2
Current Use of Bioavailability in the Management of Contaminated Soil and Sediment

Cleanup of contaminated soil and sediment in the United States follows a risk-based paradigm that takes into account individual exposure pathways linking sources to potential receptors. Typical pathways include contaminant leaching from soil to groundwater, contaminant release from sediments to overlying water, ingestion of contaminated sediments or soils, direct dermal contact with sediments or soils, inhalation of particulate matter or vapors containing contaminants, and ingestion of food items that have accumulated contaminants from soils or sediments. Risk management decisions for soils or sediments focus on identifying relevant pathways of exposure that pose a risk to human health or the environment and then developing appropriate remedial measures that could include treating or removing sources or cutting off pathways or both. Many of the exposure pathways discussed above are affected by the bioavailability processes shown in Figure 1-1. Thus, bioavailability processes are an integral part of risk assessment and risk-based management of contaminated soils and sediments, although their consideration is not always obvious or explicit.

Risk-based cleanup approaches typically are characterized by a tiered methodology, in which a screening-level step is used initially to assess site conditions and potential actions, followed by one or more levels of site-specific assessment. The states have set many guidance values for use at the screening-level step. For example, there are state and federal soil screening levels for the protection of human health (that often differentiate between residential and industrial land use), the protection of groundwater, and the protection of ecological receptors. Sediment guidelines for protection of ecological receptors are often used to guide



The National Academies | 500 Fifth St. N.W. | Washington, D.C. 20001
Copyright © National Academy of Sciences. All rights reserved.
Terms of Use and Privacy Statement



Below are the first 10 and last 10 pages of uncorrected machine-read text (when available) of this chapter, followed by the top 30 algorithmically extracted key phrases from the chapter as a whole.
Intended to provide our own search engines and external engines with highly rich, chapter-representative searchable text on the opening pages of each chapter. Because it is UNCORRECTED material, please consider the following text as a useful but insufficient proxy for the authoritative book pages.

Do not use for reproduction, copying, pasting, or reading; exclusively for search engines.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications 2 Current Use of Bioavailability in the Management of Contaminated Soil and Sediment Cleanup of contaminated soil and sediment in the United States follows a risk-based paradigm that takes into account individual exposure pathways linking sources to potential receptors. Typical pathways include contaminant leaching from soil to groundwater, contaminant release from sediments to overlying water, ingestion of contaminated sediments or soils, direct dermal contact with sediments or soils, inhalation of particulate matter or vapors containing contaminants, and ingestion of food items that have accumulated contaminants from soils or sediments. Risk management decisions for soils or sediments focus on identifying relevant pathways of exposure that pose a risk to human health or the environment and then developing appropriate remedial measures that could include treating or removing sources or cutting off pathways or both. Many of the exposure pathways discussed above are affected by the bioavailability processes shown in Figure 1-1. Thus, bioavailability processes are an integral part of risk assessment and risk-based management of contaminated soils and sediments, although their consideration is not always obvious or explicit. Risk-based cleanup approaches typically are characterized by a tiered methodology, in which a screening-level step is used initially to assess site conditions and potential actions, followed by one or more levels of site-specific assessment. The states have set many guidance values for use at the screening-level step. For example, there are state and federal soil screening levels for the protection of human health (that often differentiate between residential and industrial land use), the protection of groundwater, and the protection of ecological receptors. Sediment guidelines for protection of ecological receptors are often used to guide

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications cleanup. Because they are initial screening levels, they are typically developed to be conservative (i.e., to overestimate most exposures). Although there is continued debate about whether they are conservative enough, it is undisputed that the development of such screening levels requires that assumptions be made about certain bioavailability processes. In most cases, this has involved selecting default conditions or parameters regarding the environmental fate of the chemical as well as how it might enter a human or an ecological receptor. Examples include default assumptions about the relative amount of chemical that is absorbed via dermal contact or incidental ingestion, or the manner and degree to which an organic compound in sediment is bound to organic carbon. For some screening levels (in particular empirical sediment guidelines) bioavailability processes have not been explicitly considered but probably play a role. Understanding how bioavailability processes have been considered at a screening-level stage is an important first step for evaluating how site-specific information might be used to refine exposure and risk assessments and reduce the uncertainties inherent in their outcomes. In some cases, this might involve developing site-specific information for a particular process that can be inserted into a risk equation. As discussed below, there has been considerable work in generating site-specific information on association/dissociation and absorption (bioavailability processes A and D in Figure 1-1) for certain metals in animal models that are applicable to humans. Another type of refinement could involve making site-specific measurements of contaminant release from soils. Still other site-specific estimates—such as those encountered in ecological risk assessments— could involve measurements of available contaminant pools or tissue levels in organisms. This information can be used to both refine a risk assessment calculation and help develop models of bioavailability processes that can be used at other sites. This chapter first describes human health risk assessment to illustrate how bioavailability processes are considered in that arena, followed by an overview of the use of bioavailability processes in ecological risk assessment. The two sections describe the current state of the practice but do not represent an endorsement by the committee. Finally, the chapter describes how “bioavailability” is considered within legal and regulatory frameworks. As will become clear, the legal and even regulatory view of what is meant by “bioavailability” is narrower than the processes illustrated in Figure 1-1, in that the primary focus has been on absorption (particularly systemic absorption for humans) and thus on direct contact with soils via the oral and dermal pathways. This underscores the significance of semantic issues discussed in Chapter 1. What should be clear from this chapter is that bioavailability processes are an integral part of risk-based management of contaminated sites. They may be considered either implicitly or explicitly, and they may be dealt with either by using default values in risk assessment equations or by using site-specific data and information.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications USE OF BIOAVAILABILITY IN RISK ASSESSMENT Because bioavailability processes influence exposure of humans and ecological receptors to chemicals in soils and sediments, and because exposure is one aspect of risk assessment, measuring or modeling bioavailability is consistent with prevailing U.S. Environmental Protection Agency (EPA) and state risk assessment paradigms. The general framework used by EPA for human health risk assessments has four major components derived from NRC (1983): Hazard Identification is a systematic planning stage that identifies the major factors considered in the assessment and establishes its goals, breadth, and focus. It is essentially a scoping activity and is fundamental to the success of all subsequent components in the risk assessment. It consists of stating the objectives, developing the conceptual model, selecting and characterizing receptors, and identifying the endpoints of the assessment. Exposure Assessment estimates the magnitude of actual or potential human or ecological exposure to a contaminant of concern, the frequency and duration of exposure, and the pathways of exposure. Incorporation of bioavailability information often influences estimates of exposure. Dose-Response Assessment is “the process of characterizing the relation between the dose of an agent administered or received and the incidence of an adverse health effect.” This step estimates the probability that an individual will be adversely affected by a given chemical dose, relying primarily on data obtained from animal studies. Information on bioavailability processes may influence measures of toxicity and other effects. Risk Characterization integrates the exposure assessment and dose-response assessment into a quantitative and qualitative expression of risk. This may include deterministic calculations, probabilistic methods, and professional judgement using various lines of evidence. These four steps are similar in ecological risk assessment, with the following differences (EPA, 1992a; NRC, 1993). The first step is termed problem formulation, which determines the focus and scope of the assessment. Hazard identification and dose–response assessment are combined into an ecological effects assessment phase. And finally, dose–response is replaced with stressor–response to emphasize that physical changes make cause harm to ecosystems as well as chemicals (although for the purposes of this report, the focus is on chemical contaminants). Although bioavailability processes can be considered explicitly in both human health and ecological risk assessments, there are some important differences. Unlike human health risk assessment, assessments of exposure and risk to ecological receptors consider various species ranging from invertebrates and plants to fish and wildlife. Some of these species are in intimate contact with soils

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Both direct exposure via soil ingestion and indirect exposure via fish consumption are affected by contaminant bioavailability. Human health risk assessment often quantifies direct ingestion of soil (top photo), while ecological risk assessment frequently considers bioaccumulation of contaminants in animal tissues (bottom photo).

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications or sediments. Many are also exposed to contaminants exchanged from soils or sediments to the dissolved phase or through eating organisms that have accumulated contaminants from these media. Therefore, there are many exposure pathways and a larger number of bioavailability processes that may require simultaneous evaluation during ecological risk assessment as compared to human health risk assessment, where it is more feasible to evaluate one pathway at a time. A manifestation of this difference is that human health risk assessment often involves distinct exposure equations for the direct pathways of ingestion, dermal contact, and inhalation, within which a variable is included to account for absolute or relative bioavailability. This discrete consideration of bioavailability for individual exposure media and exposure routes is driven by the fact that human exposures can often be separated in time and space. For example, vegetables may be grown in a different section of a garden from where children play, and not all receptors have gardens. In contrast, in ecological risk assessment, at least for many receptors there is obligatory simultaneous exposure via multiple pathways and routes. Thus, ecological risk assessments include equations for some of the direct exposure pathways for wildlife (although this knowledge is not well-developed for most species) as well as many other types of measures and exposure models that differ from what is commonly employed in human health assessments. For ecological risk assessment, it is often not be possible to quantify bioavailability processes associated with each of these pathways separately, which is a primary reason for focusing on measures of bioaccumulation as an overall indicator of bioavailability. A second important factor to consider is the acceptability of making measurements on organisms such as earthworms, plants, fish, and wildlife compared to humans. As described in Chapter 4, such measurements include toxicity tests as well as uptake or accumulation tests (determination of tissue residues of contaminants)—tests that for ethical reasons cannot be conducted in humans. Thus, there are more tools for quantifying bioavailability processes and the sum of multiple exposure routes using the actual receptor of interest during ecological risk assessment. This is not the case in human health risk assessment, where greater reliance is placed on default values and where it can be difficult to modify defaults on a site-specific basis. Regardless of whether humans or ecological receptors are the concern at a particular site, some general criteria are useful when attempting to more explicitly consider bioavailability processes during risk assessment (Menzie et al., 2000). First, it is imperative to determine (as best as possible) the usefulness of incorporating new information on bioavailability in terms of altered outcomes at a site. Chapter 5 discusses the chemical and environmental settings for which bioavailability assessments are most likely to make a difference in site management. Second, a conceptual model of exposure for the site is critical to any bioavailability assessment. Because it is known that soils and sediments can alter contaminant bioavailability, relevant soil factors should be identified early. Fi-

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications nally, data on bioavailability processes should be collected using measures or models that are compatible with the risk assessment and risk management framework being used at the site. Human Health Risk Assessment In most situations, a quantitative assessment of risk to humans from exposure to contaminants in soils or sediments involves a comparison of the estimated magnitude of exposure with the measured toxicity of the chemical(s) in question. Bioavailability processes play a variety of important roles in these risk calculations. Although risk calculations for contaminated soils and sediments can sometimes be complex, there are three fundamental types of inputs: (1) the concentration of the chemical in soil or sediment at the point of contact with the individual, (2) variables related to the nature and extent of exposure (e.g., exposure frequency, amount of soil ingested, body weight), and (3) toxicity values for the chemical. Bioavailability processes can be reflected in all three types of inputs. Soil concentration: Bioavailability processes A, B, and C in Figure 1-1 can influence the concentration of chemical reaching the exposed individual from its point of release or residence in the environment. Typically, these bioavailability processes are addressed either through direct measurement of soil concentration at the point of contact or through environmental fate and transport modeling. Exposure variables: Numerical adjustments to account for bioavailability processes related to entry of soil or sediment contaminants into the body are typically included among the exposure variables. This is the usual means by which “bioavailability adjustments” are made in human health risk calculations. Clearly, the primary focus here is on bioavailability processes A and D (association/dissociation and absorption or uptake across a membrane) and to a lesser extent process E if systemic circulation is a measured endpoint. Toxicity values: Toxic potency estimates are based on one or more critical studies which offer information on the relationship between dose of the chemical and toxic effects. Most toxicity values, in the form of cancer potency estimates or acceptable daily intake rates, are based on applied rather than absorbed doses. As a result, the toxicity value is a function, in part, of the rate and extent of absorption that occurred in the critical study. This bioavailability process—the absorption of the chemical into the body in the critical toxicity study—must be kept in mind when using toxicity values. Human contact with contaminants in soils or sediments can occur through three direct routes of exposure: incidental ingestion, dermal contact, or inhalation of soil-derived particulates (dusts) or chemicals volatilized from soil. All three routes are usually relevant for human exposure to soils, while ingestion and dermal contact are the most likely exposure routes for sediments (see Figure 2-1).

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications FIGURE 2-1 Major Exposure Pathways for Human Exposure to Contaminated Soils and Sediments. SOURCE: EPA Region 9 Preliminary Remediation Goals website (www.epa.gov/region09/waste/sfund/prg). In addition to these three routes, there are other indirect pathways by which contaminants in soil and sediment can reach human receptors, notably leaching to groundwater and subsequent ingestion of well water. These routes of exposure are considered below, using contaminated soil (rather than sediment) as an example. Incidental Ingestion Incidental ingestion is often an important exposure route for contaminated soils in human health risk assessments. In its basic form, the intake equation for incidental ingestion of soils is: where: Cs = chemical concentration in the soil at the point of contact IR = incidental ingestion rate of soil RAF = relative absorption factor

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications BW = body weight EF = exposure frequency ED = exposure duration AT = period over which exposure will be averaged. The chemical concentration in soil, soil ingestion rate, and body weight are used to determine the ingestion rate for the chemical per unit body weight. The exposure frequency, exposure duration, and averaging time are used to account for periods when exposure does not occur, and to develop an average intake over time. A correction for relative bioavailability can be introduced in the form of a Relative Absorption Factor (RAF). Usually, the RAF is expressed as a ratio: where Fs is the fraction of the dose of chemical absorbed from soil under circumstances of environmental exposure, and Fsm is the fraction of the dose absorbed from the study medium (e.g., food, water, or some liquid vehicle) used in the critical study upon which the toxicity value is based. The RAF may be an estimated or measured factor, and can be less than or greater than 1.0 (100 percent). If the absorption from soil is found or assumed to be the same as absorption in the critical study upon which the toxicity value is based, then the RAF is 1.0. Note that a RAF of 1.0 does not indicate that absorption is complete, but simply that absorption is known or estimated to be the same as that in the critical study. It is not uncommon for an ingestion intake equation to lack a RAF term. This simply means that the relative bioavailability is assumed to be 1.0. Under some circumstances, the oral toxicity value might be expressed as an internal dose. In this situation, the RAF would be replaced by a term for absolute bioavailability from soil in order to permit an internal dose to be calculated for comparison. Dermal Contact A general form of the equation used to calculate the internal (absorbed) dose from dermal exposure to soil is: where: Cs = chemical concentration in soil on the skin SA = skin surface area AF = soil adherence factor (how much soil covers a unit area of skin) ABS = absorption factor from the soil into the body

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications BW = body weight EF = exposure frequency ED = exposure duration AT = period over which exposure will be averaged The soil concentration, surface area, adherence factor, and body weight terms allow calculation of an amount of chemical present on the skin per unit body weight. As with exposure by ingestion, the exposure frequency, exposure duration, and averaging time terms are present to allow determination of an average exposure rate over time. Usually, the absorption factor (ABS) is intended to reflect the absolute bioavailability of the compound from soil via the dermal route (dermal bioavailability) and is used to calculate the absorbed, or internal, dose of the chemical expected to result from dermal contact. Data on dermal bioavailability from soil are extremely limited or absent for most chemicals, although default assumptions have been specified by EPA and state agencies (see later discussion). Once the intake has been determined from the equation above, it is compared with a suitable toxicity value for dermal exposure. Unfortunately, there are very few toxicity values available specifically for dermal exposure. Instead, if the toxicity is systemic in nature (i.e., doesn’t occur through direct interaction with the skin) the applied-dose toxicity value from another route is converted to an internal-dose value in order to assess risks from dermal contact—a process known as route-to-route extrapolation. This requires knowledge or an assumption regarding the extent of absorption associated with the toxicity value. For example, an oral cancer potency value for a chemical based on a dietary study in laboratory animals could be converted to an internal dose equivalent for use in assessing risks from a chemical entering through the skin. This adjustment in the oral toxicity value would require some knowledge of the gastrointestinal absorption of the chemical in the critical study upon which the oral cancer potency estimate was derived. For cancer potency factors (such as EPA cancer slope factors), the adjustment is made by dividing the oral toxicity value by the known or inferred absolute bioavailability of the chemical from the gut in the critical cancer study. Thus, risks from dermal exposure commonly must rely on estimates of both dermal and oral absolute bioavailability of a chemical, with little supporting data for either. An alternative approach is to compare dermal intake with an oral or inhalation toxicity value without adjustment of the toxicity value to an internal dose form. If this approach is used, the ABS term has a different meaning. Instead of representing the absolute bioavailability of the chemical through the skin, ABS is instead a relative bioavailability term, in this case quantifying the expected difference in absorption from the dermal route versus the absorption implicit in the toxicity value. If the toxicity value for comparison is based on the oral route, then the comparison point is the gastrointestinal absorption of the chemical in the

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications critical oral toxicity study. The example shown in Box 2-1 uses this approach. Similarly, if an inhalation toxicity value is used to assess dermal risks, then the ABS value will be based upon differences in dermal versus inhalation exposure to the chemical. Rarely are experiments conducted to generate these ABS numbers; rather they are the products of best professional judgment. Inhalation Calculating exposure from inhalation of contaminants from soils can be accomplished by measuring or estimating the associated concentration of the chemical in air. A simple form of inhalation intake equation is: where: Ca = chemical concentration in inspired air INR = inhalation rate BW = body weight EF = exposure frequency ED = exposure duration AT = period over which exposure will be averaged This equation calculates the average amount of chemical entering the respiratory tract per unit time and per unit body weight over a specified exposure interval. This intake value is in the form of an applied dose, and is analogous to chemicals entering the gastrointestinal tract after ingestion or coming in contact with the skin during dermal exposure. For exposure to chemicals in soils, the inhalation intake equation often uses the soil concentration and incorporates a model to calculate the corresponding air concentration of the chemical. This model can be viewed as representing the bioavailability processes that make a chemical in soil accessible to its site of entry into the body, which in this case is the lungs. As with ingestion, risks from inhalation exposure are typically assessed through the use of estimates of applied doses resulting from exposure and of toxicity values based on applied doses. Unlike ingestion, however, both the doses and the toxicity values are often expressed in terms of concentration in air, rather than an amount of chemical per unit body weight. For example, a toxicity value for non-cancer health effects by inhalation exposure may be simply a safe concentration limit for the chemical in air. For estimating cancer risks from inhalation exposure, cancer potency can be expressed in reciprocal concentration terms, such that multiplication with the exposure concentration in air yields an excess cancer risk estimate (e.g., EPA inhalation unit risk values). In theory, if differ-

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications ences in pulmonary bioavailability are known to exist between the exposure situation and the critical study used to develop the inhalation toxicity value, this can be addressed through the use of a relative bioavailability or RAF term, as with exposure by ingestion. However, there are few obvious examples of situations where such an adjustment is required, and consequently it is rare in risk assessments. Instead, the implicit assumption is that the relative bioavailability associated with environmental exposure is 100 percent—that is, the pulmonary absorption of the chemical under environmental exposure conditions is equivalent to the pulmonary absorption that existed in the critical study used to derive the inhalation toxicity value. Leaching to Groundwater Leaching from soil to groundwater is another common pathway by which humans can be exposed to contaminants (see Figure 2-2). The calculation requires an estimate of the contaminant concentration in the infiltrating water and a determination of the dilution by mixing with underlying groundwater. Estimation of a soil concentration that will be “protective” of groundwater is achieved by working backward from the desired water concentration at the groundwater well (usually a water quality standard), via the dilution attenuation factor (DAF). The following equation for DAF is meant to account for the dilution by mixing with underlying groundwater: where Qgw is groundwater discharge per unit aquifer thickness over the mixing depth in the aquifer (d); Ql is the leaching recharge [L3L−2T–1]. The Qgw depends upon the aquifer hydraulic conductivity (K), hydraulic gradient (i) and mixing depth (d). The Ql depends upon the area covered by the contaminated soil (L) and infiltration rate (I). The protective soil concentration for this pathway, Cs, is estimated by assuming equilibrium partitioning between the soil- and aqueous-phase contaminant concentrations in the soil pore water using the following equation: where Cw is the water quality standard at the receptor (such as a maximum contaminant level or MCL); Kd is the sorption distribution coefficient for the contaminant; θw and θa are the volumetric air and water contents, ρb is the soil bulk density, and H′ is the dimensionless form of the Henry’s law constant or partitioning coefficient between the air and water phases at a specified temperature. Cs is then compared to the levels of soil contamination at a specific site to determine what actions should be taken next. Unlike the previous three pathways

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications biological effects are different. NOAA researchers have developed an empirical, statistical approach for screening sediment quality that does not explicitly address bioavailability processes at all. EPA has taken a more theoretical approach by developing criteria for protecting ecosystems from sediment toxicity using equilibrium partitioning theory to explain how certain sediment characteristics are thought to affect bioavailability. The USACE uses an experimental approach that tests the toxicity of every sediment (for disposal of dredge spoils), and thereby implicitly considers bioavailability on a sediment-by-sediment basis. The following sections discuss three approaches for setting sediment quality criteria, which refer to recommended concentrations of contaminants in a sediment sample. All three methods take certain bioavailability processes into account, particularly association/dissociation and absorption. Sediment quality criteria are basically analogous to standards for water quality; however, sediment quality criteria are not legally enforceable. Possible exceptions are the Great Lakes sediment criteria, which are used to set enforceable water quality standards. Section 118(c)(2) of the CWA (as amended by the Great Lakes Critical Programs Act of 1990) requires EPA to publish guidance on minimum water quality standards, antidegradation policies, and implementation procedures for the Great Lakes. The resulting guidance (EPA, 1995c) incorporates bioaccumulation factors into the derivation of sediment quality criteria and values to protect human health and wildlife. EPA Approach EPA has formulated sediment quality criteria to be consistent with previously established standards for water quality using an approach referred to as equilibrium partitioning. Recently, these sediment quality criteria have been renamed equilibrium partitioning sediment guidelines (ESGs) by EPA (see EPA, 2001c). The approach assumes that contaminants partition between the aqueous and solid phase as a function of sediment composition and contaminant type. Sediment contamination above a concentration that results in an aqueous phase level greater than water quality standards is not acceptable and thus determines the value of the ESG. The water quality standards are based upon toxicity bioassays with benthic invertebrates, dissolved contaminants, and aqueous conditions that maximize uptake. Thus, the ESGs are described as EPA’s best estimate of the concentrations of a substance that may be present in sediment and still protect benthic organisms from direct toxicity in that sediment. EPA has conducted efforts to develop and publish ESGs for some of the 65 toxic pollutants or toxic pollutant categories (EPA, 2000, 2001c). ESGs incorporate research that identified some of the chemical factors that influence partitioning from sediments to the dissolved phase, and thus directly address an important bioavailability process (particularly A in Figure 1-1). ESGs

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications can be used to address site-specific issues, and they are quantitative. Note that ESGs do not protect against synergistic or antagonistic effects of contaminants or bioaccumulative effects of contaminants, and they are not protective of wildlife or human health endpoints. ESGs are not regulations and do not impose legally binding requirements. In addition, EPA does not recommend the use of ESGs as stand-alone, pass-fail criteria for sediments. Instead, they are intended for use as screening levels. As for other strengths and limitations, it is worth noting here that the guidelines may not be applicable where digestive uptake is a significant exposure pathway, where multiple pollutants occur, or where the bioavailability is controlled by physicochemical factors not considered in the EqP approach. The measures are more applicable for instances of acute toxicity or toxicity via direct contact to gills or surfaces of the organism as compared to chronic toxicity. Given such limitations, these measures are best used for those cases where extreme sediment contamination immediately kills fauna or flora. NOAA Approach In contrast to the ESGs used by EPA, numerical sediment quality guidelines (SQGs) have been suggested by researchers at NOAA (Long et al., 1995, 1998). In this approach, contaminant concentrations in sediment are correlated with large data sets of observed biological effects (usually done with toxicity tests using field-collected sediments). To date, corresponding chemical and biological effects data have been compiled from one thousand or more studies. A SQG is defined as a range between a lower and upper concentration limit. The lower limits are intended to represent concentrations below which adverse effects were not frequently expected. The upper limit values are the concentrations above which effects had a high probability of occurrence. The range between the lower and higher values can vary but is typically between factors of 2 and 10. One limitation of this approach is that sediments often contain more than one contaminant, but the majority of studies showing biological effects were conducted by evaluating contaminants individually. Bioavailability processes are not explicitly considered in this approach at an individual site, but they have implicit influences. For example, results from sediment toxicity tests can encompass both sediment chemistry and species-specific effects. The authors state that “numerical values were not intended as regulatory criteria” (Long et al., 2000). Nevertheless, this approach is sometimes used at the local and state level, at least informally, to screen or characterize sediment contamination problems, perhaps because the guidelines are simple to apply. Because this approach is confounding with respect to bioavailability processes, it is not suggested even for screening-level assessments of sites.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications USACE Approach A third approach to sediment quality criteria is that of the USACE for managing dredge spoils. As part of its navigation dredging, the USACE disposes of contaminated sediments in confined disposal facilities for inland sites or in ocean disposal sites. The law dictating Corps activities in this regard (The Marine Protection, Research and Sanctuaries Act) has been interpreted in EPA and Corps’ guidance documents as requiring bioaccumulation testing and other bioassays for purposes of determining which materials are environmentally acceptable for ocean dumping (EPA and USACE, 1998). Using this approach, dredged soils are first tested for bulk chemical concentrations (Tier I). If these tests indicate contamination, then the sediment elutriate is tested for concentration and toxicity via benthic bioassays (Tier II). One of the assumptions of the Tier II water column analysis is that all contaminants present in the sediment will be released to the water column during disposal and emplacement, although it is acknowledged that this assumption is highly conservative due to the tendency of many contaminants to remain associated with the sediment. Third and fourth tier testing involve advanced site-specific toxicity and bioaccumulation experiments and bioassays with a deposit feeding bivalve. Tests are conducted during a 28-day exposure time for organisms that are selected based on their ability to metabolize the target analyte and to survive the exposure test. The first two tiers of this approach are widely used; the bioaccumulation tests in Tiers III and IV are less frequently needed because the earlier tests are pass–fail. The approach of USACE is empirical, site-specific, and more biologically based than are the other two approaches. The tiered treatment of site-specific sediments considers different bioavailability processes at different times. However, the measures cannot be extrapolated to other circumstances, nor are the relative influences of different bioavailability processes quantified. *** The methodology and approaches used for sediment quality criteria differ among the three agencies in fundamental ways, not the least in how certain bioavailability processes are assessed and taken into consideration. These differences could serve as a point of confusion for practitioners hoping to better quantify the risks involved in various sediment management scenarios, and they reflect the lack of consensus among environmental managers about how to deal with bioavailability processes.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications CONCLUSIONS AND RECOMMENDATIONS Considering the processes that influence bioavailability is entirely within the human health and ecological risk framework. Bioavailability processes should not be considered as “something new” that falls outside of the basic risk-based approach to hazardous waste cleanup that has been adopted in the United States. The goal of bioavailability analysis is to reduce uncertainty in exposure estimates and thus improve the accuracy of the risk assessment. Although consideration of bioavailability processes is inherent to risk assessment, usually only some of the relevant bioavailability processes are considered explicitly, and assumptions made about the other processes are not transparent. All risk assessments contain implicit, and usually conservative, assumptions about many bioavailability processes. However, different users have chosen different processes to consider explicitly. For example, EPA has focused on the absorption aspect of bioavailability (through the use of default values for dermal and oral relative bioavailability and BSAF values) while many of the other processes have been less explicitly examined. Because of this variability, it is important to use parameters containing the word “bioavailability” (such as absolute bioavailability and relative bioavailability) only with very clear definition of the parameter and its role in the entire spectrum of bioavailability processes. The lack of mechanistic understanding and description during risk assessment precludes the development of technically sound exposure models, especially those that could incorporate temporal changes in physical, chemical, and biological factors. Explicit consideration of bioavailability processes is more common in ecological risk assessment than in human health risk assessment. This is because it is easier and more acceptable to make measurements on ecological receptors (e.g., worms, small mammals, birds, fish) than it is on humans, and because risk managers are usually willing to manage uncertainty in ecological risk assessments (including the incorporation of bioavailability processes) differently. In addition, during ecological risk assessment there is a greater focus on how bioavailability processes influence bioaccumulation into various wildlife food items. The burden of proof is often higher for adjusting exposure estimates for human receptors than it is for ecological receptors. There is a misconception that the default values representing bioavailability processes in risk assessment are protective and appropriate for all circumstances. The tendency to standardize regulatory risk assessment has led to the use of certain default factors (e.g., equilibrium partitioning for organic chemicals, relative bioavailability values, dilution attenuation factors) typically considered to have wide applicability across a variety of sites. Although determining these default values required explicit consideration of bioavailability processes

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications with adoption of a conceptual model and incorporation of quantitative assumptions, the values are sometimes based on only a few studies and may not be applicable to the site of interest. Thus, replacing default values with site-specific information should be encouraged. It should be noted that consideration of site-specific information on bioavailability processes may result in an increase or decrease compared to the “default value.” At present there is no legal recognition of “bioavailability” in soil cleanup, although bioavailability concepts are emerging for sediment management, and they have been embraced more fully for biosolids management and disposal. The fact that the term “bioavailability” does not appear in the laws and regulations, but does appear in the informal comments to regulations and guidance documents, necessarily leads to confusion and even conflict over the acceptability of the concept. More formal recognition of “bioavailability” in state and federal regulatory contexts would eliminate at least some of the hesitancy and confusion on the part of risk assessors and managers. The lack of clear authorization or guidance on using bioavailability in site-specific risk assessments from EPA has generally led to the perception that the approach is not favored. There is no clear regulatory guidance or scientific consensus about the level and lines of evidence needed for comprehensive bioavailability process assessment. That is, it is not clear what threshold of knowledge is sufficient to be able to replace default assumptions about bioavailability with site-specific measurements. All of the decisions made at the limited number of case histories have been unique and variable. Regulatory guidance from EPA is needed that addresses what information must be included in a bioavailability process assessment, its scientific validity, acceptable models of exposure, and other issues. This may help to guide research efforts that will further mechanistic understanding of bioavailability processes. REFERENCES Ankley G. T., G. L. Phipps, E. N. Leonard, D. A. Benoit, V. Mattson, P. A. Kosian, A. M. Cotter, J. R. Dierkes, D. J. Hansen, and J. D. Mahony. 1991a. Acid-volatile sulfide as a factor mediating cadmium and nickel bioavailability in contaminated sediments. Environ. Toxicol. Chem. 10:1299-1307. Ankley G. T., M. K. Schubauer-Berigan, and J. R. Dierkes. 1991b. Predicting the toxicity of bulk sediments to aquatic organisms with aqueous test fractions: pore water vs. elutriate. Environ. Toxicol. Chem. 10:1359-1366. Ankley, G. T., P. M. Cook, A. R. Carlson, D. J. Call, J. A. Swenson, H. F. Corcoran, and R. A. Hoke. 1992. Bioaccumulation of PCBs from sediments by oligochaetes and fishes: comparison of laboratory and field studies. Canadian Journal of Fisheries and Aquatic Sciences 49:2080-2085. Axtmann, E. V., and S. N. Luoma. 1991. Large-scale distribution of metal contamination in the fine-grained sediments of the Clark Fork River, Montana. Applied Geochemistry 6:75-88.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Axtmann, E. V., D. J. Cain, and S. N. Luoma. 1997. Effect of tributary inflows on the distribution of trace metals in fine-grained bed sediments and benthic insects of the Clark Fork River, Montana. Environ. Sci. Technol. 3:750-758. Beyer, W. N., D. Day, A. Morton, and Y. Pachepsky. 1998a. Relation of lead exposure to sediment ingestion in mute swans on the Chesapeake Bay, USA. Environ. Toxicol. Chem. 17(11):2298-2301. Beyer, W. N., D. J. Audet, A. Morton, J. K. Campbell, and L. LeCaptain. 1998b. Lead exposure of waterfowl ingesting Coeur d’Alene River basin sediments. J. Environ. Qual. 27(6):1533-1538. Beyer, W. N., E. E. Connor, and S. Gerould. 1994. Estimates of soil ingestion by wildlife. J. Wildl. Manage. 58:375-382. Beyer, W. N., J. Spann, and D. Day. 1999. Metal and sediment ingestion by dabbling ducks. The Science of the Total Environment 231:235-239. Beyer, W. N., L. J. Blus, C. J. Henny, and D. Audet. 1997. The role of sediment ingestion in exposing wood ducks to lead. Ecotoxicology 6:181-186. Boese, B. L., and H. Lee II. 1992. Synthesis of methods to predict bioaccumulation of sediment pollutants. ERL-N No. N232. Narraganset, RI: EPA Environmental Research Laboratory. Boese B. L., M. Winsor, H. Lee, S. Echols, J. Pelletier, and R. Randall. 1995. PCB congeners and hexachlorobenzene biota sediment accumulation factors for Macoma nasuta exposed to sediments with different total organic carbon contents. Environ. Toxicol. Chem. 14:303-310. Cain, D. J., S. N. Luoma, J. L. Carter, and S. V. Fend. 1992. Aquatic insects as bioindicators of trace element contamination in cobble-bottom rivers and streams. Can. J. Fish-Aquat. Sci. 49:2141-2154. California EPA. 1993. Documentation for the CalEPA CALTOX model. Casteel, S. W., L. D. Brown, M. E. Dunsmore, C. P. Weis, G. M. Henningsen, W. J. Hoffman Brattin, and T. L. Hammon. 1996. Bioavailability of lead in soil samples from the New Jersey Zinc NPL Site, Palmerton, Pennsylvania. Report to EPA Region III from EPA Region VIII study team, C. P. Weis, Team Leader. Centers for Disease Control (CDC). 1991. Preventing lead poisoning in children: statements by the Centers for Disease Control. Atlanta, GA: Centers for Disease Control and Prevention. Chaney, R. L, S. L. Brown, and J. S. Angle. 1998. Soil-root interface: ecosystem health and human food-chain protection. Pp. 279-312 In: Soil chemistry and ecosystem health. P.M. Huang (ed.). Madison, WI: Soil Science Society of America. Chaney, R. L., S. B. Sterett, M. C. Morella, and C. A. Lloyd. 1982. Effect of sludge quality and rate, soil pH and time on heavy metal residues in leafy vegetables. Pp. 444-458 In: Proceedings of the Annual Conference on Applied Research Practices on Municipal Industrial Waste, September 22-24, 1982. Madison, WI: Univ. of Wisconsin Exension. Clark, T., K. Clark, S. Paterson, D. Mackay and R. J. Norstrom. 1988. Wildlife monitoring, modeling, and fugacity. Environ. Sci. Technol. 22:120-127. Cook, P. M., R. J. Erickson, R. L. Spehar, S. P. Bradbury, and G. T. Ankley. 1993. Interim report on the assessment of 2,3,7,8-tetrachlorodibenzo-p-dioxin risk to aquatic life and associated wildlife. Washington, DC: EPA. DiToro, D. M., J. D. Mahony, D. J. Hansen, K. J. Scott, M. B. Hicks, S. M. Mayr, and M. S. Redmond. 1990. Toxicity of cadmium in sediments: the role of acid volatile sulfides. Environ. Toxicol. Chem. 9:1487-1502. DiToro, D. M., C. S. Zarba, D. J. Hansen, W. J. Berry, R. C. Swartz, C. E. Cowan, S. P. Pavlou, H. E. Allen, N. A. Thomas, and P. R. Paquin. 1991. Technical basis for establishing sediment quality criteria for nonionic organic chemicals using equilibrium partitioning. Environ. Toxicol. Chem. 10:1541-1583. Drouillard, K. G., J. J. H. Ciborowski, R. Lazar, and G. D. Haffner. 1996. Estimation of the uptake of organochlorines by the mayfly (Hexagenia limbata; Ephemeroptera; Ephemeridae). J. Great Lakes Res. 22:26-35.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Eadie, B. J., W. R. Faust, N. R. Morehead, and P. F. Landrum. 1985. Factors affecting bioconcentration of PAH (polynuclear aromatic hydrocarbon) by the dominant benthic organisms of the Great Lakes. Polynucl. Aromat. Hydrocarbons, [Pap. Int. Symp.], 8th. 363-377. Environmental Protection Agency (EPA). 1989a. Risk assessment guidance for Superfund. Volume 1. Human Health Evaluation Manual (Part A). Interim Final. EPA/540/1-89/002. Washington, DC: EPA Office of Emergency and Remedial Response. EPA. 1989b. Development of risk assessment methodology for land application and distribution and marketing of municipal sludge. EPA/600/6-89/001. Washington, DC: EPA. EPA. 1990. Corrective action for solid waste management units at hazardous waste management facilities. Federal Register 55(145):30865-30867. EPA. 1991. A guide to principal threat and low level threat wastes. OSWER 9380.3-06FS. Washington, DC: EPA. EPA. 1992a. Ecological risk assessment guidance for Superfund: process for designing and conducting ecological risk assessments. EPA 540-R-97-006. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1992b. Technical support document for land application of sewage sludge. NTIS PB93-110575. Springfield, VA: EPA Office of Water. EPA. 1992c. Monitoring guidance for the National Estuary Program. EPA 842-B-92-004. Washington, DC: EPA Office of Water. EPA. 1993. Standards for the use or disposal of sewage sludge: final rules. Federal Register 58(32):9248-9415. EPA. 1994a. Guidance manual for the Integrated Exposure Uptake Biokinetic Model for lead in children. EPA/540/R-93/081, PB93-963510. Washington, DC: EPA. EPA. 1994b. Baseline risk assessment for lead. Priority soils operable unit, Silver Bow Creek/Butte Area NPL Site, Butte, Montana. Prepared by CDM Federal Programs Corporation, Golden, Colorado. Denver, CO: EPA Region VIII. EPA. 1994c. CWA Section 403: procedural and monitoring guidance. EPA 842-B-94-003. Washington, DC: EPA Office of Water. EPA. 1995a. A guide to the biosolids risk assessments for the EPA Part 503 Rule. EPA 832-B-93-005. Washington, DC: EPA Office of Wastewater Management. EPA. 1995b. Guidance for assessing chemical contaminant data for use in fish advisories. Volume 1, fish sampling and analysis. 2nd Ed. EPA 823-R-95-007. Washington, DC: EPA Office of Water. EPA. 1995c. Final water quality guidance for the Great Lakes System. Federal Register 60(56):15366-15425. EPA. 1996a. Soil screening guidance: technical background document. EPA/540/R-95/128, PB96-963502. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1996b. Recommendations of the technical review workgroup for an interim approach to assessing risks associated with adult exposures to lead in soil. Washington, DC: EPA Technical Review Workgroup for Lead. EPA. 1997a. Toxicological review of chlordane (technical) (CAS No. 12789-03-6) in support of summary information on the Integrated Risk Information System, December, 1997. EPA. 1997b. The incidence and severity of sediment contamination in surface waters of the United States, Volume 1: the national sediment quality survey. EPA 823-R-97-006. Washington, DC: EPA Office of Science and Technology. EPA. 1999a. IEUBK model bioavailability variable. EPA 540-F-00-006, OSWER 9285.7-32. Washington, DC: EPA Office of Emergency and Remedial Response. EPA. 1999b. Clark Fork River ecological risk assessment. Denver, CO: EPA Region 8. 365pp. EPA. 2000. Bioaccumulation testing and interpretation for the purpose of sediment quality assessment: status and needs. Washington, DC: EPA Offices of Water and Solid Waste and Emergency Response.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications EPA. 2001a. Risk assessment guidance for Superfund. Volume 1: human health evaluation manual (Part E, supplemental guidance for dermal risk assessment), Interim Guidance. Washington, DC: EPA. EPA. 2001b. User’s guide for the Integrated Exposure Uptake Biokinetic Model for lead in children (IEUBK). EPA 9285.7-42. Washington, DC: Office of Emergency and Remedial Response. EPA. 2001c. Draft implementation framework for the use of equilibrium partitioning sediment guidelines. Guidance for using equilibrium partitioning sediment guidelines (ESGs) in water quality programs. Washington, DC: EPA Office of Water. EPA and U.S. Army Corps of Engineers. 1991. Evaluation of dredged material proposed for ocean disposal—testing manual. EPA 503-8-91-001. Washington, DC: EPA/USACE. EPA and U.S. Army Corps of Engineers. 1998. Evaluation of dredged material proposed for discharge in waters of the U. S.: Inland testing manual. EPA-823-B-98-004. Washington, DC: EPA/USACE. EPA Region 3. 1998. Risk-based concentration table. Prepared by EPA Region 3 Superfund Technical Support Section. http://www.epa.gpv/reg3hwmd/risk/riskmenu.htm. EPA Region 4. 2000. Human health risk assessment bulletins—supplement to RAGS. Washington, DC: EPA Region 4 Waste Management Division. EPA Region 5. 1996. Basic Brownfields fact sheet. Washington, DC: EPA Office of Pub. Affairs. Farag, A. M., M. A. Stansbury, C. Hogstrand, E. MacConnell, and H. L. Bergmann. 1995. The physiological impairment of free-ranging brown trout exposed to metals in the Clark Fork River, Montana. Can. J. Fish. Aquat. Sci. 52:2038-2050. Freeman, G. B., J. D. Johnson, J. M. Killinger, S. C. Liao, P. I. Feder, A. O. Davis, M. V. Ruby, R. L. Chaney, S. C. Lovre, and P. D. Bergstrom. 1992. Relative bioavailability of lead from mining waste soil in rats. Fund. Appl. Tox. 19:388–398. Froese, K. L., D. A. Verbrugge, G. T. Ankley, G. J. Niemi, C. P. Larsen, and J. P. Giesy. 1998. Bioaccumulation of polychlorinated biphenyls from sediments to aquatic insects and tree swallow eggs and nestlings in Saginaw Bay, Michigan, USA. Environ. Toxicol. Chem. 17:484-492. Hani, H. 1996. Soil analysis as a tool to predict effects on the environment. Commun. Soil Sci. Plant. Anal. 27:289-306. Hansen, D. J., J. D. Mahony, W. J. Berry, S. J. Benyi, J. M. Corbin, S. D. Pratt, D. M. DiToro, and M. B. Abel. 1996. Chronic effect of cadmium in sediments on colonization by benthic marine organisms: an evaluation of the role of interstitial cadmium and acid-volatile sulfide in biological availability. Environ. Toxicol. Chem. 15:2126-2137. Hare, L., R. Carignan, and M. A. Huerta-Diaz. 1994. A field study of metal toxicity and accumulation by benthic invertebrates: implications for the acid-volatile sulfide (AVS) model. Limnol. Oceanogr. 39:1653-1668. Hawker, D. W., and D. W. Connell. 1985. Relationships between partition coefficient, uptake rate constant, clearance rate constant and time to equilibrium for bioaccumulation. Chemosphere 14:1205-1219. Hillman, T. W., D. W. Chapman, T. S. Hardin, S. E. Jensen, and W. S. Platts. 1995. Assessment of injury to fish populations: Clark Fork River NPL sites, Montana. Appendix G In: Aquatics resources injury assessment report, Upper Clark Fork River basin. J. Lipton, et al. (eds.). Report to the State of Montana Natural Resource Damage Program, Helena, Montana. Hoke, R. A., G. T. Ankley, A. M. Cotter, T. Goldenstein, P. A. Kosian, G. L. Phipps, and F. M. VanderMeiden. 1994. Evaluation of equilibrium partitioning theory for predicting acute toxicity of field-collected sediments contaminated with DDT, DDE and DDD to the amphipod Hyalella azteca. Environ. Toxicol. Chem. 13:157-166. Hope, B. 2001. A case study comparing point estimate and spatially explicit ecological exposure analysis methods. Risk Analysis 21(6):1001-1010. Illinois EPA. 1996. Tiered approach to corrective action objectives. 35 ILL. Adm. Code 742. Springfield, IL: Illinois Pollution Control Board.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Ioven, D., and J. Hubbard. 2000. Letter from D. Ioven and J. Hubbard, EPA Region III, to L. Ehlers, National Research Council, November. Jenne, E. A., and S. N. Luoma. 1977. Forms of trace elements in soils, sediments, and associated waters: an overview of their determination and biological availability. Pp. 110-143 In: Biological implications of metals in the environment. R. E. Wildung and H. Drucker (eds). Johnson, H. E., and C. L. Schmidt. 1988. Clark Fork Basin Project: status report and action plan. Report to the Office of the Governor, Helena, Montana. Karickhoff, S. W., D. S. Brown, and T. A. Scott. 1979. Sorption of hydrophobic pollutants on natural sediments. Water Res. 13:241-248. Kemble, N. E., W. D. Brumbaugh, E. Brunson, J. D. Dwyer, C. G. Ingersoll, D. P. Monda, and D. F. Woodward. 1994. Toxicity of metal contaminated sediments from the Upper Clark Fork River, Montana, to aquatic invertebrates in laboratory exposures. Environ. Toxicol. Chem. 13(12):1985-1997. Lake, J. L., N. I. Rublinstein, H. H. L. Lee, C. A. Lake, J. Heltshe, and S. Pavignano. 1990. Equilibrium partitioning and bioaccumulation of sediment-associated contaminants by infaunal organisms. Environ. Toxicol. Chem. 9:1095-1106. Landrum, P. F., and R. Poore. 1988. Toxicokinetics of selected xenobiotics in Hexagenia limbata. J. Great Lakes Res. 14:427-437. Landrum, P. F., W. R. Faust, and B. J. Eadie. 1989. Bioavailability and toxicity of a mixture of sediment-associated chlorinated hydrocarbons to the amphipod Pontoporeia Hoyl. Aquatic toxicology and hazard assessment: 12th Volume. U. M. Cowgill and L. R. Williams (eds). Philadelphia, PA: ASTM. Landrum, P. F., H. Lee II, and M. J. Lydy. 1992. Toxicokinetics in aquatic systems: model comparisons and use in hazard assessment. Environ. Toxicol. Chem. 11:1709-1725. Lang, W. L. 1988. Pp. 130 In: The last best place. W. Kittredge and A. Smith (eds.). Missoula, MT: Mont. Hist. Soc. Press. Lee, J.-S., B.-G. Lee, S. N. Luoma, H.-J. Choi, C.-H. Koh, and C. L. Brown. 2000a. Influence of acid volatile sulfides and metal concentrations on metal partitioning in contaminated sediments. Environ. Sci. Technol. 34:4511-4516. Lee, B.-G., S. B. Griscom, J.-S. Lee, H. J. Choi, C.-H. Koh, S. N. Luoma, and N. S. Fisher. 2000b. Influence of dietary uptake and reactive sulfides on metal bioavailability from aquatic sediments. Science 287:282-284. Lee, H. 1992. Models, muddles and mud: predicting bioaccumulation of sediment-associated pollutants. Pp. 267-294 In: Sediment toxicity assessment. Chelsea, MI: Lewis Publishers. Ling, H., M. Diamond, and D. Mackay. 1993. Application of the QWASI fugacity/equivalence model to assessing sources and fate of contaminants in Hamilton Harbor. J. Great Lakes Res. 19:582-602. Long, E. R., D. D. Macdonald, S. L. Smith, and F. D. Calder. 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environ. Mgt. 19:81-97. Long, E. R., L. J. Field, and D. D. MacDonald. 1998. Predicting toxicity in marine sediments with numerical sediment quality guidelines. Environ. Toxicol. Chem. 17:714-727l. Long, E. R., D. D. MacDonald, C. G. Severn, and C. B. Hong. 2000. Classifying probabilities of acute toxicity in marine sediments with empirically derived sediment quality guidelines. Environ. Toxicol. Chem. 19(10):2598-2601. Luoma, S. N., and K.-T. Ho. 1993. Appropriate uses of marine and estuarine sediment bioassays. Chapter 11 In: The handbook of ecotoxicology, Volume 1. P. Calow (ed.). Luoma, S. N., and Jenne, E. A. 1977. The availability of sediment-bound cobalt, silver, and zinc to a deposit-feeding clam. Pp. 213-230 In: Biological implications of metals in the environment. R. W. Wildung and H. Drucker (eds.).

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Luoma, S. N. 1995. Prediction of metal toxicity in nature from bioassays: limitations and research needs . Pp. 609-646 In: Metal speciation and bioavailability in aquatic systems. A. Tessier and D. Turner (eds.). London: John Wiley and Sons, Ltd. Mackay, D., and S. Paterson. 1991. Evaluating the multimedia fate of organic chemicals: a level III fugacity model. Environ. Sci. Technol. 25:427-436. Maddaloni, M. 2000. Presentation to the NRC Committee on Bioavailability of Contaminants in Soils and Sediments. October 13, 2000. Woods Hole, MA. Magee, B., P. Anderson, and D. Burmaster. 1996. Absorption adjustment factor (AAF) distributions for polycyclic aromatic hydrocarbons (PAHs). Human Ecol. Risk Assess. 2(4):841-873. Malone, L. A. 2002. Environmental regulation of land use. Section 9.03(5). St. Paul, MN: West Publishing. Massachusetts Department of Environmental Protection Office of Research and Standards. 1992. Documentation for the risk assessment short form—residential scenario. Policy #WSC/ORS-142-92. Boston, MA: Commonwealth of Massachusetts Office of the Secretary of State. Menzie, C. A., D. E. Burmaster, J. S. Freshman, and C. A. Callahan. 1992. Assessment of methods for estimating ecological risk in the terrestrial component: a case study at the Baird & McGurie Superfund site in Holbrook, Massachusetts. Environ. Toxicol. Chem. 11:245-260. Menzie, C. A., A. M. Burke, D. Grasso, M. Harnois, B. Magee, D. McDonald, C. Montgomery, A. Nichols, J. Pignatello, B. Price, R. Price, J. Rose, J. Shatkin, B. Smets, J. Smith, and S. Svirsky. 2000. An approach for incorporating information on chemical availability in soils into risk assessment and risk-based decision making. Human and Ecological Risk Assessment 6(3):479-510. MPCA. 1999. Risk-based guidance for soil-human health pathway. Volume 2. Technical support document. St. Paul, MN: Minnesota Pollution Control Agency. Available at: http://hubble.pca.state.mn.us/cleanup/pubs/srv3_99.pdf. Moore, J. N., and S. N. Luoma. 1990. Hazardous wastes from large scale metal extraction: A case study. Environ. Sci. Technol. 24:1279-1285. Neely, W. B., and D. Mackay. 1982. Evaluative model for estimating environmental fate. Pp. 127-144 In: Modeling the fate of chemicals in the aquatic environment. K. L. Dickson, A. W. Maki, and J. Cairns Jr. (eds.). Ann Arbor, MI: Ann Arbor Science Publishers. Nichols, J. W., C. P. Larsen, M. E. McDonald, G. J. Niemi, and G. T. Ankley. 1995. Bioenergetics-based model for accumulation of PCBs by nesting tree swallows, Tachycineta bicolor. Environ. Sci. Technol. 29:604-612. Novick, S. 2002. The law of environmental protection. Section 13.05(3)(G). St. Paul, MN: West Publishing. National Research Council (NRC). 1983. Risk assessment in the federal government: managing the process. Washington, DC: National Academy Press. NRC. 1993. Issues in risk assessment. Washington, DC: National Academy Press. O’Conner, G. A., D. Kiehl, G. A. Eiceman, and J. A. Ryan. 1990. Plant uptake of sludge-borne PCBs. J. Environ. Qual. 19:113-118. O’Connor, G., R. M. Brobst, R. L. Chaney, R. L. Kincaid, L. R. McDowell, G. M. Pierzynski, A. Rubin, and G. G. Van Riper. 2001. A modified risk assessment to establish molybdenum standards for land application of biosolids. J. Environ. Qual. 30:1490-1507. Ohio Department of Commerce. 1992. Risk assessment guidance document. Columbus, OH: Division of State Fire Marshall, Bureau of Underground Storage Tank Regulations. OIS. 1998. Commentary on the ordinance of 1 July 1998 relating to impacts on the soil. Swiss Agency for the Environment, Forests and Landscape (SAEFL). Order number VU-4809-E. Oklahoma DEQ. 1994. Technical memorandum: preliminary remediation goals for the National Zinc Site, Bartlesville, Oklahoma. Prepared by Oklahoma Department of Environmental Quality, Oklahoma City, Oklahoma.

OCR for page 52
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Page, A. L., T. J. Logan, and J. A. Ryan (ed.). 1989. W-170 Peer review committee analysis of the proposed 503 Rule on sewage sludge. CSRS Technical Committee W-170, Univ. of California, Riverside. Parkerton, T. F., J. P. Connolly, R. V. Thomann, and C. G. Uchrin. 1993. Do aquatic effects or human health end points govern the development of sediment-quality criteria for nonionic organic chemicals? Environ. Toxicol. Chem. 12:507-523. Rand, G. M. 1995. Fundamentals of aquatic toxicology: effects, environmental fate, and risk assessment. 2nd Edition. New York: Hemisphere Publishing Corporation. 1125 pp. Rhoads, D. C., and L. F. Boyer. 1983. The effects of marine benthos on physical properties of sediments: a successional perspective. Chapter 1 In: Animal-sediment relations: the biogenic alteration of sediments. P. L. McCall and M. J. S. Tevesz (eds.). New York: Plenum Press. Ruby, M. V., R. Schoof, W. Brattin, M. Goldade, G. Post, M. Harnois, D. E. Mosby, S. W. Casteel, W. Berti, M. Carpenter, D. Edwards, D. Cragin, and W. Chappell. 1999. Advances in evaluating the oral bioavailability of inorganics in soil for use in human health risk assessment. Environ. Sci. Technol. 33(21):3697-3705. Sablijc, A., H. Gutsen, H. Verhaar, and J. Hermens. 1995. QSAR modeling of soil sorption: improvements and systematics of log Koc vs. log Kow corrections. Chemosphere 21:4489-4514. Starodub, M. E., P. A. Miller, G. M. Ferguson, J. P. Giesy, and R. F. Willis. 1996. The use of risk assessment techniques to develop a protocol for determining acceptable concentrations of bioaccumulative chlorinated organic chemicals in sediments for the protection of piscivorous wildlife. Toxicol. Environ. Chem. 54:243-259. Szumski, M. J. 1998. The effects of mining related metals contamination on piscivorous mammals along the Upper Clark Fork River, MT. Ph D Thesis, Univ. Wyoming, Laramie. 148pp. Tracey, G. A., and D. J. Hansen. 1996. Use of biota-sediment accumulation factors to assess similarity of nonionic organic chemical exposure to benthically coupled organisms of differing trophic mode. Arch. Environ. Contam. Toxicol. 30:467-475. Velleux, M., and D. Endicott. 1994. Development of mass balance model for estimating PCB export from the lower Fox River to Green Bay. J. Great Lakes Res. 20:416-434. VSBo. 1986. Verordnung über Schadstoffgehalt im Boden. In: Swiss ordinance on pollutants in soils No. 814. 014, Publ. Eidg. Dricksachen und Materialzentrale, Bern, Switzerland. WVDEP. 1999. Guidance manual for the West Virginia Voluntary Remediation Program, Version 1-1. Charleston, WV: West Virginia Division of Environmental Protection. Wester, R. C., H. I. Maibach, D. A. W. Bucks, L. Sedik, J. Melendres, C. L. Laio, and S. DeZio. 1990. Percutaneous absorption of [14C]DDT and [14C]benzo(a)pyrene from soil. Fund. Appl. Toxicol. 15:510-516. Wester, R. C., and H. I. Maibach. 1996. Percutaneous absorption of hazardous substances from soil and water. Pp. 325-335 In: Dermatology, 5th Edition. Marzulli and Maibach (eds.). Wong, C. S., P. D. Capel, and L. H. Nowell. 2001. National-scale, field-based evaluation of the biota-sediment accumulation factor model. Environ. Sci. Technol. 35:1709-1715.