem. 8:1035-1048.
DeWitt, T.H., R.J. Ozretich, R.C. Swartz, J.O. Lamberson, D.W. Shults, G.R. Ditsworth,
J.K.P. Jones, L. Hoselton, and L.M. Smith. (In Press). The effect of organic matter
7-1
-------
quality on the toxicity and partitioning of sediment-associated fluoranthene to the infaunal
marine amphipod, Rhepoxynius abronius. 36 pp.
Di Toro, D.M., 1985. A particle interaction model of reversible organic chemical sorption.
Chemosphere. 14(10): 1503-1538.
Di Toro, D.M., C. Zarba, D.J. Hansen, R.C. Swartz, C.E. Cowan, H.E. Allen, N.A.
Thomas, P.R. Paquin, and W.J. Berry. 1991. Technical basis for establishing sediment
quality criteria for non-ionic organic chemicals using equilibrium partitioning. Ann.
Rev. Environ. Chem. (In Press).
EG&G, Bionomics. 1979. Preliminary Research Report. Acute and chronic toxicity of
fluoranthene to mysid shrimp (Mvsidopsis bahja). Submitted to U.S. EPA. Contract No.
68-01-4646.
Heitmuller, P.T., T.A. Hollister and P.R. Parrish. 1981. Acute toxicity of 54 industrial
chemicals to sheepshead minnows (Cyprinodon variegatus). Bull. Environ. Contain.
Toxicol. 27:596-604.
Hoke, R. and G.T. Ankley. 1991. Results of dieldrin sediment spiking study conducted in
support of USEPA development of sediment quality criteria. Memorandum to D.
Hansen and D. Di Toro. June 18, 1991. 9 pp.
Home, J.D. and B.R. Oblad. 1983. Aquatic toxicity studies of six priority pollutants. Final
Report Task 0. U.S. EPA. Contract No. 68-01-6201.
Kagan, J., E.D. Kagan, I.S. Kagan, P.A. Kagan, and Susan Quigley. 1985. The phototoxicity
of Non-carcinogenic polycyclic aromatic hydrocarbons in aquatic organisms.
Chemosphere. 14(11/12): 1829-1834.
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.
Landrum, P.P., S.R. Nihart, B.J. Eadie and W.S. Gardner. 1984. Reverse-phase separation
method for determining pollutant binding to Aldrich humic acid and dissolved organic
carbon in natural waters. Environ. Sci. Technol. 18:187-192.
LeBlanc, G.A. 1980. Arnte tmririty of priority pollutants tn water flea fPaphnifl magna). Bull.
Environ. Contam. Toxicol. 24:684-691.
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.
Massey, F.J. 1951. The distribution of the maximum deviation between two sample cumulative
step functions. Annals Math. Stat. 22:125-128.
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.
7-2
-------
Newsted, J.L and J.P. Giesy. 1987. Predictive models for photoinduced acute toxicity of
polycyclic aromatic hydrocarbons to Daphnia magna. Strauss (Cladocera, Crustacea).
Environ. Toxicol. Chem. 6:445-461.
NOAA. 1991. National Status and Trends Program - Second summary of data on chemical
contaminants' in sediments from the National Status and Trends Program. NOAA
Technical Memorandum NOS OMA 59. NOAA Office of Oceanography and Marine
Assessment, Rockville, MD. 29 pp 4- appendices.
Oris, J.T., R.W. Winner, and M.V. Moore. 1991. A four-day survival and reproduction
toxicity test for Ceriodaphnia dubifl. Environ. Toxicol. Chem. 10:217-224.
Rossi, S.S. and J.M. Neff. 1978. Toxicity of polynuclear aromatic hydrocarbons to the
polycheate Neanthes arenaceodentata.. Marr. Pollut. bull. 9:220-223.
Schuytema, G.A., A.V. Nebeker, W.L. Griffis, and C.E. Miller. 1989. Effects of freezing
on toxicity of sediments contaminated with DDT and endrin. Environ. Toxicol. and
Chem. 8(10):883-891.
Stephan, C.E., D.I. Mount, D.J. Hansen, J.H. Gentile, G.A. Chapman, and W.A. Brungs.
1985. Guidelines for deriving numerical national water quality criteria for the protection
of aquatic organisms and their uses. PB85-227049. National Technical Information
Service, Springfield, VA. 98 pp.
Suedel, B.C. 1989. Sediment characteristics and bioavailability of sorbed neutral organic
compounds. M.S. Thesis. University of North Texas. 144 pp.
Swartz, R.C. 1991. Acenaphthene and penanthrene files. Memorandum to David J. Hansen,
June 26, 1991. 160pp.
Swartz, R.C., G.R. Ditsworth, D.W. Schults, and LO. Lamberson. 1985. Sediment toxicity
to a marine infaunal amphipod: Cadmium and its interaction with sewage sludge. Mar.
Envir. Res. 18:133-153.
Swartz, R.C., D.W. Schults, T.H. DeWitt, G.R. Ditsworth, and J.O. Lamberson. 1990.
Toxicity of fluoranthene in sediment to marine amphipods: A test of the equilibrium
partitioning approach to sediment quality criteria. Environ. Toxicol. Chem. 9(8): 1071-
1080.
Thursby, G.B. 1991. Near-term corrections to fluoranthene FAV, ACR and FCV.
Memorandum to David J. Hansen. December 12, 1991. Ip.
U.S. Environmental Protection Agency. 1978. In-depth studies on health and environmental
impacts of selected water pollutants. U.S. EPA. Contract No. 68-01-4646.
U.S. Environmental Protection Agency. 1980. Ambient water quality criteria for fluoranthene.
Office of Water Regulations and Standards, Criteria and Standards Division. U.S. EPA,
Washington, D.C. EPA 440/5-80-049.
U.S. Environmental Protection Agency. 1985. Appendix B - Response to public comments on
"Guidelines for deriving numerical national water quality criteria for the protection of
aquatic organisms and their uses." July 19, 1985. Fed. Regist. 50:30793-30796.
7-3
-------
U.S. Environmental Protection Agency. 1989a. Sediment classification methods compendium.
Watershed Protection Division, U.S. EPA. 280 pp.
U.S. Environmental Protection Agency. 1989b. Handbook: Water Quality Control Information
System, STORET. Washington, D.C., 20406.
U.S. Environmental Protection Agency. 1989c. Briefing Report to the EPA Science Advisory
Board on the Equilibrium Partitioning Approach to Generating Sediment Quality Criteria.
Office of Water Regulations and Standards, Criteria and Standards Division, 132 pp.
U.S. Environmental Protection Agency. 1992a. Guidelines for deriving numerical national
sediment quality criteria for non-ionic chemicals for the protection of benthic organisms.
(In Preparation).
U.S. Environmental Protection Agency. 1992b. Guidance on the application of sediment quality
criteria for the protection of aquatic life. U.S. EPA, Office of Science and Technology,
Health and Ecological Criteria Division. 33 pp.
Verschueren, K. 1983. Handbook of Environmental Data on Organic Chemicals, Second
Edition. Van Nostrand Reinhold Company, New York. 1310 pp.
Windhoiz, M., S. Budavari, R.F. Blumetti, E.S. Otterbein, eds. 1983. The Merck Index, An
Encyclopedia of Chemicals, Drugs, and Biologicals, Tenth Edition. Merck & Co. Inc.,
Rahway, NJ. 1463 pp + appendices.
7-4
-------
Appendix A. - Fluoranthene (cont'd).
COMMON/SCI. NAME
Sheepshead minnow
Cvprinodon varieoatus
Inland silverside,
Menidia bervllina
LIFE* HAB-b CONCEN-
STAQB ITAT METHOD0 TRATION*
J B,W
J B,W
J W
J W
J W
S,UV
S.SL
S.D
S,UV
S,SL
U
U
U
U
U
LC50/e
HMAV
OVERALL.*1
1 BC50 SPECIES* GENUS9 GMAV
pg/L
158.8
172
616.6
30.29
13.05
pg/L
158.8
172
616.6
30.29
13.05
j»g/L
158.8
172
616.6
30.29
13.05
Hg/L
158.8
172
616.6
30.29
13.05
REFERENCES
Champlin and Poucher, 1991
Champlin and Poucher, 1991
*Lifestage: A - adult, J » juvenile, L ° larvae, B = embryo, U = lifeatage and habitat unknown, X = lifeatage unknown
but habitat known.
^Habitat: I - infauna, B - epibenthic, W » water column
> "Method: FT = flow-through
w S = static
UV = ultraviolet (UVB = 98 - 934pW/cm )
D = dark (UVB = 5/iW/cm2)
^Concentration: M = chemical measured U = chemical unmeasured
*Acute value: 96-hour LC50 or BC50
£HMAV species: Habitat Mean Acute Value - Species is the geometric mean of acute values by species for benthic and
water column lifestages.
9HMAV genus: Geometric mean of HMAV for species within a genus.
hOverall GMAV: Geometric mean of acute values across species, habitats and lifestages within the genus.
iNot used in calculation of species mean acute value.
-------
Appendix A. - Fluoranthene: Summary of acute values for freshwater and saltwater species.
COMMON/SCI. NAME
LIFE*
STAGS
HAB-b CONCEN- LC50/e
ITAT METHOD0 TRATION** EC50 i
HMAV
SPECIES'
ftg/lt M9/k
GENUS9
pg/L
OVERALL*1
GMAV REFERENCED
*9/L
FRESHWATER SPECIES
Hydra,
Hydra americana
Annelid,
Lumbriculua varieoatus
Snail,
Phvsa heterostrooha
Snail.
p, Phvsella virgata
H
Cladoceran,
Daphnia magna
Amphipod,
Gamma rus pseudol imneaus
Amphipod,
Gamma rua minus
Stonefly,
Peltooerla rearia
Dragonfly,
Qphiogomphus sp.
Rainbow trout.
OncorhviichuB mykiss
J
J
A
A
X
A
A
J
J
J
J
A
A
X
N
N
X
J
J
B
B
I
I
B
B
B
W
W
W
W
B
B
E
B
B
W
W
W
FT.D
FT.UV
FT.D
FT.UV
S,D
FT.D
FT.UV
S,D
S,D
S,D
S,UV
FT.D
S.D
S,D
FT,D
FT.UV
S.D
FT.D
FT.UV
M
M
M
M
U
M
M
M
M
O
M
M
U
0
M
M
M
M
M
70.06
2.16
>178.5
1.15
137.0
>178.5
82.0
45.0
102.8
320, OOO1
0.97
>116.6
32.0
135.0
>178.5
>109.7
187.0
>90.5
7.70
70.
2.
>178
1.
137
>178
82
-
-
68
0.
-
32
135
>178
>109
.06
.16
.5
15
.0
.5
.0
.0
97
.0
.0
.5
.7
70.06
2.16
>178.5
1.15
137.0
>178.5
82.0
-
-
68.0
0.97
-
32.0
135.0
>178.5
>109.7
70.06
2.16
>178.5
1.15
137.0
>178.5
82.0
-
-
68.0
0.97
-
32.0
135.0
>178.5
>109.7
Brooke ,
Brooke ,
1991
1991
Home and Oblad, 1983
Brooke ,
Oris et
Brooke ,
1991
al., 1991
1991
Le Blanc, 1980
Brooke ,
Brooke ,
1991
1991
Home and Oblad, 1983
Home and Oblad, 1983
Brooke,
1991
Home and Oblad, 1983,
>90
.5
7.70
>90.S
7.70
>90.5
7.70
Brooke ,
Brooke ,
1991
1991
-------
Appendix A. - Fluoranthene (cont'd).
COMMON/SCI. NAME
Fathead minnow.
pimaohalea promelaa
Bluegill,
Lapomia macrochirua
LIFE*
STAGB
J
J
L
A
J
J
HAB-b
ITAT
W
N
W
W
W
W
METHOD0
S,D
FT,D
PT.UV
FT.D
FT,D
S,D
CONCBN- LC50/4
TRATION*1 BC50
M
M
M
U
M
O
pg/L
95.0
>211.7
2.59
>1000
>116.6
4000
ro
SPECIES1
/ig/L
IAV
GENUS9
M9/L
OVERALL11
GMAV
jig/L
REFERENCES
Home and Oblad, 1983
>211.7
2.59
>1000
-
>211.7
2.59
>1000
-
>211.7
2.59
>1000
-
Brooke, 1991
Brooke, 1991
Birge et al .
Brooke, 1991
, 1982
*
>116.6 >116.6 >116.6 Buccafuaco et al . , 1981
BPA, 1978
SALTWATER SPECIES
Annelid worm.
Neanthea arenaceodentata
Coot clam,
Mulinia lateralia
Mysid,
Mysidoosis bahia
Amphipod,
Amoeliaca abdita
Grass shrimp,
PaleomonetfiB ouaio
Sea urchin,
Arbacia punctulata
Sheepshedd minnow,
Cvprinodon varieoatuB
J
J
J
J
J
J
J
J
J
J
J
J
J
B
B
J
J
I
I
B
B
B
B
B
B
B
B
'I
B
B
W
W
B,W
B,W
S,D
S.D
S.D
S.D
S.D
FT,D
FT.D
S.UV
FT.OV
S.SL
S.D
S.D
S.OV
S.D
S.UV
S.D
S.D
D
O
D
O
U
M
M
D
M
U
D
U
U
O
O
U
U
>20,000
500
10,710
58.41
40
30.53
87
2.76
0.58
1.6
66.93
142.5
21.55
>20,000
3.89
>20,000
>S60,000
-
500
10,710
-
-
-
51.54
-
0.58
1.6
66.93
142.5
21.55
>20.000
3.89
-
>20,000
-
500
10.710
-
-
-
51.54
-
0.58
1.6
66.93
142.5
21.55
>20,000
3.89
-
>20,000
.
500
10,710
-
-
-
51.54
-
0.58
1.6
66.93
142.5
21.55
>20,000
3.89
-
>20,000
Champlin and
Poucher,
1991
Roasi and Neff , 1978
Champlin and
Champlin and
BPA, 1978
Champlin and
Poucher,
Poucher,
Poucher,
1991
1991
1991
EG & G, Bionomics, 1978
Champlin and
Champlin and
Champlin and
Champlin and
Champlin and
Heitmuller et
Poucher,
Poucher,
Poucher,
Poucher,
Poucher,
1991
1991
1991
1991
1991
al., 1981
EPA, 1978
-------
APPENDIX B: THE OCTANOL-WATER PARTITION COEFFICIENT. KQW FOR
FLUORANTHENE.
B.I GENERAL INFORMATION:
Partitioning between water and natural soils, sediments, and aquifer materials is an
important process affecting transformation rates, toxicity, and the ultimate disposition of organic
chemicals in the environment. Extensive research, focusing on the partitioning of neutral
organic compounds, has shown that adsorption of these compounds generally is controlled by
hydrophobic interactions. As a result, the affinity that a natural sorbent has for neutral organic
solutes, in most cases, can be reliably estimated from the hydrophobicity of the solute and the
sorptive capacity of the sorbent. Organic carbon content has been used almost exclusively as
a measure of the sorptive capacity of natural sedimentary material. (Organic matter or volatile
solids content has also been used but not as widely.) To quantitatively characterize the
hydrophobic nature of organic compounds, researchers have used various measurable
parameters, including octanol/water partition coefficients (KoW)> water solubility (corrected for
crystal energy), reverse phase HPLC retention, and topological parameters of the compounds
such as calculated surface area. Generally, octanol/water partition coefficients have been used
more extensively, not only for estimating the partitioning of organic compounds to sedimentary
materials, but also for estimating bioaccumulation of organic compounds to aquatic organisms.
The KOW is defined as the ratio of the equilibrium concentration of a dissolved substance
in a system consisting of n-octanol and water and is ideally dependent only on temperature and
pressure:
KOW = ewe* (B-i)
where Cocr is the concentration of the substance in n-octanol and Cw is the concentration of the
substance in water. The K^ is used in estimating the organic-carbon- normalized sediment-
water partition coefficient (KoJ and is frequently reported in the form of its logarithm to base
ten as log P.
B-l
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B.2 LITERATURE DATA:
An extensive literature search was performed for fluoranthene and two standard reference
compounds, biphenyl and pyrene. Generally, problems encountered in compiling and reporting
fate constants from published data and from databases during the last several years have ranged
from retrieval of misquoted numbers to resolution of nested citations (Kollig, 1988). Some
citations were three or more authors removed from the original work or contained data that
were referenced as unpublished data or as personal communication. The same problems were
experienced during this literature search. The largest difference in misquoting numbers was six
orders of magnitude. For these reasons, ERL-Athens obtains data from the primary sources
and releases values coming only from these primary sources. Unpublished data or data which
originated through personal communication are rejected as well as data that are insufficiently
documented to determine their credibility and applicability or reliability.
Tables B-l and B-2 show the measured and estimated K^ values, respectively, retrieved
by this literature search. Each of the measured values was experimentally determined by the
researcher using one of several laboratory methods. The individual experimental methods are
not identified here. The estimated literature values were computed by the researchers by one
of several published techniques. The individual computational techniques also are not identified
here.
B.3 ERL-ATHENS MEASURED DATA:
To enhance confidence in the measured KQW values, three independent experimental
methods, [shake-centrifugation (SQ, generator column (GCol), slow-stir-flask (SSF)], were used
to determine a KOV value for fluoranthene at the U.S. EPA laboratory at Athens, Georgia. The
SC method is routinely used to measure the partitioning of compounds with KOW values on the
order of Itf to 10*. The method involves adding a layer of octanol containing the compound
of interest onto the surface of water contained in a centrifuge tube. Both phases are mutually
B-2
-------
presaturated before beginning the measurements. Equilibration is established by gentle agitation
and any emulsions formed are broken by centrifugation. The concentration in each phase is
determined usually by a chromatographic method and the KGW value calculated using Equation
B-l.
TABLE B-l. MEASURED LOGloKoW VALUES FOUND IN THE LITERATURE
Chemical
LogloKow value
Reference
rluoranthene 5.155
Siphenyl 3. 16
3.63
3.75
3.76
3.79
3.89
4.008
4.01
4.04
4.09
4.10
Pyrene 4.96
5.05
5.09
5.18
5.22
5.52
DeBruijnetal., 1989
Rogers and Cammarata, 1969
De Kock and Lord, 1987
Veithetal., 1979
Miller etal., 1984
Rapaport and Eisenreich, 1984
Woodbura et al., 1984
De Bruijn et al., 1989
Eadsforth, 1986
Banerjee etal., 1980
Ellington and Stancil, 1988
Bruggeman et al., 1982
Rapapaport and Eisenreich, 1984
Ellington and Stancil, 1988
Means etal., 1980
Karickhoff etal., 1979
Bruggeman et al., 1982
BurkhaidetaL, 1985
The original GCol method, limited to compounds with K^ values of less than 10"1 was
modified (Woodburn et al., 1984) and used to determine K<,w values up to 10*. Briefly, the
method requires the packing of a 24-cm length of tubing with silanized Chromosorb W.
B-3
-------
Octanol, containing the chemical in a known concentration, is then pulled through the dry
support by gentle suction until the octanol appears at the exit of the column. Water is then
pumped through the column at a rate of less than 2 ml per minute to allow equilibration of the
chemical between the octanol and water. The first 100 ml are discarded followed by collection
of an amount of water sufficient to determine the chemical concentration. The KQW is calculated
using Equation B-l.
The slow-stir-flask method (De Bruijn et al., 1989) achieves equilibrium of the compound
between octanol and water by gentle stirring of the phases contained in a six-liter flask. One-
liter aliquots of the phase are withdrawn at two-day intervals for the determination of the
concentration of the chemical. Equilibrium is considered to be established when the
concentration of the chemical is constant in successive samples (usually 2-6 days). Briefly, we
set up three six-liter flasks in a constant temperature room. Five liters of waqter are added to
each flask. The water is stirred with teflon-coated magnetic stir ban overnight to achieve
temperature equilibrium. Temperature equilibrated octanol, containing compound, is added very
gently along the side wall to avoid mixing of two phases. At the same time of sampling, a one
liter aqueous sample is drained from a sampling port at the base of each flask without disturbing
the octanol layer. The concentration in each phase is determined, usually, by chromatographic
method and the Kg* value calculated using equation B-l.
When repetitive measurements are made in the Athens laboratory, a protocol is established
to assure compatibility with future experiments. These protocols describe the entire
experimental scheme including planning, sample requirements, experimental set up and chemical
analysis, handling of data, and quality assurance. Only established analytical methods for solute
concentration measurement are applied and the purity and identity of the chemical are
determined by spectroscopic means. The name on the label of the chemical's container is not
proof of the identity.
B-4
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TABLE B-2. ESTIMATED LOGtoKoW VALUES FOUND IN THE LITERATURE
Chemical
LogloKoW value
Reference
Fluoranthene
Biphenyl
Pyrene
4.90
4.95
5.22
5.29
5.33
3.79
3.95
3.98
4.14
4.25
4.42
4.50
4.85
4.88
4.90
5.12
5.22
5.32
Mabey et al., 1982
U.S. EPA, Graphical Exposure
Modeling System (GEMS)'
Yalkowsky et al., 1983
Mackay etal., 1980
Callahanetal., 1979
Yalkowsky etal., 1983
Miller etal., 1985
Kamlet et al., 1988
Mackay .etal., 1980
Aibuckle, 1983
Doucette and Andren, 1987
D'Amboise and Hanai, 1982
Kamlet et al., 1988
Lyman et al., 1982
Mabey et al., 1982
Mackay etal., 1980
Yalkowsky et al., 1983
Callahan et al., 1979
"The Graphical Exposure Modeling System (GEMS) 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.
Standard reference compounds (SRCs) are tested with each experiment. SRCs are
compounds that are used as quality assurance standards and as references in inter-laboratory
generation of data. The value of the process constant(s) has been established by repetitive
measurements for an SRC and serves as baseline information for evaluating all experimental
techniques and all aspects of quality assurance. Because the SRC is taken through the entire
experimental scheme, its acceptable result will assure the experimenter that equipment and
measurement methods are functioning satisfactorily. Table B-3 shows the KQW values for
fluoranthene and the SRCs, biphenyl and pyrene, measured at the Athens laboratory by the
SC methods. The SRCs were not measured by the GCol method.
B-5
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TABLE B-3. LOGIOKoW VALUES MEASURED BY SHAKE-
CENTRIFUGATION (SQ AND GENERATOR COLUMN
(GCOL) AND SLOW-STIR-FLASK (SSF) FOR FLUORANTHENE
AND CONCURRENTLY ANALYZED STANDARD REFERENCE
COMPOUNDS.
Chemical SC GCol SSF
Fluoranthene 5.00 5.09 5.39
Biphenyl 4.06
Pyrene 5.17
The logw of the average of eight previous measurments of KQ* by the shake-centrifugation
method for biphenyl is 4.09. The logloKoW of the average of thirteen previous measurements
by the shake-centrifugation method for pyrene is 5.05. These are in good agreement with the
SQC shake-centrifugation measurements made concurrently with the fluoranthene measurements.
B.4 ESTIMATED DATA:
A promising new computational method for predicting chemical reactivity is the computer
expert system SPARC (SPARC Performs Automated Reasoning in Chemistry) being developed
by Samuel W. Karickhoff, at ERL-Athens, and other scientists at the University of Georgia
(Karickhoff et al., 1989). The system has the capability of crossing chemical boundaries to
cover all organic chemicals and uses algorithms based on fundamental chemical structural theory
to estimate parameters. Organic chemists have in the past established the types of structural
groups or atomic arrays that impart certain types of reactivity and have described, in
"mechanistic" terms, the effects on reactivity of other structural constituents appended to the
site of reaction. To encode this knowledge base, Karickhoff and his associates developed a
classification scheme that defines the role of structural constituents in affecting or modifying
reactivity. SPARC quantifies reactivity by classifying molecular structures and selecting
appropriate "mechanistic" models. It uses an approach that combines principles of quantitative
B-6
-------
structure-activity relationships, linear free energy theory (LFET), and perturbed molecular
orbital (PMO) or quantum chemistry theory. In general, SPARC utilizes LFET to compute
thermal properties and PMO theory to describe quantum effects such as delocalization energies
or polarizabilities of pi electrons.
SPARC computes KQW values from activity coefficients in the octanol( - y and water (~!»)
phases using Equation B-2.
LogioKo* = loglo (- V-lo) + log,0 (Mo/Mw) (B-2)
where MQ and Mw are solvent molecularities of octanol and water, respectively. SPARC
computes activity coefficients for any solvent/solute pair for which the structure parser can
process the structure codes. Ultimately, any solvent/solute combination can be addressed. New
solvents can be added as easily as solutes by simply providing a Simplified Molecular
Interactive Linear Entry System (SMILES) string (Anderson et al., 1987, Weiniager, 1988).
Activity coefficients for either solvent or solute are computed by solvation models that are built
from structural constituents requiring no data besides the structures.
A goal for SPARC is to compute a value that is as accurate as a value obtained
experimentally for a fraction of the cost required to measure it. Because SPARC does not
depend on laboratory operations conducted on compounds with structures closely related to that
of the solute of interest, it does not have the inherent problems of phase separation encountered
in measuring highly hydrophobic compounds (logwKov > 5). For these compounds, SPARC'S
computed value should, therefore, be more reliable than a measured one. Reliable experimental
data with good documentation are still necessary, however, for further testing and validation
of SPARC.
CLOGP (Chou and Jurs, 1979) is a computerized program that estimates the logloKo«.
based on Leo's Fragment Constant Method (Lyman et al., 1982). CLOGP provides an estimate
of logioKov using fragment constants (fj) and structural factors (FJ that have been empirically
derived for many molecular groups. The estimated loguKw is obtained from the sum of
B-7
-------
constants and factors for each of the molecular subgroups comprising the molecule using
Equation B-3.
n
Log,oKow = £ (fi + Fi) (B-3)
i = 1
The method assumes that log10KoV is a linear additive function of the structure of the solute
and its constituent parts and that the most important structural effects are described by available
factors. The structure of the compound is specified using the SMILES notation. The CLOGP
algorithm is included in the database QSAR1 located at EPA's Environmental Research
Laboratory at Duluth, Minnesota. All CLOGP values reported here were obtained through
QSAR.
Table B-4 shows the estimated log,,,!^ values that were computed with SPARC and
CLOGP.
'Quantitative Structure-Activity Relationships (QSAR) is an interactive chemical database
and hazard assessment system designed to provide basic information for the evaluation of the
fate and effects of chemicals in the environment. QSAR was developed jointly by the U.S.
EPA Environmental Research Laboratory, Duluth, Minnesota, Montana State University Center
for Data System and Analysis, and the Pomona College Medicinal Chemistry Project.
TABLE B-4. LOGuKow VALUES ESTIMATED BY SPARC AND CLOGP
Chemical
Fluoranthene
Biphenyl
Pyrene
SPARC
5.21
4.25
5.13
CLOGP
4.95
4.03
4.95
B-8
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(APPENDIX B)
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B-9
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Eadsfotth, C.V. 1986. Application of reverse-phase h.p.l.c. for the determination of partition
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B-10
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B-ll
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