3 Analysis of Exposure The first part of the case will deal with the plaintiff’s evidence which tends to show or [sic] designed to show … that the toxic materials were on the lands of the defendants, that it migrated, that it got into the water in sufficient quantity to constitute a potential hazard . … When the evidence on these issues is presented … you will be asked to make a decision as to whether the plaintiff has established all of these elements by a prepon- derance of evidence. If the answer is negative, that is the end of the case. —Judge D. J. Skinner (1986), instructions to the jury in the Woburn, MA civil action Exposure is the contact or co-occurrence of a contaminant with a receptor. The analysis of exposure estimates the magnitude of exposure of the endpoint entities to contaminants, distributed in space and time. For contaminated sites, it is necessary to consider the exposures due to current conditions and future conditions. Under current conditions the endpoint organisms are exposed to the contaminants that have been determined, through chemical or radiochemical analysis, to occur in surface water, sediments, soil, shallow groundwater, air, and food items. This is referred to as the current baseline exposure. One future condition that must be considered is the future baseline. This is the exposure that will occur if no remedial actions are taken to isolate, remove, or destroy the contaminants. In general, the future baseline is assumed to be the same as or not worse than the current baseline, so it need not be estimated. However, if expo- sures may increase in the future, resulting in increased risks, a separate estimate of future exposure must be generated. The most common reason for exposures to increase is that contaminants are migrating out of the source leading to increasing concentrations or extent of contamination. Another reason, which is seldom consid- ered, is that succession will occur on the site resulting in improved habitat quality, more species exposed, and more individuals exposed. This second reason is an issue only if the current habitat quality is poor, if succession is allowed or encouraged to occur, and if succession is rapid relative to degradation and dispersal of the contam- inants. The future baseline should also be of interest if the exposure is significantly declining. That is because, if natural processes are rapidly reducing exposures, the expense and damage caused by remediation may not be justified (Chapter 9). Another future condition that must be considered is post-remedial exposure. As part of the feasibility study (Chapter 9), it is necessary to determine that exposures will be acceptable following remediation. In most cases, this is simply a matter of assuring that the remedial goals will be met (Chapter 8). However, if the remedial actions result in changes in physical or biological conditions that may modify exposure processes, a new exposure assessment may be performed. © 2000 by CRC Press LLC Analyses of exposure should be carried out in such a way to facilitate risk characterization. That is, the exposure estimates should be appropriate for charac- terizing risks by parameterizing the exposure variables in the exposure–response models. This requires that the exposure estimates address the same forms or com- ponents of the contaminants as the effects assessment and also have concordant dimensions. For example, the estimation of effects on plants may require that concentrations of chemicals in the aqueous phase of the soil be estimated, that concentrations be averaged over the rooting depth of the plants on the site, and that the results be expressed as a median concentration and other percentiles of the empirical distribution of point concentrations. In contrast, the estimation of risks to wildlife due to ingestion may require total concentrations in surface soil, averaged over the foraging range of the species, expressed as the mean and standard deviation. In all cases, the analysis of exposure must appropriately define the intensity, the temporal dimension, and the spatial dimension of exposure (Suter, Gillett, and Norton, 1994; EPA, 1998). Intensity is usually expressed as concentration in a medium that is in contact with a receptor, but dose and dose rate are also used. Time is usually the duration of contact, but other relevant aspects of time are the frequency of episodic exposures and the timing of exposure (e.g., seasonality). The spatial dimension is usually expressed as area within which an exposure occurs or as linear distance in the case of streams. If contamination is disjunct (i.e., spotty), the spatial pattern may be important. The definition of exposure must be some measure of intensity with respect to space and time. The simplest case is an average concentra- tion over the entire site that does not vary with time. The degree of detail and conservatism in the analysis of exposure depends on the tier of the assessment. Scoping assessments need only determine qualitatively that an exposure may occur by a prescribed pathway. Screening assessments must quantify exposure but should use conservative assumptions to minimize the likeli- hood that a hazardous exposure is inadvertently screened out. Definitive assessments should be realistic and therefore should treat the estimation of exposure and uncer- tainty separately. This may be done by estimating distributions of exposure or by estimating both the most likely exposure and upper-bound exposure. This chapter begins with a discussion of the component activities that comprise an analysis of exposure. Next, it discusses issues specific to the individual environ- mental media: water, sediment, soil, and biota. It then discusses analysis of exposure for particular taxa that are exposed to multiple media. The next topic is modeling of uptake of contaminants by biota, primarily for estimation of food chain exposures. Wastes such as petroleum and its derivatives that are clearly a mélange are discussed in a separate section because they are most logically treated as a complex contam- inant rather than as a collection of chemicals that are assessed separately. Finally, presentation of the results of an analysis of exposure is discussed. 3.1 COMPONENTS OF ANALYSIS OF EXPOSURE This section discusses in general terms the activities that comprise the analysis of exposure. Specific issues in sampling, analysis, and modeling are discussed in the following medium-specific and receptor-specific sections. © 2000 by CRC Press LLC 3.1.1 S AMPLING AND C HEMICAL A NALYSIS OF M EDIA Most of the funds and effort expended on studies of contaminated sites are devoted to the collection and chemical analysis of the abiotic media: soil, water, and sediment. Similarly, most of the guidance for site studies is devoted to media sampling and analysis. These activities should be performed as specified in the analysis plan, and the quality of the data should be verified before they are used in the risk assessment (Chapter 2). The issues to be addressed here are summarization, analysis, and interpretation of those data. These issues are particularly problematical when chem- icals are detected in some, but not all, samples (Box 3.1). BOX 3.1 Handling Nondetects One perennial problem with analytical data is sets of results that include both reported concentrations (detects) and reported inability to detect the chemical (nondetects). Thus, the low end of the distribution of concentrations is censored. The problem is that nondetects do not signify that the chemical is not present, merely that it is below the method detection limit (MDL). If a chemical is detected in some samples from a site, it is likely that it is also present at low concentrations in the samples reported as nondetects. For screening assessments, this problem can be handled simply and conservatively by substituting the detection limit for the nondetect observations in order to calculate moments of the distribution, or by using the maximum measured value as the estimate of exposure. However, for definitive assessments, such conservatism is undesirable. The most appropriate solution is to estimate the complete distribution by fitting parametric distribution functions (usually lognormal) using procedures such as SAS PROC LIFEREG or UNCENSOR (SAS Institute, 1989; Newman et al., 1989, 1995). Alternatively, a nonparametric technique, the Product Limit Esti- mator, can be used to give more accurate results when data are not fit well by the parametric functions (Kaplan and Meier, 1958; Schmoyer et al., 1996). The problem of censoring is exacerbated by the EPA analytical procedures, which were based on the recommendations of the American Chemical Society (Keith, 1994). The MDLs are not actually the lowest concentration that a method can detect, but rather the lowest concentration that the EPA believes can be detected with sufficient accuracy and precision. Therefore, an analytical labo- ratory may detect a chemical at 7, 9, and 11 µ g/l in three samples, but, if the MDL is 10 µ g/L, the reported results are <MDL, <MDL, and 11 µ g/l. While the two lower concentrations in this example are more uncertain than the highest concentration, those measured values are clearly more accurate than the esti- mates generated by the methods discussed above. The best procedure from a risk assessment perspective would be to report all detected concentrations with associated uncertainties rather than allowing chemists to censor data that they deem to be too uncertain. © 2000 by CRC Press LLC At some sites, chemical concentrations in site media that were measured prior to the remedial investigation are available. The assessors must decide whether they should be used in the assessment. Although more data are generally advantageous, these encountered data may not be useful because of their age, quality, sampling techniques, or design. In general, the utility of encountered data must be determined by expert judgment. Considerations relevant to the age of the data include the rate of degradation of the contaminants, the rate of change in the rate of release of contaminants from the source, and the rate of movement of the contaminated media. Even if concentrations are declining, old data may be useful for screening assess- ments because they provide conservative estimates of current concentrations. The quality of the data and the acceptability of the sampling methods must be judged in terms of the uncertainty that is introduced relative the uncertainty due to not having the data. For example, metal analyses that were performed without clean techniques may be acceptable if the contaminants of concern occur at such high concentrations that trace contamination of the sample is inconsequential. Another important con- sideration is the detection limits. Analyses with high detection limits may create misleading results in screening as well as definitive assessments. In addition to analyses of contaminant concentrations, analyses must be per- formed of the physical and chemical characteristics of the tested media that influence toxicity. These analyses are particularly important when toxicity tests of the ambient media are performed, because the media may be unsuitable for the test organisms due to basic properties. For water, these include pH, hardness, temperature, dissolved oxygen, total dissolved solids, and total organic carbon. For sediments, they include particle size distribution, total organic carbon, dissolved oxygen, and pH. For soils, the same properties are measured, except that dissolved oxygen is omitted and water content (e.g., field capacity) and major nutrients (e.g., nitrogen, phosphorus, potas- sium, sulfer) are added. For example, differences in plant growth between contam- inated and reference soils may be due to fertility, pH, or texture rather than toxicity. Without information concerning those properties, the case for toxic effects cannot be defended. Exposure analyses for ambient media toxicity tests require analyses of chemicals of potential ecological concern (COPECs) from samples that are representative of the tested material. Therefore, results of analyses that are performed independently of the test should be used with great caution. Aqueous concentrations are highly variable over space and time. In our experience, storm events or episodic effluent releases may cause concentrations to change significantly over the course of a 7- day static replacement test, potentially making the analysis of only one of the three samples inadequate for exposure characterization. Soil samples are variable over space both vertically and horizontally. Therefore, exposures in a soil toxicity test may not be well characterized by analyses of samples that were collected “nearby” or were collected from a different range of depths. Sediments may be relatively stable in time like soils or may be mobile and therefore temporally variable. At most sites, abundant analytical data are generated which must be summarized and presented. The data summarization must meet the needs of the risk character- ization. Depending on the effects and characterization models, the data may be presented as means and variances, distribution functions, percentiles, or other forms. © 2000 by CRC Press LLC Care must be taken in statistical summarization to avoid bias. For example, because many sets of environmental data have skewed distributions that approximate the lognormal distribution, the geometric mean is commonly recommended. However, this results in an anticonservative bias when the value is used in calculations or interpretations that involve mass balance (Parkhurst, 1998). For example, if fish are exposed to varying concentrations in water, the best exposure metric for calculating their body burdens is the arithmetic mean concentration. Use of the geometric mean would improperly minimize the influence of high concentrations on uptake. In addition, the chemical data must be summarized for presentation to other members of the assessment team, risk managers, and stakeholders. The goal of these presentations should be to make important patterns in the data apparent. The best general approach to displaying relationships between parameters (e.g., stream flow and contaminant concentrations) is the conventional x – y scatterplot. Although maps are generally not as good as scatterplots for showing potentially causal relationships, they provide an important means of presenting spatially distributed data. The diffi- culty comes in converting data that are associated with points to areal representations. The simplest approach is to present the results on a map at the point where the sample was taken. The results may be in numeric form or as a glyph such as a circle with area proportional to the concentration. Alternatively, various geostatistical approaches can be used to associate concentrations with areas. These may be discrete areas (e.g., Theissen polygons), isopleths (e.g., Kriging), or gradients (e.g., polyno- mial interpolation) (Figure 3.1). Discussions of data presentation for contaminated sites can be found in Stevens et al. (1989b) and Environmental Response Team (1995a). More technical guidance may be found in Goovaerts (1997). This is an area in which a little creative thought can be useful. Good general guidance for data visualization is provided by Tufte (1983, 1990, 1997). Toxicity normalization provides a means of summarizing exposure data for numerous chemicals in an interpretable form. This is done by converting the con- centrations ( C ) to toxic units (TU), which are proportions of a standard test endpoint such as the Daphnia magna 48-h EC50. TU = C /EC50 (3.1) TU may be plotted as the value for each reach, subreach, transect, or other unit (Figure 3.2). The height of the plot is the sum of toxic units ( Σ TU) for that location. The advantage of this approach is that it displays the contaminant concentrations in units that are indicative of toxicity rather than simply mass per unit volume. There- fore, one can see which units are most likely to pose significant risks, and which chemicals are likely to be major contributors to the toxicity. The purpose of this analysis is heuristic. 3.1.2 C HEMICAL A NALYSIS OF B IOTA AND B IOMARKERS Analysis of abiotic media provides a measure of external exposure to contaminants, but not internal exposure or exposure through trophic transfers. These require esti- mates of uptake from media and transfer between biotic compartments. In the © 2000 by CRC Press LLC absence of reliable models of uptake and transfer, internal exposures and trophic transfers can be estimated by collecting and analyzing biota from the contaminated site or from laboratory exposures to contaminated media. This approach has the advantage of avoiding the use of highly variable empirical models or unvalidated mechanistic models. However, analytical chemistry is expensive, and some chemi- cals are rapidly metabolized or may not accumulate to detectable levels. Similarly, body burden analyses are not feasible for some species such as those designated as threatened and endangered. Care must be taken to ensure that the body burden analyses are relevant to the assessment. One issue is the treatment of unassimilated material. For example, if soil or sediment oligochaetes are not purged, the analysis may be dominated by chemicals in the gut contents which have not been incorporated. This may either overestimate or underestimate internal exposure of the worms and dietary exposure by vermivores, depending on whether the uptake factor (organism concentration/ FIGURE 3.1 An example of a map generated by Kriging which, given a spatial array of chemical measurements, defines areas estimated to have chemical concentrations within prescribed ranges. (Provided by Yetta Jager, ORNL) © 2000 by CRC Press LLC soil concentration) is less than or greater than 1. However, for chemicals that are rapidly depurated following assimilation, long holding times for purging may result in underestimation of exposure. Although 24 h is the standard holding time to evacuate gut contents, as little as 6 h may be sufficient (Mount et al., 1999). The issue of unassimilated material also arises with contamination of the surfaces of leaves and the fur and feathers and gut contents of wildlife. Decisions concerning this issue should be based on careful consideration of the actual mode of exposure of the endpoint organisms and of the exposure model used in the assessment. For example, if soil ingestion is included as a separate route in the exposure model, care should be taken to avoid incorporating soil into the chemical analyses of endpoint organisms or their food. Another aspect of ensuring relevance of analyses to the risk assessment is selection of appropriate species, higher taxa (e.g., insects), or assemblages (e.g., benthic invertebrates) for sampling and analysis. This depends on the purpose of the sampling. In general, the purpose is either to estimate the dietary exposure of consumers (i.e., analyzing plants to estimate exposure of herbivores) or to estimate the internal exposure of endpoint organisms. In the first case, sampling should focus on the primary food organisms and on the parts that are consumed. In the latter case, the sampling should focus on the endpoint species or, if that is impractical, on a FIGURE 3.2 A heuristic display of the toxicity normalized concentrations of metals in six reaches. The height of each bar is the sum of toxic units. © 2000 by CRC Press LLC closely related species with similar habits. If the endpoint entity is a community or higher taxon, then one may choose a representative species, representative set of species, or the entire group. To the extent that they can be identified and are relevant, the species that have the highest level of accumulation should also be selected. When other criteria are satisfied, organisms may be chosen on the basis of practical considerations such as ease of collection and body size. These issues are discussed by Phillips (1978). A third aspect of ensuring relevant analyses is selection of appropriate compo- nents of the organisms for analysis. This requires first deciding whether to analyze the whole organism, or some organ or other component. Second, one must decide which component, if not the whole organism. Once again, the primary consideration is the relationship of the analysis to the mode of exposure. If one is interested in the dietary exposure of a grazing or browsing animal, the leaves of plants would be analyzed; for beavers, the bark and cambium of small branches; for granivores, the seeds. If the analysis is performed to estimate internal exposure of an endpoint receptor, one should perform the analysis that is appropriate for the expo- sure–response model. This issue is discussed in greater detail below. It is also necessary to consider the relevance of analyses of mobile organisms. Mobile organisms collected on a site may have spent little time on that site. To the extent that it is consistent with the endpoints of the assessment, organisms that are most associated with the site should be preferred. Those are less mobile organisms and organisms with small home ranges. However, if the organisms of concern are not confined to the site, body burden analyses can still be relevant in that they realistically represent the proportional exposure of those organisms to the site and its contaminants. This rationale is applicable only if the organisms are not signifi- cantly exposed to sources of the contaminant outside the site. In some cases, analysis of organisms from the site is not practical because the site is small or highly disturbed or is not sufficiently consistent because of variability across the site. In those cases, organisms can be exposed to the contaminated site media under controlled conditions. For example, at the Concord, CA, Naval Weapons Station, plants and earthworms were exposed to site soils in the laboratory, and caged clams were exposed to site waters in the field (Jenkins et al., 1995). Similarly, earthworms were exposed in containers of soil at the Baird and McGuire Superfund site in Holbrook, MA (Menzie et al., 1992). While providing consistent bioconcen- tration data, such studies can also provide information on toxicity. An alternative to body burden analysis is analysis of biochemical biomarkers such as hepatic mixed-function oxidase enzymes (Huggett et al., 1992). Biomarkers may be detected when the contaminant cannot, and in some cases they may be measured without sacrificing the animal. For example, blood ALAD was used to estimate lead exposure in birds on the contaminated floodplain of the Coeur d’Alene River, and liver lead concentrations were determined in a subsample of birds (Johnson et al., 1999). However, biomarkers tend to be nonspecific, tend to increase nonlinearly with increasing exposure levels (e.g., to decline at high exposures due to inhibited protein synthesis), and tend to vary with extraneous variables such as the animal’s breeding cycle or nutritional state. In addition, few reliable expo- sure–response functions are available to relate biomarker levels to effects on © 2000 by CRC Press LLC organisms. For these reasons, biomarkers have been used much less than analysis of contaminant burdens in ecological risk assessments. However, one potentially important use is as bioassays (discussed below). Body burdens and biomarkers of exposure must, in most cases, be related to concentrations in media to which the organisms are exposed, because it is the media that will be remediated. The derivation of such relationships requires sampling and analysis of the exposure media colocated with the sampled biota. A series of such analyses of colocated biological and media samples can be used to develop a site- specific uptake factor or other model (Section 3.10). If the range of sites encompasses the range of contaminant levels, and if the uncertainty in the site-specific factor or model is sufficiently low, the factor or model can then be used to predict body burdens or biomarker levels at locations where media samples, but not biological samples, have been analyzed. Because media and biota concentrations may vary, samples should also be colocated in time. The acceptable interval between samples depends on the rate of variance of the biota and media, but the samples should not be taken in different seasons. A wide variety of methods is available for the collection of biota samples for residue analyses, with sampling methods generally being medium or taxon specific. Common collection methods for taxa generally of interest in risk assessments are outlined in Appendix A. General guidance on biota sampling is presented in Box 3.2. 3.1.3 B IOASSAY Bioassays are measures of biological responses that may be used to estimate the concentration or to determine the presence of some chemical or material. Bioassays are seldom used since the development of sensitive analytical chemistry. One valu- able use of bioassays is to determine the effective concentration of chemicals with a common mechanism of action. For example, the H4IIE bioassay provides a tox- icity-normalized measure of the amount of chlorinated diaromatic hydrocarbons in BOX 3.2 Rules for Sampling Biota • Take enough samples to represent the variability at the site adequately. • Sample endpoint taxa for which internal measures of exposure are useful. • Sample organisms or parts of organisms that represent the food of assessment endpoint species. • Take samples of biota and contaminated media at the same locations and at effectively the same time. • Take samples at reference and contaminated locations or on contamination gradients. • Because chemical concentrations in organisms may vary seasonally, take samples from all sites at approximately the same time. • Be aware of the information that is lost when samples are composited. © 2000 by CRC Press LLC the food of an organism (Tillit et al., 1991; Giesy et al., 1994). This use is analogous to the use of biomarkers to estimate internal exposure (Section 3.1.2), except that the goal is to estimate external response-normalized concentrations. It has also been proposed that activity of contaminant-degrading microbes be used as a bioassay for bioavailable contaminant constituents (Alexander, 1995). A weak interpretation of this bioassay is that, if biodegradation has stopped, there is no more bioavailable chemical to cause toxicity. This conclusion requires the assumption that biodegradation has stopped because the residue is unavailable rather than because it is resistant to biodegradation. A stronger interpretation would be that bioavailable concentration is a function of biodegradation rate so that one could estimate exposure from measures of degradation. This idea requires the assumption that the availability of a chemical for degradation by microbes is proportional to availability for uptake by endpoint plants and animals. The use of microbial toxicity tests as measures of bioavailability or ecological effects is beyond the current state of practice. 3.1.4 B IOSURVEY Surveys of the organisms inhabiting a site (biosurveys) are used in ecological risk assessments primarily as a means of determining effects (Chapter 4). However, they may play a role in the analysis of exposure. Specifically, they can be used to determine whether a species or taxon is present in contaminated areas, what life- stages are present, their abundance, and how long a migratory or otherwise transient species is present on the site. Without biological surveys, these presence and abun- dance parameters must be estimated using habitat models (discussed below). Bio- surveys of a contaminated site provide estimates of these parameters for the current baseline condition. Biosurveys of uncontaminated reference areas can provide esti- mates of these parameters for precontamination or postrestoration scenarios. Bio- surveys may be conducted in conjunction with the collection of organisms for chemical analysis (Section 3.1.2 and Appendix A), but care must be taken to ensure that sampling designs are adequate for both purposes. 3.1.5 T RANSPORT AND F ATE M ODELS Transport and fate models play a relatively small role in ecological risk assessment for contaminated sites, because concentrations in media can usually be determined by sampling and analysis. However, transport and fate models are needed when contamination is relatively recent, so that local media are not yet contaminated, or when sampling and analysis are not possible. Most models simulate transport and fate in a single medium. A set of transport and fate models for surface water, groundwater, soil, and air is available from the EPA through its Center for Exposure Assessment Modeling (CEAM) in Athens, GA (www.epa.gov/CEAM). However, for ecological assessments, it may be necessary to estimate concentrations in multiple media. For this purpose multimedia fate models are appropriate (Mackay, 1991; Cowan et al., 1995). Some models have been developed specifically for estimating transport and fate on contaminated sites, but in general they have focused on ground- water and other routes to human exposure (McKone, 1993). Several such models are © 2000 by CRC Press LLC [...]... Total concentrations are useful for screening assessments, but the dissolved form is appropriate for definitive risk assessments of aquatic biota for two reasons First, the form in the exposure assessment should match the form in the effects assessment Exposure–response models for aquatic biota are usually based on toxicity tests performed in clean water with highly soluble forms of the tested metal Hence,... Chemical Lead Arsenic Agricultural sites in Iran that are irrigated with wastewater Cadmium Soils east of Rocky Flats, CO Pu- 239 and Pu-240 Concentrations (mg or Bq/kg) at Various Depthsa 0–5 cm: 1400 0–10 cm: 1080 0–15 cm: 770 0–5 cm: 655 0–10 cm: 489 0–15 cm: 35 1 0–5 cm: 1.16 0–10 cm: 1.10 0–15 cm: 0.87 0–20 cm: 0.69 0 30 cm: 0.50 0–40 cm: 0 .39 0–50 cm: 0 .33 0–60 cm: 0 .30 0 3 cm: 11655 0–6 cm: 10101 0–9... depth for estimating all exposures may be necessary © 2000 by CRC Press LLC TABLE 3. 3 Rooting Profiles Relevant to Exposure Assessment for Plants in Different Biomes Biome Boreal forest Crops Desert Sclerophyllous shrubs Temperate coniferous forest Temperate deciduous forest Temperate grassland Tropical deciduous forest Tropical evergreen forest Tropical grassland savanna Tundra % Root Biomass in Upper 30 ... mechanistic and multimedia, but understanding of the processes is poor for most chemicals On the other hand, studies of mammalian toxicokinetics for human health risk assessment are abundant and should be applied to mechanistic models of wildlife exposure 3. 2 EXPOSURE TO CONTAMINANTS IN SURFACE WATER In most cases, ecological risk assessments of contaminated waters are based on measurements of aqueous concentrations... assumptions (Chapter 5) The information and associated references are more necessary for the definitive risk characterization 3. 5.1 ROOTING DEPTH The optimal depth interval for sampling soil should be the interval where most feeder roots are found for plant populations or communities that are assessment endpoint entities or are the major food sources of endpoint herbivores Ideally, the risk assessor... et al (1996) A basic set of models for the transport of petroleum constituents from contaminated terrestrial sites is presented as part of the ASTM Risk Based Corrective Action (RBCA) procedure (ASTM, 1994) When such models are used, ecological risk assessors must carefully review them to ensure that they provide the needed output for estimating ecological exposures 3. 1.6 EXPOSURE MODELS After contaminant... depuration of chemicals (Suter, 1993a) Application of PBPK models for estimation of chemical concentrations in wildlife is likely to remain limited for the foreseeable future, because these models are data intensive, requiring species-, tissue-, and contaminant-specific compartmental transfer rates and other parameters that are generally not available for wildlife species 3. 8.2 EXPOSURE BASED ON EXTERNAL... of concern In a retrospective risk assessment for a contaminated site, it is rarely necessary to model chemical transport and fate in soil An exception is the risk assessment for a future scenario in which buried waste constituents are carried to surface soil during a period of high precipitation or contaminant concentrations are reduced by degradation or other losses 3. 4.1 SAMPLING, EXTRACTION, AND... metal oxides) The EPA Office of Water has recommended that assessments of effects of aqueous metals on aquatic biota be based on dissolved metal concentrations as determined by analysis of 0.4 5- m-filtered water (Prothro, 19 93) However, some states and EPA regions still require that total concentrations be used in ecological risk assessments for the sake of conservatism In fact, even nominally dissolved... concentration (mg/l) Fc = flow of the contaminated source (l/s) Cu = upstream concentration (mg/l) Fu = flow of the receiving stream (l/s) The use of this formula is appropriate if one is concerned only about risks to biota in a receiving stream near the source, or if performing a screening assessment for a stream If the assessment must address a sufficient length of stream for processes © 2000 by CRC Press . concentrations are useful for screening assess- ments, but the dissolved form is appropriate for definitive risk assessments of aquatic biota for two reasons. First, the form in the exposure assessment should. EPA CEAM (Burns et al., 1982). 3. 3 EXPOSURE TO CONTAMINANTS IN SEDIMENT Ecological risk assessments of contaminated sediments are typically based on chem- ical concentrations in bulk sediment. useful for scoping and screening assessments but are unlikely to be acceptable for definitive assessments. Two different expressions of sediment contamination are commonly used in ecological risk assessments: