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Approaches to Ecological Risk Characterization and Management: Selecting the Right Tools for the Job

Posted on: Thursday, 20 May 2004, 06:00 CDT

ABSTRACT

Chemical-specific hazard quotient (HQ) risk characterization in ecological risk assessment (ERA) can be a value-added tool for risk management decision-making at chemical release sites, when applied appropriately. However, there is little consensus regarding how HQ results can be used for risk management decision-making at the population, community, and ecosystem levels. Furthermore, stakeholders are reluctant to consider alternatives to HQ results for risk management decisions. Chemical-specific HQs risk characterization should be viewed as only one of several approaches (i.e., tools) for addressing ecological issues; and in many situations, other quantitative and qualitative approaches will likely result in superior risk management decisions. The purpose of this paper is to address fundamental issues and limitations associated with chemical-specific HQ risk characterization in ERA, to identify when it may be appropriate, to explore alternatives that are currently available, and to identify areas that could be developed for the future. Several alternatives (i.e., compensatory restoration, performance-based ecological monitoring, ecological significance criteria, net environmental benefit analysis), including their limitations, that can supplement, augment, or substitute for HQs in ERA are presented. In addition, areas of research (i.e., wildlife habitat assessment/landscape ecology/ population biology, and field validated risk-based screening levels) that could yield new tools are discussed.

Key Words: ecological risk assessment, ecological significance criteria, hazard quotient.

INTRODUCTION

The use of the ecological risk assessment (ERA) process in environmental management programs is now common, as evidenced by the wide variety of ERA guidance documents available from diverse sources (e.g., ADEC 2000; ASTM 2002; CaIEPA 1998; EC 1997; LDEQ 1999; USACOE 1996a; USEPA 1997a, 1998, 2001a, 2001b; NZME 1999). More than 100 ERA guidance documents, procedural guidelines, or other resources-spanning just the last decade-have been published (see the annotated compilations by Sorensen 1994; Sorensen and Margolin 1998). The vast majority of these follow the U.S. Environmental Protection Agency (USEPA) general ERA framework (USEPA 1992).

A common theme and element of these resources is that the best ERAs are the ones that are appropriate for the specific risk management needs of the individual site. The risk management needs vary considerably between sites, but the overall management goals generally involve protection of populations, communities, and ecosystems (USEPA 1997b, 1999a, 200Ic). Special attention must also be given to a variety of sensitive species and habitats, where management attention is given to individuals of threatened and endangered species and their habitats because a compromised population is less capable of tolerating the loss of individuals.

Chemical-specific hazard quotient (HQ) risk characterization is essentially synonymous with the initial steps of an ERA at chemical release sites-while less attention is given to actual ecological characteristics of a site until later in the ERA process (if at all) (e.g., USEPA 1997a, 200Oa). In part, this lack of attention to the ecological components of a site has occurred because the Resource Conservation and Recovery Act (RCRA) and the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA), and other environmental protection programs require chemical measures of site conditions (i.e., concentration data on abiotic media), but do not require ecological measures of site conditions (e.g., diversity and density of potentially exposed wildlife). Also, the historic preference for basing risk management decisions on human health risk, which is exclusively based on chemical toxicity and exposure, has further encouraged this trend. As a result, data available at risk management decision points are often exclusively chemical analytical data while comparatively little information is available regarding the ecological resources at a site-the very resources the process is striving to protect. This is only further encouraged because remediation "goals" are often defined by a reduction in chemical concentration (s) rather than an improvement or restoration of ecological resources. The unfortunate dilemma is that remediation that achieves these "goals" (e.g., sediment dredging, soil removal, clay caps) can irrevocably alter or destroy the very ecosystem it was intended to "protect." Often the regulatory preference for a permanent solution leads to aggressive remedial action (causing more harm than benefit) rather than allowing the impacted area to recover naturally. In fact, when "the goal is expressed as restoration of an ecosystem, as measured by biological indicators, the conventional process cannot work" (Barnthouse 2002).

The purpose of this paper is to address issues associated with the use of chemical-specific HQ risk characterization in ERA, explore the alternatives that are available, and to identify areas that could be researched and developed for the future. Chemical- specific HQ risk characterization can be a value-added tool for risk management decision-making at chemical release sites, when applied appropriately. However, there are significant shortcomings. For example, there is little consensus regarding how HQ results are used for risk management decision-making at the population, community, and ecosystem levels. In addition, the chemical-specific HQ approach is replete with uncertainties. Moreover, chemical-specific HQs are simplistic numbers that are easy to communicate, but are often given undue merit because of this apparent precision. Finally, there is more than a decade of ERA experience to draw from, yet little effort is being focused on the large body of chemical and biological evidence available to evaluate the appropriate usefulness and/or application of chemical-specific hazard quotients.

Chemical-specific HQ risk characterization should be viewed as only one of several available approaches; and in many situations, other quantitative and qualitative approaches are likely to result in superior risk management decisions in terms of overall protectiveness to ecological resources. With all of the known limitations and contradictions in evidence, one is led to question why there is often resistance to accepting alternatives to chemical- specific HQs. One has to assume that the lack of awareness of viable alternatives to chemical-specific HQs is a significant technical and regulatory hurdle that must be overcome to expand the ecological risk assessor's and risk manager's toolbox. Several alternatives to chemical-specific HQ risk characterization are identified in this paper: compensatory restoration; performance-based ecological monitoring; ecological significance criteria; and net environmental benefit analysis. While these alternatives are not appropriate for every situation, their use to supplement, augment, or substitute for the standard chemical-specific approach will result in more flexibility for risk assessors and risk managers, thereby leading to more appropriate and superior decision-making. Research to better use wildlife habitat assessment, landscape ecology and population biology and to develop field-validated ecological risk-based screening levels offer hope for improving the ecological relevancy of ERAs.

CHEMICAL-SPECIFIC HQ RISK CHARACTERIZATION IN ERA AND ITS LIMITATIONS

Chemical-specific HQ risk characterization in ERA, as described in many of the guidance and procedural documents already cited, can be deterministic or probabilistic. Deterministic models are based on point estimates for a number of exposure and toxicity input variables, resulting in a single estimate of risk. The deterministic risk estimate is most often referred to as a HQ but is also called an ecological effects quotient, risk quotient, and toxicity quotient. The HQ is a ratio of an exposure estimate to a toxicity reference value or benchmark. Examples include simple HQs generated by the ratio of media chemical concentrations to media-specific ecotoxicity screening values (e.g., the chemical concentration of arsenic in soil divided by the ecological soil screening-level [Eco- SSL] for arsenic in soil) (USEPA 200Oa,b). HQs are also generated by food web ingestion dose modeling that is based on simplistic (and often intentionally conservative) assumptions of dietary composition, bioavailability, ingestion rates, body weight, exposure durations, and site exposure frequency. The estimated dose (exposure) is then divided by a laboratory toxicity reference value to yield the HQ (Sample el al. 1996; USEPA 1997a, 200Oa). Probabilistic models are based on a similar premise, with ranges of exposure and toxicity variables leading to ranges of possible HQs. Probabilistic models give more insight into uncertainty and variability, but suffer some of the many obstacles and limitations as the deterministic model (discussed below).

An HQ less than or equal to a value of 1 indicates that adverse impacts to ecological receptors are considered unlikely (USEPA 1997a, 200Ia). An HQ greater than 1 is an indication that "further assessmen\t" may be necessary to evaluate the potential for adverse impacts to ecological receptors, particularly at the early stages of an ERA. However, there is no clear guidance for interpreting HQs that exceed a value of 1, except that this point of departure indicates that adverse effects of some kind potentially have occurred or may occur in the future. It is also generally accepted that as HQs increase in magnitude, the potential for adverse effects (i.e., risk) increases in magnitude as well. Part of the reason for this ambiguity in interpreting results is because the HQ does not incorporate a dose-response relationship, thus the magnitude of the HQ cannot be reliably used to interpret the magnitude of the toxic response (Pastorok et al. 2002).

"Further assessment" often involves additional data collection to reduce uncertainties and may involve multiple iterations of the deterministic model or may involve probabilistic modeling (e.g., ORDEQ1999; USEPA 1997a, 1998). Others suggest that probabilistic modeling should begin in the early stages (in lieu of deterministic HQs), and focused data collection and additional probabilistic modeling should be conducted in subsequent stages (Warren-Hicks and Moore 1998). Regardless of whether a deterministic or probabilistic method is used, further assessment can and should (although it rarely does) involve some level of ecological assessment as well. In cases where ecological assessment does occur, it is often focused on gathering chemical-specific data on dietary prey (e.g., body burden data) for use in generating more HQs instead of verifying effects (e.g., local population density; nesting success) to the species of interest (Tannenbaum et al. 2003). Therefore, the approach remains tied to the chemical data and the HQ calculations.

Right (and Wrong) Occasions for Chemical-Specific HQs in ERA

It is much easier to find guidance on how to perform chemical- specific risk characterization than it is to find guidance regarding when such an approach is (or is not) appropriate. Although a placeholder for this "appropriateness" step may exist (e.g., the Problem Formulation in USEPA 1997a), there is a fundamental lack of consistency in determining what constitutes an ecologically significant and complete exposure pathway. As a result, any potentially complete exposure pathway is often regarded as ecologically significant. Another problem is that the evaluation of whether risk estimates are "ecologically significant" occurs by design after at least some (if not all) of the chemical-specific risk characterization has taken place instead of in the Problem Formulation step during development of the conceptual site model (e.g., USEPA 1994, 1997a, 200Oa, 200Ib). These very significance issues provide insight as to whether chemical-specific risk characterization (or even an ERA) provides meaningful information for risk management decision-making. Efforts that have been developed to provide a more consistent methodology for addressing these ecological significance issues at the beginning of the ERA process are discussed later as one of the potential alternatives (ecological significance criteria) to supplement, augment, or substitute for the chemical-specific risk characterization approach.

Chemical-specific risk characterization has been demonstrated to be a value-added tool for risk management decision-making at chemical release sites, when applied appropriately. For example, chemical specific risk characterization can provide useful information for sites where releases have occurred in sensitive ecological habitats, such as aquatic systems, and chemical specific risk characterization can potentially help identify which chemicals may be related to observed ecological effects on sites of large spatial extent where multiple lines of evidence have demonstrated that a problem does in fact exist. Conversely, the chemical- specific risk characterization approach may not provide useful information at sites where an acceptable remediation and/or restoration alternative has already been identified (e.g., situations encountered routinely with previous ERA results to draw from). These and other examples of appropriate and inappropriate uses of chemical-specific risk characterizations are identified in Table 1. In order for chemical-specific risk characterization to be a value-added tool, it is reasonable to suggest that all of the following criteria must be met:

* Chemical-specific risk characterization results provide meaningful information for risk management decision-making, such that the same decision could not be identified in some more cost effective or efficient manner.

* A determination has been made that non-chemical Stressors are less important risk drivers for the ecological resources being assessed.

* The effort to generate ecotoxicological and exposure data is appropriate to the scale of the risk management decision.

* Chemical-specific risk characterization results are given an appropriate risk management weigh ting (i.e., not given undue merit), especially when there are available ecological data.

Obstacles and Limitations with Chemical-Specific HQ Risk Characterization and What to Do About Them

Some of the most significant limitations with the chemical- specific HQ approach are discussed below and ideas for addressing them are provided. It must be acknowledged that site-specific and more general studies needed to address these limitations may exceed the level of effort necessary to develop new approaches or augment the alternative approaches already available.

There is IAille, Consensus Regarding How HQs Can be Used for Risk Management Decision-Making at the Population, Community, and Ecosystem Levels

Chemical-specific risk characterizations are based on potential effects to an individual, and this approach completely fails to address ecological exposure and risk at the spatial scale of populations. The first step to dealing with this issue is already in progress-i.e., the acknowledgement that a significant problem exists, that the ERA community is struggling with the issue, and that a multi-stakeholder effort is needed to lead to a consensus- based solution. This is occurring in a variety of forums. For example, this publication and others like it (e.g., Kapustka 2003a; Tannenbaum 2002, 2003; Tannenbaum et al. 2003) are bringing the issues to the forefront. In addition, a Population-Level Workgroup was formed under the Society of Environmental Toxicology and Chemistry's (SETAC) Ecological Risk Assessment Advisory Group (ERAAC) after the SETAC 2001 conference debates on "Populations" (SETAC 2002). The objective of this ERAAG workgroup is to improve the use of population-level tools in ERA. Lastly, an American Society of Testing and Materials (ASTM) Landscape Ecology Symposium in 2003 has taken on a parallel task by dealing with wildlife habitat issues at the landscape scale. These papers and group efforts are discussed in greater detail as future research that could yield new tools that supplement, augment, or substitute for the chemical-specific risk characterization approach.

Table 1. Appropriate and inappropriate occasions for chemical- specific hazard quotients in ERAs.

There are Profound Uncertainties with the Wildlife and Ecotoxicological Data Upon Which to Base the Chemical-Specific Approach

This topic alone could be the focus of an entire manuscript and has been discussed in numerous publications. Two of the papers in this journal issue discuss a variety of profound limitations for establishing ecological soil screening levels for polycyclic aromatic hydrocarbons and total petroleum hydrocarbons (Kapustka 2003b; Efromyson et al. 2003). Gross uncertainties exist in many other aspects of the toxicology relied upon for conducting ERAs, as well (Newman 1998; Lovett Doust et al. 1993). Toxicity data are only available for a limited number of species (most of them laboratory test species) under a strictly defined set of test conditions that deviate from natural conditions. In current practice, more than 95 percent of the resources in toxicology are focused toward the study of single chemicals (Cassee et al. 1998), while ecological exposures rarely occur on a single chemical basis. Further, the majority of chemical-specific laboratory studies are focused toward single species (Sample et al. 1996; Newman 1998) with little relevance to the wide variety of species encountered in nature. For example, the laboratory white rat is bred to minimize differences between individuals, thereby diminishing the genetic variability that gives mammalian wildlife in nature some capability for adaptation and tolerance (Tannenbaum 2003). Simplistic extrapolations from laboratory species to wildlife species and from laboratory conditions to field conditions are not likely to be accurate, and are rarely, if ever, validated against natural conditions (Power 1996; Tannenbaum 2003). There is also little consistency and no quantitative methodology for the consideration of the diminished chemical bioavailability (and, thereby, diminished toxicity) even though this process is well documented (e.g., Alexander and Alexander 1999; Alexander 2000). Similarly, tolerance and adaptation are not considered directly (Millward and Klerks 2002; Grant 2002). Finally, there are relatively few studies that actually evaluate the effects of toxicity on predator-prey interactions, competition, or other aspects of ecologically relevant wildlife behaviors (Atchison et al. 1995).

To move forward, there either needs to be a coordinated effort to develop an ecotoxicological database or ecotoxicity-related methodologies that effectively deal with these uncertainties, or there needs to be agreement to limit the chemicalspecific HQs to the most important (from the ERA perspective) chemicals for each site. By focusing on the few most important chemicals of ecological concern, it is likely the scaleof the risk management decision will support the generation of the necessary ecotoxicological information, if adequate ecotoxicological data for these few chemicals is lacking. The assumption would be that the risk management decisions based on HQs for these limited chemicals will be protective of ecological resources that may be exposed to the other constituents at the site, as well.

The HQs are Explicit Numbers That are Easy to Communicate and are Often Given Undue Weight

This often occurs despite obvious ecological evidence to the contrary. HQs are also often given undue risk management weight, despite seemingly unrealistic magnitudes (e.g., values suggestive that every animal should die upon acute exposure) (Tannenbaum et al. 2003). This can lead to an erroneous (or at least confounded) perception that an ecological problem exists (i.e., unacceptable risk), thereby leading to potentially destructive remediation in order to appease a problem in risk perception (as opposed to real unacceptable risk). For example, at one site, HQs of K)Os-IOOOs were calculated for lead, leading to considerable regulatory concern even though the population density studies showed this very area had significantly higher mammal density than sites with lower lead HQs and the reference location (Sorensen and Margolin 1999; Sorensen 2000). An improvement would be to only use chemical-specific risk characterization in the appropriate situations (e.g., chemical Stressors are the most important risk drivers, scope of investigations to generate needed data is appropriately scaled to risk management decision), stay focused on chemical-specific risk characterization shortcomings, and include knowledgeable ecologiste and biologists in the risk assessment and management planning process.

A Decade of Experience is Available Regarding the Chemical- Specific Risk (Characterization Approach, but There seems to be IAille Momentum to Learn from and Build on this Experience

Several recent calls have been made to reflect on lessons learned (Chapman 2002; Fairbrother 2002; Tannenbaum et al. 2003). Two significant questions were asked directly about terrestrial ERAs: "Can ecological receptors really be at risk?" (Tannenbaum 2003) and "Are we missing the forest for the trees?" (Tannenbaum 2002). Tannenbaum attempts to answer these questions by directing attention to the decade of ERA experience at Superfund sites, particularly military installations where natural resources are closely managed and monitored. he explores the reasons why so few effects have ever been observed for mammals and birds in the context of sites with long-term histories of contamination. he concludes that two possible explanations exist: (1) either toxicological effects never occurred; or, (2) the effects did occur, but the ecological system has rebounded to the point where the effects no longer exist (or are no longer measurable). Tannenbaum describes a variety of technical reasons why this may have occurred, including many of the "profound uncertainties" also discussed herein. In light of this information, Tannenbaum suggests that at least one option would be to discontinue the use of HQ calculations for terrestrial mammals and birds and use focused field efforts to identify bona fide impacts (if any). Once these lessons are learned, they should be extrapolated to sites with less available information. As always, the scope and scale of any field efforts needs to be appropriate for magnitude of the risk management decisions to be made and ensure that decisions are made in a timely and cost-effective manner.

ALTERNATIVES TO CHEMICAL-SPECIFIC HQ RISK CHARACTERIZATION

Alternatives to chemical-specific HQ risk characterization include: compensatory restoration; performance-based ecological monitoring; ecological significance criteria; and net environmental benefit analysis. Some of these are already in use, just not on a widespread basis. There is more detail for some methods than others, based on their applicability for use in ERA and current developments related to their use in ERA.

These alternatives should be considered additional tools in the risk assessor's and risk manager's toolbox, and the appropriateness for their use at any given site should be considered. These alternatives do not address or "assess" risk in its classic definition (i.e., the chance or degree of probability of an adverse effect), but neither does chemical-specific risk characterization. Instead, each of these alternatives is a tool that can be used to assess conditions at a site and develop information as to whether these conditions are protective of populations, communities, and ecosystems. Ecological services flow from intact, functioning ecosystems, which by definition include viable populations and communities. As long as the risk management decision is protective of ecological services, the requirements of CERCLA, RCRA, and other laws that require overall protection of the environment will be satisfied.

The alternatives discussed here are not appropriate for every site, just like chemical-specific HQ calculations are not appropriate for every site. An effort has been made to identify occasions where these alternatives may be most useful and when they are not. Also, each of these alternatives has limitations, including cost, technical experience required, and uncertainties. It is critical that the scope, scale, timeliness, and costs for any alternative, including chemical-specific risk characterization, be appropriate for the magnitude of the risk management decision (s) at the site. The risk assessors and risk managers for any given site should have the flexibility to evaluate the benefits and limitations of each tool, and use the one (or more) that is most effective for that site's risk management decision-making. Often the environmental setting of the site and/or site history (including future land use) will determine which alternative or set of alternatives is most appropriate for the site.

Compensatory Restoration

Compensatory restoration is drawn from practices in natural resource damage assessment where restoration is scaled to compensate for potentially lost use of ecological services (USDOI 1986, 1996). Procedures for estimating lost services and restoration alternatives vary depending on the resources impacted and the scale of the particular site, with emphasis on habitat characteristics rather than chemical impacts (e.g., NOAA 1997; USEPA 200Ob; USACOE 1996b; Yozzo et al. 1996; USFWS 1980a,b). A key tenet of this approach is that a reduction in ecological services provided by habitat in one location can be offset by a corresponding increase in services elsewhere within the same ecosystem (TNRCC 2001). Therefore, compensatory restoration projects may or may not occur at the site. The restoration effort could involve enhancement of an adjacent or nearby unaffected area, thereby increasing services in one area to offset the lost services in another. This approach is also similar to wetland banking, where wetland compensation is required as part of the wetland permitting process. Small parcels of contaminated wetlands that cannot be restored are offset by the enhancement and/ or preservation of even larger parcels of wetlands in the vicinity. Compensatory restoration can provide the meaningful increase and return of ecological services to an area that offsets historic and continuing effects at chemical release sites, allowing one to bypass substantial assessment arid remediation activities.

Figure 1. Compensatory Restoration Examples. Figure 1 shows the hypothetical restoration of impacted Area 1 compared to the enhancement of unimpacted Area 2. Note that the enhancement of Area 2 yields greater ecological services than restoration of Area 1 (i.e., 1 + 2' > 1' + 2).

Risk assessors and risk managers need to be aware that this approach can lead to a greater net benefit in terms of natural resources. For example, as illustrated on Figure 1, the restoration of a contaminated site compared to the enhancement of an uncoiilamiiialed she can result in very different changes in ecological services. In the two sites shown on Figure 1, enhancement of Area 2 yields a greater increase in overall ecological services than remediation/restoration of Area 1 (i.e., 1 + 2' > 1' + 2) because enhancement of Area 2 created greater and faster increase in services than remediation/restoration activities in Area 1. Further, it may be more cost effective to enhance Area 2, leaving a potentially adapted community of Area 1 intact. There are numerous examples of ecological communities with pollution-induced tolerance, meaning that established communities are present in areas where chemical contamination has existed for long periods of time (Millward and Klerks 2002; Grant 2002; Klerks 2002). There are often measurable differences in biological community composition, but the community itself may otherwise be stable, dynamic, and thriving. Remediation or physical perturbation to the area just to achieve chemical-specific remedial goals (as opposed to ecologically-based restoration goals) can be detrimental to these established communities that have some competitive advantage in the "impacted" environment.

The compensatory restoration approach would not be appropriate for sites or areas of concern where unacceptable human health risks are present (TNRCC 2001). For sites where significant ecological risk is present in the absence of or following remediation to address human health risk, the compensatory restoration approach provides additional flexibility. It enables the use of an appropriately scaled (determined typically by performing an ecological services analysis) on-site or off-site compensatory ecological restoration project, possibly in combination with more standard remedial actions (e.g., hot spot removal, natural attenuation), to address unacceptable ecological risk. The ec\ological services analysis would evaluate potential restoration and/or protection sites, and compare them to the ecological services afforded by the contaminated site itself across a range of alternatives from "leave it alone" to fully remediate and restore. One needs to take into account the feasibility of achieving "pre- impact" community structure, composition and function. A compensatory restoration approach may also not be appropriate if the affected habitat is rare and/or protected, as one may need to attempt on-site restoration regardless of effort required even when a low probability of success exists.

Performance-Based Ecological Moniloring (also Called Natural Remediation or Monitored Natural Recovery)

Ecological monitoring provides insight into the actual conditions present at a site. Ecological monitoring involves the systematic measurement of designated characteristics of the exposed ecosystem, such as abundance and diversity of flora and fauna. There are numerous technical resources available regarding monitoring methods and interpretation of results (e.g., USEPA 1989, 1999b; USFWS 1990; Heyer et al. 1994; Wilson et al. 1996; Reaka-Kudla et al. 1996; Droege et al. 1998). The results obtained are typically compared to the results obtained from a reference (or unaffected) area, but sometimes this is not possible. Over time, the direction of change (i.e., the trajectory) in a relevant ecological indicator (e.g., percent cover, diversity, production) can be determined (e.g., increase, decrease, or no change). Some of the benefits of ecological monitoring as a part of the risk-based restoration effort were recently highlighted in the Inside EPA's Risk Policy Report (Barnthouse 2002). In this example, remedial success is evaluated in terms of ecological performance indicators rather than chemical concentrations.

An advantage of ecological monitoring is that it directly involves the exposed populations and/or communities, so compensatory mechanisms, such as adaptation and tolerance, are considered directly. Mechanisms of adaptation and tolerance for some chemicals are well documented (e.g., Tyler et al. 1989; Marcus 1997; Newman 1998; Ross 1994; Loneragan et al. 1981; Macnair 1990), although they are generally devalued in the chemical-specific risk characterization approach because there is no agreed upon quantitative methodology to incorporate them. For example, ecotoxicological studies used for both the development of ecotoxicity screening values and food web model ecotoxicological benchmarks are based on naive organisms (i.e., organisms not previously exposed), and these organisms are known to lack the same degree of compensation as adapted populations (Loneragan et al. 1981; Macnair 1990).

Ecological monitoring can provide insight into natural remediation, as components of ecosystem services are measured directly. Ecological monitoring may indicate stable, improving, or declining conditions. Declining or stable but diminished conditions may require further action (e.g., remediation or compensatory restoration), depending on the specific management objectives for the site. Figure 2 illustrates The fundamental processes of natural remediation are illustrated in Figure 2. The general decrease in Stressor level (e.g., toxicity) over time and the increase in ecological services as the Stressor decreases are shown in Figure 2a. A variety of trends for the increase in ecological services over time is shown in Figure 2b. The rate of return to baseline conditions depends on many variables, such as chemical toxicity, species sensitivity, physical/chemical characteristics of the site, and chemical degradation rates. Also note on Figure 2b that ecological services may stabilize at a point below baseline conditions.

Figure 2. Natural Restoration/Remediation Examples. Figure 2a illustrates the inverse relationship between toxicity and ecological services over time. Figure 2b shows a variety of trends for the increase in ecological services over time. Ecological services can return to baseline conditions or return to diminished conditions (dashed line) at variable rates.

Ecological monitoring/natural remediation needs to be considered early in the problem formulation step of the ERA (Swindoll et al. 2002). Stakeholder input is necessary to ensure that appropriate ecological indicators are identified and measured (i.e., ecological performance indicators that are directly or indirectly related to ecological services to be protected or restored). It is appropriate to consider this approach for sites where there is no evidence of an imminent threat to wildlife; chemical contaminants are biodegradeable or will naturally attenuate over time; apparently viable ecosystem or community characteristics are present; sufficient time is available for monitoring or natural remediation; protecting ecosystem integrity is valued; ecological indicators can be monitored and restoration of lost ecological services are likely (Swindoll el al. 2002). A drawback of this performance-based ecological monitoring approach is the difficulty to distinguish patterns of observed change in ecological indicator(s) relative to site impacts as opposed to natural variability of the selected ecological indicator(s). This natural variability and/or ability to identify appropriately matched reference locations can make it difficult to discern significant differences in ecological indicators or to conclude an affected site has recovered. The scope and scale of ecological monitoring necessary to ensure sufficient statistical power needs to be balanced with the magnitude of the risk management decisions to be made.

Ecological Significance Criteria

The consistent use of a defined set of criteria can be used to determine if ecological issues (e.g., receptors, risk, restoration) are or are not significant at a site. The ideas behind ecological significance criteria are, in part, analogous to the ideas behind the concept of assimilative capacity. "Assimilative capacity is based on the assumption that the environment is capable of degrading and transforming anthropogenic chemical and biological wastes from harmful products to harmless products... Adverse environmental impacts arise only when the system's assimilative capacity is exceeded" (GRI 2002). Ecological significance criteria for use in environmental risk management decision-making have been developed and used in a similar fashion, although they have been referred to by a variety of names, such as exclusion criteria, screening criteria, entry criteria, and exit criteria.

Several U.S. states have also addressed this topic through the explicit use of ecological significance criteria by considering concepts such as fate and transport, scale of release, and population-level wildlife exposures (e.g., PADEP 1998; MADEP 1996; TNRCC 2001 ). Texas guidance provides the most detailed checklist available for the use of criteria (GRI 2002). It is significant that the Texas guidance was developed by a diverse stakeholder group, including representatives from the USEPA, the U.S. Fish and Wildlife Service (USFWS), the National Oceanic and Atmospheric Administration (NOAA), the Texas Natural Resource Conservation Commission (TNRCC) and a variety of industrial sectors (TNRCC 2001).

Other organizations have also discussed and developed ecological significance criteria for similar purposes. For example, the Petroleum Environment Research Forum (PERF) project 99-01, a collaborative effort of representatives and scientists from the Gas Research Institute, the American Petroleum Institute (API), U.S. Department of Energy-funded national labs (Oak Ridge National Lab, Lawrence Livermore National Lab) and the energy industry, developed ecological significance criteria specific for exploration and production sites to assess the need for further ecological assessment (PERF 2002; GRI 2002). This built on an earlier API effort that developed ecological significance criteria for underground storage tank sites (called preliminary evaluation in API 2000). Another example of an organization to address this topic is ASTM. A broad-based, multi-stakeholder ASTM workgroup debated and discussed the use of ecological significance criteria as part of the development of the Risk Based Standard Guide for the Protection of Ecological Resources, or Eco-RBCA (ASTM 2002).

Where only limited numbers of individual animals are likely to be exposed, ecological significance criteria have been used to exclude small sites with predictable migration and degradation of contaminants from having to perform chemical-specific risk characterizations (even screening-level ERAs). These small areas are likely to experience a degree of natural recovery (as illustrated in Figure 2b). Ecological significance criteria can also be used at sites with a designated land use that is incongruent with supporting wildlife populations (e.g., industrial, commercial, and residential properties). In addition, information collected to complete the ecological significance criteria checklist can provide useful information for developing the conceptual site model in the problem formulation step if it is concluded that an ERA is warranted or the checklist can provide formal documentation for sites where no further action is warranted. Ecological significance criteria may prove to be "an added step" at large sites, sites with sensitive or unique habitats, or in an area with many adjacent small releases or sites (i.e., depending on the proximity and contamination) because it is likely that these types of sites will have significant ecological issues that need to be addressed. For these situations, further ecological assessment, possibly including ERA, is likely warranted.

An apparent barrier to the consistent use of ecological significance criteria is the lack of widespread acceptance that these criteria are useful for risk management decision-maki\ng. Some regulatory agencies require the use of chemical-specific risk characterization and will not consider any substitute. The use of ecological significance criteria clearly does not avoid or bypass consideration of ecological resources that could be affected by chemical releases. Rather, it represents part of a phased approach to decision-making that focuses the effort on those situations where ecologically significant effects may occur.

Net Environmental Benefit Analysis

Net environmental benefits analysis (NEBA) is a technique to compare the benefits and disadvantages of various remedial alternatives to organisms, populations, and ecological system services (Efroymson et al. 2002). The NEBA approach is used to make risk management decisions that minimize environmental and socioeconomic harms at the site of a chemical release (IPIEGA 2000). It was originally developed to help identify remedial alternatives that could be used following catastrophic oil spills in order to avoid detrimental remedial action in time of crisis (NEBA could also be performed on selected shorelines before an oil spill as part of contingency planning). The results of the NEBA identify the remedial action that will lead to the most effective recovery of the impacted ecosystem, albeit this achievement is typically measured in terms of acceptable cleanup levels. Typical remedial actions that are compared during a NEBA involve leaving contamination in place (e.g., natural attenuation) to avoid the loss of existing ecological services, traditional active remedial actions, and improving ecological value through restoration that does not directly focus on chemical contamination (Efroymson and Stewart 2002).

An example is illustrated in Figure 3 of the comparisons that can be made using the NEBA approach to compare multiple remedial alternatives, compensatory restoration, and natural restoration/ remediation. This figure highlights that both hypothetical remedial alternatives initially diminish ecological services, but one approach clearly increases recovered services faster than the other. Also, this graph highlights that remedial alternative 1 achieves baseline conditions only slightly ahead of natural restoration/ remediation, while remedial alternative 2 does not return to baseline conditions. Risk management decision-making can be improved when approaches (cost and time to return of services) are compared in this way.

Past experience has shown that invasive remedial actions (e.g., soil excavation, sediment dredging) often are a greater source of environmental harm than the existing contamination. In these cases, a NEBA may suggest that natural attenuation (recovery), or more creative and less intrusive restoration methods (Efroymson and Stewart 2002). If a restoration alternative is deemed appropriate because remedial actions are too destructive regardless of potential ecological risks, chemical-specific risk characterization of the impacted area may not be warranted.

The NEBA approach may be recommended to supplement or substitute for the standard chemical-specific risk characterization approach at sites where substantial ecological services still remain, and especially for contamination in habitats that are difficult to restore in-situ following remedial action. A NEBA approach would be most appropriate at sites where recovery trajectories based on prior experiences and modeling are well understood. If ranges of estimated recovery times and degrees are broad, the NEBA approach may not provide sufficient discriminatory value. Broader use of a NEBA approach may lead to de-emphasizing the results of chemical- specific risk characterization at sites with sensitive environments and/or critical habitat for threatened, endangered or protected species (sites for which the chemical-specific approach was originally targeted), because the NEBA will demonstrate that invasive remedial action designed to reduce chemical exposure will cause more harm than good.

Figure 3. Net Environmental Benefit Analysis (NEBA) Examples. Illustrates examples of the comparisons that can be made using the NEBA approach.

NEBA typically focuses on benefits and debits of proposed remedial actions and/or risk management decisions at the area of contamination. Augmenting the NEBA approach with the procedures and analyses used for environmental impact analyses and statements can result in a broader and more overall measure of the benefits and detriments of proposed remedial actions, thereby resulting in the environmentally superior project being selected and implemented.

Environmental Impact Analyses (EIAs) and Environmental Impact Statements (EISs) are required in most situations involving significant development by industry and government (NEPA 1969). The results of this process can result in findings of no significant ecological impact for broad-based changes to the environment (e.g., roads, buildings, alterations oi aquatic features). Of particular value to remediation projects is that the EIA process can identify appropriate mitigation or enhancements (e.g., wetlands banking, restoration of unused areas to increase service) to counteract any significant ecological impacts from these broad-based changes. The EIA process also examines a proposed project against viable alternatives with the goal of identifying the environmentally superior project-i.e., a project where any significant adverse environmental impact has been mitigated or avoided.

Historically, the EIA process has not typically been used for making decisions regarding approval of remediation projects, although the EIA process shares many similarities with the compensatory restoration and NEBA approaches. The EIA process is one of the few tools that can effectively promote integrated decision- making and incorporate consideration of future needs into current decisions-critical elements toward moving to sustainable development (Pendergrass 2002). Although commonly used in the design and approval of new projects, EIAs can also be valuable in planning for and possibly pre-mitigating or preventing operational releases from becoming remediation problems. For example, an oil company developing a new well field could use the EIA process to pre- mitigate for small and limited-scale ecological impacts that will result from hypothetical spills in the future, thereby accounting for future remediation liability. Although the EIA process here is envisioned as an augmentation of a NEBA approach, it is likely the full use of EIA procedures will be overkill for most remediation sites, especially if the site and adjacent areas are already significantly degraded (e.g., industrial area). The cost and potential lengthening of the project schedule due to use of an EIA process would need to be balanced against the benefit of its broader scope. Remediation delays may contravene the need for "immediate response" under CERCLA and RCRA.

FUTURE RESEARCH DIRECTIONS

The following section describes two emerging areas of investigation to address limitations of the chemical-specific risk characterization approach. The objective of this research (e.g., wildlife habitat assessment/landscape ecology/population biology, and field validated risk-based screening levels) is to develop new tools that supplement, augment, or substitute for the chemical- specific risk characterization approach and to improve the ecological relevancy of ecological risk assessments and resulting risk management decisions.

Wildlife Habitat Assessment, Landscape Ecology, and Population Biology

Habitat quality, landscape ecology, and population biology must be considered in order to assess risks at the spatial scale of populations, communities, and ecosystems. What is needed, however, are tools that incorporate these concepts into the ERA, and acceptance among practitioners and regulators to use these tools (providing the tools are appropriate to the scale of the problem at a given site). ASTM, SETAC, and a variety of scientific publications have provided (or will provide) a foundation for the development and use of such tools. Some of these efforts are discussed below.

The USFWS has developed mathematical models to describe and assess habitat quality as well as a structured approach for assessing wildlife habitat. Wildlife habitat can be assessed using 157 species-specific Habitat Suitability Index (HSI) models for terrestrial, wetland, and aquatic wildlife species (NWRC 2002). These models are based on the Habitat Evaluation Procedures (USFWS 1980a). The models identify hypotheses pertaining to species- habitat relationships. These models serve as a basis for improved decision-making because they identify specific habitat variables that can be measured, and improved. HSI models may be appropriate for contaminated sites in order to predict the potential presence (or absence) of particular species, to identify appropriate ecological receptors and pathways and to evaluate potential remediation and restoration strategies. Also, these models can be used to guide remediation, restoration, or monitoring efforts associated with preservation of a particular species. However, these models in their current form do not provide statements of proven chemically-induced cause and effect relationships, so their current usefulness in ERA is limited.

An ASTM Symposium (titled "Landscape Ecology and Wildlife Habitat Assessment") has addressed the usefulness of HSI models and other tools from the field of wildlife management that may be useful tools for ERA. Efforts are also underway at SETAC, with an international Pellston Workshop focused on how to address populations in ERA. Tools from applied wildlife biology, resource management, and conservation management are being evaluated for their use in ERA. An ultimate goal is the establishment of a framework for population- level ERA that integrates workshop recommendations concerning a wide range of issues.\The need for management tools to include the interrelationship of habitat, landscape ecology, and population biology is being addressed as well. For example, Maltby el al. (2001) review the range of population-level endpoints that can be used in ERA. They discuss models and approaches that have the potential to enhance ERA, with detail regarding the advantages and limitations of each method. Finally, they discuss the linkages between individual-level effects and population-level consequences. Others have addressed similar topics, such as: the role of populations, communities, and ecosystems in contemporary management of vertebrates (Anderson 2000); contaminants as a cause of population perturbations (Fox 2000); contaminant-effect endpoints and approaches for population-level assessment in terrestrial vertebrates (Rattner el al. 2000; Albers et al. 2000; Sample et al. 2000); and the effects of contaminants on populations in spatially structured environments (McLaughlin and Landis 2000). Another group of scientists has addressed chemical effects by evaluating population models, landscape models, and ecosystems models, including how to select and use these models in ERA (Pastorok ei ai 2002). Further examples in this issue include efforts to incorporate principles and models from conservation biology into the assessment of chemical effects on wildlife populations, communities, and ecosystems (Carlsen et al. in press; Efroymson et al. in press). These authors discuss field observations from the conservation biology literature in examination of the spatial relationship of contaminants at the landscape scale and discuss the use of conservation models for the analysis of habitat loss by contamination (Carlsen et al. in press). They also discuss impacts to wildlife from the loss of reproductive habitat and reduced food availability rather than direct toxicity (Efroymson el al. in press).

These types of tools may be used to identify and refine ecological significance criteria, improving their applicability, even to small sites. Tools that can be used on a landscape scale can also be used for large sites (or multiple sites within the landscape). Although appropriate use of these tools in ERA is not clearly understood at this time, efforts noted above have provided (or will provide) a foundation for the development and use of such tools in the near future. Again, what is needed is the acceptance among practitioners and regulators to use these tools (to supplement chemical-specific risk characterization), providing they are appropriate to the scale of the problem at a given site.

Field-Validated Ecological Risk-Based Screening Levels

Risk-based screening levels (RBSLs) are chemical concentrations below which adverse impacts are unlikely to occur. However, RBSLs are often used with little to no information regarding the species upon which they were developed or the conditions of laboratory chemical toxicity testing (e.g., USEPA 2001a). As a result, laboratory-based RBSLs are unreliable, as evidenced by the vast number of RBSLs that are less than background concentrations of naturally occurring metals (Tannenbaum et al. 2003). Too often in cases where RBSLs are exceeded, the receptors expected to be absent are present, at times in abundance. Often the cause of this inconsistency is that the RBSLs were generated based on toxicity to laboratory species (e.g., alfalfa and rye), which do not reflect toxicity to native plant species (e.g., oak trees and honeysuckle) actually found on the site. Given the decade of experience now available, it is time to consider alternative methods for developing RBSLs. A new empirical approach for developing and/or validating RBSLs may come from the sites with historic releases, as discussed in Tannenbaum (2002, 2003). This could be accomplished by reviewing sites to evaluate cases where ecological observations contradict what HQs predict. One would empirically determine (or confirm) the appropriate media-based concentration threshold (field-validated RBSL) for the presence or absence of viable ecological communities. This approach can be particularly useful for classes of chemicals and sites that are encountered routinely, such as petroleum releases at exploration and production sites, explosives at firing ranges, and pesticide sites. These field-validated RBSLs will likely be superior to laboratory-based RBSLs for most sites, except those with recent releases.

A simplistic example illustrating the decrease in chemical toxicity and the increase in ecological services that occur over time is provided in Figure 4 (similar to Figures 2a and 2b). While the chemical residue in soil may not diminish, it is well documented that the toxicity of the chemical often does (Swindoll el al. 2002; Alexander and Alexander 1999; Alexander 2000). Note on Figure 2 that the RBSL remains constant regardless of these dynamic changes that are known to occur over time. This figure illustrates that the concept of "risk" is more likely to be dynamic, yet the HQ is static over time. In this particular example, the laboratory-based RBSL was relevant during the early stages of ecological exposures; however, it was not relevant as time progressed. This example highlights that the timing of an investigation is also related to the relevance of the laboratory-based RBSL. Laboratory-based RBSLs are likely to have the greatest relevance when the release first occurs. Typical remedial investigations can occur decades after the release occurred (Tannenbaum 2003). This example also highlights the lack of utility in many current RBSLs and the need for field-validated RBSLs that take into account time dependent mechanisms such as adaptation, tolerance, and ecosystem redundancy.

Tannenbaum el al. (2003) suggested that it is time to recalibrate the HQ model, and field validation of RBSLs may provide part of that recalibration. Consider this analogy (Tannenbaum et al. 2003): If one takes a person's temperature and obtains a reading of 150 Fahrenheit, you would immediately know this is impossible. The human body must be within a relatively narrow temperature range, with death above 106 Fahrenheit for most people. You would instead reason that the temperature could not be reliably measured until the thermometer was recalibrated. HQs [and RBSLs] in ERA are in similar need of recalibration.

Figure 4. Field Validated Risk Based Screening Levels (RBSLs). A simplistic example to illustrate the decrease in chemical toxicity and the related increase in ecological services that occur over time. Note that the laboratory-based RBSL remains constant regardless of these changes that are known to occur. The laboratory- based RBSL may have been relevant during the early stages of chemical release; however, it is not relevant as time progresses.

CONCLUSIONS

This paper discussed the prevalent use of chemical-specific risk characterization (i.e., HQs) in ERA, including limitations of this approach. Several available alternatives to supplement, augment, or substitute for chemical-specific HQs were identified: compensatory restoration; performance-based ecological monitoring; ecological significance criteria; and net environmental benefit analysis. Situations where these alternatives will likely result in superior risk management decisions relative to focusing on chemical-specific risk characterization are highlighted. Even though these alternatives may be useful to supplement, augment or substitute for chemical-specific risk characterization, care must still be taken to use the right approach (es) at each site. Also, it is critical that the scope and scale of any alternative used, including chemical- specific risk characterization, is appropriate for the risk management decisions at the specific site to allow remedial decisions to be made in a timely and cost-effective manner. Future research and development of tools that incorporate wildlife habitat assessment, landscape ecology, and population biology; and field- validated ecological risk-based screening levels in ERAs will also improve the ecological relevancy of risk management decisions.

The goal of this paper is to raise the level of awareness and acceptance of these alternative approaches. Doing so will facilitate the acceptance of the use of alternative approaches by risk assessors and risk managers and add much-needed tools to their toolboxes. Risk managers recognize that the ERA process is a tool used for informing decisions. However, the tool for every ecological risk management problem is not an ERA (Lackey 1996). The alternatives presented in this paper do not address or "assess" risk in its classic definition (i.e., the chance or degree of probability of an adverse effect), but neither does the chemical-specific risk characterization approach. Instead, each of these alternatives are tools that can be used to assess conditions at a site and develop information as to whether these conditions are protective of valued ecological entities: populations, communities, and ecosystems. As long as the risk management decision is protective of these valued ecological entities, the requirements of CERCLA, RCRA, and other laws that require overall protection of the environment will be satisfied. It is likely that superior risk management decisions can be made when there is more than one tool in the toolbox and risk assessors are permitted to use them.

REFERENCES

ADEC (Alaska Department of Environmental Conservation). 2000. Risk Assessment: Procedures Manual. Contaminated Sites Remediation Program. Available at http//:www.state.ak. us/dec/dspar/csites/ guidance/rapm2000.pdf

Albers P, Heinz G, and Hall R. 2000. Approaches for the assessment of terrestrial vertebrate responses to contaminants: Moving beyond individual organisms. In: Albers P, Heinz G, and Ohlendorf H (eds), Environmental Contaminants and Terrestrial Vertebrates: Effects on Populations, Communities, andEcosystems, pp. 109-148. SETAC Press, Pensacola, FL, USA

Alexander M. 2000. Aging, bioavailability, and overestimation of risk from environmental pollutants. Environ Sci Tech 34:4259-05

Alexander R and Alexander M. 1999. Genotoxicity of two polycyclic aromatic hydrocarbons declines as they age in soil. Environ Toxicol Chem 18:1140-3

Anderson S. 2000. The role of populations, communities, and ecosystems in contemporary management of vertebrates. In: Albers P, Heinz G, and Ohlendorf H (eds), Environmental Contaminants and Terrestrial Vertebrates: Effects on Populations, Communities, and Ecosystems, pp 1-18. SETAC Press, Pensacola, FL, USA

API (American Petroleum Institute). 2000. Primer for Evaluating Ecological Risk at Petroleum Release Sites. APIPUBL 4700. Revision 1. Available at api-ep.api.org/publications

ASTM (American Society of Testing and Materials). 2002. Risk- Based Corrective Action (RBCA) Standard Guide for the Protection of Ecological Resources. E2205-22. Available at http//:www.ASTM.org

Atchison GJ, Sandheinrich MB, and Bryan MD. 1995. Effects of environmental Stressors on interspecific interactions of aquatic animals. In: Newman MC and Jagoe CH (eds), Ecotoxicology-A Hierarchical Treatment, pp 319-37. CRC Press, Boca Raton, FL, USA

Barnthouse L. 2002. Guest Perspective: Risk-Based Restoration of Impaired Ecosystems. Inside EPA's Risk Policy Report. Available at http//:www.insideEPA.com, March 19.

Cal/EPA (California Environmental Protection Agency). 1998. Guidelines for Assessing Ecological Risks Posed by Chemicals. Office of Environmental Health Assessment. Available at http:// www.oehha.ca.gov/ecotox/documents/ecocrnr.html

Carlsen TM, Coty JD, Carlsen TM, et al. In press. The spatial extent of contaminants and the landscape scale: A review of field observations from the conservation biology literature. Environ Toxicol Chem

Cassee FR, GrotenJP, van Bladeren PJ, et al. 1998. Toxicological evaluation and risk assessment of chemical mixtures. Grit Rev Toxicol 28:73-101

Chapman P. 2002. Mistakes made/lessons learned. Environ Toxicol Chem 21:891-3

Droege S, Cyr A, and Larivee J. 1998. Checklists: An under-used tool for the inventory and monitoring of plants and animals. Conserv Biol 12:1134-8

EC (Environment Canada). 1997. Environmental Assessments of Priority Substances under the Canadian Environmental Protection Act. Guidance Manual Version 1.0. PS/2/CC/3E Chemicals Evaluation Division, Ottawa, ON, Canada

Efroymson RA and Stewart A. 2002. Ecological Risk Assessments Are Not just Toxicity Tests Anymore. The Fossil Report, September- October, Oak Ridge National Laboratory Fossil Energy Program, Oak Ridge, TN, USA

Efroymson R, Nicolette J, and Suter GW. 2002. Net environmental benefit analysis for remediation or ecological restoration of petroleum-contaminated sites. Presentation #304, Society for Environmental Toxicology and Chemistry 23rd Annual Meeting, Salt Lake City, UT, USA

Efroymson R, Sample B, and Peterson M. 2003. Ecotoxicity values for total petroleum hydrocarbons in soil: Plants and soil-dwelling invertebrates. Hum Ecol Risk Assess (this issue)

Efroymson R, Carlsen T, Jager H, et al. In press. Toward a framework for assessing risk to vertebrate populations from brine and petroleum spills at exploration and production sites. In: Kapustka L, Galbraith H, and Biddinger G (eds), Landscape Ecology and Wildlife Habitat Evaluation: Critical Information for Ecological Risk Assessment, Land-Use Management Activities, and Biodiversity Enhancement Practices, ASTM STP 1458. American Society for Testing and Materials, West Conshohocken, PA, USA

Fairbrother A. 2002. Risk assessment: Lessons learned. Environ Toxicol Chem 21:2261-3

Fox G. 2000. Perturbations in terrestrial vertebrate populations: Contaminants as a cause. In: Albers P, Heinz G, and Ohlendorf H (eds), Environmental Contaminants and Terrestrial Vertebrates: Effects on Populations, Communities, and Ecosystems, pp 19-60. SETAC Press, Pensacola, FL, USA

Gas Research Institute. 2002. Evaluating Potential Ecological Impacts at Exploration & Production Sites: A Compilation of Screening Criteria. GRI 02/0006. Available at http://www. gri.org

Grant A. 2002. Pollution-tolerant species and communities: Intriguing toys or invaluable monitoring tools? Hum Ecol Risk Assess 8:95-970

Heyer W, Donnelly M, McDiarmid R, et al. (eds). 1994. Measuring and Monitoring Biological Biodiversity: Standard Methods for Amphibians. Smithsonian Institution Press, Washington, DC, USA

IPIECA (International Petroleum Industry Environmental Conservation Association). 2000. Choosing Spill Response Options to Minimize Damage: Net Environmental Benefit Analysis. International Petroleum Industry Environmental Conservation Association (IPECA) Report Series, Volume 10. Available at http://www.ipieca.org/ publications

Kapustka L. 2003a. Do PAHs pose unacceptable ecological risks to terrestrial recentors at hazardous waste sites? Hum Ecol Risk Assess (this issue)

Kapustka L. 2003b. Establishing Eco-SSLs for PAHs: Lessons revealed from a review of literature on exposure and effects to terrestrial receptors. Hum Ecol Risk Assess (this issue)

Klerks P. 2002. Adaptation, ecological impacts, and risk assessment: Insights from research at Foundry Cove, Bayou Trepangnier, and Pass Foundation. Hum Ecol Risk Assess 8:971-952

Lackey B. 1996. Ecological risk analysis. In: Molak V (ed), Fundamentals of Risk Analysis and Risk Management, pp 87-97. CRC Press, Boca Raton, FL, USA

LDEQ (Louisiana Department of Environmental Quality). 1999. Risk Based Corrective Action Program. Corrective Action Group. Available at http://www.deq.state.la.us/technology/ recap/2000/index.htm

Loneragan J, Robson A, and Graham R (eds). 1981. Copper in Soils and Plants. Academic Press, New York, NY, USA

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