EPA/600/R-02/012 | December 2012
www.epa.gov/ord
United States
Environmental Protection
Agency
Equilibrium Partitioning Sediment
Benchmarks (ESBs) for the
Protection of Benthic Organisms:
Procedures for the Determination of
the Freely Dissolved Interstitial Water
Concentrations of Nonionic Organics
1-:.
Office of Research and Development
National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division
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EPA/600/R-02/012 | December 2012
www.epa.gov/ord
Equilibrium Partitioning Sediment Benchmarks (ESBs) for the
Protection of Benthic Organisms: Procedures for the
Determination of the Freely Dissolved Interstitial Water
Concentrations of Nonionic Organics
Robert M. Burgess
National Health and Environmental Effects Research Laboratory
Atlantic Ecology Division
Narragansett, RI
Susan B. Kane Driscoll
Exponent, Inc.
Maynard, MA
Robert J. Ozretich
National Health and Environmental Effects Research Laboratory
Western Ecology Division
Corvallis, OR
David R. Mount
National Health and Environmental Effects Research Laboratory
Mid-Continent Ecology Division
Duluth, MN
Mary C. Reiley
Office of Water
Washington, DC
U.S. Environmental Protection Agency
Office of Research and Development
National Health and Environmental Effects Research Laboratory
Atlantic Ecology Division, Narragansett, RI 02882
Western Ecology Division, Corvallis, OR 97333
Mid-Continent Ecology Division, Duluth, MN 55804
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Notice
Notice
The Office of Research and Development (ORD) has produced this ESB document to provide
procedures for the determination of the freely dissolved concentrations of nonionic organic chemicals
for deriving sediment interstitial water toxic units (IWTUs). ESBs may be useful as a complement to
existing sediment assessment tools. This document should be cited as:
U.S. EPA. 2012. Equilibrium Partitioning Sediment Benchmarks (ESBs) for the Protection of
Benthic Organisms: Procedures for the Determination of the Freely Dissolved Interstitial
Water Concentrations of Nonionic Organics. EPA-600-R-02-012. Office of Research and
Development, Washington, DC 20460
This document, and the other ESB documents, can also be found in electronic format at the
following web address:
http://www.epa.gov/nheerl/publications.html
The information in this document has been funded wholly by the U.S. Environmental Protection
Agency.
It has been subject to the Agency's peer and administrative review, and it has been approved for
publication as an EPA document. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Abstract
This document describes procedures to determine the concentrations of nonionic organic
chemicals in sediment interstitial waters. In previous ESB documents, the general equilibrium
partitioning (EqP) approach was chosen for the derivation of sediment benchmarks because it
accounts for the varying bioavailability of chemicals in different sediments and allows for the
incorporation of the appropriate biological effects concentration. This provides for the derivation of
benchmarks that are causally linked to the specific chemical, applicable across sediments, and
appropriately protective of benthic organisms.
In contrast to the previous ESB documents, the emphasis of this ESB document is to provide a
summary of procedures for determining the freely dissolved concentrations of nonionic organic
chemicals for deriving sediment interstitial water toxic units (IWTUs). In the last ten years,
technologies have been developed allowing for the accurate estimation and measurement of the
concentrations of nonionic organic chemicals in sediment interstitial waters. When the general EqP
model (i.e., one-carbon model) was first proposed for deriving ESBs, methods for directly measuring
interstitial water concentrations of nonionic organic chemicals were often overly technically difficult,
cost prohibitive, or simply not available. The procedures described here are an alternative or
complement to using the one-carbon general model for deriving ESBs. The one-carbon general model
estimates the bioavailability of nonionic organic contaminants based on their measured sediment
concentrations and sediment organic carbon content. The new technologies and resulting procedures
described in this document include a two-carbon model incorporating black carbon along with natural
organic carbon for making EqP-based predictions, direct measurements of interstitial water
contaminants adjusted for dissolved organic carbon, and passive samplers to measure interstitial
water concentrations directly or via the sediment. These procedures allow for the more accurate
determination of the freely dissolved and potentially bioavailable concentrations of nonionic organic
chemicals. These concentrations along with the final chronic values (FCVs), secondary chronic
values (SCVs), or other relevant water-only toxicity values are used to derive IWTUs. Depending
upon the toxicological endpoint, if the IWTUs are greater than one, benthic organisms may not be
protected and adverse effects may result.
This document is not intended as a methods manual but rather provides an overview of
procedures for determining freely dissolved concentrations of nonionic organic chemicals.
Throughout this document, the scientific literature cited provides greater methodological detail.
ESB documents have been developed for two pesticides (endrin, dieldrin), polycyclic aromatic
hydrocarbon (PAH) mixtures, metal mixtures, and a selection of 32 nonionic organic chemicals.
The ESBs do not intrinsically consider the antagonistic, additive or synergistic effects of other
sediment contaminants in combination with the individual nonionic organic chemicals discussed in
this document or the potential for bioaccumulation and trophic transfer of these chemicals to aquatic
life, wildlife or humans. However, for narcotic chemicals, additivity can be used to sum toxic effects.
IV
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Foreword
Foreword
Under the Clean Water Act (CWA), the U.S. Environmental Protection Agency (EPA) and the
States develop programs for protecting the chemical, physical, and biological integrity of the Nation's
waters. To support the scientific and technical foundations of the programs, EPA's Office of
Research and Development has conducted efforts to develop and publish equilibrium partitioning
sediment benchmarks (ESBs) for some of the 65 toxic pollutants or toxic pollutant categories. Toxic
contaminants in bottom sediments of the Nation's lakes, rivers, wetlands, and coastal waters create
the potential for continued environmental degradation, even where water column contaminant levels
meet applicable water quality standards. In addition, contaminated sediments can lead to water
quality impacts, even when direct discharges to the receiving water have ceased.
The ESBs and associated methodologies presented in this document provide a means to estimate
the concentrations of a substance that may be present in sediment while still protecting benthic
organisms from the effects of that substance. These benchmarks are applicable to a variety of
freshwater and marine sediments because they are based on the biologically available concentration
of the substance in the sediments. These ESBs are intended to provide protection to benthic
organisms from direct toxicity resulting from this substance under site conditions. The ESBs do not
intrinsically consider the antagonistic, additive, or synergistic effects of other sediment contaminants
in combination with the nonionic organic chemicals discussed in this document or the potential for
bioaccumulation and trophic transfer of these chemicals to aquatic life, wildlife, or humans.
However, in some cases, the additive toxicity for specific classes of toxicants (e.g., poly cyclic
aromatic hydrocarbon mixtures and other narcotic organic chemical) is addressed.
ESBs may be useful as a complement to existing sediment assessment tools, to help evaluate the
extent of sediment contamination, to identify chemicals causing toxicity, and to serve as targets for
pollutant loading control measures. This document provides technical information to EPA Program
Offices, including Superfund, Regions, States, the regulated community, and the public. Decisions
about risk management are the purview of individual regulatory programs, and may vary across
programs depending upon the regulatory authority and goals of the program. For this reason, each
program will have to decide whether the ESB approach is appropriate to that program and, if so, how
best to incorporate this technical information into the assessment process. While it was necessary to
choose specific parameters for the purposes of this document and other ESB documents, it is
important to realize that the basic science underlying this document can be adapted to a range of risk
management goals by adjusting the input parameters. At the same time, the ESBs do not substitute
for the CWA or other EPA regulations, nor are they regulation. Thus, they cannot impose legally
binding requirements on EPA, States, or the regulated community. EPA and State decision makers
retain the discretion to adopt approaches on a case-by-case basis that differ from this technical
information where appropriate. It is recommended that the ESBs not be used alone but with other
sediment assessment methods to make informed management decisions based on a weight of
evidence approach. EPA may change this technical information in the future. This document has been
reviewed by EPA's Office of Research and Development (Atlantic Ecology Division, Narragansett,
Rhode Island), undergone an external peer review, and has been approved for publication.
This is contribution AED-02-049 of the Office of Research and Development National Health and
Environmental Effects Research Laboratory's Atlantic Ecology Division.
Front cover image provided by Wayne R. Davis and Virginia Lee.
v
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Contents
Contents
Notice iii
Abstract iv
Foreword v
Tables ix
Figures ix
Acknowledgements x
Executive Summary xi
Glossary of Abbreviations xiii
Section 1
Introduction
1.1 General Information 1-1
1.2 Review of the General Equilibrium Partitioning Approach 1-1
1.3 Rationale for Development of Procedures for Determining
Freely Dissolved Concentrations 1-2
1.4 Freely Dissolved Concentration Procedures 1-2
1.5 Data Quality and Uncertainties 1-3
1.6 Overview 1-3
Section 2
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
2.1 Introduction 2-1
2.1.1 Rationale 2-1
2.2 Using a Two-Carbon Model for Determining Freely Dissolved
Interstitial Water Concentrations 2-3
2.2.1 Two-Carbon Model 2-4
2.2.2 Estimation of KBC 2-4
2.3 Direct Measurement of Interstitial Water Concentrations 2-5
2.3.1 Direct Collection of Interstitial Water by Centrifugation 2-5
2.3.2 Calculating the Freely-Dissolved Concentration 2-6
2.4 Use of Passive Samplers for Determining Freely Dissolved
Interstitial Water Concentrations 2-7
2.4.1 Types of Passive Samplers 2-8
2.4.2 Procedures for Whole Sediments 2-10
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
2.4.2.1 Calculation of Freely Dissolved Concentrations
using Passive Samplers 2-10
2.4.3 Procedures for Interstitial Waters 2-11
2.5 Derivation of Interstitial Water Toxic Units 2-12
Section 3
Example Calculations of ESBTUFcv and IWTUFcv
3.1 Introduction 3-1
3.2 Estimates of Freely Dissolved Contaminants in Sediment Interstitial Water 3-1
3.2.1 One-Carbon Model 3-1
3.2.2 Two-Carbon Model 3-2
3.3 Measurement of Freely Dissolved Contaminants in Sediment Interstitial Water 3-3
3.3.1 Direct Measurement of Interstitial Water 3-3
3.3.2 Passive Sampling of Interstitial Water 3-4
3.4 Considerations for Non-Planar Contaminants 3-4
3.5 Summary 3-5
Section 4
Implementation of Freely Dissolved Interstitial Water Concentrations
4.1 Introduction 4-1
4.2 Implementation of Freely Dissolved Concentrations 4-2
4.3 Research Needs 4-3
Section 5
References 5-1
Vlll
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Contents
Tables
Table 2-1. Provisional partition coefficients for selected nonionic organic contaminants 2-13
Table 2-2. Literature and calculated partition coefficients 2-15
Table 2-3. Solutions to Equation 2-7 using KDOC values calculated from Equation 2-8 2-15
Table 2-4. Advantages and disadvantages of selected approaches for determining 2-16
Table 3-1. Example calculations of ESBTUpcv and IWTUpcv for PAH mixtures: 3-6
Table 3-2. Example calculations of ESBTUpcv and IWTUpcv for PAH mixtures: 3-11
Figures
Figure 2-1. Magnified and exploded view of different types of sediment particle 2-2
Figure 2-2. Photographs of selected passive samplers, including SPME, PE, and POM 2-8
Figure 4-1. Schematic of proposed tiered approach for implementing the use 4-3
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Acknowledgements
Coauthors
Robert M. Burgess*'** U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Susan B. Kane Driscoll Exponent, Inc., Maynard, MA
Robert J. Ozretich U. S. EPA, NHEERL, Western Ecology Division, Corvallis, OR
David R. Mount U.S. EPA, NHEERL, Mid-Continent Ecology Division, Duluth, MN
Mary C. Reiley U.S. EPA, Office of Water, Washington, DC
Significant Contributors to the Development of the Approach and Supporting Science
Dominic M. Di Toro University of Delaware, Newark, DE; HydroQual, Inc., Mahwah, NJ
David J. Hansen formerly with U.S. EPA
Monique M. Perron National Research Council, U.S. EPA, NHEERL, Atlantic Ecology Division,
Narragansett, RI
Christopher S. Zarba U.S. EPA, Office of Research and Development, Washington, DC
Technical Support and Document Review
Sungwoo Ahn Exponent, Inc., Bellevue, WA
Lawrence Burkhard U.S. EPA, NHEERL, Mid-Continent Ecology Division, Duluth, MN
Patricia DeCastro SRA International Inc., Narragansett, RI
Upal Ghosh University of Maryland, Baltimore County, Baltimore, MD
Kay Ho U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Joseph LiVolsi U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Keith Maruya Southern California Coastal Water Research Project, Costa Mesa, CA
Wayne Munns U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Thomas Parkerton Exxon Mobil Biomedical Sciences, Inc., Annandale, NJ
Monique Perron U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Jaana Pietari Exponent, Inc., Maynard, MA
Lisa Portis U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Richard Pruell U.S. EPA, NHEERL, Atlantic Ecology Division, Narragansett, RI
Danny Reible University of Texas, Austin, TX
*Principal U.S. EPA contact
** Series Editor
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Executive Summary
Executive Summary
The purpose of this document is to provide guidance on procedures to determine the freely
dissolved concentrations of nonionic organic chemicals in sediment interstitial waters. These data
when combined with FCVs, SCVs or other relevant water-only toxicity data can be used to derive
interstitial water toxic units (IWTUs). This methodology is issued in support of the published ESBs
for endrin and dieldrin (U.S. EPA, 2003b,c), PAHs mixtures (U.S. EPA, 2003d), and other nonionic
organic chemicals (U.S. EPA, 2008). The procedures used to determine the freely dissolved
concentrations of nonionic organic chemicals are intended to supplement the procedures described
for calculated ESBs based on the general equilibrium partitioning (EqP) theory as described in the
ESB Technical Basis Document (U.S. EPA, 2003a).
The EqP approach was chosen because it accounts for the varying biological availability of
chemicals in different sediments and allows for the incorporation of the appropriate biological effects
concentration (Di Toro et al., 1991; U.S. EPA, 2003a). This provides for the derivation of
benchmarks that are causally linked to the specific chemical, applicable across sediments, and
appropriately protective of benthic organisms.
General EqP theory holds that a nonionic chemical in sediment partitions between sediment
organic carbon, interstitial (pore) water, and benthic organisms. At equilibrium, if the concentration
in any one phase is known, then the concentrations in the others can be predicted. The ratio of the
concentration in water to the concentration in organic carbon is termed the organic carbon-water
partition coefficient (K0c), which is expected to be a constant for each chemical. The ESB Technical
Basis Document (U.S. EPA, 2003a) demonstrates that biological responses of benthic organisms to
nonionic organic chemicals in sediments are different across sediments when the sediment
concentrations are expressed on a dry weight basis, but similar when expressed on a jig chemical/g
organic carbon basis (jig/goc). Similar responses were also observed across sediments when
interstitial water concentrations were used to normalize biological availability. The Technical Basis
Document (U.S. EPA, 2003a) further demonstrates that if the effect concentration in water is known,
the effect concentration in sediments on a |ig/goc basis can be accurately predicted by multiplying the
effect concentration in water by the chemical's K0c-
The U.S. Environmental Protection Agency (EPA) currently recognizes that the ESBs may be
under- or overprotective when differences occur in the bioavailability of the chemical in the site
sediment because of alternate partitioning phases (e.g., black carbon). In such cases, the
bioavailability of chemicals can be influenced by the site-specific partitioning behavior of sediment
carbon that may be substantially different than for typical diagenic organic carbon. The procedures
described in this document assume that the true concentration of bioavailable chemical can be
reasonably measured or estimated from the concentration of freely dissolved chemical in interstitial
water. This assumption does not imply that exposure occurs only from interstitial water, rather that
the freely dissolved concentration of NOCs in interstitial water is a better surrogate than the bulk
concentration for the fraction of chemical in the sediment that is available to partition into interstitial
water and into organisms. In the last ten years, technologies have been developed allowing for the
accurate estimation and measurement of the concentrations of nonionic organic chemicals in
XI
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
sediment interstitial waters. When the EqP model was first proposed for deriving ESBs, methods for
directly measuring interstitial water concentrations of nonionic organic chemicals were often overly
technically difficult, cost prohibitive, or simply not available. This document includes examples that
demonstrate the calculation of interstitial water toxic units using various approaches including: a
"two-carbon" model that estimates the concentrations of chemical in interstitial water by taking into
account the influence of black carbon, direct measurement of chemical in isolated samples of
interstitial water, and deploying passive samplers in interstitial water and whole sediment. This
document concludes with a proposed tiered implementation framework that may be useful in a
weight of evidence application of this guidance.
The ESBs, based on the one-carbon general model, can be used to calculate ESBs for any toxicity
endpoint for which there are water-only toxicity data; it is not limited to any single effect endpoint.
ESBs have been calculated using FCVs from water quality criteria (U.S. EPA, 2003b,c), SCVs
derived from existing toxicological data (U.S. EPA, 2008), and from narcosis theory (U.S. EPA,
2003d). The FCVs, SCVs and other relevant water-only toxicity data can be used to derive interstitial
water toxic units (IWTUs).
These values are intended to be the concentration of each chemical in water that is protective of
the presence of aquatic life. The ESBs should be interpreted as a chemical concentration below which
adverse effects are not expected. At concentrations above the ESB (i.e., > 1.0 toxic unit), assuming
equilibrium between phases, effects may occur with increasing severity as the degree of exceedance
increases. This document is intended to provide guidance for determining the freely dissolved
interstitial water concentrations of NOCs for deriving IWTUs. The document is not intended to be a
methods manual; whenever possible, relevant scientific literature is cited that provides greater
methodological detail. Further, especially for the passive samplers, as they are used more frequently,
standardized manuals for the procedures discussed here are likely to be available in the near future. A
sediment-specific site assessment (e.g., toxicity testing) would provide further information on
bioavailability and the expectation of toxicity relative to the ESB along with associated uncertainties.
The procedures in this document are intended to complement such sediment-specific assessments. In
general, the ESBs apply only to sediments having > 0.2% total organic carbon by dry weight and
nonionic organic chemicals with log KOWS > 2.
The ESBs do not intrinsically consider the antagonistic, additive, or synergistic effects of other
sediment contaminants in combination with the nonionic organic chemicals discussed in this
document or the potential for bioaccumulation and trophic transfer of these chemicals to aquatic life,
wildlife, or humans. However, for narcotic chemicals, ESB values may be used in a framework to
evaluate the potential effects of chemical mixtures. Consistent with the recommendations of EPA's
Science Advisory Board, publication of these documents does not imply the use of ESBs as stand-
alone, pass-fail criteria for all applications; rather, ESB exceedances could be used to trigger the
collection of additional assessment data. Similarly, ESBs are supportive of recommendations by
Wenning et al. (2005) to apply a weight of evidence approach when evaluating contaminated
sediments.
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Glossary of Abbreviations
Glossary of Abbreviations
ASTM American Society for Testing and Materials
BC Black carbon
Cig Octadecyl matrix used in solid chromatography
Cd Freely-dissolved interstitial water concentration of contaminant
CdjAHijcvi Freely-dissolved interstitial water effect concentration of a specific PAH
CDOC Chemical concentration associated with dissolved organic carbon
Ciw Total interstitial water concentration of contaminant
Coc Chemical concentration in sediments on an organic carbon basis
COC.PAHI PAH-specific chemical concentration in sediment on an organic carbon basis
COC,PAHI,FCVI Effect concentration of a specific PAH in sediment on an organic carbon basis
calculated from the product of its FCV and Koc
Coc,pAHi,Maxi Maximum solubility limited PAH-specific concentration in sediment on an organic
carbon basis
CPS Passive sampler concentration of contaminant
dd Total dissolved concentration of a contaminant in interstitial water
CHN Carbon, hydrogen and nitrogen elemental analyzer
CWA Clean Water Act
DDTs Dichlorodiphenyltrichloroethane and degradation products
DOC Dissolved organic carbon
EPA United States Environmental Protection Agency
EqP Equilibrium partitioning
ESB Equilibrium partitioning Sediment Benchmark; for nonionic organic contaminants, this
term usually refers to a value that is organic carbon-normalized (more formally ESB0c)
unless otherwise specified
ESBTU Equilibrium Partitioning Sediment Benchmark Toxic Units
ESBTUpcv Equilibrium Partitioning Sediment Benchmark Toxic Units based on the Final Chronic
Value
FCV Final chronic value
fee Fraction of black carbon in sediment
fNsoc Fraction of natural sedimentary organic carbon
foe Fraction of organic carbon in sediment
GC/MS Gas chromatograph/mass spectrometer
goc Gram organic carbon
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
IW Interstitial water (also known as pore water)
IWTU Interstitial water toxic units
IWTUpcv Interstitial water toxic units based on the Final Chronic Value
IWTUscv Interstitial water toxic units based on the Secondary Chronic Value
KBC Black carbon-water partition coefficient
KDOC Dissolved organic carbon-water partition coefficient
KOC Organic carbon-water partition coefficient
KQW Octanol-water partition coefficient
KP Sediment-water partition coefficient
KPDMS Poly dimethyl siloxane-water partition coefficient
KPED Polyethylene device-water partition coefficient
KPOM Polyoxymethylene-water partition coefficient
Kps-d Passive sampler-water partition coefficient
NAPL Non-aqueous phase liquid
NSOC Natural sedimentary organic carbon
NOC Nonionic organic chemical
OC Organic carbon
ORD U.S. EPA, Office of Research and Development
PAH Polycyclic aromatic hydrocarbon
PCB Polychlorinated biphenyls
PDMS Polydimethylsiloxane
FED Polyethylene device
POM Polyoxymethylene
PRC Performance reference compound
SCV Secondary chronic value
SIM Selected ion mode in analyses using GC/MS
SPE Solid phase extraction
SPMD Semi-permeable membrane device
SPME Solid phase microextraction
TIE Toxicity Identification Evaluation
TNT Trinitrotoluene
TOC Total organic carbon
TU Toxic Unit
WQC Water Quality Criteria
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Section 1
Introduction
Introduction
1.1 General Information
The purpose of this document is to provide
guidance on procedures that can be used to
determine the freely dissolved interstitial water
concentrations of nonionic organic chemicals to
derive interstitial water toxic units, reflective of
environmental conditions. The procedures are
intended to be used with any water-only toxicity
values (e.g., FCVs, SCVs, other relevant water-
only data) and are not limited to the equilibrium
partitioning sediment benchmarks for endrin and
dieldrin (U.S. EPA, 2003b,c), mixtures of
poly cyclic aromatic hydrocarbons (PAHs) (U.S.
EPA, 2003d), and a selection of nonionic
organic chemicals (U.S. EPA, 2008) discussed
here (see Table 2.1 for a list of selected nonionic
organic contaminants).
A thorough understanding of the "Technical
Basis for the Derivation of Equilibrium
Partitioning Sediment Benchmarks (ESBs) for
the Protection of Benthic Organisms: Nonionic
Organics" (U.S. EPA, 2003a), the ESB
documents for endrin and dieldrin (U.S. EPA,
2003b,c), as well as documents for mixtures of
PAHs (U.S. EPA, 2003d), and selected nonionic
organic chemicals (U.S. EPA, 2008), and
"Guidelines for Deriving Numerical National
Water Quality Criteria for the Protection of
Aquatic Organisms and their Uses" (Stephan et
al., 1985) is recommended. Importantly, it is
strongly suggested that these procedures for
determining the sediment interstitial water
concentrations should be used with other
sediment assessment lines of evidence (e.g.,
Toxicity Identification Evaluations (TIEs) (U.S.
EPA, 2007)), benthic community surveys,
sediment toxicity testing) as well as risk
assessment procedures.
1.2 Review of the General Equilibrium
Partitioning Approach
The general EqP approach assumes that (1)
the partitioning of the nonionic organic chemical
between natural sedimentary organic carbon and
interstitial water is at or near equilibrium; (2) the
concentration in the phases can be predicted
using appropriate partition coefficients and the
measured concentration in the other phases
(assuming the freely-dissolved interstitial water
concentration can be accurately measured); (3)
organisms receive equivalent exposure from
water-only exposures or from any equilibrated
phase: either from interstitial water via
respiration, from sediment via ingestion or other
sediment-integument exchange, or from a
mixture of exposure routes; (4) for nonionic
chemicals, effect concentrations in sediments on
a normalized basis can be predicted using the
partition coefficients and effects concentrations
in water; (5) the FCV or SCV concentration (or
other relevant water-only value) is an
appropriate effects concentration for freely-
dissolved chemical in interstitial water; and (6)
ESBs derived as the product of a partition
coefficent and FCV or SCV are protective of
benthic organisms.
ESB concentrations presented in previous
documents (e.g., U.S. EPA, 2003b,c,d, 2008) are
expressed as jig chemical/g sediment organic
carbon (jig/goc) and not on an interstitial water
basis because (1) interstitial water was
considered too difficult to sample and (2)
significant amounts of the dissolved chemical
may be associated with dissolved organic carbon
(DOC); thus, total concentrations in interstitial
water may overestimate exposure.
1-1
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
As discussed in Section 1.3 and Section 2, in
the last several years, the first assumption used
in the one-carbon general model that nonionic
organic contaminants always partition between
only two phases (i.e., natural sedimentary
organic carbon and interstitial water) has been
demonstrated to not be entirely true in all
sediments and that other sedimentary
partitioning phases occur in sediments (Luthy et
al., 1997; Cornelissen et al., 2005a). Further,
because of advances in technology, some
contaminated sediment measurements like
interstitial water contaminant concentrations and
assessing the effects of DOC on contaminant
partitioning can now be made more accurately.
While the one-carbon general model has been
shown to operate successfully in many
applications (e.g., Swartz et al.; 1990, DeWitt et
al., 1992; Hoke et al., 1994), the recognition of
multiple sedimentary phases and recent advances
in interstitial water measurements make the
direct derivation of interstitial water toxic units
feasible.
1.3 Rationale for Development of
Procedures for Determining Freely
Dissolved Concentrations
As noted above, current ESB documents use
a one-carbon EqP model which assumes organic
contaminants partition between the aqueous and
natural sedimentary organic carbon phases (see
Section 2). Under some environmental
conditions and ESB applications, these
assumptions may be inaccurate. ESBs may be
under- or overprotective if the sediment or
chemical quality characteristics at the site alter
the bioavailability and, consequently, the
toxicity of the sediment-bound chemical relative
to that predicted by the one-carbon EqP theory.
Therefore, it is appropriate and more accurate
that the ESBs be used with directly determined
freely dissolved concentrations of nonionic
organic chemicals to derive interstitial water
toxic units. Further, in recent years, technologies
have been developed using passive sampling to
determine these freely dissolved interstitial water
concentrations of chemicals instead of
estimating them from sediment associated
concentrations as is performed in the one-carbon
EqP approach (Maruya et al., 2012).
1.4 Freely Dissolved Concentration
Procedures
The reason for using the various
bioavailability-based procedures described in
this document is that although testing of various
sediments has demonstrated the applicability of
the one-carbon EqP approach (U.S. EPA,
2003a), EqP theory based on a one-carbon
model may not accurately predict contaminant
partitioning for certain sediments and sites.
Unique sediment phases (i.e., the mixture of
pyrogenic carbon called black carbon), chemical
speciation, or chemical form may make the
chemical more or less bioavailable than EqP
predicts, thereby altering the toxicity of the
sediment. For example, in some sediments, the
partitioning of PAHs cannot be explained by
EqP based on natural sedimentary organic
carbon (Maruya et al., 1996; McGroddy et al.,
1996). Instead, accurate predictions of
partitioning behavior may require the use of both
an organic carbon-water partition coefficient
(Koc) and a black carbon-water partition
coefficient (KBc) (see Section 2) (Gustafson et
al., 1997; Cornelissen et al., 2005a). Further, to
derive accurate interstitial water toxic units
based on existing water-only toxicity data (e.g.,
FCV), quantification of partitioning at these sites
may require direct measurement of the freely
dissolved concentration of the nonionic organic
chemical in interstitial water (see Section 2).
Application of these ESB procedures may
indicate improvements to the one-carbon general
model that will require implementation over
time. Further, because these procedures can be
technically complex and sometimes costly, it is
important that they be conducted only by those
who are well qualified and experienced, and
potentially only as a second-tier assessment
approach (see Section 4).
1-2
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Introduction
This document focuses on black carbon as an
important alternate sedimentary phase. However,
other phases may be present in sediments
including incompletely degraded petroleum
(Jonker et al., 2003). The effects of petroleum
and other non-aqueous phase liquids (NAPLs)
on contaminant bioavailability are not
considered in this document due to the lack of
approaches, at this time, for accurately
addressing their effects.
1.5 Data Quality and Uncertainties
Data sources and manipulations used to
generate black carbon-water and passive
sampler-water partition coefficients (i.e., KBC,
Kps-d) presented in this document are discussed
in detail in Section 2. Due in part to the
relatively recent development and application of
many of the passive sampling technologies as
well as black carbon partitioning for estimating
bioavailability, the magnitude of the accuracy,
precision and uncertainties associated with these
partition coefficients is not well known. Recent
intensive evaluations of partition coefficients for
solid phase microextraction (SPME),
polyoxymethylene (POM), and polyethylene
(PE) (DiFilippo and Eganhouse, 2010; Endo et
al., 2011; Lohmann, 2012) are good examples of
the types of analyses needed to parameterize
these data quality assurance measures in the
future. There is also a need to encourage the
organization of expert workshops and funding of
quality assurance-related research to provide
guidance on these issues. For example,
determining when contaminants have achieved
equilibrium between the passive samplers and
the dissolved phase is currently a critical
challenge in the use of passive samplers. Further,
as the number of values for KBC and Kps-d
increase in the scientific literature some values
may need to be retired and replaced with values
that are more scientifically-sound and robust.
Similarly, as new toxicological data and models
become available (e.g., Di Toro et al., 2007;
McGrath and Di Toro, 2009), older data and
models may need to be reassessed or removed
from the data base. Further, the relationship
between predicted toxicological effects and
physicochemical parameters like KOW, may also
need to be reassessed. For example, McGrath
and Di Toro (2009), recently suggested to not
use log KOW values greater than 6.4 to predict
toxicological effects using the target lipid model,
frequently used with narcotic chemicals, because
of the uncertainties in the model's predictions
above that K0w value. Such a cut-off would
affect five of the chemicals specifically
discussed in this document (i.e., high molecular
weight PAHs). At this time, this guidance does
not recommend users to apply this cut-off but
does want to make users aware of this type of
discussion in the scientific literature. In contrast
to the passive samplers, black carbon, and
toxicological models, the accuracy, precision
and uncertainties associated with other aspects
of the procedures discussed in this document;
such as, sediment and interstitial water
instrumental analysis for contaminants and
sediment characteristics (e.g., DOC) are well
understood and have been discussed in detail
elsewhere.
This document was reviewed as part of a
formal external peer review coordinated at the
U.S. EPA National Health and Environmental
Effects Research Laboratory, Research Triangle
Park, North Carolina, and Atlantic Ecology
Division, Narragansett, Rhode Island. Any
detected errors of omission, substance or
calculation discovered during the peer review
process were corrected.
1.6 Overview
This document presents procedures for
determining the freely dissolved concentration of
nonionic organic chemicals for calculating
interstitial water toxic units. Section 2 of the
document provides background and guidance on
the procedures. Section 3 illustrates examples of
the use of the procedures. The implementation of
the procedures is discussed in Section 4. Section
5 lists the references for this document. Finally,
1-3
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
the focus of this document is to provide the
reader with an overview of the current
approaches for determining the freely dissolved
concentrations of nonionic organic chemicals in
sediment interstitial waters. The document is not
intended to serve as a methods manual. In the
different sections of the document, relevant
scientific literature is cited to provide the reader
with more in-depth information.
1-4
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
Section 2
Procedures for Determining Freely
Dissolved Interstitial Water
Concentrations
2.1 Introduction
Current ESB documents for the nonionic
organics endrin, dieldrin, PAH mixtures, and
compendium chemicals (U.S. EPA, 2003
a,b,c,d, 2008) use a two-phase EqP model to
derive ESB values. This model assumes
sediment contaminants partition between
natural sedimentary organic carbon and the
freely dissolved phase in interstitial water. In
this document, this model is called the "one-
carbon" model. The procedures in this
document are intended for determining the
freely dissolved concentration of chemicals in
sediment interstitial waters. These
concentrations can then be used with current
ESBs and other relevant water-only toxicity
values. For example, as recently discussed by
Di Toro et al. (2007) and McGrath and Di
Toro (2009), the target lipid model used to
calculate the FCVs for PAHs (U.S. EPA,
2003d) can also be modified to calculate
"mode of action" based FCVs. The mode of
action FCV considers the 5th percentile of the
distribution of the critical target lipid body
burdens using the target lipid model, a
chemical class variable, and an empirically-
derived geometric mean acute to chronic ratio.
When applied in combination, the freely
dissolved interstitial water concentration and
water-only toxicity value are used to derive
interstitial water toxic units that more directly
consider contaminant bioavailability. The
procedures capture the partitioning of organic
contaminants to specific phases in the
sediment in addition to natural sedimentary
organic carbon. These alternate phases may
include, but are not limited to, interstitial DOC
and different forms of BC. These phases are
discussed below. The objective of this
document is to generate an assessment of
sediment toxicity that is more accurate in
terms of environmental bioavailability. The
focus of this document is on the performance
of assessments for nonionic organic
contaminants.
2.1.1 Rationale
As noted above, U.S. EPA's sediment
guidelines or benchmarks (ESBs) for nonionic
organic chemicals are based on the one-carbon
general equilibrium partitioning model. The
general model uses a two-phase approach:
particulate-associated chemical and dissolved
interstitial water chemical, where the total
concentration in the sediment matrix equals
the concentration in the particulate phase plus
the concentration freely-dissolved in
interstitial water under equilibrium conditions.
With this model, the dissolved phase
concentration (Cd) (ng/L) of a nonionic
organic contaminant can be calculated as
follows:
Cp
KP
(2-1)
where, Cp is the particulate contaminant
concentration (|ig/Kg dry) and KP is the
sediment-water partition coefficient (L/Kg).
2-1
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Substituting KP with the product of the
fraction of natural sediment organic carbon
(foe) (Kg organic carbon (OC) /Kg dry) and
organic carbon-water partition coefficient
(Koc) (L/Kg OC) (for more discussion of
selecting Koc, see U.S. EPA, 2003b,c,d,
2008), Equation 2-1 can be rewritten as:
O, =
cp
focKoc
(2-2)
Using Equation 2-2, the conventional ESB
can be determined. For example, for PAH
mixtures (U.S. EPA, 2003d), in this one-
carbon model, the estimated Cd for each of 34
PAHs are divided by water quality criteria
(WQC) final chronic values (FCVs),
secondary chronic values (SCVs) or any other
relevant water-only value to derive the ESB
toxic units (ESBTUs) (see Section 2.5 for
more discussion on how to calculate interstitial
water toxic units). It should be noted that in
the conventional ESB procedure using the one-
carbon general model, ESBTUs are often
derived from the quotient of the contaminant
organic carbon normalized sediment
concentration (|ig/Kg OC) and the organic
carbon normalized toxicity value (jig/Kg OC).
See U.S. EPA (2003b,c,d, 2008) for more
discussion of the conventional ESB procedures
and selection of water-only toxicity values.
If additional sorbing phases exist in
sediment, it is possible that the EqP model
(i.e., Equation 2-2) for Cd may not always be
accurate. In these cases, the toxicity of the
sediment may not be accurately predicted by
the one-carbon model because, in addition to
natural sedimentary organic carbon, black
carbon, or other properties of the sediments
may alter bioavailability (Figure 2-1). To
accurately consider the effect of these phases
and properties in a sediment assessment, use of
the procedures described in this document is
warranted.
' Black carbon
combustion
residue
2 Natural
sedimentary
organic
carbon
3 Dissolved
organic
carbon
DOC3 structure based on structure proposed
byZafiriouetal.(1984)
Figure 2-1. Magnified and exploded view of
different types of sediment particle
phases.
2-2
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
In this section, approaches are presented to
estimate or measure the bioavailable, freely
dissolved interstitial water concentration (i.e.,
Cd), which can be compared to the WQC FCV,
SCV, or any relevant water-only toxicity
value. See U.S. EPA (2003d, 2008) for more
discussion of the ESB procedures and the
selection of water-only toxicity values. FCVs
and SCVs for many nonionic chemicals can be
found in U.S. EPA (2003b,c,d, 2008), Suter
and Mabrey (1994), Suter and Tsao (1996),
and Great Lakes Water Quality Initiative
(1995) as well as other sources (Di Toro et al.,
2007; McGrath and Di Toro, 2009). The
approaches EPA recommends in this
document for estimating or measuring the
freely dissolved chemical concentration in
interstitial water require procedures
appropriate for obtaining and chemically
analyzing interstitial water or sampling the
interstitial water or whole sediments with
passive samplers. The approaches assume that
the contaminant is distributed into multiple
phases: freely dissolved, DOC-associated,
natural sedimentary organic carbon, and BC.
Recent research and technological advances
have made the measurement or collection of
these samples feasible. Further, the freely
dissolved concentrations can be determined in
various ways: (1) estimated using a two-
carbon model that takes into account the
association of contaminants with BC,
(2) extracted directly from interstitial waters,
(3) estimated by passive sampling of whole
sediments, and (4) estimated by passive
sampling of interstitial waters. The procedures
presented below employ the best technologies
available at the time this document was
prepared for obtaining interstitial water,
chemically analyzing interstitial water
contaminant concentrations, passive sampling
interstitial water and whole sediment, and
estimating or measuring the freely dissolved
concentration of contaminants. The last part of
this section (2.5) describes an approach for
using the freely dissolved concentrations
collected with the procedures listed above to
derive interstitial water toxic units (IWTUs). If
the IWTUs are greater than one, benthic organisms
may not be protected and adverse effects may
result.
2.2 Using a Two-Carbon Model for
Determining Freely Dissolved
Interstitial Water Concentrations
The comparison of dissolved contaminant
concentrations derived from carbon-
normalized concentrations in bulk sediment to
FCVs or SCVs as described in Section 2.1
(i.e., Equation 2-2), may be inaccurate at some
sites if the characteristics of the sediment or
the contaminant reduces the partitioning into
the interstitial water, thereby reducing
bioavailability and toxicity. For example,
several studies have demonstrated that
partitioning of PAHs cannot always be
explained by the conventional two phase "one-
carbon" EqP model (Equation 2-2)
(McGroddy et al., 1995; Maruya et al., 1996).
Additional studies suggest that PAHs that are
occluded in or sorbed to forms of black carbon
exhibit reduced partitioning (Gustafsson et al.,
1997; Bucheli and Gustafsson, 2000; Accardi-
Dey and Gschwend, 2002; Arp et al., 2009)
and limited bioaccumulation by benthic
invertebrates (Vinturella et al., 2004a; Rust et
al., 2004) which suggests bioavailability is
being reduced.
Highly condensed forms of pyrogenic
carbon (e.g., soot) and residues of incomplete
combustion (e.g., charcoal), commonly termed
BC, are ubiquitous in the aquatic environment.
It is estimated that BC constitutes
approximately 10% of sedimentary organic
carbon in ordinary sediments (Middelburg et
al., 1999). In sediments from contaminated
sites, the contribution of BC may exceed 50%
due to the fossil-fuel related residues of
historic industrial activity. The sorption of
nonionic organic contaminants to BC has been
observed to be up to 10-1,000 times stronger
2-3
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Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
than the sorption to natural sedimentary
organic carbon (NSOC), which includes
diagenic organic carbon, such as plant material
(Burgess et al., 2004). The sorption to BC is
often nonlinear following a Freundlich
isotherm, with the effect of BC being strongest
at low contaminant concentrations (Accardi-
Dey and Gschwend, 2002).
Concentrations of BC in sediment are
commonly measured using a chemothermal
oxidation (CTO) method (Gustafsson et al.,
1997, 2001; Accardi-Dey and Gschwend,
2003). The method involves quantification of
BC and total organic carbon (TOC) using a
carbon, hydrogen, nitrogen elemental analyzer
(CHN): (1) removal of inorganic carbonates
via acidification; (2) removal of NSOC in a
furnace under air flow (375°C, 24 hours); and
finally; (3) quantification of remaining carbon
as BC using a CHN. Other methods are
available for measuring BC but are less
commonly used with contaminated sediments
(e.g., chromic acid digestion, microscopic
inspection).
2.2.1 Two- Carbon Model
Unlike the model in Equation 2-2, the
freely dissolved concentration of nonionic
organic contaminants in interstitial water can
be estimated using a "two-carbon" model that
accounts for association of nonionic organic
contaminants with the fraction of BC (fee) in
sediment and the fraction of NSOC (fNsoc). A
two-carbon model accounts for linear
absorption into the NSOC in sediment and
nonlinear adsorption to BC. The two-carbon
model can be used to calculate the freely
dissolved concentration of each nonionic
organic contaminant in interstitial water using
the following relationship:
cp
fNsoc KQC+ /BC KBC <
(2-3)
where, fNsoc is the weight fraction of NSOC in
sediment (Kg NSOC/Kg dry), calculated from
the difference between TOC and BC, fee is the
weight fraction of BC in sediment (g BC/g
dry), KBC is the BC to water partition
coefficient (L/Kg BC), and n is the Freundlich
exponent, which accounts for nonlinear
sorption behavior (n = 0.6) (Accardi-Dey and
Gschwend, 2002). The value of n will vary
depending on the nonionic organic
contaminant. For example, to date, 0.6 has
been used for PAHs with log K0ws of
approximately 4.00 to 5.50. Because Cd
appears on both sides of the equation, an
iterative approach must be used to solve for
Cd. Computer-based statistical protocols such
as the "Goal Seek" function in Excel®
(Microsoft, Seattle, WA, USA) are available
for this purpose.
2.2.2 Estimation of KBC
Values of KBC have been reported for
several PAHs in spiked sediments (Accardi-
Dey and Gschwend, 2003; Burgess et al.,
2004) as well as for PAHs and chlorinated
compounds in native sediments (Lohmann et
al., 2004, 2005; Vinturella et al., 2004b;
Hawthorne et al., 2007). Because KBc values
are not available in the literature for many
nonionic organic contaminants, one study
developed a linear regression relationship
between the octanol-water partition coefficient
(Kow) and experimentally-derived values of
KBC for 17 PAHs (Accardi-Dey and
Gschwend, 2003) to estimate KBC values for
PAHs for which experimental data was not
available (Kane Driscoll and Burgess, 2007).
Use of estimated values of KBC in a two-
carbon model thus far has been successful in
predicting interstitial water concentrations
(e.g., Armitage et al., 2008; Accardi-Dey and
Gschwend, 2003) or in screening-out
sediments that were not toxic to aquatic
invertebrates (e.g., Driscoll et al., 2009);
however more data are needed. For example,
another study of 114 sediments reported that
2-4
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Procedures for Determining Freely Dissolved Interstitial Water Concentrations
the two-carbon model showed no significant
improvement over the one-carbon general
model for predicting the distribution of PAHs
between sediment and interstitial water with
KBC ranging over three orders of magnitude
(Hawthorne et al., 2007). Additional studies
have demonstrated that various types of
carbonaceous materials, such as coal tar pitch,
exhibit a range of partitioning behavior.
Further, the size of the black carbon particle
affects the magnitude of the KBC (Hong et al.,
2003; Ghosh et al., 2003; Khalil et al., 2006;
Ghosh, 2007; Ghosh and Hawthorne, 2010).
As a result of the lack of empirical values
and current level of uncertainty associated
with predicted KBC values, provisionally, this
document uses the relationship developed by
Driscoll et al. (2009) for 17 PAHs:
Log KBC = 3.41 + 0.54 LogKow (2-4)
KOW values for a range of nonionic organic
contaminants can be found in Mackay et al.
(1992a,b), Karickhoff and Long (1995), and
U.S. EPA (2003d, 2008). Other predictive
relationships are also available (e.g., Koelmans
et al., 2006; van Noort, 2003).
Table 2-1 provides a list of calculated
provisional log KBc values for several
nonionic organic contaminants based on
Equation 2-4. As noted in Section 1.5, as the
number of available empirical KBc values
increases, Equation 2-4 should be updated.
Further, KBC should only be used for nonionic
organic contaminants that are planar and not
non-planar chemicals unless the KBcS were
derived specifically for those non-planar
chemicals (see discussion in Section 4).
2.3 Direct Measurement of Interstitial
Water Concentrations
Over the last several decades, a variety of
methods have been developed to estimate or
measure concentrations of freely dissolved
chemicals in interstitial water. Earlier methods
calculated the freely dissolved concentration
from the difference between the total (i.e.,
freely dissolved and DOC-associated) and
DOC-associated phases. More recent methods
use passive sampling devices to directly
measure the freely dissolved concentration in
interstitial water or whole sediment.
2.3.1 Direct Collection of Interstitial Water
by Centrifugation
The problems associated with adequately
collecting and processing interstitial water
samples are well documented (Adams, 1991;
Schults et al., 1992; Ankley and Schubauer-
Berigan, 1994; ASTM, 1994; Ozretich and
Schults, 1998; Adams et al., 2003; Carr and
Nipper, 2003). Artifacts from the procedures
can preclude accurate determination of
interstitial water contaminant concentrations.
Further, in general, eliminating and/or
avoiding these artifacts when centrifuging can
be quite difficult experimentally. The
procedures cited below have been shown to
minimize artifactual effects of interstitial water
sample collection and analysis for
contaminants.
If performed with a minimum of artifacts,
centrifugation without subsequent filtration
results in an acceptable sample of interstitial
water which can be used to make an accurate
measurement of the freely dissolved
concentration of nonionic organic
contaminants in sediments (Adams et al.,
2003). Substantial artifacts include the
formation of dissolved and colloidal organic
matter during interstitial water preparation and
isolation which can result in an over-
estimation of interstitial water nonionic
organic contaminant concentrations and
potential bioavailability especially for those
contaminants with high KOWS. Another
substantial artifact is absorption and loss of
NOCs to laboratory equipment surfaces. The
objective of centrifugation is to obtain
2-5
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
interstitial water containing contaminants
operationally defined as freely dissolved.
Therefore, any combination of gravitational
force and time that settles the particles is
acceptable. For example, a procedure applied
by Lee et al. (1994) and Swartz et al. (1994)
on marine sediments was shown to effectively
reduce losses of organic contaminants to
laboratory equipment surfaces (Ozretich and
Schults, 1998). The procedure also allowed for
the chemical analysis of DOC (U.S. EPA,
2000) and total contaminant concentrations
(i.e., freely dissolved fraction plus the fraction
associated to DOC). Conversely, the total
interstitial water can be sub-sampled for direct
measurement of the DOC-bound contaminants
(see Section 2.3.2).
Centrifugation of the sediment and sub-
sampling of the interstitial water should be
performed within two hours of each other to
avoid complications from the potential
formation of new artificial particles caused by
oxidation of reduced iron. It is clear that
cleanly sampled interstitial water is important,
as the presence of a particle of sediment, as
noted above, could result in erroneously high
concentrations; on the other hand, if the time
periods before extractions are extended, by
filtering and excessive sample handling,
erroneously low concentrations would result
because of contaminant sorption to laboratory
equipment surfaces.
2.3.2 Calculating the Freely-Dissolved
Concentration
Regardless of the extraction method used,
it is critical that the instrumental analysis can
detect contaminant concentrations below the
FCV, SCV or other relevant water-only effect
concentrations (i.e., -0.01 |ig/L for a great
deal of the toxic nonionic organic chemicals).
With the interstitial water concentration data
for total contaminant and DOC concentrations,
the freely dissolved interstitial water
concentration of a nonionic organic chemical
can be determined in the following three ways:
1. It can be assumed that the measured total
interstitial water concentration (Ciw) for a
nonionic organic chemical with a low to
intermediate \ogKow value (i.e., 2.5 to 4.0)
is equivalent to the dissolved concentration
(Cd); that is, the freely dissolved interstitial
water concentration equals the total
measured interstitial water concentration.
However, this approach is problematic and
is not recommended because high
concentrations of DOC can be present in
isolated interstitial water. Even low K0w
nonionic organic chemicals are known to
associate with this material, causing a
reduction in their bioavailability. Therefore,
contaminant concentrations measured in
interstitial water isolated by centrifugation
would contain both the freely dissolved and
the DOC-associated chemical, over-
estimating the true bioavailability of the
nonionic organic chemicals. The magnitude
of the over-estimation would depend on the
concentration of the DOC and affinity of the
DOC for the chemicals of interest. This
affinity is represented by the dissolved
organic carbon partition coefficient (
KDOC —
-DOC
(2-5)
where, CDOC is the contaminant
concentration associated with the DOC.
2. It can be determined that the freely
dissolved interstitial water concentration is
the difference between the interstitial water
concentration and the DOC-associated
concentration. For example, solid phase
extraction (SPE) using Ci8 selectively
isolates the freely dissolved chemical on the
column while the DOC-associated chemical
passes through the column media. The
freely dissolved chemical can then be eluted
from the column with an organic solvent
2-6
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Procedures for Determining Freely Dissolved Interstitial Water Concentrations
(Landrum et al., 1984; Ozretich et al.,
1995). The application of this method
depends on the DOC-associated
concentration being operationally defined as
the chemical passing through the column.
However, use of this procedure doubles the
number of samples that need to be analyzed,
and may require monitoring of DOC
retention by the column (Ozretich et al.,
1995). In a similar procedure, both the
DOC-associated chemical and freely-
dissolved chemical concentrations can be
directly measured (Burgess et al., 1996)
rather than by being determined by
difference from the interstitial water
chemical concentration. This approach
should be used only if acceptable mass
balances (approximately 90% or greater) of
the DOC-associated, freely dissolved, and
total chemical are demonstrated.
3. Using Equations 2-6 and 2-7 below, Cd can
be calculated from the measured total
interstitial water concentration (Ciw), the
DOC concentration, and the contaminant
KDOC'-
(2-6)
(DOC KDOC+ i)
as can the percentage of the total
contaminant that is freely dissolved (%
(DOC KDOC+ i)
100
(2-7)
This method depends on determination
of DOC (kg/L) and KDOC. Determining the
concentration of DOC in water is a routine
analysis (see above) (U.S. EPA, 2000), and
KDOC values can be found in Burkhard
(2000). Burkhard (2000) derived the
following expression based on the analysis
of several interstitial water studies:
Log KDOC = -0.88 + 0.99 Log Kow (2-8)
As noted earlier, K0w values for a range
of nonionic organic contaminants can be
found in Mackay et al. (1992 a,b),
Karickhoff and Long (1995), and U.S. EPA
(2003d, 2008).
As an example, using Equation 2-8,
KDoc values from the endrin and dieldrin
ESB documents (U.S. EPA, 2003a,b) were
compared with KOC values (Table 2-2), and
the percentage of the total compound that is
freely-dissolved calculated using Equation
2-7 for a range of DOC concentrations
likely to be encountered in interstitial water
(Table 2-3). In this example, the greatest
percentage of endrin or dieldrin that would
be associated with DOC using KDOC is
approximately 51% and 34%, respectively
at 70 mg DOC/L. An example of using this
procedure is also presented in Section 3.
2.4 Use of Passive Samplers for
Determining Freely Dissolved
Interstitial Water Concentrations
Recently, a variety of passive sampling
methods have been developed to directly
measure the concentrations of freely dissolved
chemical in contaminated sediments (Figure
2-2). Some methods sample interstitial water
generated by centrifugation while others
sample directly from sediment matrix, either in
the laboratory or in the field. Laboratory
experiments have been conducted by tumbling
sediments with passive samplers (Lohmann et
al., 2005; Fernandez et al., 2009a,b;
Hawthorne et al., 2009, 2011) or by static
placement of the passive sampler into
sediment (Lohmann et al., 2005; Fernandez et
al., 2009a,b). All methods are similar in that an
organic polymer is used to absorb nonionic
organic contaminants from sediment and
interstitial waters. Once the contaminant
achieves equilibrium between the polymer,
sediment, and interstitial water, partition
coefficients can be used to calculate the
interstitial water dissolved phase
2-7
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
concentrations (Cd) of contaminants of
interest. Chemical analysis of passive samplers
starts with a simple organic solvent extraction
similar to a sediment extraction (e.g., U.S.
EPA Method 3540). Because of their small
size, solid phase microextraction fibers
(SPMEs) can also be directly injected with
thermal desorption into the analytical
Figure 2-2. Photographs of selected passive
samplers, including SPME, PE,
and POM.
instrument (i.e., GC/MS). The freely dissolved
concentration can then be used to calculate the
interstitial water toxic units associated with the
sediment (see Section 2.5).
2.4.1 Types of Passive Samplers
Several of the more commonly used
passive samplers in North America are
discussed below:
Semi-permeable membrane devices
(SPMDs) are composed of flat, low-density
polyethylene (LDPE) tubing that contains a
thin film of a pure, high molecular weight
synthetic lipid (triolein). The polymer structure
of the LDPE allows for the diffusion of
nonionic organic chemicals within and through
the tubing, which are then sequestered in both
the lipid and LDPE phases. SPMD is an
established method for assessing freely
dissolved concentrations in water (Huckins et
al., 1993; Huckins et al., 2006) that has more
recently been used to measure concentrations
of freely dissolved organic contaminants in
sediments, soils and ground water (Macrae and
Hall, 1998; Gustavson and Harkin, 2000;
Rantalainen et al., 2000; Wells and Lanno,
2001; Williamson et al., 2002; Leppanen and
Kukkonen, 2000, 2006; Schubauer-Berigan et
al., 2012).
Polyethylene devices (PEDs) consist of flat
strips of LDPE lacking the inner triolien layer
used in SPMDs (Lohmann et al., 2004; Adams
et al., 2007; Tomaszewski and Luthy, 2008).
The thickness of PEDs varies from 25 jim
(Fernandez et al., 2009a,b) to > 100 |im, and
strips up to 50 cm in length have been used
(Booij et al., 2003a). PEDs can reach
equilibrium faster than SPMDs due to their
smaller sorption capacity (Booij et al., 2002).
Conversely, PEDs have greater contaminant
capacity than some other passive samplers
(e.g., solid phase microextraction (SPME)) but
require a longer time to reach equilibrium.
Performance reference compounds (PRC)
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
incorporated into the FED (as well as
polyoxymethylene (POM) and SPMDs) before
deployment can be used to estimate the extent
to which equilibrium is reached during
deployment, and to estimate adjusted
equilibrium concentrations (Fernandez et al.,
2009b). Biofouling of FED can be a concern
as a barrier to exchange and equilibration but
PRCs offer correction for this effect. Because
they are inexpensive, robust, and easily
deployable, PEDs have been used to measure
interstitial water concentrations in laboratory
and field applications, (Lohmann et al., 2004,
2005; Tomaszewski and Luthy, 2008;
Fernandez et al., 2009a,b; Gschwend et al.,
2011). Further, FED accumulation of PAHs
and polychlorinated biphenyls (PCBs) has also
shown good correlation to bioaccumulation by
a benthic polychaete (Vinturella et al., 2004a;
Friedman et al., 2009). In static sediment
deployments, PEDs may deplete the
surrounding interstitial water of contaminants
if too much PE is used.
Solid phase microextraction (SPME)
devices are composed of fused silica fibers
that are coated with a layer of absorbing
polymer. Polydimethylsiloxane (PDMS),
which is typically used as a coating material, is
thermally stable, and absorbed contaminants
can be either thermally desorbed or extracted
with solvent. Polyacrylate coatings have also
been used to sample TNT (Conder et al.,
2004). PDMS SPME fibers reach equilibrium
rapidly in water, although their small capacity
can result in elevated detection limits in
comparison to other passive samplers. Time to
reach equilibrium when deployed in sediment
can be longer, ranging from 14 to 110 days
(Maruya et al., 2009). The fibers are fragile,
but can be protected for deployment in the
field (Maruya et al., 2009 ) and used to
determine vertical profiles of contaminants in
sediment (Lu et al., 2011). SPME has been
used to measure interstitial water
concentrations in several laboratory and field
studies (Mayer et al., 2000a,b; Hawthorne et
al., 2006; Maruya et al., 2009; Gschwend et
al., 2011). Freely dissolved concentrations
determined using SPME have been shown to
be good predictors of sediment toxicity
(Kreitinger et al., 2007; Xu et al., 2007) as
well as bioaccumulation (Kraaij et al., 2003).
Polyoxymethylene (POM) is like the FED
but is a harder plastic polymer with strong
partitioning and greater capacity than PDMS.
Studies have shown strong and reproducible
partitioning of nonionic organic contaminants
to POM, with sorption of contaminants being
similar to the polymer coatings used for
SPMEs (Jonker and Koelmans, 2001; Jonker
et al., 2003; Cornellisen et al., 2008;
Hawthorne et al., 2009, 2011). One advantage
is that the surface of POM is hard and smooth,
which allows any particulate matter
accumulated on the sampler during the
deployment to be physically wiped off after
recovery (Jonker and Koelmans, 2001). Like
PEDs, POM may deplete interstitial water
concentration in static sediment deployments;
consequently, the ratio of sampler to sediment
organic carbon may need to be limited.
As mentioned earlier, passive samplers
discussed here represent some of the more
common devices used in the North America.
Recently, published comparisons of some of
these samplers provide very useful information
for selecting which type of passive sampler is
most appropriate for a given application (e.g.,
Gschwend et al., 2011; Oen et al., 2011; U.S.
EPA, 2012). Because of their limited use in
sediments, SPMDs will not be discussed
further. Other samplers not discussed here can
be found described in the cited literature (e.g.,
Stuer-Lauridsen, 2005; Vrana et al., 2005;
Ouyang and Pawliszyn, 2007; Seethapathy et
al., 2008; Allan et al., 2009; Rusina et al.,
2010; Lohmann et al., 2012).
2-9
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
2.4.2 Procedures for Whole Sediments
Passive samplers can be exposed to whole
sediments by a variety of methods. For
example, laboratory exposures have been
conducted by tumbling small pieces of pre-
cleaned FED with a sediment slurry, with time
to reach equilibrium ranging from 1 to 60 days
for chemicals with log K0w < 7 (Booij et al.,
2003a; Lohmann et al., 2005; Hawthorne et
al., 2009; Gschwend et al., 2011). Passive
samplers can also be exposed to static
sediments, either in the laboratory or in the
field. Time to reach equilibrium in static
sediments is expected to be longer than in
tumbled sediment, due to the decrease in
transport resistance associated with tumbled
sediment slurries (Booij et al., 2003a).
If equilibrium is not achieved during
deployment, information on the uptake
kinetics or the extent of equilibrium is required
in order to estimate equilibrium concentra-
tions. As noted above, in these instances,
PRCs can be incorporated into the sampler
prior to deployment to provide information on
equilibrium status (Fernandez et al., 2009b).
Experimental evidence indicates that the
compound-specific rate at which PRCs
dissipate from the passive sampler to the
environment is related to the rate of uptake of
chemically similar target compounds from the
environment (Huckins et al., 2002; Booij et al.,
2002; Thomaszewski and Luthy, 2008;
Fernandez et al., 2009b). Thus, the concentra-
tions of PRCs in the sampler before and after
deployment can be used to estimate the
equilibrium concentration of target chemicals
from non-equilibrium concentrations measured
in the sampler after retrieval. Guidance on the
use of PRCs is still being developed but for
losses of PRC less than 10% during a
deployment, because of the uncertainties
potentially associated with such low losses,
those PRC data should not be used to adjust
the equilibrium status of target contaminants.
2.4.2.1 Calculation of Freely Dissolved
Concentrations using Passive
Samplers
Starting with the measured concentration
of contaminants in the passive sampler based
on chemical analysis, the dissolved phase
concentration is calculated as follows:
CPS
Kps-d
(2-9)
where, CPS is the passive sampler
concentration of a chemical (|ig/Kg passive
sampler) and KPS.d is the passive sampler -
water partition coefficient (L/Kg passive
sampler). As discussed in Section 2.4.2, there
may be the need to adjust CPS if the passive
sampler deployment was insufficient in
duration to achieve equilibrium conditions
(Fernandez et al., 2009b). Values for KPS-d are
passive sampler specific and can be found in
the literature (Jonker and Koelmans, 2001;
Leslie et al., 2002; Booij et al., 2003b; Zeng et
al., 2004; Lohmann et al., 2005; Adams et al.,
2007; Cornelissen et al., 2008; Fernandez et
al., 2009a,b; Maruya et al., 2009; Perron et al.,
2009; DiFilippo and Eganhouse, 2010;
Lohmann and Muir, 2010; Endo et al., 2011;
Lohmann, 2012). However, whenever
possible, laboratory confirmation of literature-
based Kps-d values for a given polymer is
recommended highly. Table 2-1 provides
calculated provisional KPS-d values for PEDs,
PDMS, and POM based on the following
relationships from Lohmann and Muir (2010),
DiFilippo and Eganhouse (2010) and Endo et
al. (2011), respectively:
Log KPED = -0.59 + 1.05 Log Kow
(2-10)
Log KPDMS = 0.07 + 0.83 Log Kow
(2-11)
Log KPOM = - 0.60 + 1.01 Log Kow
(2-12)
2-10
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
Several authors have reported that as K0w
increases, the relationship between KPS.d and
KOW begins to demonstrate curvilinear
behavior often at log K0ws of greater than 6.5
or 7.0. The result is that the KPS-d decreases in
value. Our understanding of why this behavior
occurs is incomplete and research to better
understand this phenomena is underway.
Conversely, for chemicals in Table 2-1 with
partition coefficients less than log 2.00, the use
of PEDs, PDMS and POM-based passive
samplers may not be effective because of weak
partitioning to the polymers. For chemicals
demonstrating this relatively low level of
hydrophobicity, direct extraction and analysis
of the interstitial water may be more effective.
2.4.3 Procedures for Interstitial Waters
A standardized and commercially-
available method for using SPME to isolate
PAHs from interstitial water is available
(Hawthorne et al., 2005; ASTM, 2010). The
method was developed to analyze 24 PAHs,
consisting of the two- to four-ring parent and
alkylated PAHs which contributed 95% of the
ESBTUs measured in 120 samples of
interstitial water from uncontaminated and
contaminated sediments (Hawthorne et al.,
2006). In this method, approximately 40 mL of
sediment is centrifuged for 30 min at 1,000 g.
Dissolved organic carbon (DOC) which
interferes with the analysis, is removed from
the produced interstitial water by flocculation
with the addition of aluminum potassium
sulfate (i.e., alum) followed by sodium
hydroxide (Ghosh et al., 2000). Two rounds of
flocculation and centrifugation are conducted
no more than 24 hours prior to extraction of
interstitial water for nonionic organic
contaminant analysis. Immediately after
flocculation, deuterated-PAH internal
standards are mixed with a 1.5-mL aliquot of
the interstitial water, which is then extracted
for 30 mins. using a commercially available
PDMS SPME fiber. Under these conditions,
30 minutes is sufficient to depletively sample
the interstitial water sample. The internal
standards are used to quantify the target PAHs
and compensate for incomplete extraction in
the same way as a liquid-liquid extraction.
Following the sorption period, the SPME fiber
is immediately thermally desorbed in a
GC/MS. The PAHs are detected and quantified
using the selected ion monitoring (SIM) mode.
The SPME fiber can be reused after cleaning
for fifteen minutes to one hour (for heavily
contaminated samples) under helium at
elevated temperatures (Hawthorne et al.,
2006).
Because this method removes some, but
not all, of the DOC, both target PAHs and
internal standards partition between the
interstitial water and the DOC before
extraction and analysis. The concentration of a
target PAH is determined on the basis of its
deuterated PAH internal standard and
represents a "total dissolved" PAH
concentration (Cxd) that includes both the
freely dissolved PAHs and some PAHs
associated with DOC. Because contaminant
KDOC increases with hydrophobicity, this over-
estimate of the freely dissolved concentration
is much greater for four- to six-ring PAHs than
for the two- and three-ring PAHs. For
example, the "total dissolved" concentrations
of five- and six-ring PAHs were as much as
7-fold higher than the freely dissolved
concentration, whereas differences for two-
and three-ring PAHs were insignificant
(Hawthorne et al., 2005). However, because
the lower molecular weight PAHs are present
in interstitial water at much higher
concentrations than the higher molecular
weight PAHs, there was no significant
difference in the sum of the ESBTUs
regardless of whether "total dissolved" or
freely dissolved PAH concentrations were
determined. Using the dd value, interstitial
water toxic units can be calculated.
2-11
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
This method is currently available for
PAHs but the fundamental approach is viable
with other nonionic organic contaminants. For
example, Friedman et al. (2011) used a similar
approach to calculate interstitial water
concentrations of PCBs sampled with PEDs
and adjusted for DOC.
Based on the discussion in this section,
Table 2-4 summarizes a selection of the
advantages and disadvantages of each
approach for determining freely dissolved
interstitial water concentrations.
2.5 Derivation of Interstitial Water Toxic
Units
In the previous three subsections (2.2, 2.3,
2.4), approaches for determining organic
contaminants were described. In this
subsection, that data will be used to derive the
interstitial water toxic units. The Cd for the
freely dissolved concentrations of each
nonionic organic contaminant is divided by its
corresponding FCV, SCV or relevant water-
only toxicity value to derive interstitial water
toxic units (IWTUs) (Equations 2-13, 2-14).
Cd values can be generated from the use of the
two-carbon model (Section 2.2.1), direct
measurement of interstitial waters after
adjustment for DOC (Section 2.3), or passive
samplers (Section 2.4). In deriving the IWTUs,
the estimated interstitial water concentration of
each contaminant (Cd) is also compared to the
limit of water solubility for that contaminant in
deionized water. If Cd is less than the limit of
water solubility, then Cd is divided by the
corresponding FCV (or SCV) to calculate the
IWTUpcv (or IWTUscv) for that contaminant.
If Cd exceeds the available limit of water
solubility, then the limit of water solubility is
divided by the corresponding FCV (or SCV) to
derive an IWTU value for that contaminant.
Water solubility values for several nonionic
organic contaminants can be found in Mackay
et al., (1992a,b) as well as for PAHs in U.S.
EPA (2003d, reported in Appendix E). For
chemicals with a narcosis mode of action, the
IWTUs for all contaminants may be summed
to derive the £IWTU (see U.S. EPA, 2003d,
2008 for more discussion of narcosis).
Using the FCV or SCV, the interstitial
water toxic units are determined as follows:
FCV —
FCV
or
IWTUSCV =
(2-13)
(2-14)
If the IWTU for a nonionic organic
contaminant or £IWTU for a mixture of
narcotic nonionic organic contaminants is less
than or equivalent to 1.0, the concentration of
the contaminant or mixture of contaminants,
respectively, in the sediment is acceptable for
the protection of benthic organisms from
chronic toxic effects. Conversely, if the
IWTUs exceed 1.0, benthic organisms are not
protected and adverse effects may occur. See
Section 3 for examples of these calculations
and their interpretation.
2-12
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
Table 2-1. Provisional partition coefficients for selected nonionic organic contaminants
Class
PAHs
Contaminants
Naphthalene
Cl -naphthalenes
Acenaphthylene
Acenaphthene
C2-naphthalenes
Fluorene
C3 -naphthalenes
Anthracene
Phenanthrene
Cl-fluorenes
C4-naphthalenes
C 1 -phenanthrene/anthracenes
C2-fluorenes
Pyrene
Fluoranthene
C2-Phenanthrene/anthracenes
C3-fluorenes
C 1 -pyrene/fluoranthenes
C3 -phenanthrene/anthracenes
Benz(a)anthracene
Chrysene
C4-Phenanthrenes/anthracenes
C 1 -Benzanthracene/chrysenes
Benzo(a)pyrene
Perylene
Benzo(e)pyrene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
C2-benzanthracene/chrysenes
Benzo(ghi)perylene
C3-benzanthracene/chrysenes
Indeno( 1 ,2,3 -cd)pyrene
Dibenz(a,h)anthracene
C4-benzanthracene/chrysenes
Log
Ka
OW
3.356
3.8
3.223
4.012
4.3
4.208
4.8
4.534
4.571
4.72
5.3
5.04
5.2
4.922
5.084
5.46
5.7
5.287
5.92
5.673
5.713
6.32
6.14
6.107
6.135
6.135
6.266
6.291
6.429
6.507
6.94
6.722
6.713
7.36
Log
T^ b
J^BC
5.22
5.46
5.15
5.58
5.73
5.68
6.00
5.86
5.88
5.96
6.27
6.13
6.22
6.07
6.16
6.36
6.49
6.26
6.61
6.47
6.50
6.82
6.73
6.71
6.72
6.72
6.79
6.81
6.88
6.92
7.16
7.04
7.04
7.38
Log
KpED
2.93
3.40
2.79
3.62
3.93
3.83
4.45
4.17
4.21
4.37
4.98
4.70
4.87
4.58
4.75
5.14
5.40
4.96
5.63
5.37
5.41
6.05
5.86
5.82
5.85
5.85
5.99
6.02
6.16
6.24
6.70
6.47
6.46
7.14
Log
T^ d
•k-FDMS
2.86
3.22
2.75
3.40
3.64
3.56
4.05
3.83
3.86
3.99
4.47
4.25
4.39
4.16
4.29
4.60
4.80
4.46
4.98
4.78
4.81
5.32
5.17
5.14
5.16
5.16
5.27
5.29
5.41
5.47
5.83
5.65
5.64
6.18
Log
KpOM
2.79
3.24
2.66
3.45
3.74
3.65
4.25
3.98
4.02
4.17
4.75
4.49
4.65
4.37
4.53
4.91
5.16
4.74
5.38
5.13
5.17
5.78
5.60
5.57
5.60
5.60
5.73
5.75
5.89
5.97
6.41
6.19
6.18
6.83
2-13
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 2-1. Continued
Class
Other
Chemicals
Contaminants
Benzene
Delta-BHC
Gamma-BHC, Lindane
Biphenyl
4-Bromophenyl phenyl ether
Butyl benzyl phthalate
Chlorobenzene
Diazinon
Dibenzofuran
1 ,2-Dichlorobenzene
1 , 3 -Dichlorobenzene
1 ,4-Dichlorobenzene
Di-n-butyl phthalate
Dieldrin
Diethyl phthalate
Endosulfan mixed isomers
Alpha-Endosulfan
Beta-Endosulfan
Endrin
Ethylbenzene
Hexachloroethane
Malathion
Methoxychlor
Pentachlorobenzene
1 , 1 ,2,2-Tetrachloroethane
Tetrachloroethene
Tetrachloromethane
Toluene
Toxaphene
Tribromomethane (Bromoform)
1, 2, 4-Trichlorobenzene
1, 1, 1-Trichloroethane
Trichloroethene
m-Xylene
Log
KOW
2.13
3.78
3.73
3.96
5.00
4.84
2.86
3.70
4.07
3.43
3.43
3.42
4.61
5.37
2.50
4.10
3.83
4.52
5.06
3.14
4.00
2.89
5.08
5.26
2.39
2.67
2.73
2.75
5.50
2.35
4.01
2.48
2.71
3.20
Log
V b
J^BC
4.56
-
-
5.55
-
-
4.95
-
5.61
5.26
5.26
5.26
-
-
-
-
-
-
-
5.11
-
-
-
6.25
-
-
-
4.90
-
-
5.58
-
-
5.14
Log
KpED
1.65
3.38
3.33
3.57
4.66
4.49
2.41
3.30
3.68
3.01
3.01
3.00
4.25
5.05
2.04
3.72
3.43
4.16
4.72
2.71
3.61
2.44
4.74
4.93
1.92
2.21
2.28
2.30
5.19
1.88
3.62
2.01
2.26
2.77
Log
v d
J^PDMS
1.84
3.21
3.17
3.36
4.22
4.09
2.44
3.14
3.45
2.92
2.92
2.91
3.90
4.53
2.15
3.47
3.25
3.82
4.27
2.68
3.39
2.47
4.29
4.44
2.05
2.29
2.34
2.35
4.64
2.02
3.40
2.13
2.32
2.73
Log
KpOM
1.55
3.22
3.17
3.40
4.45
4.29
2.29
3.14
3.51
2.86
2.86
2.85
4.06
4.82
1.93
3.54
3.27
3.97
4.51
2.57
3.44
2.32
4.53
4.71
1.81
2.10
2.16
2.18
4.96
1.77
3.45
1.90
2.14
2.63
aFrom corresponding ESB documents.
b Derived using equation 2-4. KBc values not derived for non-planar molecules.
0 Derived using equation 2-10.
d Derived using equation 2-11.
e Derived using equation 2-12.
2-14
-------
Procedures for Determining Freely Dissolved Interstitial Water Concentrations
Table 2-2. Literature and calculated partition coefficients
Compound
Endrin
Dieldrin
Log K0wa
5.06
5.37
Log Koca'b
4.97
5.28
Log KDoca'c
4.13
4.44
a From corresponding ESB documents (U.S. EPA 2003b,c).
b From corresponding ESB documents using: Log Koc = 0.00028 + 0.983 x LogKow
(U.S. EPA2003b,c).
0 Derived using Equation 2-8.
Table 2-3. Solutions to Equation 2-7 using KDOC values calculated from Equation 2-8
DOC
(mg/L)
0
5
10
15
20
25
30
40
50
60
70
Endrin
(% freely dissolved)
100
94
88
83
79
75
71
65
60
55
51
Dieldrin
(% freely dissolved)
100
88
78
71
64
59
55
48
42
38
34
2-15
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 2-4. Advantages and disadvantages of selected approaches for determining freely
dissolved interstitial water concentrations of nonionic organic contaminants
Approach
Advantages
Disadvantages
One-Carbon
EqP Model
• Standard methods and guidance document
available to commercial laboratories
• Partition coefficients (e.g., K0w, K0c) and
uncertainties available in the scientific
literature
Simplistic model of NOC partitioning in
sediments
May substantially over-estimate Cd
Two-Carbon
EqP Model
• More complete and sophisticated model of
NOC partitioning in sediment
• Method for fBC determination for
commercial laboratories is available
• Data available for PAHs
Limited information for partition
coefficients (e.g., KBC) and their
uncertainties available in the scientific
literature (especially for non-planar
NOCs)
Uncertainty in the measurement of fBc
May under-estimate Cd
Direct
Measurement
• Direct determination of Cd
• Partition coefficient (e.g., KDOc) and related
uncertainties available in the scientific
literature
Large amounts of sediment required
Physical manipulation during
centrifugation of sediment may result in
artifacts that alter Cd
Methods are not standardized or
available to commercial laboratories
Passive Samplers
Passive Sampler:
Polyethylene
Devices
(PED)
• Inexpensive and rugged sampling
technology
• Growing acceptance in scientific literature
and in practical use
• Laboratory and field deployments possible
Increasing but limited amount of
information on partition coefficients
(e.g., KPED) and their uncertainties
available in the scientific literature
Determination of sampler equilibrium is
currently an area of research
Passive Sampler:
Solid Phase
Microextraction
(SPME)
Inexpensive sampling technology
Established acceptance in scientific
literature and in practical use
Laboratory and field deployments possible
Increasing amount of information on
partition coefficients (e.g., KPDMS) and
their uncertainties available in the
scientific literature
Determination of sampler equilibrium is
currently an area of research
Fibers are fragile and require protective
covering when used
Passive Sampler:
Polyoxymethylene
(POM)
• Inexpensive and rugged sampling
technology
• Growing acceptance in scientific literature
and in practical use
• Laboratory and field deployments possible
Increasing but limited amount of
information on partition coefficients
(e.g., KPQM) and their uncertainties
available in the scientific literature
Determination of sampler equilibrium is
currently an area of research
2-16
-------
Example Calculations of ESBTUFcv and IWTUFcv
Section 3
Example Calculations of ESBTUFcv
and IWTUFcv
3.1 Introduction
To assist users of this document, example
calculations for deriving interstitial water toxic
units are provided in Tables 3-1 and 3-2.
Although the examples provided are for PAHs,
they illustrate how chemical measurements
can be used to evaluate the acceptability of a
mixture of nonionic organic chemicals within
the technical framework of the EqP approach
(U.S. EPA, 2003a,d, 2008).
In these examples, the following values
represent analytically measured
concentrations: PAHs in sediment (|ig/g dry
weight), PAHs in interstitial water that was
generated by centrifugation (|ig/L), PAHs
absorbed to a passive sampler (|ig/g), TOC
(%) and BC (%) in sediment, and DOC in
interstitial water (mg/L). All other values were
calculated. The 34 PAHs presented in this
example constitute what is defined as "total
PAH" in U.S. EPA (2003d). Also listed are the
FCVs expressed on an organic carbon
normalized basis (Coc,pAHi,Fcvi) for each of the
34 PAHs from U.S. EPA (2003d). The
sediment sample for Example A is from a
manufactured gas plant site. The concentration
of total (34) PAHs is 257 |ig/g dry weight, the
TOC content is 3.75%, the BC content is
1.2%, and the interstitial water DOC is 11
mg/L. The fraction OC (foe) and BC (fee) are
calculated by dividing by 100 (i.e., %
TOC/100 or %BC/100, respectively). The
sediment sample for Example B is from a
nearby location. The concentration of total
PAHs is 39.4 |ig/g dry weight, the TOC is 3.2
% (foc = 0.032), the BC is 0.3 % (fBC = 0.003),
and interstitial water DOC is 5 mg/L.
3.2 Estimates of Freely Dissolved
Contaminants in Sediment Interstitial
Water
As discussed in Section 2, two equilibrium
partitioning models can be used to estimate the
concentration of freely dissolved chemical in
interstitial water based on the measured
concentration of nonionic organic chemicals in
bulk sediment. The first model, the one-carbon
model, considers contaminant partitioning
between interstitial water and TOC and
generates the conventional ESB (i.e., U.S.
EPA, 2003d, 2008). The second model, the
two-carbon model, considers partitioning
between interstitial water and two forms of
carbon: NSOC and BC. Because partitioning
to BC is greater than to NSOC, concentrations
of freely dissolved chemical in interstitial
water estimated using the two-carbon model
are lower than estimates from the one-carbon
model. Examples are provided for two
sediments, one with relatively higher
concentrations of BC in sediment and DOC in
interstitial water.
3.2.1 One- Carbon Model
In the first step of this approach (see U.S.
EPA, 2003d, 2008 for specifics), the dry
weight concentration of each PAH was divided
by the fraction of organic carbon to convert to
the organic carbon normalized concentration
(Coc, Hg PAH/g organic carbon). Second, the
organic-carbon-normalized concentration of
each PAH in the sediment was divided by its
organic carbon normalized PAH-specific FCV
(Coc,FAHi,Fcvi) to derive the conventional
ESBTUpcvi (U.S. EPA, 2003d). In both
sediments, none of the measured Coci exceed
3-1
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
the corresponding Coc, pAHi,Maxi, so solubility
constraints did not affect the calculation of
ESBTUpcvi (see solubility values for PAHs
Table 3-4 in U.S. EPA, 2003d). Next, the
ESBTUpcvi for the 34 PAHs were added to
derive the IESBTU for the 34 PAHs
(lESBTUFcv,34):
ZESBTUFCv,34 =
coci
COC,PAHi,PCVi
(3-1)
The £ESBTUFcv,34is 9.3 for Sediment A
(Table 3-1) and 1.7 for Sediment B (Table
3-2). A £ESBTUFcv,34 value greater than 1.0
indicates that the concentration of PAHs may
be non-protective to sensitive benthic
organisms. Based on this one-carbon model
assessment of toxic units, concentrations of
PAHs in both sediments could be chronically
toxic to sensitive benthic species. Note,
chemicals which have others modes of toxicity
that are more potent than narcosis (e.g.,
pesticides, phthalates) should be assessed on
an individual basis, it is not appropriate to
"sum" such chemicals unless there is evidence
that they exhibit additive toxicity.
As noted above, the approach for using the
one-carbon model followed the convention
described in the earlier U.S. EPA ESB
documents (U.S. EPA, 2003d; 2008) in which
the dry weight PAH concentrations were
converted to the organic carbon normalized
concentrations (Coc) and then divided by the
organic carbon normalized PAH-specific FCV
to derive the ESBTUpcv- An alternative
approach discussed in Section 2.1.1 and one
more similar to the approach used in the two-
carbon model discussed in the next section, is
to calculate the Cd using Equation 2-2:
r — ^
LH — ————
voc
this equation is equivalent to:
r = c°c
d ~ KQC
(3-2)
The Cd value calculated for each PAH would
then be divided by the CdpAHiFcvi from Table
3-1 rather than the Coc,pAHi,Fcvi to derive the
ESBTUpcvS. The two approaches will result in
the same number of interstitial water toxic
units for a given sediment but the approach
using Coc,FAHi,Fcviis currently more commonly
applied in assessing sediments when using the
one-carbon model.
3.2.2 Two-Carbon Model
Like the conventional XESBTUpcv,34
calculated using the one-carbon model in
Section 3.2.1, the sum of interstitial water
toxic units (£IWTUpcv) can also be calculated
using a two-carbon model that accounts for
association of PAHs with the fraction of BC
(fBc) and the fraction of NSOC (fNsoc) in
sediment. The elevated concentration of BC in
Sediment A (1.2% of total dry weight) is
within the range of levels observed at other
manufactured gas plant sites (Driscoll et al.,
2009). The concentration of BC in Sediment B
(0.3% of total dry weight) is representative of
levels observed near urban sources (-10% of
TOC) (Gustaffson and Gschwend, 1998;
Middelburg et al., 1999). As described in
Section 2.2.1, the two-carbon model is used to
calculate the freely dissolved concentration of
each PAH (Cd,pAHi) in interstitial water using
Equation 2-3:
C.r,
C, =
/wsoc KOC + /BC KBC
-n- 1
As discussed in Section 2, provisional
values for KBC used in this example were
derived from the relationship developed by
Driscoll et al. (2009).
3-2
-------
Example Calculations of ESBTUFcv and IWTUFcv
Because Cd appears on both sides of
Equation 2-3, an iterative approach must be
used to solve for Cd. Statistical protocols such
as the "Goal Seek" function in Excel® are
available for this purpose. The estimated
concentration of each PAH in interstitial water
(Cd,pAHi) was divided by its PAH-specific FCV
(Cd,pAHi,pcvi) to derive the IWTUFcvi (Equation
2-10). Note that the CdpAHiFcvi is a different
value than the CocpAHiFcvi used for calculating
the one-carbon model based toxic units. The
CdpAHiFcvi is a water-only based toxicity value
and can be a FCV, SCV, or any other relevant
water-only toxicity value, although toxic units
are additive only for contaminants with similar
modes of toxic action (e.g., narcosis for
nonionic organic contaminants). For PAHs,
CdpAHiFcvi values for the 34 compounds can be
found in U.S. EPA (2003d). In these examples,
none of the measured Cdi exceed the
corresponding Cd, pAHi,Maxi, so solubility
constraints did not affect the derivation of
IWTUpcvi for these sediments (U.S. EPA,
2003d). Based on the two-carbon model, the
total concentrations of dissolved PAHs in
Sediment A and B were 19 and 5.51 |ig/L,
respectively. The IWTUFCvi for the 34 PAHs
were added to derive the £IWTU for the 34
PAHs (ZIWTUFCV,34):
I,iWTUFCVi34 =
Jl Cd,PAHi,FCVi
(i o\
3-3)
The £IWTUFCV,34 for Sediment A was 1.4
(Table 3-1) and 0.5 for Sediment B (Table 3-
2). Although the £IWTUFCV,34 for Sediment A
(1.4) is less than the XIWTUFcv,34 calculated
using the one-carbon model (i.e., 9.3), the two-
carbon model still predicts that this sediment
could be chronically toxic to sensitive benthic
species. For Sediment B, the XIWTUFcv,34 of
0.5 indicates that the concentrations of PAHs
in this sediment are acceptable for the
protection of benthic organisms.
3.3 Measurement of Freely Dissolved
Contaminants in Sediment Interstitial
Water
This section presents two approaches that
can be used to directly measure concentrations
of nonionic organic contaminants in interstitial
water. In the first approach, interstitial water
produced by centrifugation is extracted with
organic solvent and resulting concentrations
are corrected for the fraction of total
contaminants associated with DOC. The
second approach uses a passive sampler placed
in sediment to measure concentrations of
freely dissolved contaminant in interstitial
water.
3.3.1 Direct Measurement of Interstitial
Water
In this approach, the £IWTU is calculated
from the total measured concentration of freely
dissolved PAH in interstitial water (Crw), after
correction for the fraction of the total
interstitial water concentration associated with
DOC. Total measured concentrations of
interstitial water PAHs were 6 and 8.8 |ig/L
for Sediments A and B, respectively (Tables
3-1, 3-2). As described in Section 2.3, the
percentage of the total interstitial water
concentration for each PAH that is freely-
dissolved (% Cd) is calculated using Equation
2-7:
- 100
Using the % Cd value derived for each
PAH, the freely dissolved concentration of
each PAH is calculated as:
Cri= C,
iw
\ 100 /
(3-4)
3-3
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Next, as shown in Section 3.2.2, the freely
dissolved concentration of each PAH (Cdi) is
divided by its PAH-specific FCV (Cd,pAHi,Fcvi)
to derive the IWTUFCvi In these examples,
none of the measured Cdi exceed the
corresponding Cd, pAHi,Maxi, so solubility
constraints do not affect the calculation of
IWTUpcvi for these sediments (U.S. EPA
2003d). The IWTUFCvi for the 34 PAHs were
added to derive the £IWTU for the 34 PAHs
(£IWTUFCv,34). The IWTUFCvi is 0.6 for
Sediment A (Table 3-1) and 0.7 for Sediment
B (Table 3-2). A £IWTUFCV,34 value less than
or equivalent to 1.0 indicates that
concentrations of PAHs in these sediments are
acceptable for the protection of benthic
invertebrates (U.S. EPA, 2003d).
3.3.2 Passive Sampling of Interstitial Water
As discussed in Section 2.4, various
passive samplers can be used to measure
concentrations of freely dissolved contaminant
in interstitial water. In the current examples,
data for concentrations of PAHs associated
with FED are used to determine the
concentration of freely dissolved PAHs in
interstitial water. For this example, the use of
PRCs loaded in the FED demonstrated the
target contaminants had achieved 100%
equilibrium between the sampler and the freely
dissolved phase. For actual deployments, it is
critical to understand the equilibrium status of
the samplers before performing the
calculations below. First, the concentrations
for each PAH associated with the FED (|ig/g
FED) is divided by it corresponding KPS-d, to
estimate the freely dissolved concentration in
interstitial water (|ig/L) (i.e., Equation 2-9):
a, =
-PS
The provisional KPS.d or KPED values used
in these examples were derived from the
relationship developed by Lohmann and Muir
(2010) for polyethylene devices:
Log KPED = -0.59 + 1.05 Log Kow (3-5)
Second, the freely dissolved concentration
of each PAH is divided by its PAH-specific
FCV (Cd,pAHi,Fcvi) to derive the IWTUFCvi. In
these sediments, none of the measured Cdi
exceed the corresponding Cd, pAHi,Maxi, so
solubility constraints do not affect the
calculation of IWTUFCVi (U.S. EPA, 2003d).
The IWTUFcvi for the 34 PAHs were then
added together to derive the ^IWTU for the 34
PAHs (XIWTUFCV,34). The XIWTUFCV,34 were
0.5 for Sediment A (Table 3-1) and 0.7 for
Sediment B (Table 3-2). IIWTUFCV,34 values
less than or equivalent to 1.0 indicate that the
concentrations of PAHs in these sediments are
acceptable for the protection of benthic
invertebrates (U.S. EPA, 2003d).
For both sediments, the direct
measurement of interstitial water, as well as
determinations of interstitial water
concentrations based on concentrations in a
polyethylene passive sampler, predicted a
similar number of total toxic units: 0.6 and 0.5,
respectively, for Sediment A, and 0.8 and 0.7,
respectively, for Sediment B. For both
approaches and sediments, concentrations of
PAHs are protective and not predicted to be
chronically toxic to sensitive benthic
organisms.
3.4 Considerations for Non-Planar
Contaminants
The two examples discussed above were
performed using PAHs. As planar compounds,
PAHs are expected to partition to black carbon
to a substantially higher degree than non-
planar compounds (Jonker and Koelmans,
2002; Cornelissen et al., 2005b). This is
because planar compounds can better interact
with the planar conformation of black carbon
and form stronger intermolecular associations.
This understanding is the basis for the two
carbon model described in Section 2.2.1. The
relationship between log K0w and log KBc
3-4
-------
(Equation 2-4) used to estimate KBc was
derived with PAHs. Consequently, at this time
Equation 2-4 should not be used for non-
planar nonionic organic chemicals; for
example, DDTs and their degradation
products, and ortho-substituted PCBs (also see
Table 2-1). To this end, log KBC values for
non-planar compounds are not included in
Table 2-1. If considering using the two-carbon
model with non-planar compounds, it is
recommended that compound-specific KBC
values be empirically determined.
3.5 Summary
These examples demonstrate the utility of
the procedures for determining the freely
dissolved concentrations of nonionic organic
chemicals used for deriving interstitial water
toxic units. The initial use of the one-carbon
general model to predict the likelihood of
Example Calculations of ESBTUFcv and IWTUFcv
toxicity resulted in both sediments being
designated as chronically toxic to benthic
organisms. Use of the approaches described in
this document results in a consistent reduction
in expected chronic toxicity from both
sediments. This downward trend in expected
toxicity concluded with a designation that both
sediments are protective of benthic organisms
and neither would cause chronic toxicity.
These examples also illustrate that deriving
interstitial water toxic units can be a useful
sediment assessment tool. However, IWTU
derivation is not inexpensive or particularly
simple and, as noted earlier, require scientific
expertise. In the next section of this document,
a proposed strategy for implementing the use
of IWTUs is discussed that relates the benefits
of using these approaches to the scientific
robustness and financial costs.
3-5
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 3-1. Example calculations of ESBTUpcv and IWTUFcv for PAH mixtures: Sediment A
PAH,
Naphthalene
Cl -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthene s/Pyrenes
B enz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghi]perylene
Total
ESB-Final Chronic Values a
ESB-FCV
COG, PAHi, FCVi
(MS/Soc)
385
444
510
581
657
452
491
538
611
686
769
596
594
670
746
829
913
707
697
770
841
844
929
1,008
1,112
1,214
979
981
965
967
967
1,115
1,123
1,095
-
ESB-FCV
Cd, PAHi, FCVi
(ng/L)
193.55
81.69
30.24
11.1
4.048
306.9
55.85
39.3
13.99
5.305
1.916
19.13
20.73
7.436
3.199
1.256
0.5594
7.109
10.11
4.887
2.227
2.042
0.8557
0.4827
0.1675
0.07062
0.6774
0.6415
0.9573
0.9008
0.9008
0.275
0.2825
0.4391
-
ESB-Maxi
COG, PAHi, Maxi
(Hg/goc)
61,700
-
-
-
-
24,000
33,400
26,000
-
-
-
34,300
1,300
-
-
-
-
23,870
9,090
-
4,153
826
-
-
-
-
2,169
1,220
3,840
431
4,300
-
2,389
648
-
ESB-Maxi
Cd, PAHi, Maxi
(ng/L)
30,995
-
-
-
-
16,314
3,800
1,900
-
-
-
1,100
45
-
-
-
-
239.9
131.9
-
11
2
-
-
-
-
1.501
0.7999
3.810
0.4012
4.012
-
0.6012
0.2600
-
3-6
-------
Example Calculations of ESBTUFcv and IWTUFcv
Table 3-1. Continued
PAH,
Naphthalene
C 1 -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthene s/Pyrenes
B enz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghijperylene
Total
One-Carbon Model
Cp^AHi
Measured
Sediment Cone
(Hg/g dry wt)
5.0
1.0
1.6
1.7
1.1
4.9
2.4
4.4
2.3
1.8
1.0
18.0
10.0
16.0
8.6
2.6
1.1
28.0
23.0
23.0
17.0
16.0
7.7
2.9
1.1
0.8
14.0
8.3
13.0
2.8
7.9
4.2
1.4
2.7
257
CoC.PAHi
Measured
Sediment Cone
(Hg/goc)
133
27
43
45
29
131
64
117
61
48
26
480
267
427
229
69
29
747
613
613
453
427
205
77
29
22
373
221
347
75
211
112
37
72
-
ESBTUFCVl
0.3
0.06
0.08
0.08
0.04
0.3
0.1
0.2
0.1
0.07
0.03
0.8
0.4
0.6
0.3
0.08
0.03
1.1
0.9
0.8
0.5
0.5
0.2
0.08
0.03
0.02
0.4
0.2
0.4
0.08
0.2
0.1
0.03
0.07
9.3
3-7
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 3-1. Continued
PAH,
Naphthalene
C 1 -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthene s/Pyrenes
B enz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghijperylene
Total
Two Carbon Model
LogKoca
(L/Kg NSOC)
3.299
3.736
4.227
4.719
5.21
3.168
3.944
4.137
4.64
5.112
5.603
4.494
4.457
4.955
5.367
5.82
6.213
4.998
4.839
5.197
5.577
5.616
6.036
6.32
6.822
7.235
6.16
6.184
6.003
6.031
6.031
6.608
6.599
6.397
-
LogKBCb
(L/Kg BC)
5.24
5.48
5.75
6.02
6.29
5.16
5.59
5.70
5.98
6.24
6.51
5.90
5.88
6.15
6.38
6.63
6.85
6.18
6.09
6.29
6.50
6.52
6.75
6.91
7.19
7.41
6.82
6.83
6.73
6.75
6.75
7.07
7.06
6.95
-
Cd.PAHi
Estimated
Freely Dissolved
IWConc
(ng/L)
4.07
0.12
0.09
0.03
0.01
5.21
0.31
0.55
0.066
0.016
0.002
2.42
1.04
0.75
0.12
0.006
0.0007
1.61
1.68
0.77
0.215
0.180
0.023
0.0027
0.00019
0.00005
0.05
0.019
0.056
0.0046
0.024
0.0026
0.0004
0.0020
19
IWTUFCVI
0.02
0.001
0.003
0.003
0.001
0.02
0.01
0.01
0.005
0.003
0.001
0.1
0.05
0.1
0.04
0.01
0.001
0.2
0.2
0.2
0.1
0.09
0.03
0.01
0.001
0.001
0.07
0.03
0.06
0.01
0.03
0.01
0.002
0.005
1.4
3-8
-------
Example Calculations of ESBTUFcv and IWTUFcv
Table 3-1. Continued
PAH,
Naphthalene
C 1 -Naphthalenes
C2 -Naphthalenes
C3 -Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrene s/Anthracenes
C2 -Phenanthrene s/Anthracene s
C3 -Phenanthrene s/Anthracene s
C4 -Phenanthrene s/Anthracene s
Fluoranthene
Pyrene
C 1 -Fluoranthenes/Pyrenes
Benz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghijperylene
Total
Directly Measured Interstitial Water
Ciw, PAHi
Total Measured
IW Cone
(ng/L)
0.25
0.080
0.270
0.170
0.099
0.043
1.300
1.100
0.120
0.073
0.065
0.790
0.180
0.200
0.140
0.099
0.081
0.370
0.300
0.160
0.110
0.093
0.006
0.006
0.006
0.006
0.063
0.031
0.056
0.011
0.052
0.020
0.006
0.021
6
Log KDOcc
(L/Kg DOC)
2.44
2.88
3.38
3.87
4.37
2.31
3.09
3.29
3.79
4.27
4.76
3.65
3.61
4.11
4.53
4.98
5.38
4.15
3.99
4.35
4.74
4.78
5.20
5.48
5.99
6.41
5.32
5.35
5.17
5.19
5.19
5.77
5.77
5.56
-
Fraction
Freely
Dissolved
0.997
0.992
0.974
0.924
0.796
0.998
0.987
0.979
0.936
0.831
0.611
0.954
0.957
0.876
0.731
0.487
0.276
0.865
0.902
0.801
0.625
0.604
0.365
0.229
0.085
0.034
0.302
0.290
0.383
0.368
0.368
0.132
0.135
0.200
-
Cd.PAHi
Estimated
Freely
Dissolved
IWConc
(ng/L)
0.249
0.079
0.263
0.157
0.079
0.043
1.283
1.077
0.112
0.061
0.040
0.753
0.172
0.175
0.102
0.048
0.022
0.320
0.271
0.128
0.069
0.056
0.002
0.001
0.0005
0.0002
0.019
0.009
0.021
0.004
0.019
0.003
0.001
0.004
-
IWTUFcv,
0.001
0.001
0.009
0.01
0.02
0.0001
0.02
0.03
0.008
0.01
0.02
0.04
0.008
0.02
0.03
0.04
0.04
0.05
0.03
0.03
0.03
0.03
0.002
0.003
0.003
0.003
0.03
0.01
0.02
0.004
0.02
0.01
0.003
0.01
0.6
3-9
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 3-1. Continued
PAH,
Naphthalene
Cl -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthene s/Pyrenes
B enz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghi]perylene
Total
Passive Sampler
Passive
Sampler Cone
(Hg/g PE)
0.16
0.18
2.02
4.26
5.95
0.02
4.66
6.87
2.09
4.08
8.94
11.02
2.07
8.06
12.51
18.17
29.68
1.40
10.98
17.38
12.79
11.02
1.22
1.59
2.04
1.51
13.66
17.62
8.64
2.27
7.82
3.82
1.44
5.42
-
LogKPEDd
(L/Kg PE)
2.93
3.40
3.93
4.45
4.98
2.79
3.62
3.83
4.37
4.87
5.40
4.21
4.17
4.70
5.14
5.63
6.05
4.75
4.58
4.96
5.37
5.41
5.86
6.16
6.70
7.14
5.99
6.02
5.82
5.85
5.85
6.47
6.46
6.24
-
Cd,pAHi Estimated
Freely Dissolved
IW Cone fog/L)
0.19
0.07
0.24
0.151
0.063
0.03
1.11
1.02
0.09
0.055
0.036
0.68
0.14
0.16
0.09
0.043
0.0267
0.03
0.29
0.19
0.055
0.043
0.002
0.0011
0.00041
0.00011
0.01
0.017
0.013
0.0032
0.011
0.0013
0.0005
0.0031
-
IWTUFCVl
0.001
0.001
0.008
0.01
0.02
0.0001
0.02
0.03
0.006
0.01
0.02
0.04
0.007
0.02
0.03
0.03
0.05
0.004
0.03
0.04
0.02
0.02
0.002
0.002
0.002
0.002
0.02
0.03
0.01
0.004
0.01
0.005
0.002
0.007
0.5
Note: Characteristics of the example sediment are: TOC (3.75%), DOC (11 mg/L), and BC (1.2%).
a Values are from U.S. EPA (2003d).
b Values are derived from relationship developed by Driscoll et al. (2009).
0 Values are derived from relationship developed by Burkhard (2000).
d Values are derived from relationship developed by Lohmann and Muir (2010) (Equation 3-5).
3-10
-------
Example Calculations of ESBTUFcv and IWTUFcv
Table 3-2. Example calculations of ESBTUpcv and IWTUFcvfor PAH mixtures: Sediment B
PAH,
Naphthalene
Cl -Naphthalenes
C2-Naphthalenes
C3 -Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrene s/Anthracenes
C2 -Phenanthrene s/Anthracene s
C3 -Phenanthrene s/Anthracene s
C4 -Phenanthrene s/Anthracene s
Fluoranthene
Pyrene
C 1 -Fluoranthenes/Pyrenes
Benz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo [b] fluoranthene
Benzo[k]fluoranthene
Benzo [a]pyrene
Perylene
Benzo [e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghi]perylene
Total
ESB-Final Chronic Values a
ESB-FCV
COG, PAHi, FCVi
(HS/SOC)
385
444
510
581
657
452
491
538
611
686
769
596
594
670
746
829
913
707
697
770
841
844
929
1,008
1,112
1,214
979
981
965
967
967
1,115
1,123
1,095
-
ESB-FCV
Cd, PAHi, FCVi
(ng/L)
193.5
81.69
30.24
11.1
4.048
306.9
55.85
39.3
13.99
5.305
1.916
19.13
20.73
7.436
3.199
1.256
0.5594
7.109
10.11
4.887
2.227
2.042
0.8557
0.4827
0.1675
0.07062
0.6774
0.6415
0.9573
0.9008
0.9008
0.275
0.2825
0.4391
-
ESB-Maxi
COG, PAHi, Maxi
(Hg/goc)
61,700
-
-
-
-
24,000
33,400
26,000
-
-
-
34,300
1,300
-
-
-
-
23,870
9,090
-
4,153
826
-
-
-
-
2,169
1,220
3,840
431
4,300
-
2,389
648
-
ESB-Maxi
Cd, PAHi, Maxi
(ng/L)
30,995
-
-
-
-
16,314
3,800
1,900
-
-
-
1,100
45
-
-
-
-
239.9
131.9
-
11
2
-
-
-
-
1.501
0.7999
3.810
0.4012
4.012
-
0.6012
0.2600
-
3-11
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 3-2. Continued
PAH,
Naphthalene
C 1 -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthene s/Pyrenes
B enz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghijperylene
Total
One-Carbon Model
Cp,FAHi
Measured
Sediment Cone
(Hg/g dry wt)
0.2
0.4
0.8
0.6
0.2
0.4
0.7
0.5
0.3
0.3
0.1
3.6
1.2
2.3
1.3
0.4
0.3
3.4
5.6
3.5
2.3
2.1
1.0
0.3
0.1
0.1
1.6
0.6
2.1
0.5
1.1
0.7
0.2
0.7
39.4
CoC.PAHi
Measured
Sediment Cone
(Hg/goc)
7
14
24
18
7
14
23
15
10
9
3
113
38
72
41
11
10
106
175
109
72
66
30
8
4
3
50
19
66
15
34
22
7
21
-
ESBTUpcvi
0.02
0.03
0.05
0.03
0.01
0.03
0.05
0.03
0.02
0.01
0.004
0.2
0.06
0.1
0.05
0.01
0.01
0.2
0.3
0.1
0.09
0.08
0.03
0.01
0.003
0.003
0.05
0.02
0.07
0.02
0.04
0.02
0.01
0.02
1.7
3-12
-------
Example Calculations of ESBTUFcv and IWTUFcv
Table 3-2. Continued
PAH,
Naphthalene
Cl -Naphthalenes
C2 -Naphthalenes
C3 -Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthenes/Pyrenes
Benz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo [k] fluoranthene
Benzo[a]pyrene
Perylene
Benzo [e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghi]perylene
Total
Two-Carbon Model
Log KOC a
(L/Kg NSOC)
3.299
3.736
4.227
4.719
5.21
3.168
3.944
4.137
4.64
5.112
5.603
4.494
4.457
4.955
5.367
5.82
6.213
4.998
4.839
5.197
5.577
5.616
6.036
6.32
6.822
7.235
6.16
6.184
6.003
6.031
6.031
6.608
6.599
6.397
-
Log KBC b
(L/Kg BC)
5.24
5.48
5.75
6.02
6.29
5.16
5.59
5.70
5.98
6.24
6.51
5.90
5.88
6.15
6.38
6.63
6.85
6.18
6.09
6.29
6.50
6.52
6.75
6.91
7.19
7.41
6.82
6.83
6.73
6.75
6.75
7.07
7.06
6.95
-
Cd,pAHi Estimated
Freely Dissolved
PW Cone
(ng/L)
0.22
0.26
0.21
0.05
0.00
0.87
0.37
0.12
0.02
0.01
0.00
1.14
0.25
0.22
0.04
0.00
0.00
0.35
0.98
0.23
0.06
0.05
0.01
0.00
0.00
0.00
0.01
0.00
0.02
0.00
0.01
0.00
0.00
0.00
5.51
IWTUFcVl
0.001
0.003
0.007
0.004
0.001
0.003
0.007
0.003
0.002
0.001
0.0002
0.06
0.01
0.03
0.01
0.002
0.001
0.05
0.1
0.05
0.03
0.02
0.007
0.001
0.0003
0.0002
0.01
0.003
0.02
0.002
0.008
0.004
0.001
0.004
0.5
3-13
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
Table 3-2. Continued
PAH,
Naphthalene
Cl -Naphthalenes
C2-Naphthalenes
CS-Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1-Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3-Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1-Fluoranthenes/Pyrenes
B enz [a] anthracene
Chrysene
C 1-Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Perylene
Benzo[e]pyrene
Indeno[ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghi]perylene
Total
Directly Measured Interstitial Water
Ciw, PAHi
Measured
Total IW Cone
(ng/L)
0.92
1.30
1.10
0.48
0.15
0.02
1.90
0.46
0.20
0.11
0.01
0.44
0.18
0.16
0.14
0.06
0.04
0.19
0.36
0.16
0.05
0.06
0.03
0.01
0.01
0.01
0.04
0.02
0.03
0.07
0.03
0.02
0.01
0.03
8.8
Log KDOCC
(L/Kg
DOC)
2.44
2.88
3.38
3.87
4.37
2.31
3.09
3.29
3.79
4.27
4.76
3.65
3.61
4.11
4.53
4.98
5.38
4.15
3.99
4.35
4.74
4.78
5.20
5.48
5.99
6.41
5.32
5.35
5.17
5.19
5.19
5.77
5.77
5.56
-
Fraction
Freely
Dissolved
0.999
0.996
0.988
0.964
0.896
0.999
0.994
0.990
0.970
0.915
0.775
0.978
0.980
0.940
0.856
0.676
0.456
0.934
0.953
0.898
0.786
0.770
0.559
0.396
0.170
0.073
0.487
0.473
0.577
0.561
0.561
0.251
0.255
0.354
-
Cd,pAHi Estimated
Freely Dissolved
IWConc
(HS/L)
0.919
1.297
1.091
0.468
0.139
0.018
1.892
0.457
0.196
0.103
0.005
0.434
0.178
0.153
0.126
0.041
0.020
0.177
0.343
0.144
0.037
0.045
0.017
0.002
0.001
0.004
0.018
0.007
0.018
0.039
0.021
0.005
0.001
0.009
-
IWTUFCVI
0.005
0.02
0.04
0.04
0.03
0.0001
0.03
0.01
0.01
0.02
0.002
0.02
0.009
0.02
0.04
0.03
0.04
0.02
0.03
0.03
0.02
0.02
0.02
0.005
0.006
0.006
0.03
0.01
0.02
0.04
0.02
0.02
0.005
0.02
0.7
3-14
-------
Example Calculations of ESBTUFcv and IWTUFcv
Table 3-2. Continued
PAH,
Naphthalene
Cl -Naphthalenes
C2-Naphthalenes
C3 -Naphthalenes
C4-Naphthalenes
Acenaphthylene
Acenaphthene
Fluorene
Cl-Fluorenes
C2-Fluorenes
C3-Fluorenes
Phenanthrene
Anthracene
C 1 -Phenanthrenes/Anthracenes
C2-Phenanthrenes/Anthracenes
C3 -Phenanthrenes/Anthracenes
C4-Phenanthrenes/Anthracenes
Fluoranthene
Pyrene
C 1 -Fluoranthenes/Pyrenes
Benz [a] anthracene
Chrysene
C 1 -Benzanthracenes/Chrysenes
C2- Benzanthracenes/Chrysenes
C3- Benzanthracenes/Chrysenes
C4- Benzanthracenes/Chrysenes
Benzo[b]fluoranthene
Benzo [k] fluoranthene
Benzo[a]pyrene
Perylene
Benzo [e]pyrene
Indeno [ 1 ,2,3 -cd]pyrene
Dibenz [a,h] anthracene
Benzo [ghijperylene
Total
Passive Sampler
Passive
Sampler Cone
(Hg/g PE)
0.66
2.50
8.58
14.37
10.29
0.03
7.21
2.16
2.81
7.49
2.23
6.43
2.31
7.40
15.43
28.74
34.46
9.35
15.18
18.39
5.35
30.24
11.51
8.68
4.98
13.74
6.73
9.33
19.93
20.61
4.98
2.94
2.88
27.96
LogKPEDd
(L/Kg PE)
2.93
3.40
3.93
4.45
4.98
2.79
3.62
3.83
4.37
4.87
5.40
4.21
4.17
4.70
5.14
5.63
6.05
4.75
4.58
4.96
5.37
5.41
5.86
6.16
6.70
7.14
5.99
6.02
5.82
5.85
5.85
6.47
6.46
6.24
-
Cd,pAHi Estimated
Freely Dissolved
IW Cone (jig/L)
0.767
0.997
1.020
0.510
0.109
0.045
1.720
0.320
0.121
0.101
0.009
0.397
0.156
0.147
0.111
0.068
0.031
0.167
0.401
0.201
0.023
0.118
0.016
0.006
0.001
0.001
0.007
0.009
0.030
0.029
0.007
0.001
0.001
0.016
-
IWTUFcvi
0.004
0.01
0.03
0.05
0.03
0.0001
0.03
0.008
0.009
0.02
0.005
0.02
0.008
0.02
0.03
0.05
0.06
0.02
0.04
0.04
0.01
0.06
0.02
0.01
0.006
0.01
0.01
0.01
0.03
0.03
0.008
0.004
0.004
0.04
0.7
Note: Characteristics of the example sediment are: TOC (3.2%), DOC (5 mg/L), and BC (0.3%).
a Values are from U.S. EPA (2003d).
b Values are derived from relationship developed by Driscoll et al. (2009).
0 Values are derived from relationship developed by Burkhard (2000).
d Values are derived from relationship developed by Lohmann and Muir (2010) (Equation 3-5).
3-15
-------
-------
Implementation of Freely Dissolved Interstitial Water Concentrations
Section 4
Implementation of
Freely Dissolved Interstitial Water
Concentrations
4.1 Introduction
A typical component of assessing
contaminated sediment sites is the collection of
sediment samples for chemical analysis to
determine the likelihood that sediment
contamination will result in adverse
toxicological effects to the benthos. For several
years, the use of sediment quality guidelines,
like the ESBs (U.S. EPA, 2003b,c,d, 2008), have
been one line of evidence for performing these
assessments. Depending upon whether toxic
effects are suspected or demonstrated based on a
sediment assessment, the site may need to be
remediated via dredging, natural monitored
recovery or capping (U.S. EPA, 2005). As
discussed previously in this document, because
of technological advances, the use of the freely
dissolved concentrations of contaminants in the
interstitial water may result in a more accurate
assessment of toxic effects than the one-carbon
general model used to derive ESBs for nonionic
organic chemicals.
As discussed in Table 2-4, currently, the use
of ESBs based on the one-carbon model may
over-predict Cd and be more environmentally
conservative (i.e., protective) and less expensive
than using toxic units derived based on the freely
dissolved concentrations in sediment interstitial
water discussed in this document. The two
carbon EqP model discussed in this document
may result in less environmentally conservative
assessments and requires the measurement of
sedimentary black carbon. At this time, black
carbon measurements are not commonly
performed by most commercial environmental
chemistry laboratories limiting their practicality.
Further, the two carbon model is applicable for
all classes of nonionic organic chemicals only if
the appropriate KBc values are available for the
model calculations. Currently, many KBc values
are derived using linear free energy relationships
between KBC and KOW for planar chemicals like
PAHs but not non-planar compounds. As noted
earlier, these KBcs may not be appropriate for
use with non-planar compounds and may result
in elevated uncertainties in estimates of freely
dissolved concentrations. Similarly, the
collection of sufficient sediment interstitial water
for direct measurement of contaminants
continues to include significant artifacts and
requires additional chemical analyses that may
not always be cost effective (e.g., DOC,
interstitial water). Currently, the analysis of
whole or bulk contaminated sediments for a suite
of nonionic organic chemicals is less expensive
than performing a similar chemical analysis for
passive samplers, such as SPME, PEDs or POM.
However, the difference in cost is rapidly
decreasing. Further, the number of laboratories
capable of performing the analysis on passive
samplers is currently limited but is also growing
as the methods become more established.
Considering these advantages and disadvantages,
implementing the use of the approaches for
determining the freely dissolved concentrations
of sediment interstitial water chemicals
discussed in this document requires the
scientifically-informed balancing of
environmental protection and cost. However, the
4-1
-------
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations
additional costs incurred in improved sediment
characterization may be offset by the reduced
costs for risk management and remediation.
Further, an implementation framework for
performing sediment assessments should
seriously consider a weight of evidence
approach.
4.2 Implementation of Freely Dissolved
Concentrations
Figure 4-1 illustrates a potential tiered
approach for implementing the use of freely
dissolved interstitial water concentrations while
balancing environmental protection, risk
management and cost. The first tier involves
using the one-carbon EqP-based ESB to assess
the likelihood of toxicity to the benthos.
Sediments in which the ESB is not exceeded are
considered environmentally unimpacted and
require no further consideration based on the
ESB line of evidence. However, this is a very
important conclusion with the potential to carry
significant implications for the entire site
assessment. Consequently, it is critical that every
scientific effort (e.g., high data quality, robust
analytical chemistry) should be taken to insure
this conclusion is accurate. Sediments in which
the ESB is exceeded using the one-carbon EqP
model are considered as representing a possible
risk to the benthos and may require remediation.
However, the second tier of this implementation
approach may be performed on sediments which
exceed the conventional EqP-based ESB. In the
second tier, a passive sampler may be used to
generate interstitial water toxic units (Section
2.4). If for a given contaminant, the interstitial
water toxic units are greater than one, the
sediments are considered as representing a
possible risk to the benthos and may require
remediation. Like the first tier, if the interstitial
water toxic units in Tier 2 are not greater than
one, the sediment is considered environmentally
unimpacted and requires no further consideration
based on the ESB line of evidence. Again, like
the decision made in Tier 1, this is a very
important conclusion with the potential to carry
significant implications for the entire site
assessment. It is critical that every scientific
effort (e.g., high data quality, robust analytical
chemistry) is taken to insure this conclusion is
accurate.
In the second tier, the two carbon model can
also be used to generate interstitial water toxic
units (Section 2.2). However, given the
uncertainty around the KBc values currently
available in the scientific literature as well as the
measurement of fee, the use of passive samplers
in Tier 2 is recommended over the two carbon
model. In addition, as discussed above, because
of the continued difficulties and costs associated
with collecting interstitial water, a similar
recommendation is made for using a passive
sampler based measure to generate interstitial
water toxic units rather than a direct
measurement of interstitial water (Section 2.3).
Using the passive sampler approach, if the
IWTUs exceed 1.0, sediment toxicity testing can
be conducted in the third tier to verify the
findings of the first two tiers. The cost of testing
and analysis in the third tier is likely to be the
greatest compared to the others and may require
collection and chemical analysis of more
sediment (and possibly interstitial water).
However, the performance of sediment toxicity
testing with sensitive organisms is one of the
most data rich and accurate lines of evidence to
assess for the adverse effects of sediment
contaminants on the benthos. It should be
recognized that if a whole sediment toxicity test
finds significant toxicity, the cause or causes
may be toxic chemicals other than those
measured in Tiers 1 and 2. Finally, in a
recommended weight of evidence sediment
assessment, toxicity testing should not be used in
exclusion of other lines of evidence including
chemistry, bioaccumulation, TIE, and benthic
community analyses. It is highly recommended
that the data generated in Tier 3, as well as Tiers
1 and 2, of the proposed tiered approach be
informed by lines of evidence in addition to
acute and chronic toxicity testing. This
4-2
-------
Implementation of Freely Dissolved Interstitial Water Concentrations
consideration emphasizes the merits of a weight
of evidence approach when performing
contaminated sediment assessments.
4.3 Research Needs
As discussed in Section 1.5, there remain
several areas of research and development for
the approaches for determining interstitial water
concentrations discussed in this document. Most
of these areas involve making a better measure
of (i.e., standardizing) or reducing the
uncertainty associated with the array of partition
coefficients used in these approaches (e.g., KBc,
KPED, KPDMS, KPOM))- A second area is improving
the measurement of black carbon in sediments.
Currently, the commonly used version of this
measurement involving the removal of inorganic
and NSOC (Gustafsson et al., 1997) has been
shown to be highly variable in inter-laboratory
comparisons (Gustafsson et al., 2001). In
addition, for the passive samplers, it is critical to
develop improvements in the methods for the
determination of when equilibrium between
chemicals in the interstitial water and the passive
sampler has been established.
Figure 4-1. Schematic of proposed tiered approach for implementing the use of the freely dissolved
interstitial water concentrations of nonionic organic chemicals (based on Burgess, 2009).
Tier 3
Tierl
Assessment of
Chronic Toxicity
based on
Chronic Toxicity Testing
{e.g., Amphipod toxicity tests)
t
Assessment of
Chronic Toxicity
based on Measurement
of Freely Dissolved
Interstitial Water
Concentrations
Probable Risk
of Adverse
Effects
Interstitial Water
Toxic Units > 1.0?
Assessment of
Chronic Toxicity
based on
OneCarbon-EqP Model
Yes
Is ESB Exceeded?
No
4-3
-------
-------
Section 5
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