EPA/600/R-02/012 | December 2012

www.epa.gov/ord

Office of Research and Development

National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division

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 | 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 

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. 

iii 
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 
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., polycyclic 

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 

 

 

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 

vii 
 

 

 

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 

viii 
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 ESBTUFCV and IWTUFCV for PAH mixtures: ...................... 3-6 

Table 3-2. Example calculations of ESBTUFCV and IWTUFCV 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 

ix 
 

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 

x 
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 (KOC), 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 μg chemical/g 

organic carbon basis (μg/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 μg/gOC basis can be accurately predicted by multiplying the 

effect concentration in water by the chemical’s KOC. 

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 
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 standalone, 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. 

xii 
 

 

 

 

Glossary of Abbreviations 

Glossary of Abbreviations 

ASTM American Society for Testing and Materials 

BC Black carbon 

C18 Octadecyl matrix used in solid chromatography 

Cd Freely-dissolved interstitial water concentration of contaminant 

Cd,PAHi,FCVi 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 

CTd 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 ESBOC) 

unless otherwise specified 

ESBTU Equilibrium Partitioning Sediment Benchmark Toxic Units 

ESBTUFCV Equilibrium Partitioning Sediment Benchmark Toxic Units based on the Final Chronic 

Value 

FCV Final chronic value 

fBC Fraction of black carbon in sediment 

fNSOC Fraction of natural sedimentary organic carbon 

fOC Fraction of organic carbon in sediment 

GC/MS Gas chromatograph/mass spectrometer 

gOC Gram organic carbon 

xiii 
 

 

 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

IW Interstitial water (also known as pore water) 

IWTU Interstitial water toxic units 

IWTUFCV 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 

KOW Octanol-water partition coefficient 

KP Sediment-water partition coefficient 

KPDMS Polydimethylsiloxane-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 

PED 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 

xiv 
Introduction 

Section 1 

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 wateronly data) and are not limited to the equilibrium 

partitioning sediment benchmarks for endrin and 

dieldrin (U.S. EPA, 2003b,c), mixtures of 

polycyclic 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 freelydissolved 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 μg chemical/g sediment organic 

carbon (μg/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 
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 
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 KOW 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 
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 “onecarbon” 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 empiricallyderived 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) (μg/L) of a nonionic 

organic contaminant can be calculated as 

follows: 

(2-1)

ು

ು

௄

஼ ܥௗ ൌ

where, CP is the particulate contaminant 

concentration (μg/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 

(fOC) (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: 

஼ು ܥௗ ൌ (2-2) ௙ೀ಴௄ೀ಴

Using Equation 2-2, the conventional ESB 

can be determined. For example, for PAH 

mixtures (U.S. EPA, 2003d), in this onecarbon 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 onecarbon general model, ESBTUs are often 

derived from the quotient of the contaminant 

organic carbon normalized sediment 

concentration (μg/Kg OC) and the organic 

carbon normalized toxicity value (μg/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. 

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 twocarbon 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 carbonnormalized 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 “onecarbon” 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; AccardiDey 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 
 

 

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 (AccardiDey 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 (375oC, 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 (fBC) 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: 

஼ು (2-3) ೙ష	భ ೏ ௙ಿೄೀ಴ ௄ೀ಴ା	௙ಳ಴ ௄ಳ಴ ஼

ܥௗ ൌ	

where, fNSOC is the weight fraction of NSOC in 

sediment (Kg NSOC/Kg dry), calculated from 

the difference between TOC and BC, fBC 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 KOWs 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 (AccardiDey 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 twocarbon 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 
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: 

(2-4) ைௐ ݃݋ܮ	0.54ܭ ൅ 3.41 ൌ ஻஼ ݃݋ܮܭ

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 SchubauerBerigan, 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 overestimation 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 subsampling 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 μg/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 log KOW 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 KOW 

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, overestimating 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): 

(2-5) ஼ವೀ಴ ܭ஽ை஼ ൌ	

஼೏

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 C18 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 
 

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 freelydissolved 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) ஼಺ೈ ܥௗ ൌ	 ሺ஽ை஼ ௄ವೀ಴ା	ଵሻ

as can the percentage of the total 

contaminant that is freely dissolved (% Cd): 

100 (2-7)

ವೀ಴ା	ଵሻ

ଵ

௄ሺ஽ை஼ %	ܥௗ ൌ	

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:

As noted earlier, KOW 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 

(2-8) ைௐܭ Log 0.99 + -0.88஼ = ை஽ܭ ݃݋ܮ

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 μm 

(Fernandez et al., 2009a,b) to >100 μm, 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) 

2-8 
Procedures for Determining Freely Dissolved Interstitial Water Concentrations 

incorporated into the PED (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 PED 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, PED 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 PED 

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 precleaned PED with a sediment slurry, with time 

to reach equilibrium ranging from 1 to 60 days 

for chemicals with log KOW < 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 concentrations. 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 concentrations 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: 

(2-9) ஼ುೄ ܥௗ ൌ ௄ುೄష೏

where, CPS is the passive sampler 

concentration of a chemical (μg/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 literaturebased 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: 

ைௐܭ ݃݋ܮ	1.05 ൅ 0.59െ ൌ஽ ா௉ܭ ݃݋ܮ

(2-10) 

ைௐܭ ݃݋ܮ	0.83 ൅ 0.07 ൌ ெௌ௉஽ܭ ݃݋ܮ

 (2-11) 

ைௐܭ ݃݋ܮ	1.01 ൅ 0.60	െ ൌ ைெ௉ܭ ݃݋ܮ

 (2-12) 

2-10 
Procedures for Determining Freely Dissolved Interstitial Water Concentrations 

Several authors have reported that as KOW 

increases, the relationship between KPS-d and 

KOW begins to demonstrate curvilinear 

behavior often at log KOWs 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 commerciallyavailable 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 (CTd) that includes both the 

freely dissolved PAHs and some PAHs 

associated with DOC. Because contaminant 

KDOC increases with hydrophobicity, this overestimate 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 twoand 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 CTd 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 wateronly 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 

IWTUFCV (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: 

஼೏ ܫܹܷܶி஼௏ ൌ	

ி஼௏ (2-13) 

or 

஼೏

ൌ	 ܫܹܷܶௌ஼௏ ௌ஼௏ (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 Contaminants 

Log 

KOW a 

Log 

KBC 

b 

Log 

KPED c 

Log 

KPDMS 

d 

Log 

KPOM e 

PAHs Naphthalene 3.356 5.22 2.93 2.86 2.79 

C1-naphthalenes 3.8 5.46 3.40 3.22 3.24 

Acenaphthylene 3.223 5.15 2.79 2.75 2.66 

Acenaphthene 4.012 5.58 3.62 3.40 3.45 

C2-naphthalenes 4.3 5.73 3.93 3.64 3.74 

Fluorene 4.208 5.68 3.83 3.56 3.65 

C3-naphthalenes 4.8 6.00 4.45 4.05 4.25 

Anthracene 4.534 5.86 4.17 3.83 3.98 

Phenanthrene 4.571 5.88 4.21 3.86 4.02 

C1-fluorenes 4.72 5.96 4.37 3.99 4.17 

C4-naphthalenes 5.3 6.27 4.98 4.47 4.75 

C1-phenanthrene/anthracenes 5.04 6.13 4.70 4.25 4.49 

C2-fluorenes 5.2 6.22 4.87 4.39 4.65 

Pyrene 4.922 6.07 4.58 4.16 4.37 

Fluoranthene 5.084 6.16 4.75 4.29 4.53 

C2-Phenanthrene/anthracenes 5.46 6.36 5.14 4.60 4.91 

C3-fluorenes 5.7 6.49 5.40 4.80 5.16 

C1-pyrene/fluoranthenes 5.287 6.26 4.96 4.46 4.74 

C3-phenanthrene/anthracenes 5.92 6.61 5.63 4.98 5.38 

Benz(a)anthracene 5.673 6.47 5.37 4.78 5.13 

Chrysene 5.713 6.50 5.41 4.81 5.17 

C4-Phenanthrenes/anthracenes 6.32 6.82 6.05 5.32 5.78 

C1-Benzanthracene/chrysenes 6.14 6.73 5.86 5.17 5.60 

Benzo(a)pyrene 6.107 6.71 5.82 5.14 5.57 

Perylene 6.135 6.72 5.85 5.16 5.60 

Benzo(e)pyrene 6.135 6.72 5.85 5.16 5.60 

Benzo(b)fluoranthene 6.266 6.79 5.99 5.27 5.73 

Benzo(k)fluoranthene 6.291 6.81 6.02 5.29 5.75 

C2-benzanthracene/chrysenes 6.429 6.88 6.16 5.41 5.89 

Benzo(ghi)perylene 6.507 6.92 6.24 5.47 5.97 

C3-benzanthracene/chrysenes 6.94 7.16 6.70 5.83 6.41 

Indeno(1,2,3-cd)pyrene 6.722 7.04 6.47 5.65 6.19 

Dibenz(a,h)anthracene 6.713 7.04 6.46 5.64 6.18 

C4-benzanthracene/chrysenes 7.36 7.38 7.14 6.18 6.83 

2-13 
 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Table 2-1. Continued 

Class Contaminants 

Log 

KOW a 

Log 

KBC 

b 

Log 

KPED c 

Log 

KPDMS 

d 

Log 

KPOM e 

Other 

Chemicals Benzene 2.13 4.56 1.65 1.84 1.55

 Delta-BHC 3.78 - 3.38 3.21 3.22 

Gamma-BHC, Lindane 3.73 - 3.33 3.17 3.17 

Biphenyl 3.96 5.55 3.57 3.36 3.40 

4-Bromophenyl phenyl ether 5.00 - 4.66 4.22 4.45 

Butyl benzyl phthalate 4.84 - 4.49 4.09 4.29 

Chlorobenzene 2.86 4.95 2.41 2.44 2.29 

Diazinon 3.70 - 3.30 3.14 3.14 

Dibenzofuran 4.07 5.61 3.68 3.45 3.51 

1,2-Dichlorobenzene 3.43 5.26 3.01 2.92 2.86 

1,3-Dichlorobenzene 3.43 5.26 3.01 2.92 2.86 

1,4-Dichlorobenzene 3.42 5.26 3.00 2.91 2.85 

Di-n-butyl phthalate 4.61 - 4.25 3.90 4.06 

Dieldrin 5.37 - 5.05 4.53 4.82 

Diethyl phthalate 2.50 - 2.04 2.15 1.93 

Endosulfan mixed isomers 4.10 - 3.72 3.47 3.54 

Alpha-Endosulfan 3.83 - 3.43 3.25 3.27 

Beta-Endosulfan 4.52 - 4.16 3.82 3.97 

Endrin 5.06 - 4.72 4.27 4.51 

Ethylbenzene 3.14 5.11 2.71 2.68 2.57 

Hexachloroethane 4.00 - 3.61 3.39 3.44 

Malathion 2.89 - 2.44 2.47 2.32 

Methoxychlor 5.08 - 4.74 4.29 4.53 

Pentachlorobenzene 5.26 6.25 4.93 4.44 4.71 

1,1,2,2-Tetrachloroethane 2.39 - 1.92 2.05 1.81 

Tetrachloroethene 2.67 - 2.21 2.29 2.10 

Tetrachloromethane 2.73 - 2.28 2.34 2.16 

Toluene 2.75 4.90 2.30 2.35 2.18 

Toxaphene 5.50 - 5.19 4.64 4.96 

Tribromomethane (Bromoform) 2.35 - 1.88 2.02 1.77 

1, 2, 4-Trichlorobenzene 4.01 5.58 3.62 3.40 3.45 

1, 1, 1-Trichloroethane 2.48 - 2.01 2.13 1.90 

Trichloroethene 2.71 - 2.26 2.32 2.14 

m-Xylene 3.20 5.14 2.77 2.73 2.63 

a From corresponding ESB documents. 

b

 Derived using equation 2-4. KBC values not derived for non-planar molecules. c

 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 Log KOW a Log KOC

a,b Log KDOCa,c 

Endrin 5.06 4.97 4.13 

Dieldrin 5.37 5.28 4.44 

a

 From corresponding ESB documents (U.S. EPA 2003b,c).
 b

 From corresponding ESB documents using: Log KOC = 0.00028 + 0.983 × Log KOW

 (U.S. EPA 2003b,c). c Derived using Equation 2-8. 

Table 2-3. Solutions to Equation 2-7 using KDOC values calculated from Equation 2-8 

DOC 

(mg/L) 

Endrin 

(% freely dissolved) 

Dieldrin 

(% freely dissolved) 

0 100 100 

5 94 88 

10 88 78 

15 83 71 

20 79 64 

25 75 59 

30 71 55 

40 65 48 

50 60 42 

60 55 38 

70 51 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., KOW, KOC) 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., KPOM) 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 (μg/g dry 

weight), PAHs in interstitial water that was 

generated by centrifugation (μg/L), PAHs 

absorbed to a passive sampler (μg/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 μg/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 (fOC) and BC (fBC) 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 μg/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, μg 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,PAHi,FCVi) to derive the conventional 

ESBTUFCVi (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 

ESBTUFCVi (see solubility values for PAHs 

Table 3-4 in U.S. EPA, 2003d). Next, the 

ESBTUFCVi for the 34 PAHs were added to 

derive the ΣESBTU for the 34 PAHs 

(ΣESBTUFCV,34): 

∑ ESBTU୊େ୚,ଷସ ൌ ∑୧

େోి౟ (3-1) େోి,ౌఽౄ౟,ౌి౒౟

The ∑ESBTUFCV,34 is 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 ESBTUFCV. An alternative 

approach discussed in Section 2.1.1 and one 

this equation is equivalent to: 

஼ೀ಴ ܥௗ ൌ (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 

ESBTUFCVs. The two approaches will result in 

the same number of interstitial water toxic 

units for a given sediment but the approach 

using COC,PAHi,FCVi is currently more commonly 

applied in assessing sediments when using the 

one-carbon model. 

3.2.2 Two-Carbon Model 

Like the conventional ∑ESBTUFCV,34 

calculated using the one-carbon model in 

Section 3.2.1, the sum of interstitial water 

toxic units (∑IWTUFCV) 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: 

more similar to the approach used in the twocarbon model discussed in the next section, is ܥௗ ൌ	

to calculate the Cd using Equation 2-2: 

௉ܥ

௡ି ଵ ݂ேௌை஼ ܭை஼ ൅	݂஻஼ ܭ ஼஻ܥௗ

ܥ ௉ܥௗ ൌ ݂ை஼ܭை஼

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,FCVi) 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 

IWTUFCVi 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 μg/L, 

respectively. The IWTUFCVi for the 34 PAHs 

were added to derive the ∑IWTU for the 34 

PAHs (∑IWTUFCV,34): 

(3-3) ஼೏

ி஼௏,ଷସ ൌ	 ∑௜ ∑ ܫܹܷܶ

஼೏,ುಲಹ೔,ಷ಴ೇ೔

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 ∑IWTUFCV,34 calculated 

using the one-carbon model (i.e., 9.3), the twocarbon model still predicts that this sediment 

could be chronically toxic to sensitive benthic 

species. For Sediment B, the ∑IWTUFCV,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 (CIW), 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 μg/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 freelydissolved (% Cd) is calculated using Equation 

2-7: 

ଵ 100 ௄ሺ஽ை஼ ವೀ಴ା	ଵሻ %	ܥௗ ൌ	

Using the % Cd value derived for each 

PAH, the freely dissolved concentration of 

each PAH is calculated as: 

%	஼೏ ቀ	 ܥௗ ൌ	ܥூௐ ଵ଴଴ ቁ	 (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 

IWTUFCVi 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 PED are used to determine the 

concentration of freely dissolved PAHs in 

interstitial water. For this example, the use of 

PRCs loaded in the PED 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 PED (μg/g 

PED) is divided by it corresponding KPS-d, to 

estimate the freely dissolved concentration in 

interstitial water (μg/L) (i.e., Equation 2-9): 

ܥ௉ௌ ܥௗ ൌ ܭ௉ௌିௗ

The provisional KPS-d or KPED values used 

(3-5) ைௐܭ ݃݋ܮ	1.05 ൅ 0.59െ ൌ஽ ா௉ܭ ݃݋ܮ

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 (∑IWTUFCV,34). The ∑IWTUFCV,34 were 

0.5 for Sediment A (Table 3-1) and 0.7 for 

Sediment B (Table 3-2). ∑IWTUFCV,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 nonplanar 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. 

in these examples were derived from the This understanding is the basis for the two

relationship developed by Lohmann and Muir carbon model described in Section 2.2.1. The 

(2010) for polyethylene devices: relationship between log KOW 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 nonplanar 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 ESBTUFCV and IWTUFCV for PAH mixtures: Sediment A 

PAHi 

ESB-Final Chronic Values a 

ESB-FCV 

COC, PAHi, FCVi 

(μg/goc) 

ESB-FCV 

Cd, PAHi, FCVi 

(μg/L) 

ESB-Maxi 

COC, PAHi, Maxi 

(μg/gOC) 

ESB-Maxi 

Cd, PAHi, Maxi 

(μg/L) 

Naphthalene 385 193.55 61,700 30,995 

C1-Naphthalenes 444 81.69 - -

C2-Naphthalenes 510 30.24 - -

C3-Naphthalenes 581 11.1 - -

C4-Naphthalenes 657 4.048 - -

Acenaphthylene 452 306.9 24,000 16,314 

Acenaphthene 491 55.85 33,400 3,800 

Fluorene 538 39.3 26,000 1,900 

C1-Fluorenes 611 13.99 - -

C2-Fluorenes 686 5.305 - -

C3-Fluorenes 769 1.916 - -

Phenanthrene 596 19.13 34,300 1,100 

Anthracene 594 20.73 1,300 45 

C1-Phenanthrenes/Anthracenes 670 7.436 - -

C2-Phenanthrenes/Anthracenes 746 3.199 - -

C3-Phenanthrenes/Anthracenes 829 1.256 - -

C4-Phenanthrenes/Anthracenes 913 0.5594 - -

Fluoranthene 707 7.109 23,870 239.9 

Pyrene 697 10.11 9,090 131.9 

C1-Fluoranthenes/Pyrenes 770 4.887 - -

Benz[a]anthracene 841 2.227 4,153 11 

Chrysene 844 2.042 826 2 

C1-Benzanthracenes/Chrysenes 929 0.8557 - -

C2- Benzanthracenes/Chrysenes 1,008 0.4827 - -

C3- Benzanthracenes/Chrysenes 1,112 0.1675 - -

C4- Benzanthracenes/Chrysenes 1,214 0.07062 - -

Benzo[b]fluoranthene 979 0.6774 2,169 1.501 

Benzo[k]fluoranthene 981 0.6415 1,220 0.7999 

Benzo[a]pyrene 965 0.9573 3,840 3.810 

Perylene 967 0.9008 431 0.4012 

Benzo[e]pyrene 967 0.9008 4,300 4.012 

Indeno[1,2,3-cd]pyrene 1,115 0.275 - -

Dibenz[a,h]anthracene 1,123 0.2825 2,389 0.6012 

Benzo[ghi]perylene 1,095 0.4391 648 0.2600 

Total - - - -

3-6 
Example Calculations of ESBTUFCV and IWTUFCV 

Table 3-1. Continued 

PAHi 

One-Carbon Model 

CP,PAHi 

Measured 

Sediment Conc 

(μg/g dry wt) 

COC,PAHi 

Measured 

Sediment Conc 

(μg/gOC) ESBTUFCVi 

Naphthalene 5.0 133 0.3 

C1-Naphthalenes 1.0 27 0.06 

C2-Naphthalenes 1.6 43 0.08 

C3-Naphthalenes 1.7 45 0.08 

C4-Naphthalenes 1.1 29 0.04 

Acenaphthylene 4.9 131 0.3 

Acenaphthene 2.4 64 0.1 

Fluorene 4.4 117 0.2 

C1-Fluorenes 2.3 61 0.1 

C2-Fluorenes 1.8 48 0.07 

C3-Fluorenes 1.0 26 0.03 

Phenanthrene 18.0 480 0.8 

Anthracene 10.0 267 0.4 

C1-Phenanthrenes/Anthracenes 16.0 427 0.6 

C2-Phenanthrenes/Anthracenes 8.6 229 0.3 

C3-Phenanthrenes/Anthracenes 2.6 69 0.08 

C4-Phenanthrenes/Anthracenes 1.1 29 0.03 

Fluoranthene 28.0 747 1.1 

Pyrene 23.0 613 0.9 

C1-Fluoranthenes/Pyrenes 23.0 613 0.8 

Benz[a]anthracene 17.0 453 0.5 

Chrysene 16.0 427 0.5 

C1-Benzanthracenes/Chrysenes 7.7 205 0.2 

C2- Benzanthracenes/Chrysenes 2.9 77 0.08 

C3- Benzanthracenes/Chrysenes 1.1 29 0.03 

C4- Benzanthracenes/Chrysenes 0.8 22 0.02 

Benzo[b]fluoranthene 14.0 373 0.4 

Benzo[k]fluoranthene 8.3 221 0.2 

Benzo[a]pyrene 13.0 347 0.4 

Perylene 2.8 75 0.08 

Benzo[e]pyrene 7.9 211 0.2 

Indeno[1,2,3-cd]pyrene 4.2 112 0.1 

Dibenz[a,h]anthracene 1.4 37 0.03 

Benzo[ghi]perylene 2.7 72 0.07 

Total 257 - 9.3 

3-7 
 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Table 3-1. Continued 

PAHi 

Two Carbon Model 

Log KOC a 

(L/Kg NSOC) 

Log KBC 

b 

(L/Kg BC) 

Cd,PAHi 

Estimated 

Freely Dissolved 

IW Conc 

(μg/L) IWTUFCVi 

Naphthalene 3.299 5.24 4.07 0.02 

C1-Naphthalenes 3.736 5.48 0.12 0.001 

C2-Naphthalenes 4.227 5.75 0.09 0.003 

C3-Naphthalenes 4.719 6.02 0.03 0.003 

C4-Naphthalenes 5.21 6.29 0.01 0.001 

Acenaphthylene 3.168 5.16 5.21 0.02 

Acenaphthene 3.944 5.59 0.31 0.01 

Fluorene 4.137 5.70 0.55 0.01 

C1-Fluorenes 4.64 5.98 0.066 0.005 

C2-Fluorenes 5.112 6.24 0.016 0.003 

C3-Fluorenes 5.603 6.51 0.002 0.001 

Phenanthrene 4.494 5.90 2.42 0.1 

Anthracene 4.457 5.88 1.04 0.05 

C1-Phenanthrenes/Anthracenes 4.955 6.15 0.75 0.1 

C2-Phenanthrenes/Anthracenes 5.367 6.38 0.12 0.04 

C3-Phenanthrenes/Anthracenes 5.82 6.63 0.006 0.01 

C4-Phenanthrenes/Anthracenes 6.213 6.85 0.0007 0.001 

Fluoranthene 4.998 6.18 1.61 0.2 

Pyrene 4.839 6.09 1.68 0.2 

C1-Fluoranthenes/Pyrenes 5.197 6.29 0.77 0.2 

Benz[a]anthracene 5.577 6.50 0.215 0.1 

Chrysene 5.616 6.52 0.180 0.09 

C1-Benzanthracenes/Chrysenes 6.036 6.75 0.023 0.03 

C2- Benzanthracenes/Chrysenes 6.32 6.91 0.0027 0.01 

C3- Benzanthracenes/Chrysenes 6.822 7.19 0.00019 0.001 

C4- Benzanthracenes/Chrysenes 7.235 7.41 0.00005 0.001 

Benzo[b]fluoranthene 6.16 6.82 0.05 0.07 

Benzo[k]fluoranthene 6.184 6.83 0.019 0.03 

Benzo[a]pyrene 6.003 6.73 0.056 0.06 

Perylene 6.031 6.75 0.0046 0.01 

Benzo[e]pyrene 6.031 6.75 0.024 0.03 

Indeno[1,2,3-cd]pyrene 6.608 7.07 0.0026 0.01 

Dibenz[a,h]anthracene 6.599 7.06 0.0004 0.002 

Benzo[ghi]perylene 6.397 6.95 0.0020 0.005 

Total - - 19 1.4 

3-8 
 

Example Calculations of ESBTUFCV and IWTUFCV 

Table 3-1. Continued 

PAHi

Directly Measured Interstitial Water 

CIW, PAHi 

Total Measured 

IW Conc 

(μg/L) 

Log KDOC c 

(L/Kg DOC) 

Fraction 

Freely 

Dissolved 

Cd,PAHi 

Estimated 

Freely 

Dissolved 

IW Conc 

(μg/L) IWTUFCVi 

Naphthalene 0.25 2.44 0.997 0.249 0.001 

C1-Naphthalenes 0.080 2.88 0.992 0.079 0.001 

C2-Naphthalenes 0.270 3.38 0.974 0.263 0.009 

C3-Naphthalenes 0.170 3.87 0.924 0.157 0.01 

C4-Naphthalenes 0.099 4.37 0.796 0.079 0.02 

Acenaphthylene 0.043 2.31 0.998 0.043 0.0001 

Acenaphthene 1.300 3.09 0.987 1.283 0.02 

Fluorene 1.100 3.29 0.979 1.077 0.03 

C1-Fluorenes 0.120 3.79 0.936 0.112 0.008 

C2-Fluorenes 0.073 4.27 0.831 0.061 0.01 

C3-Fluorenes 0.065 4.76 0.611 0.040 0.02 

Phenanthrene 0.790 3.65 0.954 0.753 0.04 

Anthracene 0.180 3.61 0.957 0.172 0.008 

C1-Phenanthrenes/Anthracenes 0.200 4.11 0.876 0.175 0.02 

C2-Phenanthrenes/Anthracenes 0.140 4.53 0.731 0.102 0.03 

C3-Phenanthrenes/Anthracenes 0.099 4.98 0.487 0.048 0.04 

C4-Phenanthrenes/Anthracenes 0.081 5.38 0.276 0.022 0.04 

Fluoranthene 0.370 4.15 0.865 0.320 0.05 

Pyrene 0.300 3.99 0.902 0.271 0.03 

C1-Fluoranthenes/Pyrenes 0.160 4.35 0.801 0.128 0.03 

Benz[a]anthracene 0.110 4.74 0.625 0.069 0.03 

Chrysene 0.093 4.78 0.604 0.056 0.03 

C1-Benzanthracenes/Chrysenes 0.006 5.20 0.365 0.002 0.002 

C2- Benzanthracenes/Chrysenes 0.006 5.48 0.229 0.001 0.003 

C3- Benzanthracenes/Chrysenes 0.006 5.99 0.085 0.0005 0.003 

C4- Benzanthracenes/Chrysenes 0.006 6.41 0.034 0.0002 0.003 

Benzo[b]fluoranthene 0.063 5.32 0.302 0.019 0.03 

Benzo[k]fluoranthene 0.031 5.35 0.290 0.009 0.01 

Benzo[a]pyrene 0.056 5.17 0.383 0.021 0.02 

Perylene 0.011 5.19 0.368 0.004 0.004 

Benzo[e]pyrene 0.052 5.19 0.368 0.019 0.02 

Indeno[1,2,3-cd]pyrene 0.020 5.77 0.132 0.003 0.01 

Dibenz[a,h]anthracene 0.006 5.77 0.135 0.001 0.003 

Benzo[ghi]perylene 0.021 5.56 0.200 0.004 0.01 

Total 6 - - - 0.6 

3-9 
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Table 3-1. Continued 

PAHi 

Passive Sampler 

Passive 

Sampler Conc 

(μg/g PE) 

Log KPED

 d 

(L/Kg PE) 

Cd,PAHi Estimated 

Freely Dissolved 

IW Conc (μg/L) IWTUFCVi 

Naphthalene 0.16 2.93 0.19 0.001 

C1-Naphthalenes 0.18 3.40 0.07 0.001 

C2-Naphthalenes 2.02 3.93 0.24 0.008 

C3-Naphthalenes 4.26 4.45 0.151 0.01 

C4-Naphthalenes 5.95 4.98 0.063 0.02 

Acenaphthylene 0.02 2.79 0.03 0.0001 

Acenaphthene 4.66 3.62 1.11 0.02 

Fluorene 6.87 3.83 1.02 0.03 

C1-Fluorenes 2.09 4.37 0.09 0.006 

C2-Fluorenes 4.08 4.87 0.055 0.01 

C3-Fluorenes 8.94 5.40 0.036 0.02 

Phenanthrene 11.02 4.21 0.68 0.04 

Anthracene 2.07 4.17 0.14 0.007 

C1-Phenanthrenes/Anthracenes 8.06 4.70 0.16 0.02 

C2-Phenanthrenes/Anthracenes 12.51 5.14 0.09 0.03 

C3-Phenanthrenes/Anthracenes 18.17 5.63 0.043 0.03 

C4-Phenanthrenes/Anthracenes 29.68 6.05 0.0267 0.05 

Fluoranthene 1.40 4.75 0.03 0.004 

Pyrene 10.98 4.58 0.29 0.03 

C1-Fluoranthenes/Pyrenes 17.38 4.96 0.19 0.04 

Benz[a]anthracene 12.79 5.37 0.055 0.02 

Chrysene 11.02 5.41 0.043 0.02 

C1-Benzanthracenes/Chrysenes 1.22 5.86 0.002 0.002 

C2- Benzanthracenes/Chrysenes 1.59 6.16 0.0011 0.002 

C3- Benzanthracenes/Chrysenes 2.04 6.70 0.00041 0.002 

C4- Benzanthracenes/Chrysenes 1.51 7.14 0.00011 0.002 

Benzo[b]fluoranthene 13.66 5.99 0.01 0.02 

Benzo[k]fluoranthene 17.62 6.02 0.017 0.03 

Benzo[a]pyrene 8.64 5.82 0.013 0.01 

Perylene 2.27 5.85 0.0032 0.004 

Benzo[e]pyrene 7.82 5.85 0.011 0.01 

Indeno[1,2,3-cd]pyrene 3.82 6.47 0.0013 0.005 

Dibenz[a,h]anthracene 1.44 6.46 0.0005 0.002 

Benzo[ghi]perylene 5.42 6.24 0.0031 0.007 

Total - - - 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).
 c 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 ESBTUFCV and IWTUFCV for PAH mixtures: Sediment B 

PAHi 

ESB-Final Chronic Values a 

ESB-FCV 

COC, PAHi, FCVi 

(μg/goc) 

ESB-FCV 

Cd, PAHi, FCVi 

(μg/L) 

ESB-Maxi 

COC, PAHi, Maxi 

(μg/gOC) 

ESB-Maxi 

Cd, PAHi, Maxi 

(μg/L) 

Naphthalene 385 193.5 61,700 30,995 

C1-Naphthalenes 444 81.69 - -

C2-Naphthalenes 510 30.24 - -

C3-Naphthalenes 581 11.1 - -

C4-Naphthalenes 657 4.048 - -

Acenaphthylene 452 306.9 24,000 16,314 

Acenaphthene 491 55.85 33,400 3,800 

Fluorene 538 39.3 26,000 1,900 

C1-Fluorenes 611 13.99 - -

C2-Fluorenes 686 5.305 - -

C3-Fluorenes 769 1.916 - -

Phenanthrene 596 19.13 34,300 1,100 

Anthracene 594 20.73 1,300 45 

C1-Phenanthrenes/Anthracenes 670 7.436 - -

C2-Phenanthrenes/Anthracenes 746 3.199 - -

C3-Phenanthrenes/Anthracenes 829 1.256 - -

C4-Phenanthrenes/Anthracenes 913 0.5594 - -

Fluoranthene 707 7.109 23,870 239.9 

Pyrene 697 10.11 9,090 131.9 

C1-Fluoranthenes/Pyrenes 770 4.887 - -

Benz[a]anthracene 841 2.227 4,153 11 

Chrysene 844 2.042 826 2 

C1-Benzanthracenes/Chrysenes 929 0.8557 - -

C2- Benzanthracenes/Chrysenes 1,008 0.4827 - -

C3- Benzanthracenes/Chrysenes 1,112 0.1675 - -

C4- Benzanthracenes/Chrysenes 1,214 0.07062 - -

Benzo[b]fluoranthene 979 0.6774 2,169 1.501 

Benzo[k]fluoranthene 981 0.6415 1,220 0.7999 

Benzo[a]pyrene 965 0.9573 3,840 3.810 

Perylene 967 0.9008 431 0.4012 

Benzo[e]pyrene 967 0.9008 4,300 4.012 

Indeno[1,2,3-cd]pyrene 1,115 0.275 - -

Dibenz[a,h]anthracene 1,123 0.2825 2,389 0.6012 

Benzo[ghi]perylene 1,095 0.4391 648 0.2600 

Total - - - -

3-11 
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Table 3-2. Continued 

PAHi 

One-Carbon Model 

CP,PAHi 

Measured 

Sediment Conc

 (μg/g dry wt) 

COC,PAHi 

Measured 

Sediment Conc 

(μg/gOC) ESBTUFCVi 

Naphthalene 0.2 7 0.02 

C1-Naphthalenes 0.4 14 0.03 

C2-Naphthalenes 0.8 24 0.05 

C3-Naphthalenes 0.6 18 0.03 

C4-Naphthalenes 0.2 7 0.01 

Acenaphthylene 0.4 14 0.03 

Acenaphthene 0.7 23 0.05 

Fluorene 0.5 15 0.03 

C1-Fluorenes 0.3 10 0.02 

C2-Fluorenes 0.3 9 0.01 

C3-Fluorenes 0.1 3 0.004 

Phenanthrene 3.6 113 0.2 

Anthracene 1.2 38 0.06 

C1-Phenanthrenes/Anthracenes 2.3 72 0.1 

C2-Phenanthrenes/Anthracenes 1.3 41 0.05 

C3-Phenanthrenes/Anthracenes 0.4 11 0.01 

C4-Phenanthrenes/Anthracenes 0.3 10 0.01 

Fluoranthene 3.4 106 0.2 

Pyrene 5.6 175 0.3 

C1-Fluoranthenes/Pyrenes 3.5 109 0.1 

Benz[a]anthracene 2.3 72 0.09 

Chrysene 2.1 66 0.08 

C1-Benzanthracenes/Chrysenes 1.0 30 0.03 

C2- Benzanthracenes/Chrysenes 0.3 8 0.01 

C3- Benzanthracenes/Chrysenes 0.1 4 0.003 

C4- Benzanthracenes/Chrysenes 0.1 3 0.003 

Benzo[b]fluoranthene 1.6 50 0.05 

Benzo[k]fluoranthene 0.6 19 0.02 

Benzo[a]pyrene 2.1 66 0.07 

Perylene 0.5 15 0.02 

Benzo[e]pyrene 1.1 34 0.04 

Indeno[1,2,3-cd]pyrene 0.7 22 0.02 

Dibenz[a,h]anthracene 0.2 7 0.01 

Benzo[ghi]perylene 0.7 21 0.02 

Total 39.4 - 1.7 

3-12 
 

Example Calculations of ESBTUFCV and IWTUFCV 

Table 3-2. Continued 

Two-Carbon Model 

PAHi 

Log KOC a 

(L/Kg NSOC) 

Log KBC 

b 

(L/Kg BC) 

Cd,PAHi Estimated 

Freely Dissolved 

PW Conc 

(μg/L) IWTUFCVi 

Naphthalene 3.299 5.24 0.22 0.001 

C1-Naphthalenes 3.736 5.48 0.26 0.003 

C2-Naphthalenes 4.227 5.75 0.21 0.007 

C3-Naphthalenes 4.719 6.02 0.05 0.004 

C4-Naphthalenes 5.21 6.29 0.00 0.001 

Acenaphthylene 3.168 5.16 0.87 0.003 

Acenaphthene 3.944 5.59 0.37 0.007 

Fluorene 4.137 5.70 0.12 0.003 

C1-Fluorenes 4.64 5.98 0.02 0.002 

C2-Fluorenes 5.112 6.24 0.01 0.001 

C3-Fluorenes 5.603 6.51 0.00 0.0002 

Phenanthrene 4.494 5.90 1.14 0.06 

Anthracene 4.457 5.88 0.25 0.01 

C1-Phenanthrenes/Anthracenes 4.955 6.15 0.22 0.03 

C2-Phenanthrenes/Anthracenes 5.367 6.38 0.04 0.01 

C3-Phenanthrenes/Anthracenes 5.82 6.63 0.00 0.002 

C4-Phenanthrenes/Anthracenes 6.213 6.85 0.00 0.001 

Fluoranthene 4.998 6.18 0.35 0.05 

Pyrene 4.839 6.09 0.98 0.1 

C1-Fluoranthenes/Pyrenes 5.197 6.29 0.23 0.05 

Benz[a]anthracene 5.577 6.50 0.06 0.03 

Chrysene 5.616 6.52 0.05 0.02 

C1-Benzanthracenes/Chrysenes 6.036 6.75 0.01 0.007 

C2- Benzanthracenes/Chrysenes 6.32 6.91 0.00 0.001 

C3- Benzanthracenes/Chrysenes 6.822 7.19 0.00 0.0003 

C4- Benzanthracenes/Chrysenes 7.235 7.41 0.00 0.0002 

Benzo[b]fluoranthene 6.16 6.82 0.01 0.01 

Benzo[k]fluoranthene 6.184 6.83 0.00 0.003 

Benzo[a]pyrene 6.003 6.73 0.02 0.02 

Perylene 6.031 6.75 0.00 0.002 

Benzo[e]pyrene 6.031 6.75 0.01 0.008 

Indeno[1,2,3-cd]pyrene 6.608 7.07 0.00 0.004 

Dibenz[a,h]anthracene 6.599 7.06 0.00 0.001 

Benzo[ghi]perylene 6.397 6.95 0.00 0.004 

Total ­ ­ 5.51 0.5 

3-13 
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Table 3-2. Continued 

PAHi 

Directly Measured Interstitial Water 

CIW, PAHi 

Measured 

Total IW Conc 

(μg/L) 

Log KDOC c 

(L/Kg 

DOC) 

Fraction 

Freely 

Dissolved 

Cd,PAHi Estimated 

Freely Dissolved 

IW Conc 

(μg/L) IWTUFCVi 

Naphthalene 0.92 2.44 0.999 0.919 0.005 

C1-Naphthalenes 1.30 2.88 0.996 1.297 0.02 

C2-Naphthalenes 1.10 3.38 0.988 1.091 0.04 

C3-Naphthalenes 0.48 3.87 0.964 0.468 0.04 

C4-Naphthalenes 0.15 4.37 0.896 0.139 0.03 

Acenaphthylene 0.02 2.31 0.999 0.018 0.0001 

Acenaphthene 1.90 3.09 0.994 1.892 0.03 

Fluorene 0.46 3.29 0.990 0.457 0.01 

C1-Fluorenes 0.20 3.79 0.970 0.196 0.01 

C2-Fluorenes 0.11 4.27 0.915 0.103 0.02 

C3-Fluorenes 0.01 4.76 0.775 0.005 0.002 

Phenanthrene 0.44 3.65 0.978 0.434 0.02 

Anthracene 0.18 3.61 0.980 0.178 0.009 

C1-Phenanthrenes/Anthracenes 0.16 4.11 0.940 0.153 0.02 

C2-Phenanthrenes/Anthracenes 0.14 4.53 0.856 0.126 0.04 

C3-Phenanthrenes/Anthracenes 0.06 4.98 0.676 0.041 0.03 

C4-Phenanthrenes/Anthracenes 0.04 5.38 0.456 0.020 0.04 

Fluoranthene 0.19 4.15 0.934 0.177 0.02 

Pyrene 0.36 3.99 0.953 0.343 0.03 

C1-Fluoranthenes/Pyrenes 0.16 4.35 0.898 0.144 0.03 

Benz[a]anthracene 0.05 4.74 0.786 0.037 0.02 

Chrysene 0.06 4.78 0.770 0.045 0.02 

C1-Benzanthracenes/Chrysenes 0.03 5.20 0.559 0.017 0.02 

C2- Benzanthracenes/Chrysenes 0.01 5.48 0.396 0.002 0.005 

C3- Benzanthracenes/Chrysenes 0.01 5.99 0.170 0.001 0.006 

C4- Benzanthracenes/Chrysenes 0.01 6.41 0.073 0.004 0.006 

Benzo[b]fluoranthene 0.04 5.32 0.487 0.018 0.03 

Benzo[k]fluoranthene 0.02 5.35 0.473 0.007 0.01 

Benzo[a]pyrene 0.03 5.17 0.577 0.018 0.02 

Perylene 0.07 5.19 0.561 0.039 0.04 

Benzo[e]pyrene 0.03 5.19 0.561 0.021 0.02 

Indeno[1,2,3-cd]pyrene 0.02 5.77 0.251 0.005 0.02 

Dibenz[a,h]anthracene 0.01 5.77 0.255 0.001 0.005 

Benzo[ghi]perylene 0.03 5.56 0.354 0.009 0.02 

Total 8.8 - - - 0.7 

3-14 
Example Calculations of ESBTUFCV and IWTUFCV 

Table 3-2. Continued 

PAHi 

Passive Sampler 

Passive 

Sampler Conc 

(μg/g PE) 

Log KPED

 d 

(L/Kg PE) 

Cd,PAHi Estimated 

Freely Dissolved 

IW Conc (μg/L) IWTUFCVi 

Naphthalene 0.66 2.93 0.767 0.004 

C1-Naphthalenes 2.50 3.40 0.997 0.01 

C2-Naphthalenes 8.58 3.93 1.020 0.03 

C3-Naphthalenes 14.37 4.45 0.510 0.05 

C4-Naphthalenes 10.29 4.98 0.109 0.03 

Acenaphthylene 0.03 2.79 0.045 0.0001 

Acenaphthene 7.21 3.62 1.720 0.03 

Fluorene 2.16 3.83 0.320 0.008 

C1-Fluorenes 2.81 4.37 0.121 0.009 

C2-Fluorenes 7.49 4.87 0.101 0.02 

C3-Fluorenes 2.23 5.40 0.009 0.005 

Phenanthrene 6.43 4.21 0.397 0.02 

Anthracene 2.31 4.17 0.156 0.008 

C1-Phenanthrenes/Anthracenes 7.40 4.70 0.147 0.02 

C2-Phenanthrenes/Anthracenes 15.43 5.14 0.111 0.03 

C3-Phenanthrenes/Anthracenes 28.74 5.63 0.068 0.05 

C4-Phenanthrenes/Anthracenes 34.46 6.05 0.031 0.06 

Fluoranthene 9.35 4.75 0.167 0.02 

Pyrene 15.18 4.58 0.401 0.04 

C1-Fluoranthenes/Pyrenes 18.39 4.96 0.201 0.04 

Benz[a]anthracene 5.35 5.37 0.023 0.01 

Chrysene 30.24 5.41 0.118 0.06 

C1-Benzanthracenes/Chrysenes 11.51 5.86 0.016 0.02 

C2- Benzanthracenes/Chrysenes 8.68 6.16 0.006 0.01 

C3- Benzanthracenes/Chrysenes 4.98 6.70 0.001 0.006 

C4- Benzanthracenes/Chrysenes 13.74 7.14 0.001 0.01 

Benzo[b]fluoranthene 6.73 5.99 0.007 0.01 

Benzo[k]fluoranthene 9.33 6.02 0.009 0.01 

Benzo[a]pyrene 19.93 5.82 0.030 0.03 

Perylene 20.61 5.85 0.029 0.03 

Benzo[e]pyrene 4.98 5.85 0.007 0.008 

Indeno[1,2,3-cd]pyrene 2.94 6.47 0.001 0.004 

Dibenz[a,h]anthracene 2.88 6.46 0.001 0.004 

Benzo[ghi]perylene 27.96 6.24 0.016 0.04 

Total - - - 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).
 c 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 fBC, 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). 

4-3 

References 

Section 5 

References 

Accardi-Dey A, Gschwend PM. 2002. Assessing 

the combined role of natural organic matter and 

black carbon as sorbents in sediments. Environ 

Sci Technol 36:21−29. 

Accardi-Dey A, Gschwend PM. 2003. 

Reinterpreting literature sorption data 

considering both absorption into organic carbon 

and adsorption onto black carbon. Environ Sci 

Technol 37:99−106. 

Adams DD. 1991. Sediment pore water 

sampling. In Mudroch A, MacKnight SD, eds, 

Handbook of Techniques for Aquatic Sediments 

Sampling. CRC Press, Boca Raton, FL, USA 

pp 171−202. 

Adams RG, Lohmann R, Fernandez LA, 

MacFarlane JK, Gschwend PM. 2007. 

Polyethylene devices: Passive samplers for 

measuring dissolved hydrophobic compounds in 

aquatic environments. Environ Sci Technol 

41:1317−1323. 

Adams WJ, Burgess RM, Gold-Bouchet G, 

LeBlanc L, Liber K, Williamson B. 2003. 

Porewater chemistry: effects of sampling, 

storage, handling, and toxicity testing. In 

Porewater Toxicity Testing: Biological, 

Chemical, and Ecological Considerations. 

Carr RS and Nipper M Eds. Society of 

Environmental Toxicology and Chemistry 

(SETAC). Pensacola, FL, USA. 

Allan IJ, Booij K, Paschke A, Vrana B, Mills 

GA, Greenwood R. 2009. Field performance of 

seven passive sampling devices for monitoring 

of hydrophobic substances. Environ Sci Technol 

43:5383−5390. 

American Society for Testing and Materials. 

1994. Standard Guide For Collection, Storage, 

Characterization, And Manipulation Of 

Sediments For Toxicological Testing. E 1391-94. 

ASTM 1994 Annual Book of Standards. 

Philadelphia, PA, USA. 

American Society for Testing and Materials. 

2010. Standard Test Method for Determination 

of Parent and Alkyl Polycyclic Aromatics in 

Sediment Pore Water Using Solid-Phase 

Microextraction and Gas Chromatography/ 

Mass Spectrometry in Selected Ion Monitoring 

Mode. D7363-07. ASTM 2010 Annual Book of 

Standards Philadelphia, PA, USA 

(www.astm.org). 

Ankley GT, Schubauer-Berigan MK. 1994. 

Comparison of techniques for isolation of 

sediment pore water for toxicity testing. Arch 

Environ Contam Toxicol 27:507–512. 

Armitage JM, Cousins IT, Persson NJ, 

Gustafsson O, Cornelissen G, Saloranta T, 

Broman D, Naes K. 2008. Black carboninclusive modeling approaches for estimating 

the aquatic fate of dibenzo-p-dioxins and 

dibenzofurans. Environ Sci Technol 

42:3697−3703. 

Arp HP, Breedveld GD, Cornelissen G. 2009. 

Estimating the in situ sediment-porewater 

distribution of PAHs and chlorinated aromatic 

hydrocarbons in anthropogenic impacted 

sediments. Environ Sci Technol 43:5576–5585. 

Booij K, Smedes F, van Weerlee EM. 2002. 

Spiking of performance reference compounds in 

low density polyethylene and silicone passive 

water samplers. Chemosphere 46:1157–1161. 

5-1 
 

 

 

 

 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Booij K, Hoedemaker J, Bakker JF. 2003a. 

Dissolved PCBs, PAHs, and HCB in pore waters 

and overlying waters of contaminated harbor 

sediments. Environ Sci Technol 37:4213–4220. 

Booij K, Hofmans H, Fischer C, Van Weerlee 

EM. 2003b. Temperature-dependent uptake rates 

of nonpolar organic compounds by semipermeable membrane devices and low-density 

polyethylene membranes. Environ Sci Technol 

37:361−366. 

Bucheli TD, Gustafsson O. 2000. Quantification 

of the soot-water distribution coefficient of 

PAHs provides mechanistic basis for enhanced 

sorption observations. Environ Sci Technol 

34:5144−5151. 

Burgess RM. 2009. Evaluating Ecological Risk 

to Invertebrate Receptors from PAHs in 

Sediments at Hazardous Waste Sites (Final 

Report). U.S. Environmental Protection Agency, 

Ecological Risk Assessment Support Center, 

EPA/600/R-06/162F. Cincinnati, OH, USA. 

Burgess RM, McKinney RA, Brown WA, Quinn 

JG. 1996. Isolation of marine sediment colloids 

and associated polychlorinated biphenyls: An 

evaluation of ultrafiltration and reverse-phase 

chromatography. Environ Sci Technol 

30:1923–1932. 

Burgess RM, Ryba SA, Perron MM, Tien R, 

Thibodeau LM, Cantwell MG. 2004. Sorption 

of 2,4'-dichlorobiphenyl and fluoranthene to a 

marine sediment amended with different types 

of black carbon. Environ Toxicol Chem 

23:2534−2544. 

Burkhard LP. 2000. Estimating dissolved 

organic carbon partition coefficients for nonionic 

organic chemicals. Environ Sci Technol 

34:4663−4668. 

Carr RS, Nipper M. 2003. Porewater toxicity 

testing: Biological, Chemical, and Ecological 

Considerations. Society of Environmental 

Toxicology and Chemistry (SETAC). Pensacola, 

FL, USA. 

Conder JM, LaPoint TW, Steevens JA, Lotufo 

GR. 2004. Recommendations for the assessment 

of TNT toxicity in sediment. Environ Toxicol 

Chem 23:141–149. 

Cornelissen G, Gustafsson Ö, Bucheli TD, 

Jonker MTO, Koelmans AA, Van Noort PCM. 

2005a. Extensive sorption of organic compounds 

to black carbon, coal, and kerogen in sediments 

and soils: Mechanisms and consequences for 

distribution, bioaccumulation, and 

biodegradation. Environ Sci Technol 

39:6881−6895. 

Cornelissen G, Haftka J, Parsons J, Gustafsson 

O. 2005b. Sorption to black carbon of organic 

compounds with varying polarity and planarity. 

Environ Sci Technol 39: 3688−3694. 

Cornelissen G, Pettersen A, Broman D, Mayer P, 

Breedveld GD. 2008. Field testing of 

equilibrium passive samplers to determine freely 

dissolved native hydrocarbon concentrations. 

Environ Toxicol Chem 27:499−508. 

DeWitt TH, Ozretich RJ, Swartz RC, Lamberson 

JO, Shults DW, Ditsworth GR, Jones JKP, 

Hoselton L, Smith LM.1992. The influence of 

organic matter quality on the toxicity and 

partitioning of sediment-associated fluoranthene. 

Environ Toxicol Chem 11:197−208. 

DiFilippo EL, Eganhouse RP. 2010. Assessment 

of PDMS-water partition coefficients: 

implications for passive environmental sampling 

of hydrophobic organic compounds. Environ Sci 

Technol 44:6917−6925. 

5-2 
 

 

References 

Di Toro DM, McGrath JA, Stubblefield WA. 

2007. Predicting the toxicity of neat and 

weathered crude oil: Toxic potential and the 

toxicity of saturated mixtures. Environ Toxicol 

Chem 26:24−36. 

Di Toro DM, Zarba CS, Hansen DJ, Berry WJ, 

Swartz RC, Cowan CE, Pavlou SP, Allen HE, 

Thomas NA, Paquin PR. 1991. Technical basis 

for establishing sediment quality criteria for 

nonionic organic chemicals by using equilibrium 

partitioning. Environ Toxicol Chem 

10:1541–1583. 

Driscoll SBK, Amos CB, McArdle ME, Menzie, 

CA, Coleman A. 2009. Predicting sediment 

toxicity at former manufactured gas plants using 

equilibrium partitioning benchmarks for PAH 

mixtures. Soil Sediment Contam 18:307–319. 

Endo S, Hale SE, Goss K-U, Arp HPH. 2011. 

Equilibrium partition coefficients of diverse 

polar and nonpolar organic compounds to 

polyoxymethylene (POM) passive sampling 

devices Environ Sci Technol 45:10124−10132. 

Fernandez LA, MacFarlane JK, Tcaciuc AP, 

Gschwend PM. 2009a. Measurement of freely 

dissolved PAH concentrations in sediment beds 

using passive sampling with low-density 

polyethylene strips. Environ Sci Technol 43:1430– 

1436. 

Fernandez LA, MacFarlane JK, Tcaciuc AP, 

Gschwend PM. 2009b. Using performance 

reference compounds in polyethylene passive 

samplers to deduce sediment porewater 

concentrations for numerous target chemicals. 

Environ Sci Technol 43:8888−8894. 

Friedman CL, Burgess RM, Perron MM, 

Cantwell MG, Ho KT, Lohmann R. 2009. 

Comparing polychaete and polyethylene uptake 

to assess sediment resuspension effects on PCB 

bioavailability. Environ Sci Technol 

43:2865−2870. 

Friedman CL, Lohmann R, Burgess RM, Perron 

MM, Cantwell MG. 2011. Resuspension of 

PCB-contaminated field sediment: Release to the 

water column and determination of site-specific 

KDOC. Environ Toxicol Chem 30:377−384. 

Ghosh U. 2007. The role of black carbon in 

influencing availability of PAHs in sediments. 

Human Ecol Risk Assess 13:276–285. 

Ghosh U, Hawthorne SB. 2010. Particle-scale 

measurement of PAH aqueous equilibrium 

partitioning in impacted sediments. Environ Sci 

Technol 44:1204–1210. 

Ghosh U, Weber AS, Jensen JN, Smith JR. 

2000. Relationship between PCB desorption 

equilibrium kinetics, and availability during land 

biotreatment. Environ Sci Technol 

34:2542−2548. 

Ghosh U, Zimmerman JR, Luthy RG. 2003. 

PCB and PAH Speciation among particle types 

in contaminated harbor sediments and effects on 

PAH bioavailability. Environ Sci Technol 

37:2209−2217. 

Great Lakes Water Quality Initiative. 1995. 

Final Water Quality Guidance for the Great 

Lakes System: Final Rule. U.S. Environmental 

Protection Agency, Federal Register, Vol. 

60:15400ff. Washington, DC, USA. 

Gschwend PM, MacFarlane JK, Reible DD, lu 

X, Hawthorne SB, Nakles DV, Thompson T. 

2011. Comparison of polymeric samplers for 

accurately assessing PCBs in pore waters. 

Environ Toxicol Chem 30:1288−1296. 

Gustafsson O, Gschwend PM. 1998. The flux of 

black carbon in surface sediments on the New 

England continental shelf. Geochim Cosmochim 

Acta 62:465–472. 

5-3 
 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Gustafsson O, Haghseta F, Chan C, MacFarlane 

J, Gschwend P. 1997. Quantification of the 

dilute sedimentary soot phase: Implications for 

PAH speciation and bioavailability. Environ Sci 

Technol 31:203−209. 

Gustafsson O, Thomas D, Bucheli TD, Zofia 

Kukulska Z, Andersson M, Largeau C, Rouzaud 

JN, Reddy CM, Eglinton TI. 2001. Evaluation of 

a protocol for the quantification of black carbon 

in sediments, Global Biogeochem Cycle 

15:881−890. 

Gustavson KE, Harkin JM. 2000. Comparison of 

sampling techniques and evaluation of semipermeable membrane devices (SPMDs) for 

monitoring polynuclear aromatic hydrocarbons 

(PAHs) in groundwater. Environ Sci Technol 

34:4445−4451. 

Hawthorne SB, Grabanski CB, Miller DJ, 

Kreitinger JP. 2005. Solid-Phase Microextraction measurement of parent and alkyl 

polycyclic aromatic hydrocarbons in milliliter 

sediment pore water samples and determination 

of KDOC values. Environ Sci Technol 

39:2795−2803. 

Hawthorne SB, Grabanski CB, Miller DJ. 2006. 

Measured partitioning coefficients for parent and 

alkyl polycyclic aromatic hydro-carbons in 114 

historically contaminated sediments: Part 1, KOC 

values. Environ Toxicol Chem 25:2901−2911. 

Hawthorne SB, Grabanski CB, Miller DJ. 2007. 

Measured partition coefficients for parent and 

alkyl polycyclic aromatic hydrocarbons in 114 

historically contaminated sediments: Part 2. 

Testing the KOCKBC two-carbon-type model. 

Environ Toxicol Chem 26: 2505−2516. 

Hawthorne SB, Jonker MTO, van der Heijden 

SA, Grabanski CB, Azzolina NA, Miller DJ. 

2011. Measuring picogram per liter 

concentrations of freely dissolved parent and 

alkyl PAHs (PAH-34), using passive sampling 

with polyoxymethylene. Anal Chem 

83:6754–6761. 

Hawthorne SB, Miller DJ, Grabanski CB. 2009. 

Measuring low picogram per liter concentrations 

of freely dissolved polychlorinated biphenyls in 

sediment pore water using passive sampling with 

polyoxymethylene. Anal Chem 81:9472–9480. 

Hoke RA, Ankley GT, Cotter AM, Goldenstein 

T, Kosian PA, Phipps GL, VanderMeiden FM. 

1994. Evaluation of equilibrium partitioning 

theory for predicting acute toxicity of fieldcollected sediments contaminated with DDT, 

DDE and DDD to the amphipod Hyalella azteca. 

Environ Toxicol Chem 13:157−166. 

Hong L, Ghosh U, Mahajan T, Zare RN, Luthy 

RG. 2003. PAH sorption mechanism and 

partitioning behavior in Lampblack impacted 

soils from former oil-gas plant sites. Environ Sci 

Technol 37:3625−3634. 

Huckins JN, Manuweera GK, Petty JD, Mackay 

D, Lebo JA. 1993. Lipid-containing semipermeable membrane devices for monitoring 

organic contaminants in water. Environ Sci 

Technol 27:2489–2496. 

Huckins JN, Petty JD, Booij K. 2006. Monitors 

of Organic Chemicals in the Environment; 

Springer: New York, NY, USA p 223. 

Huckins JN, Petty JD, Lebo JA, Almeida FV, 

Booij K, Alvarez DA, Cranor WL, Clark RC, 

Mogensen BB. 2002. Development of the 

permeability/performance reference compound 

approach for in situ calibration of semipermeable membrane devices. Environ Sci 

Technol 36:85−91. 

Jonker MTO, Koelmans AA. 2001. 

Polyoxymethylene solid phase extraction as a 

partitioning method for hydrophobic organic 

chemicals in sediment and soot. Environ Sci 

Technol 35:3742–3748. 

5-4 
References 

Jonker MTO, Koelmans AA. 2002. Sorption of 

polycyclic aromatic hydrocarbons and 

polychlorinated biphenyls to soot and soot-like 

materials in the aqueous environment: 

Mechanistic considerations. Environ Sci Technol 

36:3725−3734. 

Jonker MTO, Sinke AJC, Brils JM, Koelmans 

AA. 2003. Sorption of polycyclic aromatic 

hydrocarbons to oil contaminated sediment: 

Unresolved complex? Environ Sci Technol 

37:5197–5203. 

Kane Driscoll SB, Burgess RM. 2007. An 

overview of the development, status, and 

application of equilibrium partitioning sediment 

benchmarks for PAH mixtures. Hum Ecol Risk 

Assess 13:286–301. 

Karickhoff SW, Long JM. 1995. Internal report 

on summary of measured, calculated, and 

recommended log Kow values. Internal Report. 

U.S. Environmental Protection Agency, Office 

of Research and Development, Athens, GA, 

USA. 

Khalil MF, Ghosh U, Kreitinger JP. 2006. Role 

of weathered coal tar pitch in the partitioning of 

polycyclic aromatic hydrocarbons in 

manufactured gas plant site sediments Environ 

Sci Technol 40: 5681−5687. 

Koelmans AA, Jonker MTO, Cornelissen G, 

Bucheli TD, Van Noort PCM, Gustafsson O. 

2006. Black carbon: The reverse of its dark side. 

Chemosphere 63:365–377. 

Kraaij R, Mayer P, Busser FJM, Bolscher MVH, 

Seinen W, Tolls J. 2003. Measured pore-water 

concentrations make equilibrium partitioning 

work−A data analysis. Environ Sci Technol 

37:268–274. 

Kreitinger JP, Newhauser EF, Doherty FG, 

Hawthorne SB. 2007. Greatly reduced 

bioavailability and toxicity of polycyclic 

aromatic hydrocarbons to Hyalella azteca in 

sediments from manufactured gas plant sites. 

Environ Toxicol Chem 26:1146–1157. 

Landrum PF, Nihart SR, Eadie BJ, Gardner WS. 

1984. Reverse-phase separation method for 

determining pollutant binding to Aldrich humic 

acid and dissolved organic carbon in natural 

waters. Environ Sci Technol 18:187−192. 

Lee II H, Lincoff A, Boese BL, Cole FA, Ferraro 

SP, Lamberson JD, Ozretich RJ, Randall RC, 

Rukavina KR, Schults DW, Sercu KA, Specht 

DT, Swartz RC, Young DK. 1994. Ecological 

risk assessment of the marine sediments at the 

United Heckathorn superfund site. EPA-600/X94/029. Final Report. Environmental Research 

Laboratory, U.S. Environmental Protection 

Agency, Newport, OR, USA. 

Leppanen MT, Kukkonen JVK. 2000. Effect of 

sediment–chemical contact time on availability 

of sediment-associated pyrene and benzo 

[a]pyrene to oligochaete worms and semipermeable membrane devices. Aquat Toxicol 

49:227–241. 

Leppanen MT, Kukkonen JVK. 2006. 

Evaluating the role of desorption in 

bioavailability of sediment-associated 

contaminants using oligochaetes, semipermeable 

membrane devices, and Tenax extraction. 

Environ Pollut 140:150–163. 

Leslie HA, Terlaak TL, Busser FJM, Kraak 

MHS, Hermens JLM. 2002. Bioconcentration of 

organic chemicals: Is a solid-phase microextraction fiber a good surrogate for biota? 

Environ Sci Technol 36:5399−5404. 

Lohmann R. 2012. Critical review of lowdensity polyethylene’s partitioning and diffusion 

coefficients for trace organic contaminants and 

implications for its use as a passive sampler. 

Environ Sci Technol 46: 606−618. 

5-5 
Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

Lohmann R, Booij K, Smedes F, Vrana B. 2012. 

Use of passive sampling devices for monitoring 

and compliance checking of POP concentrations 

in water. Environ Sci Pollut Res 19:1885−1895. 

Lohmann R, Burgess RM, Cantwell MG, Ryba 

SA, MacFarlane JK, Gschwend PM. 2004. 

Dependency of polychlorinated biphenyl and 

polycyclic aromatic hydrocarbon 

bioaccumulation in Mya arenaria on both water 

column and sediment bed chemical activities. 

Environ Toxicol Chem. 23:2551−2562. 

Lohmann R, MacFarlane JK, Gschwend PM. 

2005. Importance of black carbon to sorption of 

native PAHs, PCBs, and PCDDs in Boston and 

New York Harbor sediments. Environ Sci 

Technol 39:141−148. 

Lohmann R, Muir D. 2010. Global aquatic 

passive sampling (AQUA-GAPS): Using passive 

samplers to monitor POPs in the waters of the 

world. Environ Sci Technol 44:860−864. 

Lu X, Drake B, Skwarski A, Reible D. 2011. 

Predicting bioavailability of PAHs and PCBs 

with pore water concentrations measured by 

disposable solid-phase micro-extraction fibers. 

Environ Toxicol Chem 30:1109−1116. 

Luthy RG, Aiken GR, Brusseau ML, 

Cunningham SD, Gschwend PM, Pignatello JJ, 

Reinhard M, Traina SJ, Weber WJ, Westall JC. 

1997. Sequestration of hydrophobic organic 

contaminants by geosorbents. Environ Sci 

Technol 31:3341−3347. 

Mackay D, Shiu WY, Ma KC. 1992a. Illustrated 

handbook of physical-chemical properties and 

environmental fate for organic chemicals. 

Volume II - Polynuclear aromatic hydrocarbons, 

polychlorinated dioxins and dibenzofurans. 

Lewis Publishers, Boca Raton, FL, USA. 

Mackay D, Shiu WY, Ma KC. 1992b. Illustrated 

handbook of physical-chemical properties and 

environmental fate for organic chemicals. 

Volume I - Monoaromatic hydrocarbons, 

chlorobenzenes, and PCBs. Lewis Publishers, 

Boca Raton, FL, USA. 

Macrae J, Hall KJ. 1998. Comparison of 

methods used to determine the availability of 

polycyclic aromatic hydrocarbons in marine 

sediment. Environ Sci Technol 32:3809−3815. 

Maruya KA, Landrum PF, Burgess RM, Shine 

JP. 2012. Incorporating contaminant 

bioavailability into sediment quality assessment 

frameworks. Integr Environ Assess Manage 

8:659−673. 

Maruya KA, Risebrough RW, Horne AJ. 1996. 

Partitioning of polynuclear aromatic 

hydrocarbons between sediments from San 

Francisco Bay and their porewaters. Environ Sci 

Technol 30:2942–2947. 

Maruya KA, Zeng EY, Tsukada D, Bay SM. 

2009. A passive sampler based on solid-phase 

microextraction for quantifying hydrophobic 

organic contaminants in sediment pore water. 

Environ Toxicol Chem 28:733–740. 

Mayer P, Vaes WHJ, Wijnker F, Legierse 

KCHM, Kraaij R, Tolls J, Hermens JLM. 2000a. 

Sensing dissolved sediment porewater 

concentrations of persistent and bioaccumulative 

pollutants using disposable solid-phase 

microextraction fibers. Environ Sci Technol 

34:5177–5183. 

Mayer P, Vaes WHJ, Hermens JLM. 2000b. 

Absorption of hydrophobic compounds into the 

poly(dimethylsiloxane) coating of solid-phase 

microextraction fibers: High partition 

coefficients and fluorescence microscopy 

images. Anal Chem 72:459−464. 

5-6 
 

 

 

References 

McGrath JA, Di Toro DM. 2009. Validation of 

the target lipid model for toxicity assessment of 

residual petroleum constituents: Monocyclic and 

polycyclic aromatic hydrocarbons. Environ 

Toxicol Chem 28:1130−1148. 

McGroddy SE, Farrington JW. 1995. Sediment 

porewater partitioning of polycyclic aromatic 

hydrocarbons in three cores from Boston Harbor, 

Massachusetts. Environ Sci Technol 

29:1542−1550. 

McGroddy SE, Farrington JW, Gschwend PM. 

1996. Comparison of the in situ and desorption 

sediment-water partitioning of polycyclic 

aromatic hydrocarbons and polychlorinated 

biphenyls. Environ Sci Technol 30:172–177. 

Middelburg JJ, Nieuwenhuize J, vanBruegel P. 

1999. Black carbon in marine sediments. Mar 

Chem 65:245–252. 

Oen AMP, Janssen EML, Cornelissen G, 

Breeveld GD, Eek E, Luthy RG. 2011. In situ 

measurement of PCB pore water concentration 

profiles in activated carbon-amended sediment 

using passive samplers. Environ Sci Technol 45: 

4053−4059. 

Ouyang G, Pawliszyn J. 2007. Configurations 

and calibration methods for passive sampling 

techniques. J Chromatogr A 1168:226−235. 

Ozretich RJ, Smith LM, Roberts FA. 1995. 

Reverse-phase separation of estuarine interstitial 

water fractions and the consequences of C18 

retention of organic matter. Environ Toxicol 

Chem 14:1261–1272. 

Ozretich RJ, Schults DW. 1998. A comparison 

of interstitial water isolation methods 

demonstrates centrifugation with aspiration 

yields reduced losses of organic constituents. 

Chemosphere 36:603–615. 

Perron MM, Burgess RM, Ho KT, Pelletier MC, 

Friedman CL, Cantwell MG, Shine JP. 2009. 

Development and evaluation of reverse 

polyethylene samplers for marine phase ii 

whole-sediment toxicity identification 

evaluations. Environ Toxicol Chem 28:749−758. 

Rantalainen A-L, Cretney W, Ikonomou MG. 

2000. Uptake rates of semipermeable membrane 

devices (SPMDs) for PCDDs, PCDF, and PCBs. 

Chemosphere 40:147−158. 

Rusina TP, Smedes F, Koblizkova M, Anova J. 

2010. Calibration of silicone rubber passive 

samplers: experimental and modeled relations 

between sampling rate and compound properties. 

Environ Sci Technol 44:362–367. 

Rust AJ, Burgess RM, McElroy AE, Cantwell 

MG, Brownawell BJ. 2004. Influence of soot 

carbon on the bioaccumulation of sedimentbound polycyclic aromatic hydrocarbons by 

marine benthic invertebrates: An interspecies 

comparison. Environ Toxicol Chem 

23:2594–2603. 

Schubauer-Berigan JP, Foote EA, Magar VS. 

2012. Using SPMDs to assess natural recovery 

of PCB-contaminated sediments in Lake 

Hartwell, SC: I. A field test of new in-situ 

deployment methods. Soil Sed Contam 

21:82−100. 

Schults DW, Ferraro SP, Smith LM, Roberts FA, 

Poindexter CK. 1992. A comparison of methods 

for collecting interstitial water for trace organic 

compounds and metals analyses. Water Res 

26:989–995. 

Seethapathy S, Gorecki T, Li X. 2008. Passive 

sampling in environmental analysis. J 

Chromatogr A 1184:234−253. 

Stephan CE, Mount DI, Hansen DJ, Gentile JH, 

Chapman GA, Brungs WA. 1985. Guidelines for 

deriving numerical national water quality 

5-7 
 

Equilibrium Partitioning Sediment Benchmarks (ESBs): Freely Dissolved Concentrations 

criteria for the protection of aquatic organisms 

and their uses. PB85-227049. National 

Technical Information Service (NTIS), 

Springfield, VA, USA. 

Stuer-Lauridsen F. 2005. Review of passive 

accumulation devices for monitoring organic 

micropollutants in the aquatic environment. 

Environ Pollut 136:503−524. 

Suter SW, Mabrey JM. 1994. Toxicological 

benchmarks for screening potential 

contaminants of concern for effects on aquatic 

biota: 1994 revision. ES/ER/TM-96/RI. Oak 

Ridge National Laboratory. Environmental 

Sciences Division, Oak Ridge, TN, USA. 

(http://www.hsrd.ornl.gov/ecorisk/tm96r2.pdf). 

Suter SW, Tsao CL. 1996. Toxicological 

benchmarks for screening potential 

contaminants of concern for effects on aquatic 

biota: 1996 revision. ES/ER/TM-96/R2. Oak 

Ridge National Laboratory. Environmental 

Sciences Division, Oak Ridge, TN, USA. 

(http://www.esd.ornl.gov/programs/ecorisk/documen 

ts/tm96r2.pdf). 

Swartz RC, Cole FA, Lamberson JD, Ferraro SP, 

Schults DW, Deben WA, Lee II H, Ozretich RJ. 

1994. Sediment toxicity, contamination, and 

amphipod abundance at a DDT and dieldrin 

contaminated site in San Francisco Bay. Environ 

Toxicol Chem 13:949–962. 

Swartz RC, Schults DW, DeWitt TH, Ditsworth 

GR, Lamberson JO. 1990. Toxicity of 

fluoranthene in sediment to marine amphipods: 

A test of the equilibrium partitioning approach to 

sediment quality criteria. Environ Toxicol Chem 

9:1071−1080. 

Tomaszewski JE, Luthy RG. 2008. Field 

deployment of polyethylene devices to measure 

PCB concentrations in pore water of 

contaminated sediment. Environ Sci Technol 

42:6086–6091. 

United States Environmental Protection Agency. 

2000. Methods for collection, storage and 

manipulation of sediments for chemical and 

toxicological analyses:technical manual. EPA823-B-01-002. Office of Water, Washington, 

DC, USA. 

United States Environmental Protection Agency. 

2003a. Technical basis for the derivation of 

equilibrium partitioning sediment benchmarks 

(ESBs) for the protection of benthic organisms: 

nonionic organics. EPA-600-R-02-014. Office 

of Research and Development, Washington, DC, 

USA. (draft). 

United States Environmental Protection Agency. 

2003b. Procedures for the derivation of 

equilibrium partitioning sediment benchmarks 

(ESBs) for the protection of benthic organisms: 

Endrin. EPA-600-R-02-009. Office of Research 

and Development, Washington, DC, USA. 

United States Environmental Protection Agency. 

2003c. Procedures for the derivation of 

equilibrium partitioning sediment benchmarks 

(ESBs) for the protection of benthic organisms: 

Dieldrin. EPA-600-R-02-010. Office of 

Research and Development, Washington, DC, 

USA. 

United States Environmental Protection Agency. 

2003d. Procedures for the derivation of 

equilibrium partitioning sediment benchmarks 

(ESBs) for the protection of benthic organisms: 

PAH mixtures. EPA-600-R-02-013. Office of 

Research and Development, Washington, DC, 

USA. 

United States Environmental Protection Agency. 

2005. Contaminated Sediment Remediation 

Guidance for Hazardous Waste Sites. 

EPA/540/R-05/012. Office of Solid Waste and 

Emergency Response. Washington, DC, USA. 

5-8 
 

References 

United States Environmental Protection Agency. 

2007. Sediment Toxicity Identification 

Evaluation (TIE) Phases I, II and III: 

Guidance Document. EPA-600-R-07-080. 

Office of Research and Development. 

Washington, DC, USA. 

United States Environmental Protection Agency. 

2008. Procedures for the derivation of 

equilibrium partitioning sediment benchmarks 

(ESBs) for the protection of benthic organisms: 

Compendium of tier 2 values for nonionic 

organics. EPA-600-R-02-016. Office of 

Research and Development. Washington, DC, 

USA. 

United States Environmental Protection Agency. 

2012. Guidelines for Using Passive Samplers to 

Monitor Organic Contaminants at Superfund 

Sediment Sites. 9200.1-110 FS. Office of 

Superfund Remediation and Technology 

Innovation and Office of Research and 

Development. Washington, DC, USA. 

van Noort PCN. 2003. A thermodynamics-based 

estimation model for adsorption of organic 

compounds by carbonaceous materials in 

environmental sorbents. Environ Toxicol Chem 

22:1179−1188. 

Vinturella AE, Burgess RM, Coull BA, 

Thompson KM, Shine JP. 2004a. The use of 

passive samplers to mimic uptake of polycyclic 

aromatic hydrocarbons by benthic polychaetes. 

Environ Sci Technol 38:1154−1160. 

Vinturella AE, Burgess RM, Coull BA, 

Thompson KM, Shine JP. 2004b. Importance of 

black carbon in distribution and bioaccumulation 

models of polycyclic aromatic hydrocarbons in 

contaminated marine sediments. Environ Toxicol 

Chem 23:2578–2586. 

Vrana B, Mills GA, Allan IJ, Dominiak E, 

Svensson K, Knutsson J, Morrison G, 

Greenwood R. 2005. Passive sampling 

techniques for monitoring pollutants in water. 

Trend Anal Chem 24:845−868. 

Wells JB, Lanno RP. 2001. Passive sampling 

devices (PSDs) as biological surrogates for 

estimating the bioavailability of organic 

chemicals in soil. In: Environmental Toxicology 

and Risk Assessment: Science, Policy, and 

Standardization - Implications for 

Environmental Decisions: Tenth Volume, 

ASTM STP 1403, B.M. Greenberg, R.N. Hull, 

M.H. Roberts, Jr., and R.W. Gensemer, Eds., 

American Society for Testing and Materials, 

West Conshohocken, PA, USA p. 253−270. 

Wenning RJ, Batley GE, Ingersoll CG, Moore 

DW. 2005. Use of Sediment Quality Guidelines 

and Related Tools for the Assessment of 

Contaminated Sediments. Society of 

Environmental Toxicology and Chemistry Press, 

Pensacola, FL, USA. 

Williamson KS, Petty JD, Huckins JN, Lebo JA, 

Kaiser EM. 2002. Sequestration of priority 

pollutant PARs from sediment pore water 

employing semipermeable membrane devices. 

Chemosphere 49:717−729. 

Xu Y, Spurlock F, Wang Z, Gan J. 2007. 

Comparison of five methods for measuring 

sediment toxicity of hydrophobic organic 

contaminants. Environ Sci Technol 

41:8394−8399. 

Zafiriou OC, Joussot-Dubien J, Zepp RG, Zika 

RG. 1984. Photochemistry of natural waters. 

Environ Sci Technol 18:358A−371A. 

Zeng EY, Tsukada D, Diehl DW. 2004. 

Development of a solid phase microextraction– 

based method for sampling of persistent 

chlorinated hydrocarbons in an urbanized coastal 

environment. Environ Sci Technol 

38(21):5737−5743 

5-9 
Office of Research and Development 

National Health and Environmental

 Effects Research Laboratory

Atlantic Ecology Division

27 Tarzwell Drive

Narragansett, RI 02882

www.epa.gov/ord

Official Business

Penalty for Private Use

$300

PRESORTED STANDARD

POSTAGE & FEES PAID

EPA

PERMIT NO. G-35

Please make all necessary changes on the below label, 

detach or copy, and return to the address in the upper 

left-hand corner.

If you do not wish to receive these reports CHECK HERE :

detach, or copy this cover, and return to the address in the 

upper left-hand corner.

Recycled/Recyclable

Printed with vegetable-based ink on 

paper that contains a minimum of 

50% post-consumer fiber content

processed chlorine free.