5
Moving Forward with Bioavailability in Decision-Making

Soils and sediments are the ultimate sink for many persistent organic and inorganic contaminants and have the potential to impact human and environmental health for a long time. Remediation and management of contaminated soils and sediments is often technically difficult and can be very expensive when there are large volumes of contaminated material. To more rationally allocate limited environmental management and remediation resources, there is a need to improve risk assessment by including more explicit consideration of bioavailability processes.

Inadequate scientific understanding has hampered the widespread consideration of bioavailability processes in remedial decision making to date. Uncertainty in the relationship between total contaminant concentrations in soils and sediments and risk has often resulted in a conservative approach to exposure assessment in which the total contaminant present in a particular material is assumed to be available for uptake by possible receptors. Other assumptions (of relative bioavailability being less than 100 percent or about relevant exposure pathways for ecological receptors) may have led to situations where risk was underestimated. All assumptions have important implications with respect to the amount of material that must be treated and to the selection of a technology capable of reaching treatment goals. Explicitly incorporating bioavailability routinely and rigorously into the risk assessment process would offer the possibility of demonstrating in some cases that only a fraction of a contaminant’s total mass contained in a soil or sediment actually has the potential to enter potential receptors. In other cases, better understanding of bioavailability processes can lead to



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 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications 5 Moving Forward with Bioavailability in Decision-Making Soils and sediments are the ultimate sink for many persistent organic and inorganic contaminants and have the potential to impact human and environmental health for a long time. Remediation and management of contaminated soils and sediments is often technically difficult and can be very expensive when there are large volumes of contaminated material. To more rationally allocate limited environmental management and remediation resources, there is a need to improve risk assessment by including more explicit consideration of bioavailability processes. Inadequate scientific understanding has hampered the widespread consideration of bioavailability processes in remedial decision making to date. Uncertainty in the relationship between total contaminant concentrations in soils and sediments and risk has often resulted in a conservative approach to exposure assessment in which the total contaminant present in a particular material is assumed to be available for uptake by possible receptors. Other assumptions (of relative bioavailability being less than 100 percent or about relevant exposure pathways for ecological receptors) may have led to situations where risk was underestimated. All assumptions have important implications with respect to the amount of material that must be treated and to the selection of a technology capable of reaching treatment goals. Explicitly incorporating bioavailability routinely and rigorously into the risk assessment process would offer the possibility of demonstrating in some cases that only a fraction of a contaminant’s total mass contained in a soil or sediment actually has the potential to enter potential receptors. In other cases, better understanding of bioavailability processes can lead to

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications more protective risk estimates, for example by refining a default relative bioavailability factor or identifying an important exposure pathway that was overlooked. Consideration of bioavailability processes could also be used to improve evaluation of remediation technologies. For example, dredging is a common remediation technology applied to contaminated sediments. In certain cases, natural burial processes have isolated the contamination to the extent that contact between sensitive species and the contaminated matrix is not possible (a situation that can be evaluated through the use of coring studies). Dredging may promote the release of contaminants to the water column, possibly resulting in an increase in mobility and hence bioavailability. In such cases, decision-makers need to consider whether an increase in bioavailability is consistent with the goals of site remediation. This chapter examines the developments needed in both science and decision-making approaches to promote better consideration of bioavailability processes in remediation and management of contaminated soil and sediment. The chapter examines limitations in our current understanding of bioavailability processes and their implications and what can be done to overcome these limitations. Scenarios in which consideration of bioavailability processes has the greatest potential to impact decision-making are identified, with the hope of focusing science and technology development efforts on these situations. The chapter concludes by recommending specific steps that can be taken to move forward with consideration of bioavailability processes at individual sites, in regulation and decision-making, and in scientific research. CURRENT LIMITS OF KNOWLEDGE As demonstrated in Chapter 3, bioavailability of contaminants in soils and sediments to human and ecological receptors is governed by a wide range of physical, chemical, and biological processes. Qualitative and quantitative understanding of some of these processes is substantial, but for other processes there is much to be learned. For example, there is much about contaminant–solid interactions that is only weakly understood. While conceptual models exist for many kinds of contaminant–solid interactions, tools to test these models are often inadequate or nonexistent. As a result, there is significant uncertainty in the models used to describe contaminant–solid interactions and in the parameter values employed in these models. As some description of contaminant–solid interaction will usually be needed for assessment of risk associated with contaminated soils and sediments, the model and parameter uncertainty will transfer directly to the exposure assessment in a risk analysis. All models and parameters used in exposure assessment have a certain degree of uncertainty associated with them, including those used in bioavailability process considerations. In screening-level assessments for contaminated soils and sediments, this uncertainty is often recognized and dealt with by assuming

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications that all the contaminant mass is readily available. In practical terms this means that no special adjustments are made to account for bioavailability processes when exposures are estimated. If explicit consideration of bioavailability processes is to become more frequent, the uncertainties inherent in their measurements must be addressed and reduced, if possible. Some general sources of uncertainty associated with bioavailability processes include: a lack of knowledge about how physical, chemical, and biological processes acting at the level of soil and sediment particles influence the binding and release of chemicals; variations in soil and sediment characteristics at various spatial scales; a lack of knowledge about how biota modify bioavailability of chemicals in soils and sediments that come into contact with external membranes (e.g., skin) or that are taken into the body (e.g., digestive systems), and whether information obtained for one species is representative of another; variations in chemical form or properties (e.g., redox state of metals or diffusive rates for organics); physical, chemical, or biological changes that might, at some point in the future, change the bioavailability of a chemical. Given these multiple sources of uncertainty, regulatory agencies have been cautious about moving away from default assumptions concerning bioavailability processes in risk estimates. It is not clear whether there is too much uncertainty associated with bioavailability tools for regulatory agencies to feel comfortable about more explicitly incorporating their results into exposure estimates. Input received by the committee indicates that there is disagreement over this issue. An individual who has a strong precautionary stance might argue against replacing certain default assumptions (e.g., of 100 percent availability) to account for site-specific bioavailability processes. On the other hand, someone who sees large trade-offs among alternatives that hinge on bioavailability considerations would likely support their inclusion in specific situations. Risk assessment practitioners well versed in uncertainty and probabilistic analyses might argue that the uncertainties could be identified and taken into account, thereby providing more complete information to the risk manager. Explicit incorporation of information on bioavailability processes has occurred in ecological and human health risk assessments for particular types of problems and chemicals where the uncertainty has been relatively low due to extensive testing of certain contaminants and processes. Examples include exposure of humans to lead in soils (oral), and to polychlorinated biphenyls (PCBs) in soils (dermal); leaching of soil contaminants to groundwater; exposure of benthic invertebrates to non-polar organic chemicals (e.g., polyaromatic hydrocarbons or PAHs) in sediments; and site-specific determinations of bioavailability via up-

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications take studies from soils or sediments to benthic invertebrates, sediment invertebrates, plants, and wildlife (see Table 2-3). Clearly, the inclusion of site-specific bioavailability information has been judged to be important in a number of cases, and uncertainties were addressed at a level appropriate to risk-based decision making. There have been many other cases, however, in which the level of uncertainty has been judged to be too high for bioavailability measurements to replace default assumptions. A prominent example is the case of the Times Beach, Missouri, Superfund site, where large amounts of dioxin-contaminated soil were excavated and incinerated (see Box 5-1). There was a limited, generic consideration of bioavailability processes in determining the dioxin action levels for soil to be excavated and treated. However, site-specific assessments of bioavailability processes were not used to guide remediation decision-making, at least in part due to uncertainty in the bioavailability process measurements. WHY THESE LIMITATIONS AND UNCERTAINTIES MATTER The limitations in our understanding of bioavailability processes and the large uncertainties associated with their measurement have important ramifications for site management. The most obvious is that a lack of knowledge may inadvertently support poor decisions regarding exposure assessment, which has implications for how much contamination should be cleaned up and at what cost. For example, site managers working with incomplete information may be inclined to excavate a contaminated site even if the contaminants are not bioavailable. This could present myriad problems, including increasing the bioavailability of the material and potentially the risk to other receptors, such as wildlife, that were not originally the receptors of concern. Our lack of understanding of bioavailability processes also has important implications for the remediation of hazardous waste in situ. With regard to remedy selection, a large number of treatment and containment technologies rely on biological processes that are partially controlled by bioavailability, such as the transformation reactions of microorganisms. Without a better understanding of bioavailability processes, it is difficult to choose among technologies or to know if they are effective. (Although many might agree with the conceptual model of bioavailability processes outlined in Figure 1-1, there is little consensus on how to identify and quantify the dominant processes relevant for a specific situation.) This is aggravated by the plethora of different bioavailability tools and measurements used, many of which do not actually test a relevant endpoint. Additionally, site managers may not be cognizant of when treatment technologies unintentionally affect bioavailability. Especially for technologies that have yet to be fully tested, like phytoremediation, there may be unanticipated “side effects” that result in undesirable changes in bioavailability to certain receptors. Finally, in the last several years, approaches using simple tests to assess bioavailability at hazardous

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications BOX 5-1 Times Beach Superfund Site: How Uncertainty Influenced Decision-Making about Bioavailability The remediation of Times Beach, Missouri, has been one of the largest Superfund projects in the nation after hazardous levels of dioxin were found throughout eight square miles of the small agricultural and residential town in 1982. Waste oil used to spray the roads for dust control in 1972 and 1973 contained dioxin (2,3,7,8 TCDD). After the waste oil application to the roads, animal mortality and human illness were observed. Almost immediately a toxic chemical in the oil treatment was suspected. The U.S. Environmental Protection Agency (EPA) tested soil samples from the town’s unpaved roads and right of ways, revealing dioxin levels ranging from 1 ppb to 127 ppb. The entire Times Beach site is situated within the floodplain of the Meramec River. Shortly after the discovery of dioxin, the Meramec River flooded the city, which spread the contamination. Times Beach was evacuated in February of 1983, and the federal government used $33 million from Superfund to buy the dioxin site and relocate the residents. The Centers for Disease Control and Prevention (CDC) evaluated the health implications of dioxins in the soil at the site (Kimbrough et al., 1984)—one of the earliest examples of explicitly including bioavailability information in an assessment. CDC investigators noted that “regarding dermal absorption, there is some evidence that TCDD binds to soil and would not be as easily available for absorption.” They considered three routes of exposure: dermal contact, incidental ingestion, and inhalation. In their estimates of exposure, Kimbrough et al. (1984) used the available literature values for relative bioavailability—1 percent to estimate dermal uptake and 30 percent to estimate absorption in the digestive system. Bioavailability was not included in the estimate of inhaled dose. Interestingly, in discussing the implications of their assessment for management of the soils at Times Beach, Kimbrough et al. (1984) state: “The precise bioavailability of TCDD from soil is not known. Such bioavailability may vary with the soil type. It has been recently established that TCDD-contaminated soil from Missouri is toxic to guinea pigs and rats, if given orally. It was estimated that the [relative] bioavailability was waste sites have become popular. Some of these approaches do not seek to better understand underlying bioavailability processes such that their widespread application may become problematic. Technologies Developed with the Intent to Decrease Bioavailability A number of treatment technologies have been reported that “decrease bioavailability”—that is, treatment that impedes transfer of a contaminant from the soil or sediment matrix to a living organism. Although institutional controls and containment remedies would theoretically be encompassed by this definition, this discussion focuses on in situ treatments that aim to either (1) remove the labile fraction of contaminants (e.g., by microbial or plant mineralization), (2) convert

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications 30–50 percent or more [compared to ingestion of TCDD in corn oil] (McConnell et al., 1984.)” As a result of the Kimbrough et al. (1984) study, CDC recommended a 1 ppb TCDD action level for residential areas and 20 ppb level for industrial areas. Site-specific assessments of relative bioavailability performed later (Umbreit et al., 1986a, b, 1987, 1988a, b; Shu et al., 1988), which would probably have changed the cleanup goals by a factor of about 2, were not used to guide remedial actions because: There apparently was little communication early in the decision-making process concerning the role that site-specific bioavailability information might have in guiding remediation. Regulatory agencies prefer to err on the side of health protectiveness. Given the uncertainties in the bioavailability information derived from the Umbreit et al. studies, the regulators chose not to apply a bioavailability adjustment in the risk assessment. The Umbreit studies were controversial because the controls used conditions that were dissimilar from the critical toxicity study from which the reference dose for TCDD is derived (Kociba et al., 1978). There was a lack of an accepted framework for incorporating the measurements of site-specific bioavailability processes into risk estimates. Roads and affected areas at Times Beach containing dioxin levels over 1 ppb were excavated to a depth of four feet of contaminated soil and stored. A 50,000 cubic yard concrete tank with a flood-proof covering was used as a storage facility for the excavated soil, which was subsequently treated via incineration. Contaminated soil from 26 other dioxin sites was also brought to Times Beach to be incinerated—a fifteen-month process resulting in 265,000 tons of waste material. The incinerators ceased operation in June 1997, and the site was declared fully recovered. the labile fraction to a stable fraction (e.g., by the precipitation of metals), or (3) increase the mass transfer resistance of pollutants (e.g., by modifying the physical structure of the geosorbent). Examples of such technologies include biostabilization (the use of bioremediation to reduce contaminant mobility and toxicity of contaminated soils and sediments); sediment capping (reducing the ability of a bottom dwelling organism to get to the contaminant, and increasing mass transfer distance); vitrification or solidification (decreasing contaminant mobility by vastly increasing mass transfer resistance out of the solid matrix); and chemical alteration (e.g., converting a compound to a low solubility redox state via an amendment). Biostabilization relies on the microbial degradation of contaminants serving as carbon or energy sources or as electron acceptors. It consists of an initial active

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications and often engineered bioremediation phase (that may last months) to remove or transform those compounds that are more bioavailable, followed by a passive bioremediation phase (lasting years) to ensure that there is no chemical migration away from the actively treated material. The concept of the second phase is that intrinsic biodegradation rates equal or exceed the rate at which low solubility compounds become available. Box 5-2 discusses the characteristic desorption curves for PAH-contaminated solids, which are a frequent target of biostabilization efforts. One limitation of biostabilization is that the organic compounds may not meet threshold concentrations needed to drive microbial metabolism. Threshold concentrations of compounds are thought to play a role in energy maintenance and microbial enzyme induction (e.g., Schmidt et al., 1985) and they are experimentally manifested as residual concentrations of pollutants in various biodegradation tests (Bosma et al., 1996; Tros et al., 1996a, b). For a given contaminant, the value of the threshold concentration is determined by the efficiency of microbial metabolism (e.g., the relative values of specific uptake rates versus maintenance coefficients). Thus, thresholds can be affected by external mass transfer limitations, which often occur with aged pollutants in soils and sediments (Bosma et al., 1997). The existence of threshold values may be irrelevant when these values are far below concentrations that present risk. However, when these microbial threshold concentrations are above values deemed to represent a risk, biostabilization may not be a suitable remedial technology. Other remediation approaches use isolation to reduce bioavailability by employing capping or burial to remove access of a contaminant to the biosphere. In a physically active waterbody, however, capping will not permanently remove contaminants from the bioaccessible or bioavailable location if the sedimentary environment is erosional. To evaluate the potential success of isolation techniques, it is important to take sediment cores and evaluate their sedimentation regimes. Several technologies to reduce bioavailability of metals in soil, sediment, or other contaminated matrices rely on amending the solid phase to alter the redox or acid–base status of metals or sulfur species (NRC, 1997a). Certain metals (e.g., chromium or uranium) may have highly unavailable (low solubility) species depending on redox conditions, which can be imposed by specific technologies. This has been demonstrated at the Department of Energy (DOE) Hanford Site in Washington, where groundwater hexavalent chromium levels have been reduced from 0.060 mg/L to below detection limits (0.008 mg/L). The zone of reduction was created by injecting reagents that reduce iron naturally present in the aquifer sediments from Fe(III) to surface-bound and structural Fe(II) species, which concomitantly reduces the hexavalent chromium. Other metals may not have such speciation, but they can be precipitated as phosphates or sulfides, and hence the reduction of oxidized sulfur species can reduce their bioavailability (Benner et al., 1999). This strategy is exemplified by the case study presented in Box 5-3,

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications BOX 5-2 Biostabilization of PAH-laden Soils or Sediments Biostabilization generally refers to the situation where biological processes alone—intrinsic or stimulated—are deemed sufficient to reduce the risk associated with contaminants in soils and sediments. Although an awkward term, stabilization alludes to the fact that the labile fractions of the total contaminant are being reduced in size. This remedy has been suggested extensively for soils and sediments contaminated with PAHs, many of which have been documented to undergo microbial mineralization under various redox conditions (Kanaly and Harayama, 2000). Documenting the success of biostabilization typically requires demonstrating not only a decrease in total contaminant mass but also a decrease in the labile fraction of the contaminant pool between the onset and the end of the examined stabilization period. Popular tests to make these measurements examine the “rate of release” of contaminants using infinite sorption sinks or different extraction solvents (e.g., Cornelissen et al., 1998; Hawthorne and Grabanski, 2000), or they use toxicological endpoints (Loehr and Webster, 1997). Typical results for desorption data are shown in Figure 5-1 for two compounds in contaminated sediment. Detailed studies that directly inspect the soil and sediment phase to determine the stabilization mechanism are rare, and the actual contribution of microbial metabolic activity is only sporadically demonstrated (Ringelberg et al., 2001). FIGURE 5-1 (A) Desorption of fluoranthene, a compound amenable to microbial degradation, before (triangles) and after (squares) bioremediation. Total fluoranthene concentration dropped from approximately 170 mg/kg to 20 mg/kg over four months of active bioremediation. The shape of the desorption curves are very different before and after bioremediation. The rapidly desorbing fraction (obtained from curve fits shown in figure) dropped from 67 percent ± 3 to 10 percent ± 4 after bioremediation. This drop in rapidly desorbing fraction was observed for all the compounds that were biodegraded, suggesting a decrease in their labile fraction, and hence biostabilization. (B) Desorption results for the non-degraded compound benzo(ghi)perylene indicating very similar shapes of the desorption curves before and after the bioremediation. SOURCE: Reprinted, with permission, from Cornelissen et al. (1998). © (1998) American Chemical Society.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications BOX 5-3 Soil Amendments to Reduce Lead Bioavailability at the Joplin Superfund Site Joplin, Missouri, was included on the National Priorities List because of soil contamination from the smelting of locally mined lead (Pb) and zinc (Zn) ores. As part of the remedial action undertaken for the site, 2,600 homes in Joplin have had their soil replaced with clean material. In conjunction with this cleanup, a field site was established to test the ability of different in situ soil amendments to reduce the bioavailability of soil Pb to children. This project was undertaken by the Inplace Inactivation and Natural Ecosystem Restoration Team (IINERT) of EPA’s Remedial Technology Development Forum, whose stated mission was to identify in situ technologies that could chemically and physically inactivate hazardous metals in soils by reducing the metal’s solubility and bioavailability. This box focuses on human health because of the urban focus of the risk assessment. However, there are adjacent lots at the site where soil amendments are also being tested and where ecological receptors (including plants, herbivores, and insectivores) are the primary receptors of concern. Background Several lines of evidence suggested that soil amendments, including different sources of phosphorus, high iron materials, and biosolids compost, might be successful in reducing Pb availability in situ. The solubilities of different Pb species are known to vary in relation to the mineral form (Nriagu, 1984). In the presence of phosphorus, lead can form chloropyromorphite, which has a very low Ksp (10–84.4), such that the compound is likely to be stable under most soil and gastric systems. Thus, amendments that would promote formation of this mineral became the focus of research. Controlled environment studies demonstrated that it was possible to alter the mineral form of Pb in both pure and soil systems (Ma et al., 1993, 1994a, b). Field validation of these technologies was determined to be the appropriate next phase of research, for which hypotheses were developed. The initial phase of research focused on defining an appropriate animal surrogate to measure changes in bioavailability and on determining what extractions or in vitro tests can potentially substitute for animal feeding studies. Identifying the mechanisms that are responsible for the observed reduction in bioavailability and the appropriate tools to measure changes in speciation was also a goal. Animal surrogates and in vitro testing Initial results from the field site showed that additions of both H3PO4 and biosolids compost in situ are capable of reducing Pb bioavailability in juvenile swine, and in weanling and adult rats (Casteel et al., 2001; Maddaloni et al., 2001). However, although animal feeding studies have consistently shown reduced lead bioavailability as a function of treatment, the reductions are not consistent across groups or over time after treatment (see Casteel et al., 2001 for details). A second goal of the field study was to determine whether an in vitro extraction test could substitute for in vivo trials to assess reduction in Pb bioavailability. For the Joplin site soils, the in vitro test results at an extraction pH of 2.3 were comparable to the results from the swine studies (Ruby et al., 2001). Mineral Form The final goal of the field trial was to determine the mechanisms responsible for the observed reduction in bioavailability. Using X-ray adsorption spectroscopy (XAS) and

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications comparing field samples to known mineral forms, the formation of chloropyromorphite in the treatments that included phosphorus addition has been confirmed (Scheckel and Yang, 2001). In addition, the portion of total Pb present in this mineral phase increased over time. Mineral forms in the unamended soils have remained constant over time. There also seems to be a relationship between the observed decrease in bioavailability and the presence of pyromorphite [although the correlation is weak—r2 = 0.5546 (Ryan and Berti, 2001)]. For other treatments, results are even less clear. Although compost addition resulted in reduced bioavailability as measured by in vitro and in vivo (weanling rats) studies, XAS was not able to quantify the formation of a new mineral phase. A shift was observed from carbonate- and S-associated Pb in the control soils to what was identified as adsorbed Pb in the compost amended soils. Clearly, more information will be required before this shift can be accepted as the cause of the observed decrease in bioavailability. Conclusions On many levels, the preliminary research at the Joplin field site has been a success. It should be noted that this is the first time that feeding studies on animals have used treated soils. Thus, the methods are clearly a work in progress. All in vivo (human, pig, and rat) and in vitro studies (data unpublished) have shown that soil amendments are able to reduce the portion of total soil Pb that is bioavailable. In addition, it has been demonstrated that when P is added to the soil, the mineral form of Pb shifts, at least in part, to pyromorphite. This mineral shift appears to weakly correlate with the observed decrease in bioavailability. The stability of this mineral phase also suggests that the observed decrease in bioavailability will persist over time. During the limited sampling time since treatment addition, increasing pyromorphite concentrations have been observed for select treatment. However, this field site also illustrates some of the complexities involved in the measurement of bioavailability to assess risks posed by Pb in soil. Although all indices used in this study show decreases in bioavailability, they also show considerable variability. At this time, it is not clear if a single, appropriate index can be identified. The initial results from this field site indicate that, while it is possible to reduce the bioavailability of Pb in situ, it is not clear how to interpret or utilize these observed reductions in the regulatory arena. Plots at the Joplin Superfund Site being subjected to soil amendment in order to reduce metal bioavailability to residents.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications in which different soil amendments were tested for their ability to reduce the bioavailability of lead in soil to children. Both bioassays (feeding studies) and physicochemical tests (x-ray spectroscopy) were conducted to determine the effectiveness of the soil amendments. Environmentally Acceptable Endpoints. Of specific relevance in this discussion (particularly for biostabilization) is the increasing popularity of environmentally acceptable endpoints (EAE). The EAE concept is based on the observation that many organic contaminants become less “available” as they age within soil or as the soils undergo treatment, due to changes in the way soils and sediments encapsulate chemicals over time (Alexander, 1995). It has been proposed that this reduced availability should have an impact on cleanup levels and remediation goals and should be incorporated in site-specific risk assessment (Stroo et al., 2000). In some cases, this may involve modifying the default assumptions to reflect bioavailability limitations. This reduced availability, which has been described for organic contaminants by such mechanisms as sequestration and entrapment, has largely been inferred from the behavior of persistent hydrophobic compounds (mainly PAHs) in the field. After an initially rapid rate of chemical degradation, a period follows with little or no change in chemical concentrations. In the case where the considered chemicals are known to be biodegradable, the lack of continued decline—all other things remaining favorable for microbial activity—suggests that the chemicals themselves have largely become the limiting factor to microbial biodegradation, probably because of reduced availability. It is postulated, but rarely confirmed, that reduced availability to microorganisms relates to reduction in risk posed by the contaminants. Concomitant reductions in toxicity to other more relevant receptors has only occasionally been demonstrated (Salanitro et al., 1997; Olivera et al., 1998). Although plausible, the lack of availability of contaminants in soils or sediments to resident microorganisms does not suffice to characterize the suite of possible bioavailability processes. As an analogy, consider the fact that exchange of metals from sediments to pore water declines as the metal–sediment association ages (Schlekat et al., 2002). While the risk to water column species may decline with contaminant aging in sediments, there will not necessarily be a change in the risk to species whose food web is connected to ingestion of the sediments themselves. Hence, the evidence on which environmentally acceptable endpoints are based (microbial availability) may be insufficient, unless multiple exposure pathways and multiple receptors are considered. The challenge to all bioavailability assessment is to quantify the relevant bioavailability processes at work in a given situation, which requires an understanding of the importance of all exposure routes and receptors.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications ensure that all important bioavailability processes relevant to a particular site are studied. At the present time, individual bioavailability tools are frequently applied, producing information that is difficult to interpret in isolation, that is extrapolated to the field without adequate scientific justification, or that is not relevant to the key bioavailability processes at a site. Chapter 3 stressed the need for better understanding bioavailability processes at the field-scale. As a corollary, field tests are critical to determining whether proposed measurement techniques and models can accurately describe and predict bioavailability process performance at relevant scales. There has been limited investment in well-designed field experiments in which the complexity of environmental conditions can be accurately represented. Because these studies are expensive, priority should be given to selected important, recurring soil and sediment contamination problems. To provide more regulatory confidence, these studies could be conducted strictly in a pilot context before adopting the techniques widely. In addition to providing the most rigorous scientific test platform for a bioavailability measurement or modeling tool, field testing also enables realistic assessment of implementation costs and regulatory and public acceptance of the results obtained. Funding for Bioavailability Research Significant advances in understanding of bioavailability processes will have to come from new research. There are several potential avenues for funding of this research by federal agencies with research missions and responsibilities for managing environmental contamination. These agencies include the National Science Foundation (NSF), the National Institutes of Health, EPA, DOE, and the Department of Defense (DoD). NSF has funded a variety of studies of bioavailability processes, principally those related to interactions between environmental contaminants and media and the movement of chemicals in the environment. The National Institutes of Health, through the Superfund Basic Research Program administered by the National Institute for Environmental Health Sciences, funds a few bioavailability process studies, as does DoD, principally through the Strategic Environmental Research and Development Program. DOE is conducting research on methods for assessment of bioavailability processes as they affect remediation. Among federal agencies, the greatest commitment to bioavailability research has been made by EPA. Over the last decade, EPA has supported nearly 100 studies on bioavailability processes through its National Center for Environmental Research. The vast majority of these research projects have involved mobility of chemicals in the environment, uptake relevant to assessing ecological risks, and bioavailability processes that might affect bioremediation. Despite this research investment, progress in understanding these bioavailability processes is quite limited. For example, the number of bioavailability field trials or mechanis-

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications tic studies from EPA’s Superfund program is surprising low. Bioavailability studies at complex hazardous waste sites could be instrumental in designing improved risk management at those sites. Recently, EPA has evaluated research needs and prioritized research topics (EPA, 1999); bioavailability in human health risk assessment emerged as a high priority. For example, for soils the topic with the highest research priority was “Estimating Human Exposure and Delivered Dose.” This topic included focus points such as “evaluating the bioavailability of contaminants in various soil matrices,” “deriving dermal absorption factors for common soil contaminants,” and “developing biotransfer and bioaccumulation factors for contaminants to facilitate estimates of exposure via the food chain.” Despite this high priority, however, very little in the way of sponsored research on this topic is being funded by the agency. In fact, most of what is known about the potential oral bioavailability of contaminants from soil matrices, for example, comes not from agency-sponsored research projects, but rather from studies conducted by EPA Regions, states, and responsible parties on bioavailability of lead and arsenic from contaminated sites (e.g., EPA, 1996b; Casteel et al., 1997, 2001; Freeman et al., 1992, 1993, 1995; Roberts et al., 2002). These studies offer valuable observations regarding the absorption of contaminants from soils in specific situations, and some inferences on general behavior of absorption from soils might be gained from looking at these studies collectively. However, they are not an effective substitute for directed research because they have a different objective. The purpose of these studies was to obtain empirical measurements of relative bioavailability to support a human health risk assessment. For understandable reasons, this objective does not include an exploration of factors that might influence bioavailability processes, and therefore it is difficult to determine the extent to which these observations can be generalized or used to predict the results that might be obtained at different sites or under different conditions. Unless a greater commitment is made to fund bioavailability process studies from more of a research perspective, progress in developing information that can be utilized to advance human health risk assessments will be slow. OVERARCHING CONCLUSIONS AND RECOMMENDATIONS Bioavailability process considerations are not uniformly or widely embraced by scientists, regulators, or the public because of a lack of scientific and technical understanding. Explicit consideration of bioavailability processes and modeling in risk assessment would help to adjust cleanup goals by more accurately identifying that fraction of contaminant total mass that has the potential to enter receptors. Also, bioavailability process understanding would help guide the selection of appropriate remediation technologies. It is clear that more numerous validated tools and models are needed and that there should be reliance on an integrated suite of tools that lead to mechanistic understanding rather than on a single tool or

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications wholly empirical approaches. Ultimately, bioavailability process considerations are likely to make a difference where less than an order of magnitude adjustments in chemical concentrations are sought compared to proposed action levels, and where investments in assessment of bioavailability processes lead, over time, to familiarity with specific issues. Where site-specific consideration of bioavailability processes leads to more contaminated material remaining on site, long-term monitoring is needed to assess treatment performance, validate models, and demonstrate that contaminant bioavailability is not increasing over time. The following overarching conclusions and recommendations summarize our current understanding of processes that affect whether chemical contaminants in soils and sediments are bioavailable to humans, animals, microorganisms, and plants. Bioavailability processes are defined as the individual physical, chemical, and biological interactions that determine the exposure of plants and animals to chemicals associated with soils and sediments. First, in the broadest sense, bioavailability processes describe a chemical’s ability to interact with the biological world. Second, bioavailability processes are quantifiable through the use of multiple tools. Third, bioavailability processes incorporate a number of steps not all of which are applicable for all contaminants or all settings. Fourth, there are barriers that change exposure at each step. Thus, bioavailability processes modify the amount of chemical in soil or sediment that is actually absorbed and available to cause a biological response. Bioavailability processes are embedded within human health and ecological risk frameworks. The goal of bioavailability analysis is to reduce uncertainty in exposure estimates and thus improve the accuracy of the risk assessment. However, today “bioavailability” is commonly thought of in relation to one process only—absorption efficiency—such that a single “bioavailability” factor is used as an adjustment to applied dose. Most of the other bioavailability processes are hidden within the risk assessment process, and assumptions made about these processes are not clear. The knowledge base underlying many default assumptions about bioavailability processes is weak. Mechanistic understanding of bioavailability processes is ultimately needed to improve the scientific basis of risk assessment. Thus, tools for measuring bioavailability processes that further mechanistic understanding and promote predictive model development are preferred over conventional empirical approaches. In the short term, empirical approaches are useful in generating site-specific information—provided that their results are analyzed using a weight-of-evidence approach and with an understanding that they will be replaced with more mechanistic tools as they are developed. At any given site, a suite of tools will be necessary to describe bioavailability processes in soils or sediments.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications The potential for the consideration of bioavailability processes to influence risk-based decision-making is greatest when certain chemical, environmental, and regulatory factors align. Consideration of bioavailability processes is most likely to impact decision-making when the contaminant is, and is likely to remain, the risk driver; when the default assumptions made for a particular site are inappropriate; when significant change to remedial goals is likely (e.g., because large amounts of contaminated soil or sediment are involved); when conditions present at the site are unlikely to change substantially over time; and where regulatory and public acceptance is high. These factors should be evaluated before committing the resources needed for a detailed consideration of bioavailability processes. Moving bioavailability concepts further into the hazardous waste arena will require specific actions at individual sites, further scientific research on critical bioavailability processes, and large-scale, coordinated testing of bioavailability tools and techniques at pilot sites. At individual sites, assessment of bioavailability processes must be accompanied by uncertainty analysis, process-based long-term monitoring to ensure that present assessments of bioavailability remain accurate and acceptable, and community involvement beginning at the early stages of remediation planning. Although bioavailability is not a unique risk communication problem, experience has demonstrated that communities often have concerns about consideration of bioavailability processes during risk assessments. In order to demonstrate the utility of explicitly considering bioavailability processes and to test new models and tools, adaptive management should be applied to select pilot bioavailability test sites. Adaptive management applies findings from carefully monitored experiments to the adjustment of future management and policy decisions in light of changing conditions and new knowledge. REFERENCES Achtnich, C., E. Fernandes, J. M. Bollag, H. J. Knackmuss, and H. Lenke. 1999. Covalent binding of reduced metabolites of [N-15(3)]TNT to soil organic matter during a bioremediation process analyzed by 15N NMR spectroscopy. Environ. Sci. Technol. 33(24):4448-4456. Achtnich, C., H. Lenke, U. Klaus, M. Spiteller, and H. J. Knackmuss. 2000. Stability of immobilized TNT derivatives in soil as a function of nitro group reduction. Environ. Sci. Technol. 34(17):3698-3704. Alexander, M. 1995. How toxic are toxic chemicals in soil? Environ. Sci. Technol. 29:2713-2716. Ashford, N. A., and K. M. Rest. 1999. Public participation in contaminated communities. Cambridge, MA: Center for Technology, Policy, and Industrial Development, Massachusetts Institute of Technology. Benner, S. G., D. W. Blowes, W. D. Gould, R. B. Herbert, and C. J. Ptacek. 1999. Geochemistry of a permeable reactive barrier for metals and acid mine drainage. Environ. Sci. Technol. 33:2793-2799. Bernstein, J. 1991. Report from Aspen. The New Yorker. November 25. Pp. 121-136.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Blaylock, M. J., D. E. Salt, S. Dushenkov, O. Zakharova, C. Gussman, Y. Kapulnik, B. D. Ensley, and I. Raskin. 1997. Enhanced accumulation of Pb in Indian mustard by soil-applied chelating agents. Environ. Sci. Technol. 31:860-865. Bosma, T. N. P., E. M. W. Ballemans, N. K. Hoekstra, R. A. G. Welscher, J. Smeenk, G. Schraa, and A. J. B. Zehnder. 1996. Biotransformation of organics in soil columns and an infiltration area. Ground Water 34:49-56. Bosma, T. N. P., P. J. M. Middeldorp, G. Schraa, and A. J. B. Zehnder. 1997. Mass transfer limitation of biotransformation: quantifying bioavailability. Environ. Sci. Technol. 31:248-252. Bricker, T. J., J. Pichtel, H. J. Brown, and M. Simmons. 2001. Phytoextraction of Pb and Cd from a Superfund soil: effects of amendments and croppings. Journal of Environmental Science and Health, Part A—Toxic/Hazardous Substances & Environmental Engineering 36:1597-1610. Brown, S., R. Chaney, and J. S. Angle. 1997. Subsurface liming and metal movement in soils amended with lime-stabilized biosolids. J. Environ. Qual.26:724-732. Burmaster, D. E., and J. C. Wilson. 1998. Risk assessment for chemicals in the environment. In: The encyclopedia of biostatistics. New York: Wiley. Burton, G. A., P. M. Chapman, and E. P. Smith. 2002a. Weight-of-evidence approaches for assessing ecosystem impairment. Human and Ecological Risk Assessment 8:1657-1674. Burton, G. A., G. E. Batley, P. M. Chapman, V. E. Forbes, E. P. Smith, T. Reynoldson, C. E. Schlekat, P. J. den Besten, A. J. Bailer, A. S. Green, and R. L. Dwyer. 2002b. A weight of evidence framework for assessing sediment (or other) contamination: improving certainty in the decision-making process. Human and Ecological Risk Assessment 8(7):1675-1696. Casteel, S. W., R. P. Cowart, C. P. Weis, et al. 1997. Bioavailability of lead to juvenile swine dosed with soil from the Smuggler Mountain NPL site of Aspen, Colorado. Fund. Appl. Toxicol. 36:177–187. Casteel, S., T. J. Evans, J. Yang, and D. Moseby. 2001. Effects of treatments on soil-lead bioavailability in swine. Agronomy abstracts 2001. Annual meeting October 21-25, Charlotte, NC. Chapman, P. M., B. G. McDonald, and G. S. Lawrence. 2002. Weight of evidence issues and frameworks for sediment quality (and other) assessments. Human and Ecological Risk Assessment 8: 1489-1516. Chen, H., and T. Cutright. 2001. EDTA and HEDTA effects on Cd, Cr, and Ni uptake by Helianthus annuus. Chemosphere 45:21-28. Cornelissen, G., H. Rigterink, M. M. A. Ferdinandy, and P. C. M. VanNoort. 1998. Rapidly desorbing fractions of PAHs in contaminated sediments as a predictor of the extent of bioremediation. Environ. Sci. Technol. 32:966-970. Cotter-Howells, J., and I. Thornton. 1991. Sources and pathways of environmental lead to children in a Derbyshire mining village. Environ. Geochem. Health 13:127-135. Cullen, A. C., and H. C. Frey. 1999. Probabilistic techniques in exposure assessment. New York: Plenum Press. Dasappa, S. M., and R. C. Loehr. 1991. Toxicity reduction in contaminated soil bioremediation processes. Water Research 25:1121-1130. Daun, G., H. Lenke, M. Reuss, and H.-J. Knackmuss. 1998. Biological treatment of TNT-contaminated soil. 1. Anaerobic cometabolic reduction and interaction of TNT and metabolites with soil components. Environ. Sci. Technol. 32:1956-1963. Davis, A., M. V. Ruby, and P. D. Bergstrom. 1992. Bioavailability of arsenic and lead in soils from the Butte, Montana, mining district. Environ. Sci. Technol. 26:461-468. Deshpande, S., L. Wesson, D. Wade, D. A. Sabatini, and J. H. Harwell. 2000. DOWFAX surfactant components for enhancing contaminant solubilization. Water Research 34:1030-1036. Dwarakanath, V., K. Kostarelos, G. A. Pope, D. Shotts, and W. H. Wade. 1999. Anionic surfactant remediation of soil columns contaminated by nonaqueous phase liquids. Journal of Contaminant Hydrology 38:465-488.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Elder, M. 1997. The process of community involvement—a case study: the Bartlesville, Oklahoma, Lead Project. Toxicol. Ind. Health 13(2/3):395–400. Environmental Protection Agency (EPA). 1996a. Mercury study report to Congress. EPA 600-P-94-002Ab. External Review Draft. Washington, DC: EPA Office of Research and Development. EPA. 1996b. Bioavailability of arsenic and lead in environmental substrates: results of an oral dosing study of immature swine. EPA 910/R-96-002. Seattle, WA: EPA Region 10. EPA. 1997a. Policy for use of probabilistic analysis in risk assessment at the U.S. Environmental Protection Agency. Available at http://www.epa.gov/ncea/mcpolicy.htm. EPA. 1997b. Guiding principles for Monte Carlo Analysis. EPA/630/R-97/001. Washington, DC: EPA. EPA. 1999. Waste research strategy. EPA/600/R-98/154. Washington, DC: EPA Office of Research and Development. EPA. 2000a. Stressor identification guidance document. EPA/822/B-00/025. Washington, DC: EPA Office of Research and Development. EPA. 2000b. Review of an integrated approach to metals assessment in surface waters and sediments. EPA-SAB-EPEC-00-005. Washington, DC: EPA Science Advisory Board Ecological Processes and Effects Committee. EPA. 2001a. Methods for the collection, storage, and manipulation of sediments for chemical and toxicological analysis: technical manual. EPA-823-B-01-002. Washington, DC: EPA Office of Water. EPA. 2001b. Risk assessment guidance for Superfund. Volume 3 Part A: process for conducting probabilistic risk assessment. EPA-540-R-02-002. Washington, DC: EPA OSWER. EPA. 2001c. Whitman decides to dredge Hudson River. Press Release, U.S. Environmental Protection Agency, Washington, DC, August 1. Erickson, D. C., R. C. Loehr, and E. F. Neuhauser. 1993. PAH loss during bioremediation of manufactured gas plant site soils. Water Research 27:911-919. Francis, A. J., and C. J. Dodge. 1998. Remediation of soils and wastes contaminated with uranium and toxic metals. Environ. Sci. Technol. 32:3993-3998. 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. Toxicol. 19:388-398. Freeman, G. B., J. D. Johnson, J. M. Killinger, S. C. Liao, A. O. Davis, M. V. Ruby, R. L. Chaney, S. C. Lovre, and P. D. Bergstrom. 1993. Bioavailability of arsenic in soil impacted by smelter activities following oral administration in rabbits. Fund. Appl. Toxicol. 21:83-88. Freeman, G. B., R. A. Schoof, M. V. Ruby, A. O. Davis, J. A. Dill, S. C. Liao, C. A. Lapin, and P. D. Bergstrom. 1995. Bioavailability of arsenic in soil and house dust impacted by smelter activities following oral administration in cynomolgus monkeys. Fund. Appl. Toxicol. 28:215-222. Grcman, H., S. Velikonja-Bolta, D. Vodnik, B. Kos, and D. Lestan. 2001. EDTA enhanced heavy metal phytoextraction: metal accumulation, leaching and toxicity. Plant and Soil 235:105-114. Gunderson, L. H., C. S. Holling, and S. S. Light (eds.). 1995. Barriers and bridges to the renewal of ecosystems and institutions. New York: Columbia University Press. 245pp. Harwell, J. H., D. A. Sabatini, and R. C. Knox. 1999. Surfactants for ground water remediation. Colloids and Surfaces: Physicochemical and Engineering Aspects 151:255-268. Hawthorne, S. B., D. G. Poppendieck, C. B. Grabanski, and R. C. Loehr. 2001. PAH release during water desorption, supercritical carbon dioxide extraction, and field bioremediation. Environ. Sci. Technol. 35:4577-4583. Hawthorne, S. B., and C. B. Grabanski. 2000. Correlating selective supercritical fluid extraction with bioremediation behavior of PAHs in a field treatment plot. Environ. Sci. Technol. 34:4103-4110.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Huang, J. W. W., M. J. Blaylock, Y. Kapulnik, and B. D. Ensley. 1998. Phytoremediation of uranium contaminated soils: role of organic acids in triggering uranium hyperaccumulation in plants. Environ. Sci. Technol. 32:2004-2008. Kanaly, R. A., and S. Harayama. 2000. Biodegradation of high-molecular-weight polycyclic aromatic hydrocarbons by bacteria. J. Bacteriol. 182:2059-67. Kim, J. Y., C. Cohen, and M. L. Shuler. 2000. Use of amphiphilic polymer particles for in situ extraction of sorbed phenanthrene from a contaminated aquifer material. Environ. Sci. Technol. 34:4133-4139. Kimbrough, R. D., H. Falk, and P. Stehr. 1984. Health implications of 2,3,7,8-tetrachloro-dibenzodioxin (TCDD) contamination of residential soil. J. Toxicol. Environ. Health 14:47-93. Kociba, R. J., D. G. Keyes, J. E. Beyer, R. M. Carreon, C. E. Wade, D. A. Dittenber, R. P. Kalnins, L. E. Frauson, C. N. Park, S. D. Barnard, R. A. Hummel, and C. G. Humiston. 1978. Results of a two-year chronic toxicity and oncogenicity study of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. Toxicol. Appl. Pharmacol. 46:279-303. Labieniec, P. A., D. A. Dzombak, and R. L. Siegrist. 1997. Evaluation of uncertainty in a site-specific risk assessment. J. Environ. Eng. 123:234-243. Lee, K. 1993. Compass and gyroscope: integrating science and politics for the environment. Covelo, CA: Island Press. Lee, K. N. 1999. Appraising adaptive management. Conservation Ecology 3(2):3. Lenke, H., J. Warrelmann, G. Daun, K. Hund, U. Sieglen, U. Walter, and H.-J. Knackmuss. 1998. Biological treatment of TNT-contaminated soil. 2. Biologically induced immobilization of the contaminants and full-scale application. Environ. Sci. Technol. 32:1964-1971. Liu, Z., A. M. Jacobson, and R. G. Luthy. 1995. Biodegradation of naphthalene in aqueous nonionic surfactant systems. Appl. Environ. Microbiol. 61(1):145-151. Loehr, R., and M. T. Webster. 1997. Effect of treatment on contaminant availability, mobility, and toxicity. Pp. 138-386 In: Environmentally acceptable endpoints in soil: risk-based approach to contaminated site management based on availability of chemicals in soil . D. G. Linz et al. (eds.). Annapolis, MD: American Academy of Environmental Engineers. Löser, C., H. Seidel, P. Hoffmann, and A. Zehnsdorf. 1999. Bioavailability of hydrocarbons during microbial remediation of a sandy soil. Applied Microbiology and Biotechnology 51:105-111. Ma, Q. Y., S. J. Traina, T. J. Logan, and J. A Ryan. 1993. In situ lead immobilization by apatite. Environ. Sci. Technol. 27(9):1803-1810. Ma, Q. Y., S. J. Traina, T. J. Logan, and J. A Ryan. 1994a. Effects of NO3–, Cl–, F–, SO42–, and CO32– on Pb2+ immobilization by hydroxyapatite. Environ. Sci. Technol. 28(3):408-418. Ma, Q. Y., S. J. Traina, T. J. Logan, and J. A Ryan. 1994b. Effects of aqueous Al, Cd, Cu, Fe(II), Ni, and Zn on Pb immobilization by hydroxyapatite. Environ. Sci. Technol. 28(7):1219-1228. Maddaloni, M. A., N. Loiacono, C. Blum, S. Chilirud, and J. Graziano. 2001. Effects of treatments on soil-lead bioavailability (human studies). Agronomy abstracts 2001. Annual meeting October 21-25, Charlotte, NC. McConnell, E. E., G. W. Lucier, R. C. Rumbaugh, P. W. Albro, D. J. Harvan, J. R. Hass, and M. W. Harris. 1984. Dioxin in soil: bioavailability after ingestion by rats and guinea pigs. Science 223:1077–1079. McGrath, R., and I. Singleton. 2000. Pentachlorophenol transformation in soil: a toxicological assessment. Soil Biology and Biochemistry 32:1311-1314. Menzie, C. A., M. H. Henning, J. Cura, K. Finkelstein, J. Gentile, J. Maughan, D. Mitchell, S. Petron, B. Potocki, S. Svirsky, and P. Tyler. 1996. Special report of the Massachusetts weight-of-evidence workgroup: a weight-of-evidence approach for evaluating ecological risks. Human and Ecological Risk Assessment 2(2):277-304.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications 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. Morgan, M. G., and M. Henrion. 1990. Uncertainty: a guide to dealing with uncertainty in quantitative policy analysis. Cambridge, UK: Cambridge University Press. Morgan, M. G., B. Fischhoff, A. Bostrom, L. Lave, and C. J. Atman. 1992. Communicating risk to the public. Environ. Sci. Technol. 26(11):2049–2056. NACCHO. 1995. Don’t hazard a guess: addressing community health concerns at hazardous waste sites. Washington, DC: National Association of County & City Health Officials. National Research Council (NRC). 1996. Upstream: salmon and society in the Pacific Northwest. Washington, DC: National Academy Press. NRC. 1997a. Innovations in soil and ground water cleanup: from concept to commercialization. Washington, DC: National Academy Press. NRC. 1997b. Contaminated sediments in ports and waterways: cleanup strategies and technologies. Washington, DC: National Academy Press. NRC. 1999. Downstream: adaptive management of Glen Canyon Dam and the Colorado River ecosystem. Washington, DC: National Academy Press. NRC. 2001a. A risk management strategy for PCB-contaminated sediments. Washington, DC: National Academy Press. NRC. 2001b. Aquifer storage and recovery in the comprehensive Everglades restoration plan: a critique of the pilot projects and related plans for ASR in the Lake Okeechobee and Western Hillsboro areas. Washington, DC: National Academy Press. NRC. 2003. Environmental cleanup at Navy facilities: adaptive site management. Washington, DC: National Academy Press. NEPI. 2000. Assessing the bioavailability of metals in soil for use in human health risk assessment. Washington, DC: National Environmental Policy Institute. Nriagu, J. O. 1984. Phosphate minerals. J. O. Nriagu and P. Moore (eds.). New York: Springer Verlag. Olivera, F. L., R. C. Loehr, B. C. Coplin, H. Eby, and M. T. Webster. 1998. Prepared bed land treatment of soils containing diesel and crude oil hydrocarbons. J. Soil Cont. 7:657-674. Preuß, A., and P.-G. Rieger. 1995. Anaerobic transformation of 2,4,6-trinitrotoluene and other nitroaromatic compounds. In: Biodegradation of nitroaromatic compounds. J. C. Spain (ed.). New York and London: Plenum Press. Ramaswami, A., and R. G. Luthy. 1997. Measuring and modeling physicochemical limitations to bioavailability and biodegradation. Pp. 721-729 In: Manual of environmental microbiology. C. J. Hurst (ed.). Washington, DC: American Society for Microbiology. Riefler, R. G., and B. F. Smets. 2000. Enzymatic reduction kinetics of 2,4,6-trinitrotoluene and related nitroarenes: kinetics linked to one-electron redox potentials. Environ. Sci. Technol. 34 (18):3900-3906. Rijnaarts, H. H. M., A. Bachman, J. C. Jumelet, and A. J. B. Zehnder. 1990. Effect of desorption and intraparticle mass transfer on the aerobic mineralization of α-hexaclorocyclohexane in contaminated calcereous soil. Environ. Sci. Technol. 24:1349-1354. Ringelberg, D. B., J. W. Talley, E. J. Perkins, S. G. Tucker, R. G. Luthy, E. J. Bouwer, and H. L. Fredrickson. 2001. Succession of phenotypic, genotypic, and metabolic community characteristics during in vitro bioslurry treatment of polycyclic aromatic hydrocarbon-contaminated sediments. Appl. Environ. Microbiol. 67:1542-50. Roberts, S. M., W. R. Weimar, J. T. R. Vinson, J. W. Munson, and R. J. Bergeron. 2002. Measurement of arsenic bioavailability in soil using a primate model. Toxicological Sciences 67:303-310.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Romkens, P., L. Bouwman, J. Japenga, and C. Draaisma. 2002. Potentials and drawbacks of chelate-enhanced phytoremediation of soils. Environ. Pollut. 116:109-121. Ruby, M. V., A. Davis, J. H. Kempton, J. W. Drexler, and P. D. Bergstrom. 1992. Lead bioavailability—dissolution kinetics under simulated gastric conditions. Environ. Sci. Technol. 26:1242-1248. Ruby, M. V., S. L. Brown, K. G. Scheckel, and D. Allen. 2001. Effects of treatments on soil-lead bioavailability: implications of in vitro extraction testing. Agronomy abstracts 2001. Annual meeting October 21-25, Charlotte, NC. Ryan, J. A., and W. R. Berti. 2001. Field evaluation of in situ treatment to reduce soil-lead bioavailability. Agronomy abstracts 2001. Annual meeting October 21-25, Charlotte, NC. Salanitro, J. P, P. B. Dorn, M. H. Huesemann, K. O. Moore, I. A. Rhodes, L. M. Rice-Jackson, T. E. Vipond, M. M. Western, and H. L. Wisniewski. 1997. Crude oil hydrocarbon bioremediation and soil ecotoxicity assessment. Environ. Sci. Technol. 31:1769-1776. Sandman, P. 1986. Explaining environmental risk. Washington, DC: EPA. Sandman, P. 1996. Responding to community outrage: strategies for effective risk communication. Fairfax, VA: American Industrial Hygiene Association. Scheckel, K. G., and J. Yang. 2001. Effect of phosphorus treatment on lead mineralogy. Agronomy abstracts 2001. Annual meeting October 21-25, Charlotte, NC. Schlekat, C. E., B.-G. Lee, and S. N. Luoma. 2002. Dietary metals exposure and toxicity to aquatic organisms: implications for ecological risk assessment. Pp. 151-188 In: Coastal and estuarine risk assessment. M. C. Newman, M. H. Roberts Jr., and R. C. Hale (eds.). Boca Raton, FL: Lewis Publishers. Schmidt, S. K., M. Alexander, and M. L. Shuler. 1985. Predicting threshold concentrations of organic substrates for bacterial growth. J. Theor. Biol. 114:1-8. Shu, H., D. Paustenbach, F. J. Murray, et al. 1988. Bioavailability of soil-bound TCDD: oral bioavailability in the rat. Fund. Appl. Toxicol. 10:648–654. Sopper, W. E. 1993. Municipal sludge use in land restoration. Boca Raton, FL: Lewis Publishers. Steele, M. J., B. D. Beck, B. L. Murphy, and H. S. Strauss. 1990. Assessing the contribution from lead in mining wastes to blood. Reg. Toxicol. Pharm. 11:158-190. Stroo, H. F., R. Jensen, R. C. Loehr, D. V. Nakles, A. Fairbrother, and C. B. Liban. 2000. Environmentally acceptable endpoints for PAHs at a manufactured gas plant site. Environ. Sci. Technol. 34(18):3831-3836. Taylor, B., L. Kremsater, and R. Ellis. 1997. Adaptive management of forests in British Columbia. Victoria, BC: British Columbia Ministry of Forests, Forest Practices Branch. 93 pp. Tiehm, A., M. Stieber, P. Werner, and F. H. Frimmel. 1997. Surfactant-enhanced mobilization and biodegradation of polycyclic aromatic hydrocarbons in manufactured gas plant soil. Environ. Sci. Technol. 31:2570-2576. Tros, M., G. Schraa, and A. Zehnder. 1996a. Transformation of low concentrations of 3-chlorobenzoate by Pseudomonas sp. Strain B13: kinetics and residual concentrations. Appl. Environ. Microbiol. 62:437-442. Tros, M. E., T. N. Bosma, G. Schraa, and A. J. Zehnder. 1996b. Measurement of minimum substrate concentration (Smin) in a recycling fermentor and its prediction from the kinetic parameters of Pseudomonas strain B13 from batch and chemostat cultures. Appl. Environ. Microbiol. 62:3655-61. Umbreit, T. H., E. J. Hesse and M. A. Gallo. 1986a. Bioavailability of dioxin in soil form a 2,4,5-T manufacturing site. Science 232:497-499. Umbreit, T. H., E. J. Hesse and M. A. Gallo. 1986b. Comparative toxicity of TCDD contaminated soil from Times Beach, Missouri, and Newark, New Jersey. Chemosphere 15:2121-2124. Umbreit, T. H., E. J. Hesse and M. A. Gallo. 1987. Differential bioavailability of 2,3,7,8-tetrachlorodibenzo-p-dioxin from contaminated soils. ACS Symposium Series 10(338):131-139.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications Umbreit, T. H., E. J. Hesse and M. A. Gallo. 1988a. Reproductive studies of C57B/6 male mice treated with TCDD-contaminated soils from 2,4,5-trichlorophenoxyacetic acid manufacturing site. Arch. Environ. Contam. Toxicol. 17:145-150. Umbreit, T. H., E. J. Hesse and M. A. Gallo. 1988b. Bioavailability and cytochrome p-450 induction from 2,3,7,8-tetrachlorodibenzo-p-dioxin contaminated soils from Times Beach, Missouri, and Newark, New Jersey. Drug Chem. Toxicol. 11:405-418. Walters, C. 1997. Challenges in adaptive management of riparian and coastal ecosystems. Conservation Ecology 1(2):1. Weis, C. 2000. Presentation to the NRC Committee on Bioavailability of Contaminants in Soils and Sediments. September 13, 2000. White, J. C., and B. D. Kottler. 2002. Citrate-mediated increase in the uptake of weathered 2,2-bis(p-chlorophenyl)1,1-dichloroethylene residues by plants. Environ. Toxicol. Chem. 21:550-556. Yang, Y., D. Ratté, B. F. Smets, J. J. Pignatello, and D. Grasso. 2001. Mobilization of soil organic matter by metal ion complexing agents and implications for polycyclic aromatic hydrocarbon desorption. Chemosphere 43:1013-1021. Zehnder, A. J. B. 1991. Mikrobiologische Reinigung eines kontaminierten Bodens, Beitrag des 9. Dechema-Fachgespräches Umweltschutz, Frankfurt/Main. Pages 68-72. Zhenbin, L., J. A. Ryan, J. L. Chen, and S. R. Al-Abed. 2001. Adsorption of cadmium on biosolidsamended soils. J. Environ. Qual. 30:903-911.

OCR for page 356
Bioavailability of Contaminants in Soils and Sediments: Processes, Tools, and Applications This page in the original is blank.