<IP_ADDRESS> Apportionment Decisions If the answer to the question in Box 11 is no ( there is not more than one relevant medium specific criterion/ regulatory action), then the recommended method for setting a healthbased water quality criterion is to utilize a subtraction calculation ( Box 12). Specifically, appropriate intake values for each exposure source other than the source of concern are subtracted out. EPA will rely on average values commonly used in the Agency for food ingestion and inhalation rates, combined with mean contaminant concentration values, for calculating RSC estimates to subtract. Alternatively, contaminant concentrations could be selected based on the variability associated with those concentrations for each source. This implies that a case by case determination of the variability and the resulting intake chosen would be made, as each chemical evaluated can be expected to have different variations in concentration associated with each source of intake. However, EPA anticipates that the available data for most contaminants will not allow this for determination ( based on past experience). Guidance addressing this possibility is addressed in the Exposure Assessment TSD. EPA does not recommend that high end intakes be subtracted for every exposure source, since the combination may not be representative of any actually exposed population or individual. The subtraction method would also include an 80 percent ceiling and a 20 percent floor. 4 14 If the answer to the question in Box 11 is yes ( there is more than one medium specific criterion/ regulation relevant), then the recommended method for setting health based water quality criteria is to apportion the RfD or POD/ UF among those sources for which health based criteria are being set ( Box 13). This is done via a percentage approach ( with a ceiling and floor). This simply refers to the percentage of overall exposure contributed by an individual exposure source. For example, if for a particular chemical, drinking water were to represent half of total exposure and diet were to represent the other half, then the drinking water contribution ( or RSC) would be 50 percent. The health based criteria would, in turn, be set at 50 percent of the RfD or POD/ UF. This method also utilizes an appropriate combination of intake values for each exposure source based on values commonly used in the Agency for food ingestion and inhalation rates, combined with mean contaminant concentration values.

Finally, if the levels of exposure for the chemical in question are near ( currently defined as greater than 80 percent), at, or in excess of the RfD or POD/ UF ( i. e., the answer in Box 9 is yes), then the estimates of exposures and related uncertainties, recommended apportionment ( either box 12 or 13), toxicity related information, control issues, and other information are to be presented to managers for a decision ( Box 10). The high levels referred to in Box 9 may be due to one source contributing that high level ( while other sources contribute relatively little) or due to more than one source contributing levels that, in combination, approach or exceed the RfD or POD/ UF. Management input may be necessary due to the control issues ( i. e., cost and feasibility concerns), especially when multiple criteria are at issue. In practice, risk managers are routinely a part of decisions regarding regulatory actions and will be involved with any recommended outcome of the Exposure Decision Tree or, for that matter, any alternative to the Exposure Decision Tree. However, because exposures approach or exceed the RfD or POD/ UF and because the feasibility of controlling different sources of exposure are complicated issues, risk managers will especially need to be directly involved in final decisions in these circumstances.

It is emphasized here that the procedures in these circumstances are not different than the procedures when exposures are not at or above the RfD ( or POD/ UF). Therefore, in these cases, estimates should be performed as with Boxes 11, 12, and 13. The recommendation should be made based on health based considerations only, just as when the chemical in question was not a Box 10 situation. If the chemical is relevant to one health criterion or regulatory action only, the other sources of exposure could be subtracted from the RfD or POD/ UF to determine if there is any leftover amount for setting the criterion. If the chemical is a multiple media criteria issue, then an apportionment should be made, even though it is possible that all sources would need to be reduced. Regardless of the outcome of Box 9, all apportionments made ( via the methods of Boxes 12 or 13) should include a presentation of the uncertainty in the estimate and in the RfD or POD/ UF for a more complete characterization.

The process for a Box 10 situation ( versus a situation that is not) differs in that the presentations for Boxes 12 and 13 are based on apportionments ( following the review of available information and a determination of appropriate exposure parameters) that must address additional control issues and may result in more selective reductions. With Box 10, one or several criteria possibilities (" scenarios") could be presented for comparison along with implications of the effects 4 15 of various control options. It is appropriate to present information in this manner to risk managers given the complexity of these additional control issues.

4.2.3 Additional Points of Clarification on the Exposure Decision Tree Approach for Setting AWQC As with Box 9, if a determination is made in Box 8A ( i. e., information is available to characterize exposure) that exposures are near, at, or above the RfD ( or POD/ UF) based on the available information, the apportionments made need to be presented to risk managers for decision. If information is lacking on some of the multiple exposure sources, then EPA would use a default of 20 percent of the RfD or POD/ UF ( Box 8B).

Results of both Boxes 12 and 13 rely on the 80 percent ceiling and 20 percent floor. The 80 percent ceiling was implemented to ensure that the health based goal will be low enough to provide adequate protection for individuals whose total exposure to a contaminant is, due to any of the exposure sources, higher than currently indicated by the available data. This also increases the margin of safety to account for possible unknown sources of exposure. The 20 percent floor has been traditionally rationalized to prevent a situation where small fractional exposures are being controlled. That is, below that point, it is more appropriate to reduce other sources of exposure, rather than promulgating standards for de minimus reductions in overall exposure.

If it can be demonstrated that other sources and routes of exposure are not anticipated for the pollutant in question ( based on information about its known/ anticipated uses and chemical/ physical properties), then EPA would use the 80 percent ceiling. EPA qualifies this policy with the understanding that as its policy on cumulative risk assessment continues to develop, the 80 percent RSC may prove to be underprotective.

In the cases of pollutants for which substantial data sets describing exposures across all anticipated pathways of exposure exist, and probabilistic analyses have been conducted based on those data, consideration will be given to the results of those assessments as part of the Exposure Decision Tree approach for setting AWQC.

For many chemicals, the rate of absorption from ingestion can differ substantially from absorption by inhalation. There is also available information for some chemicals that demonstrates appreciable differences in gastrointestinal absorption depending on whether the chemical is ingested from water, soil, or food. For some contaminants, the absorption of the contaminant from food can differ appreciably for plant compared with animal food products. Regardless of the apportionment approach used, EPA recommends using existing data on differences in bioavailability between water, air, soils, and different foods when estimating total exposure for use in apportioning the RfD or POD/ UF. The Agency has developed such exposure estimates for cadmium ( USEPA, 1994). In the absence of data, EPA will assume equal rates of absorption from different routes and sources of exposure. 4 16 4.2.4 Quantification of Exposure When selecting contaminant concentration values in environmental media and exposure intake values for the RSC analysis, it is important to realize that each value selected ( including those recommended as default assumptions in the AWQC equation) may be associated with a distribution of values for that parameter. Determining how various subgroups fall within the distributions of overall exposure and how the combination of exposure variables defines what population is being protected is a complicated and, perhaps, unmanageable task, depending on the amount of information available on each exposure factor included. Many times, the default assumptions used in EPA risk assessments are derived from the evaluation of numerous studies and are considered to generally represent a particular population group or a national average. Therefore, describing with certainty the exact percentile of a particular population that is protected with a resulting criteria is often not possible.

By and large, the AWQC are derived to protect the majority of the general population from chronic adverse health effects. However, as stated above in Section <IP_ADDRESS>, States and authorized Tribes are encouraged to consider protecting population groups that they determine are at greater risk and, thus, would be better protected using alternative exposure assumptions. The ultimate choice of the contaminant concentrations used in the RSC estimate and the exposure intake rates requires the use of professional judgment. This is discussed in greater detail in the Exposure Assessment TSD.

4.2.5 Inclusion of Inhalation and Dermal Exposures EPA intends to develop policy guidelines to apply to this Methodology for explicitly incorporating inhalation and dermal exposures. When estimating overall exposure to pollutants for AWQC development, EPA believes that the sources of inhalation and dermal exposures considered should include, on a case by case basis, both non oral exposures from water and other inhalation and dermal sources ( e. g., ambient or indoor air, soil). When the policy guidelines are completed, this Methodology will be refined to include that guidance.

A number of drinking water contaminants are volatile and thus diffuse from water into the air where they may be inhaled. In addition, drinking water is used for bathing and, thus, there is at least the possibility that some contaminants in water may be dermally absorbed. Volatilization may increase exposure via inhalation and decrease exposure via ingestion and dermal absorption. The net effect of volatilization and dermal absorption upon total exposure to volatile drinking water contaminants is unclear in some cases and varies from chemical to chemical. Dermal exposures are also important to consider for certain population groups, such as children and other groups with high soil contact.

With regard to additional non water related exposures, it is clear that the type and magnitude of toxicity produced via inhalation, ingestion, and dermal contact may differ; that is, the route of exposure can affect absorption of a chemical and can otherwise modify its toxicity. For example, an inhaled chemical such as hydrogen fluoride may produce localized effects on the 4 17 lung that are not observed ( or only observed at much higher doses) when the chemical is administered orally. Also, the active form of a chemical ( and principal toxicity) can be the parent compound and/ or one or more metabolites. With this Methodology, EPA recommends that differences in absorption and toxicity by different routes of exposure be determined and accounted for in dose estimates and applied to the exposure assessment. EPA acknowledges that the issue of whether the doses received from inhalation and ingestion exposures are cumulative ( i. e., toward the same threshold of toxicity) is complicated. Such a determination involves evaluating the chemical's physical characteristics, speciation, and reactivity. A chemical may also exhibit different metabolism by inhalation versus oral exposure and may not typically be metabolized by all tissues. In addition, a metabolite may be much more or much less toxic than the parent compound. Certainly with a systemic effect, if the chemical absorbed via different routes enters the bloodstream, then there is some likelihood that it will contact the same target organ. Attention also needs to be given to the fact that both the RfD and RfC are derived based on the administered level. Toxicologists generally believe that the effective concentration of the active form of a chemical( s) at the site( s) of action determines the toxicity. If specific differences between routes of exposure are not known, it may be reasonable to assume that the internal concentration at the site from any route contributes as much to the same effect as any other route. A default of assuming equal absorption has often been used. However, for many of the chemicals that the Agency has reviewed, there is a substantial amount of information already known to determine differences in rates of absorption. For example, absorption is, in part, a function of blood solubility ( i. e., Henry's Constant) and better estimations than the default can be made.

The RSC analyses that accompany the 2000 Human Health Methodology accommodate inclusion of inhalation exposures. Even if different target organs are involved between different routes of exposure, a conservative policy may be appropriate to keep all exposures below a certain level. A possible alternative is to set allowable levels ( via an equation) such that the total of ingestion exposures over the ingestion RfD added to the total of inhalation exposures over the inhalation RfC is not greater than 1 ( Note: the RfD is typically presented in mg/ kg day and the RfC is in mg/ m3). Again, EPA intends to develop guidance for this Methodology to explicitly incorporate inhalation and dermal exposures, and will refine the Methodology when that guidance is completed.

4.3 EXPOSURE FACTORS USED IN THE AWQC COMPUTATION This section presents values for the specific exposure factors that EPA will use in the derivation of AWQC. These include human body weight, drinking water consumption rates, and fish ingestion rates.

When choosing exposure factor values to include in the derivation of a criterion for a given pollutant, EPA recommends considering values that are relevant to population( s) that is ( are) most susceptible to that pollutant. In addition, highly exposed populations should be considered when setting criteria. In general, exposure factor values specific to adults and relevant to lifetime exposures are the most appropriate values to consider when determining criteria to protect against effects from long term exposure which, by and large, the human health criteria are 4 18 derived to protect. However, infants and children may have higher rates of water and food consumption per unit body weight compared with adults and also may be more susceptible to some pollutants than adults ( USEPA, 1997a). There may be instances where acute or subchronic developmental toxicity makes children the population group of concern. In addition, exposure of pregnant women to certain toxic chemicals may cause developmental effects in the fetus ( USEPA, 1997b). Exposures resulting in developmental effects may be of concern for some contaminants and should be considered along with information applicable to long term health effects when setting AWQC. ( See Section 3.2 for further discussion of this issue.) Short term exposure may include multiple intermittent or continuous exposures occurring over a week or so. Exposure factor values relevant for considering chronic toxicity, as well as exposure factor values relevant for short term exposure developmental concerns, that could result in adverse health effects are discussed in the sections below. In appropriate situations, EPA may consider developing criteria for developmental health effects based on exposure factor values specific to children or to women of childbearing age. EPA encourages States and Tribes to do the same when health risks are associated with short term exposures.

EPA believes that the recommended exposure factor default intakes for adults in chronic exposure situations are adequately protective of the population over a lifetime. In providing additional exposure intake values for highly exposed subpopulations ( e. g., sport anglers, subsistence fishers), EPA is providing flexibility for States and authorized Tribes to establish criteria specifically targeted to provide additional protection using adjusted values for exposure parameters for body weight, drinking water intake, and fish consumption. The exposure factor values provided for women of childbearing age and children would only be used in the circumstances indicated above.

Each of the following sections recommends exposure parameter values for use in developing AWQC. These are based on both science policy decisions that consider the best available data, as well as risk management judgments regarding the overall protection afforded by the choice in the derivation of AWQC. These will be used by EPA to derive new, or revise existing, 304( a) national criteria.

4.3.1 Human Body Weight Values for Dose Calculations The source of data for default human body weights used in deriving the AWQC is the third National Health and Nutrition Examination Survey ( NHANES III). NHANES III represents a very large interview and examination endeavor of the National Center for Health Statistics ( NCHS) and included participation from the Centers for Disease Control ( CDC). The NHANES III was conducted on a nationwide probability sample of over 30,000 persons from the civilian, non institutionalized population of the United States. The survey began in October 1988 and was completed in October 1994 ( WESTAT, 2000; McDowell, 2000). Body weight data were taken from the NHANES III Examination Data File. Sampling weights were applied to all persons examined in the Mobile Examination Centers ( MECs) or at home, as was recommended by the NHANES data analysts ( WESTAT, 2000). 4 19 The NHANES III survey has numerous strengths and very few weaknesses. Its primary strengths are the national representativeness, large sample size, and precise estimates due to this large sample size. Another strength is its high response rate; the examination rate was 73 percent overall, 89 percent for children under 1 year old, and approximately 85 percent for children 1 to 5 years old ( McDowell, 2000). Interview response rates were even higher, but the body weight data come from the NHANES examinations; that is, all body weights were carefully measured by survey staff, rather than the use of self reported body weights. The only significant potential weakness of the NHANES data is the fact that the data are now between 6 and 12 years old. Given that there were upward trends in body weight from NHANES II to NHANES III, and that NCHS has indicated the prevalence of overweight people increased in all age groups, the data could underestimate current body weights if that trend has continued ( WESTAT, 2000).

The NHANES III collected standard body measurements of sample subjects, including height and weight, that were made at various times of the day and in different seasons of the year. This technique was used because one's weight may vary between winter and summer and may fluctuate with recency of food and water intake and other daily activities ( McDowell, 2000).

As with the other exposure assumptions, States and authorized Tribes are encouraged to use alternative body weight assumptions for population groups other than the general population and to use local or regional data over default values as more representative of their target population group( s).

<IP_ADDRESS> Rate Protective of Human Health from Chronic Exposure EPA recommends maintaining the default body weight of 70 kg for calculating AWQC as a representative average value for both male and female adults. As previously indicated, exposure factor values specific to adults are recommended to protect against effects from longterm exposure. The value of 70 kg is based on the following information. In the analysis of the NHANES III database, median and mean values for female adults 18 74 years old are 65.8 and 69.5 kg, respectively ( WESTAT, 2000). For males in the same age range, the median and mean values are 79.9 and 82.1 kg, respectively. The mean body weight value for men and women ages 18 to 74 years old from this survey is 75.6 kg ( WESTAT, 2000). This mean value is higher than the mean value for adults ages 20 64 years old of 70.5 kg from a study by the National Cancer Institute ( NCI) which primarily measured drinking water intake ( Ershow and Cantor, 1989). The NCI study is described in the subsection on Drinking Water Intake Rates that follows ( Section 4.3.2). The value from the NHANES III database is also higher than the value given in the revised EPA Exposure Factors Handbook ( USEPA, 1997b), which recommends 71.8 kg for adults, based on the older NHANES II data. The Handbook also acknowledges the commonly used 70 kg value and encourages risk assessors to use values which most accurately reflect the exposed population. However, the point is also made that the 70 kg value is used in the derivation of cancer slope factors and unit risks that appear in IRIS. Consistency is advocated between the dose response relationship and exposure factors assumed. Therefore, if a value higher than 70 kg is used, the assessor needs to adjust the dose response relationship as described in the Appendix to Chapter 1, Volume 1 of the Handbook ( USEPA, 1997b). 4 20 <IP_ADDRESS> Rates Protective of Developmental Human Health Effects As noted above, pregnant women may represent a more appropriate population for which to assess risks from exposure to chemicals in ambient waters in some cases, because of the potential for developmental effects in fetuses. In these cases, body weights representative of women of childbearing age may be appropriate to adequately protect offspring from such health effects. To determine a mean body weight value appropriate to this population, separate body weight values for women in individual age groups within the range of 15 to 44 years old were analyzed from the NHANES III data ( WESTAT, 2000). The resulting median and mean body weight values are 63.2 and 67.3 kg, respectively. Ershow and Cantor ( 1989) present body weight values specifically for pregnant women included in the survey; median and mean weights are 64.4 and 65.8 kilograms, respectively. Ershow and Cantor ( 1989), however, do not indicate the ages of these pregnant women. Based on this information for women of childbearing age and pregnant women, EPA recommends use of a body weight value of 67 kg in cases where pregnant women are the specific population of concern and the chemical of concern exhibits reproductive and/ or developmental effects ( i. e., the critical effect upon which the RfD or POD/ UF is based). Using the 67 kg assumption would result in lower ( more protective) criteria than criteria based on 70 kg.

As discussed earlier, because infants and children generally have a higher rate of water and food consumption per unit body weight compared with adults, a higher intake rate per unit body weight may be needed when comparing estimated exposure doses with critical doses when RfDs are based on health effects in children. To calculate intake rates relevant to such effects, the body weight of children should be used. As with the default body weight for pregnant women, EPA is not recommending the development of additional AWQC ( i. e., similar to drinking water health advisories) that focus on acute or short term effects, since these are not seen routinely as having a meaningful role in the water quality criteria program. However, there may be circumstances where the consideration of exposures for these groups is warranted. Although the AWQC generally are based on chronic health effects data, they are intended to also be protective with respect to adverse effects that may reasonably be expected to occur as a result of elevated shorter term exposures. EPA acknowledges this as a potential course of action and is, therefore, recommending these default values which EPA would consider in an appropriate circumstance and for States and authorized Tribes to utilize in such situations.

EPA is recommending an assumption of 30 kg as a default child's body weight to calculate AWQC to provide additional protection for children when the chemical of concern indicates health effects in children are of predominant concern ( i. e., test results show children are more susceptible due to less developed immune systems, neurological systems, and/ or lower body weights). The value is based on the mean body weight value of 29.9 kg for children ages 1 to14 years old, which combines body weight values for individual age groups within this larger group. The mean value is based on body weight information from NHANES III for individual year age groups between one and 14 years old ( WESTAT, 2000). A mean body weight of 28 kg is obtained using body weight values from Ershow and Cantor ( 1989) for five age groups within this range of 0 14 years and applying a weighting method for different ages by population percentages from the U. S. Bureau of the Census. The 30 kg assumption is also consistent with the age range 4 21 for children used with the estimated fish intake rates. Unfortunately, fish intake rates for finer age group divisions are not possible due to the limited sampling base from the fish intake survey; there is limited confidence in calculated values ( e. g., the mean) for such fine age groups. Given this limitation, the broad age category of body weight for children is suitable for use with the default fish intake assumption.

Given the hierarchy of preferences regarding the use of fish intake information ( see Section 4.3.3), States may have more comprehensive data and prefer to target a more narrow, younger age group. If States choose to specifically evaluate toddlers, EPA recommends using 13 kg as a default body weight assumption for children ages 1 to 3 years old. The median and mean values of body weight for children 1 to 3 years old are 13.2 and 13.1 kg, respectively, based on an analysis of the NHANES III database ( WESTAT, 2000). The NHANES III median and mean values for females between 1 and 3 years old are 13.0 and 12.9 kg, respectively, and are 13.4 and 13.4 kg for males, respectively. Median and mean body weight values from the earlier Ershow and Cantor ( 1989) study for children ages 1 to 3 years old were 13.6 and 14.1 kg, respectively. Finally, if infants are specifically evaluated, EPA recommends a default body weight of 7 kg based on the NHANES III analysis. Median and mean body weights for both male and female infants ( combined) 2 months old were 6.3 and 6.3 kg, respectively, and for infants 3 months old were 7.0 and 6.9 kg, respectively. With the broader age category of males and females 2 to 6 months old, median and mean body weights were 7.4 and 7.4 kg, respectively. The NHANES analysis did not include infants under 2 months of age. Although EPA is not recommending body weight values for newborns, the NCHS National Vital Statistics Report indicates that, for 1997, the median birth weight ranged from 3 to 3.5 kg, according to WESTAT ( 2000).

Body weight values for individual ages within the larger range of 0 14 years are listed in the Exposure Assessment TSD for those States and authorized Tribes who wish to use body weight values for these individual groups. States and Tribes may wish to consider certain general developmental ages ( e. g., infants, pre adolescents, etc.), or certain specific developmental landmarks ( e. g., neurological development in the first four years), depending on the chemical of concern. EPA encourages States and authorized Tribes to choose a body weight intake from the tables presented in the TSD, if they believe a particular age subgroup is more appropriate.

4.3.2 Drinking Water Intake Rates The basis for the drinking water intake rates ( also for the fish intake rates presented in Section 4.3.3) is the 1994 96 Continuing Survey of Food Intake by Individuals ( CSFII) conducted by the U. S. Department of Agriculture ( USDA, 1998). The CSFII survey collects dietary intake information from nationally representative samples of non institutionalized persons residing in United States households. Households in these national surveys are sampled from the 50 states and the District of Columbia. Each survey collects daily consumption records for approximately 10,000 food codes across nine food groups. These food groups are ( 1) milk and milk products; ( 2) meat, poultry, and fish; ( 3) eggs; ( 4) dry beans, peas, legumes, nuts, and 4 22 seeds; ( 5) grain products; ( 6) fruit; ( 7) vegetables; ( 8) fats, oils, and salad dressings; and ( 9) sweets, sugars, and beverages. The survey also asks each respondent how many fluid ounces of plain drinking water he or she drank during each of the survey days. In addition, the CSFII collects household information, including the source of plain drinking water, water used to prepare beverages, and water used to prepare foods. Data provide " up to date information on food intakes by Americans for use in policy formation, regulation, program planning and evaluation, education, and research." The survey is " the cornerstone of the National Nutritional Monitoring and Related Research Program, a set of related federal activities intended to provide regular information on the nutritional status of the United States population" ( USDA, 1998).

The 1994 96 CSFII was conducted according to a stratified, multi area probability sample organized using estimates of the 1990 United States population. Stratification accounted for geographic location, degree of urbanization, and socioeconomics. Each year of the survey consisted of one sample with oversampling for low income households.

Survey participants provided two non consecutive, 24 hour days of dietary data. Both days' dietary recall information was collected by an in home interviewer. Interviewers provided participants with an instructional booklet and standard measuring cups and spoons to assist them in adequately describing the type and amount of food ingested. If the respondent referred to a cup or bowl in their own home, a 2 cup measuring cup was provided to aid in the calculation of the amount consumed. The sample person could fill their own bowl or cup with water to represent the amount eaten or drunk, and the interviewer could then measure the amount consumed by pouring it into the 2 cup measure. The Day 2 interview occurred three to 10 days after the Day 1 interview, but not on the same day of the week. The interviews allowed participants " three passes" through the daily intake record to maximize recall ( USDA, 1998). Proxy interviews were conducted for children aged six and younger and sampled individuals unable to report due to mental or physical limitations. The average questionnaire administration time for Day 1 intake was 30 minutes, while Day 2 averaged 27 minutes.

Two days of dietary recall data were provided by 15,303 individuals across the three survey years. This constitutes an overall two day response rate of 75.9 percent. Survey weights were corrected by the USDA for nonresponse.

All three 1994 96 CSFII surveys are multistage, stratified cluster samples. Sample weights, which project the data from a sampled individual to the population, are based on the probability of an individual being sampled at each stage of the sampling design. The sample weights associated with each individual reporting two days of consumption data were adjusted to correct for nonresponse bias.

The 1994 96 CSFII surveys have advantages and limitations for estimating per capita water ( or fish) consumption. The primary advantage of the CSFII surveys is that they were designed and conducted by the USDA to support unbiased estimation of food consumption across the population in the United States and the District of Columbia. Second, the survey is designed to record daily intakes of foods and nutrients and support estimation of food consumption. 4 23 One limitation of the 1994 96 CSFII surveys is that individual food consumption data were collected for only two days   a brief period which does not necessarily depict " usual intake." Usual dietary intake is defined as " the long run average of daily intakes by an individual." Upper percentile estimates may differ for short term and longer term data because short term food consumption data tend to be inherently more variable. It is important to note, however, that variability due to duration of the survey does not result in bias of estimates of overall mean consumption levels. Also, the multistage survey design does not support interval estimates for many of the subpopulations of interest because of sparse representation in the sample. Subpopulations with sparse representation include Native Americans on reservations and certain ethnic groups. While these individuals are participants in the survey, they are not present in sufficient numbers to support consumption estimates.

Despite these limitations, the CSFII is considered one of the best sources of current information on consumption of water and fish containing foods. The objective of estimating per capita water and fish consumption by the United States population is compatible with the statistical design and scope of the CSFII survey.

<IP_ADDRESS> Rate Protective of Human Health from Chronic Exposure EPA recommends maintaining the default drinking water intake rate of 2 L/ day to protect most consumers from contaminants in drinking water. EPA believes that the 2 L/ day assumption is representative of a majority of the population over the course of a lifetime. EPA also notes that there is comparatively little variability in water intake within the population compared with fish intake ( i. e., drinking water intake varies, by and large, by about a three fold range, whereas fish intake can vary by 100 fold). EPA believes that the 2 L/ day assumption continues to represent an appropriate risk management decision. The results of the 1994 96 CSFII analysis indicate that the arithmetic mean, 75th, and 90th percentile values for adults 20 years and older are 1.1, 1.5, and 2.2 L/ day, respectively ( USEPA, 2000a). The 2 L/ day value represents the 86th percentile for adults. These values can also be compared to data from an older National Cancer Institute ( NCI) study, which estimated intakes of tapwater in the United States based on the USDA's 1977 78 Nationwide Food Consumption Survey ( NFCS). The arithmetic mean, 75th, and 90th percentile values for adults 20  64 years old were 1.4, 1.7, and 2.3 L/ day, respectively ( Ershow and Cantor, 1989). The 2 L/ day value represents the 88th percentile for adults from the NCI study.

The 2 L/ day assumption was used with the original 1980 AWQC National Guidelines and has also been used in EPA's drinking water program. EPA believes that the newer studies continue to support the use of 2 L/ day as a reasonable and protective consumption rate that represents the intake of most water consumers in the general population. However, individuals who work or exercise in hot climates could have water consumption rates significantly above 2 L/ day, and EPA believes that States and Tribes should consider regional or occupational variations in water consumption. 4 24 <IP_ADDRESS> Rates Protective of Developmental Human Health Effects Based on the 1994 96 CSFII study data, EPA also recommends 2 L/ day for women of childbearing age. The analysis for women of childbearing age ( ages 15 44) indicate mean, 75th, and 90th percentile values of 0.9, 1.3, and 2.0 L/ day, respectively. These rates compare well with those based on an analysis of tapwater intake by pregnant and lactating women by Ershow et al. ( 1991), based on the older USDA data, for women ages 15 49. Arithmetic mean, 75th and 90th percentile values were 1.2, 1.5, and 2.2 L/ day, respectively, for pregnant women. For lactating women, the arithmetic mean, 75th and 90th percentile values were 1.3, 1.7, and 1.9 L/ day, respectively.

As noted above, because infants and children have a higher daily water intake per unit body weight compared with adults, a water consumption rate measured for children is recommended for use when RfDs are based on health effects in children. Use of this water consumption rate should result in adequate protection for infants and children when setting criteria based on health effects for this target population. EPA recommends a drinking water intake of 1 L/ day to, again, represent a majority of the population of children that consume drinking water. The results of the 1994 96 CSFII analysis indicate that for children from 1 to 10 years of age, the arithmetic mean, 75th, and 90th percentile values are 0.4, 0.6, and 0.9 L/ day, respectively ( USEPA, 2000a). The 1 L/ day value represents the 93rd percentile for this group. The arithmetic mean, 75th, and 90th percentile values for smaller children, ages 1 to 3 years, are 0.3, 0.5, and 0.7 L/ day, respectively. The 1 L/ day value represents the 97th percentile of the group ages 1 to 3 years old. For the category of infants under 1 year of age, the arithmetic mean, 75th, and 90th percentile values are 0.3, 0.7, and 0.9 L/ day, respectively. These data can similarly be compared to those of the older National Cancer Institute ( NCI) study. The arithmetic mean, 75th, and 90th percentile values for children 1 to 10 years old were 0.74, 0.96, and 1.3 L/ day, respectively. The mean, 75th, and 90th percentile values for children 1 to 3 years old in the NCI study were 0.6, 0.8, and 1.2 L/ day, respectively. Finally, the mean, 75th, and 90th percentile values for infants less than 6 months old were 0.3, 0.3, and 0.6 L/ day, respectively ( Ershow and Cantor, 1989).

<IP_ADDRESS> Rates Based on Combining Drinking Water Intake and Body Weight As an alternative to considering body weight and drinking water intake rates separately, EPA is providing rates based on intake per unit body weight data ( in units of ml/ kg) in the Exposure Assessment TSD, with additional discussion on their use. These rates are based on selfreported body weights from the CSFII survey respondents for the 1994 96 data. While EPA intends to derive or revise national default criteria on the separate intake values and body weights, in part due to the strong input received from its State stakeholders, the ml/ kg BW/ day values are provided in the TSD for States or authorized Tribes that prefer their use. It should be noted that in their 1993 review, EPA's Science Advisory Board ( SAB) felt that using drinking water intake rate assumptions on a per unit body weight basis would be more accurate, but did not believe this change would appreciably affect the criteria values ( USEPA, 1993). 4 25 4.3.3 Fish Intake Rates The basis for the fish intake rates is the 1994 96 CSFII conducted by the USDA, and described above in Section 4.3.2.

<IP_ADDRESS> Rates Protective of Human Health from Chronic Exposure EPA recommends a default fish intake rate of 17.5 grams/ day to adequately protect the general population of fish consumers, based on the 1994 to 1996 data from the USDA's CSFII Survey. EPA will use this value when deriving or revising its national 304( a) criteria. This value represents the 90th percentile of the 1994 96 CSFII data. This value also represents the uncooked weight estimated from the CSFII data, and represents intake of freshwater and estuarine finfish and shellfish only. For deriving AWQC, EPA has also considered the States' and Tribes' needs to provide adequate protection from adverse health effects to highly exposed populations such as recreational and subsistence fishers, in addition to the general population. Based on available studies that characterize consumers of fish, recreational fishers and subsistence fishers are two distinct groups whose intake rates may be greater than the general population. It is, therefore, EPA's decision to discuss intakes for these two groups, in addition to the general population.

EPA recommends default fish intake rates for recreational and subsistence fishers of 17.5 grams/ day and 142.4 grams/ day, respectively. These rates are also based on uncooked weights for fresh/ estuarine finfish and shellfish only. However, because the level of fish intake in highly exposed populations varies by geographical location, EPA suggests a four preference hierarchy for States and authorized Tribes to follow when deriving consumption rates that encourages use of the best local, State, or regional data available. A thorough discussion of the development of this policy method and relevant data sources is contained in the Exposure Assessment TSD. The hierarchy is also presented here because EPA strongly emphasizes that States and authorized Tribes should consider developing criteria to protect highly exposed population groups and use local or regional data over the default values as more representative of their target population group( s). The four preference hierarchy is: ( 1) use of local data; ( 2) use of data reflecting similar geography/ population groups; ( 3) use of data from national surveys; and ( 4) use of EPA's default intake rates.

The recommended four preference hierarchy is intended for use in evaluating fish intake from fresh and estuarine species only. Therefore, to protect humans who additionally consume marine species of fish, the marine portion should be considered an other source of exposure when calculating an RSC for dietary intake. Refer to the Exposure Assessment TSD for further discussion. States and Tribes need to ensure that when evaluating overall exposure to a contaminant, marine fish intake is not double counted with the other dietary intake estimate used. Coastal States and authorized Tribes that believe accounting for total fish consumption ( i. e., fresh/ estuarine and marine species) is more appropriate for protecting the population of concern may do so, provided that the marine intake component is not double counted with the RSC estimate. Tables of fish consumption intakes based on the CSFII in the TSD provide rates for fresh/ estuarine species, marine species, and total ( combined) values to facilitate this option for 4 26 States and Tribes. Throughout this section, the terms " fish intake" or " fish consumption" are used. These terms refer to the consumption of finfish and shellfish, and the CSFII survey includes both. States and Tribes should ensure that when selecting local or regionally specific studies, both finfish and shellfish are included when the population exposed are consumers of both types.

EPA's first preference is that States and authorized Tribes use the results from fish intake surveys of local watersheds within the State or Tribal jurisdiction to establish fish intake rates that are representative of the defined populations being addressed for the particular waterbody. Again, EPA recommends that data indicative of fresh/ estuarine species only be used which is, by and large, most appropriate for developing AWQC. EPA also recommends the use of uncooked weight intake values, which is discussed in greater detail with the fourth preference. States and authorized Tribes may use either high end values ( such as the 90th or 95th percentile values) or average values for an identified population that they plan to protect ( e. g., subsistence fishers, sport fishers, or the general population). EPA generally recommends that arithmetic mean values should be the lowest value considered by States or Tribes when choosing intake rates for use in criteria derivation. When considering geometric mean ( median) values from fish consumption studies, States and authorized Tribes need to ensure that the distribution is based on survey respondents who reported consuming fish because surveys based on both consumers and nonconsumers can often result in median values of zero. If a State or Tribe chooses values ( whether the central tendency or high end values) from studies that particularly target high end consumers, these values should be compared to high end fish intake rates for the general population to make sure that the high end consumers within the general population would be protected by the chosen intake rates. EPA believes this is a reasonable procedure and is also consistent with the recent Great Lakes Water Quality Initiative ( known as the " GLI") ( USEPA, 1995). States and authorized Tribes may wish to conduct their own surveys of fish intake, and EPA guidance is available on methods to conduct such studies in Guidance for Conducting Fish and Wildlife Consumption Surveys ( USEPA, 1998). Results from broader geographic regions in which the State or Tribe is located can also be used, but may not be as applicable as results from local watersheds. Since such studies would ultimately form the basis of a State or Tribe's AWQC, EPA would review any surveys of fish intake for consistency with the principles of EPA's guidance as part of the Agency's review of water quality standards under Section 303( c).

If surveys conducted in the geographic area of the State or Tribe are not available, EPA's second preference is that States and authorized Tribes consider results from existing fish intake surveys that reflect similar geography and population groups ( e. g., from a neighboring State or Tribe or a similar watershed type), and follow the method described above regarding target values to derive a fish intake rate. Again, EPA recommends the use of uncooked weight intake values and the use of fresh/ estuarine species data only. Results of existing local and regional surveys are discussed in greater detail in the TSD.

If applicable consumption rates are not available from local, State, or regional surveys, EPA's third preference is that States and authorized Tribes select intake rate assumptions for different population groups from national food consumption surveys. EPA has analyzed one such 4 27 national survey, the 1994 96 CSFII. As described in Section 4.3.2, this survey, conducted annually by the USDA, collects food consumption information from a probability sample of the population of all 50 states. Respondents to the survey provide two days of dietary recall data. A detailed description of the combined 1994 96 CSFII survey, the statistical methodology, and the results and uncertainties of the EPA analyses are provided in a separate EPA report ( USEPA, 2000b). The Exposure Assessment TSD for this Methodology presents selected results from this report including point and interval estimates of combined finfish and shellfish consumption for the mean, 50th ( median), 90th, 95th, and 99th percentiles. The estimated fish consumption rates are by fish habitat ( i. e., freshwater/ estuarine, marine and all habitats) for the following population groups: ( 1) all individuals; ( 2) individuals age 18 and over; ( 3) women ages 15 44; and ( 4) children age 14 and under. Three kinds of estimated fish consumption rates are provided: ( 1) per capita rates ( i. e., rates based on consumers and nonconsumers of fish from the survey period   refer to the TSD for further discussion); ( 2) consumers only rates ( i. e., rates based on respondents who reported consuming finfish or shellfish during the two day reporting period); and ( 3) per capita consumption by body weight ( i. e., per capita rates reported as milligrams of fish per kilogram of body weight per day).

EPA's fourth preference is that States and authorized Tribes use as fish intake assumptions the following default rates, based on the 1994 96 CSFII data, that EPA believes are representative of fish intake for different population groups: 17.5 grams/ day for the general adult population and sport fishers, and 142.4 grams/ day for subsistence fishers. These are risk management decisions that EPA has made after evaluating numerous fish intake surveys. These values represent the uncooked weight intake of freshwater/ estuarine finfish and shellfish. As with the other preferences, EPA requests that States and authorized Tribes routinely consider whether there is a substantial population of sport fishers or subsistence fishers when developing sitespecific estimates, rather than automatically basing them on the typical individual. Because the combined 1994 96 CSFII survey is national in scope, EPA will use the results from this survey to estimate fish intake for deriving national criteria. EPA has recognized the data gaps and uncertainties associated with the analysis of the 1994 96 CSFII survey in the process of making its default recommendations. The estimated mean of freshwater and estuarine fish ingestion for adults is 7.50 grams/ day, and the median is 0 grams/ day. The estimated 90th percentile is 17.53 grams/ day; the estimated 95th percentile is 49.59 grams/ day; and the estimated 99th percentile is 142.41 grams/ day. The median value of 0 grams/ day may reflect the portion of individuals in the population who never eat fish as well as the limited reporting period ( 2 days) over which intake was measured. By applying as a default 17.5 grams/ day for the general adult population, EPA intends to select an intake rate that is protective of a majority of the population ( again, the 90th percentile of consumers and nonconsumers according to the 1994 96 CSFII survey data). Trophic level breakouts are: TL2 = 3.8 grams/ day; TL3 = 8.0 grams/ day; and TL4 = 5.7 grams/ day. EPA further considers 17.5 grams/ day to be indicative of the average consumption among sport fishers based on averages in the studies reviewed, which are presented in the Exposure Assessment TSD. Similarly, EPA believes that the assumption of 142.4 grams/ day is within the range of average consumption estimates for subsistence fishers based on the studies reviewed. Experts at the 1992 National Workshop that initiated the effort to revise this Methodology acknowledged that the national survey high end values are representative of 4 28 average rates for highly exposed groups such as subsistence fishermen, specific ethnic groups, or other highly exposed people. EPA is aware that some local and regional studies indicate greater consumption among Native American, Pacific Asian American, and other subsistence consumers, and recommends the use of those studies in appropriate cases, as indicated by the first and second preferences. Again, States and authorized Tribes have the flexibility to choose intake rates higher than an average value for these population groups. If a State or authorized Tribe has not identified a separate well defined population of high end consumers and believes that the national data from the 1994 96 CSFII are representative, they may choose these recommended rates.

As indicated above, the default intake values are based on the uncooked weights of the fish analyzed. There has been some question regarding whether to use cooked or uncooked weights of fish intake for deriving the AWQC. Studies show that, typically, with a filet or steak of fish, the weight loss in cooking is about 20 percent; that is, the uncooked weight is approximately 20 percent higher ( Jacobs et al., 1998). This obviously means that using uncooked weights results in a slightly higher intake rate and slightly more stringent AWQC. In researching consumption surveys for this proposal, EPA has found that some surveys have reported rates for cooked fish, others have reported uncooked rates, and many more are unclear as to whether cooked or uncooked rates are used. The basis of the CSFII survey was prepared or as consumed intakes; that is, the survey respondents estimated the weight of fish that they consumed. This was also true with the GLI ( which was specifically based on studies describing consumption rates of cooked fish) and, by and large, cooked fish is what people consume. However, EPA's Guidance For Assessing Chemical Contaminant Data For Use In Fish Advisories recommends analysis and advisories based on uncooked fish ( USEPA, 1997a). EPA considered the potential confusion over the fact that the uncooked weights are used in the fish advisory program. Further, the measures of a contaminant in fish tissue samples that are applicable to compliance monitoring and the permitting program are related to the uncooked weights. The choice of intakes is also complicated by factors such as the effect of the cooking process, the different parts of a fish where a chemical may accumulate, and the method of preparation.

After considering all of the above ( in addition to public input received), EPA will derive its national default criteria based on the uncooked weight fish intakes. The Exposure Assessment TSD provides additional guidance on site specific modifications. Specifically, an alternate approach is described for calculating AWQC with the as consumed weight   which is more directly associated with human exposure and risk   and then adjusting the value by the approximate 20 percent loss to an uncooked equivalent ( thereby representing the same relative risk as the as consumed value). This approach results in a different AWQC value ( than using the uncooked weights) and represents a more direct translation of the as consumed risk to the uncooked equivalent. However, EPA understands that it is more scientifically rigorous and may be too intensive of a process for States and Tribes to rely on. The option is presented in the TSD to offer States and authorized Tribes greater flexibility with their water quality standards program.

The default fish intake values also reflect specific designations of species classified in accordance with information regarding the life history of the species or based on landings information form the National Marine Fisheries Service. Most significantly, salmon has been 4 29 reclassified from a freshwater/ estuarine species to a marine species. As marine harvested salmon represents approximately 99 percent of salmon consumption in the 1994 96 CSFII Survey, removal reduces the overall fresh/ estuarine fish consumption rate by 13 percent. Although they represent a very small percentage of freshwater/ estuarine intake, land locked and farm raised salmon consumed by 1994 96 CSFII respondents are still included. The rationale for the default intake species designations is explained in the Exposure Assessment TSD. Once again, EPA emphasizes the flexibility for States and authorized Tribes to use alternative assumptions based on local or regional data to better represent their population groups of concern.

<IP_ADDRESS> Rates Protective of Developmental Human Health Effects Exposures resulting in health effects in children or developmental effects in fetuses may be of primary concern. As discussed at the beginning of this section on exposure factors used, in a situation where acute or sub chronic toxicity and exposure are the basis of an RfD ( or POD/ UF), EPA will consider basing its national default criteria on children or women of childbearing age, depending on the target population at greatest risk. EPA recommends that States and authorized Tribes use exposure factors for children or women of childbearing age in these situations. As stated previously, EPA is not recommending the development of additional AWQC but is acknowledging that basing a criterion on these population groups is a potential course of action and is, therefore, recommending the following default intake rates for such situations.

EPA's preferences for States and authorized Tribes in selecting values for intake rates relevant for children is the same as that discussed above for establishing values for average daily consumption rates for chronic effects; i. e., in decreasing order of preference, results from fish intake surveys of local watersheds, results from existing fish intake surveys that reflect similar geography and population groups, the distribution of intake rates from nationally based surveys ( e. g., the CSFII), or lastly, the EPA default rates. When an RfD is based on health effects in children, EPA recommends a default intake rate of 156.3 grams/ day for assessing those contaminants that exhibit adverse effects. This represents the 90th percentile consumption rate for actual consumers of freshwater/ estuarine finfish and shellfish for children ages 14 and under using the combined 1994 to 1996 results from the CSFII survey. The value was calculated based on data for only those children who ate fish during the 2 day survey period, and the intake was averaged over the number of days during which fish was actually consumed. EPA believes that by selecting the data for consumers only, the 90th percentile is a reasonable intake rate to approximate consumption of fresh/ estuarine finfish and shellfish within a short period of time for use in assessments where adverse effects in children are of primary concern. As discussed previously, EPA will use a default body weight of 30 kg to address potential acute or subchronic effects from fish consumption by children. EPA is also providing these default intake values for States and authorized Tribes that choose to provide additional protection when developing criteria that they believe should be based on health effects in children. This is consistent with the rationale in the recent GLI ( USEPA, 1995) and is an approach that EPA believes is reasonable. Distributional information on intake values relevant for assessing exposure when health effects to children are of concern is presented in the Exposure Assessment TSD. 4 30 There are also cases in which pregnant women may be the population of most concern, due to the possibility of developmental effects that may result from exposures of the mother to toxicants. In these cases, fish intake rates specific to females of childbearing age are most appropriate when assessing exposures to developmental toxicants. When an RfD is based on developmental toxicity, EPA proposes a default intake rate of 165.5 grams/ day for assessing exposures for women of childbearing age from contaminants that cause developmental effects. This is equivalent to the 90th percentile consumption rate for actual consumers of freshwater/ estuarine finfish and shellfish for women ages 15 to 44 using the combined 1994 to1996 results from the CSFII survey. As with the rate for children, this value represents only those women who ate fish during the 2 day survey period. As discussed previously, EPA will use a default body weight of 67 kg for women of childbearing age.

<IP_ADDRESS> Rates Based on Combining Fish Intake and Body Weight As with the drinking water intake values, EPA is providing values for fish intake based on a per unit body weight basis ( in units of mg/ kg) in the Exposure Assessment TSD. These rates use the self reported body weights of the 1994 96 CSFII survey. Again, while EPA intends to derive or revise national default criteria on the separate intake values and body weights, the mg/ kg BW/ day values are provided in the TSD for States or authorized Tribes that prefer their use.

4.4 REFERENCES FOR EXPOSURE Ershow A. G., Brown L. M. and Cantor K. P. 1991. Intake of tapwater and total water by pregnant and lactating women. Am. J. Public Health. 81: 328 334.

Ershow A. G. and K. P. Cantor. 1989. Total Water and Tap Water Intake in the United States: Population based Estimates of Quantities and Sources. National Cancer Institute. Bethesda, MD. Order # 263 MD 810264.

Gibbons, J. D. 1971. Nonparametric Statistical Inference. Chapter 2: Order Statistics. McGraw Hill, Inc. New York, NY.

Jacobs, H. L., H. D. Kahn, K. A. Stralka, and D. B. Phan. 1998. Estimates of per capita fish consumption in the U. S. based on the continuing survey of food intake by individuals ( CSFII). Risk Analysis: An International Journal 18( 3).

McDowell, M. 2000. Personal communication between Denis R. Borum, U. S. Environmental Protection Agency, and Margaret McDowell, Health Statistician, National Health and Nutrition Examination Survey, National Center for Health Statistics. March 24, 2000.

USDA. 1998. U. S. Department of Agriculture. 1994   1996 Continuing Survey of Food Intakes by Individuals and 1994   1996 Diet and Health Knowledge Survey. Agricultural Research Service, USDA. NTIS CD   ROM, accession number PB98   500457. 4 31 [ Available from the National Technical Information Service, 5285 Port Royal Road, Springfield, VA 22161. Phone: ( 703) 487   4650.]

USEPA. 1993. Review of the Methodology for Developing Ambient Water Quality Criteria for the Protection of Human Health. Prepared by the Drinking Water Committee of the Science Advisory Board. EPA SAB DWC USEPA. 1994. Reference dose ( RfD) for oral exposure for cadmium. Integrated Risk Information System ( IRIS). Online. ( Verification date 02/ 01/ 94.) Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office. Cincinnati, OH.

USEPA. 1995. Great Lakes Water Quality Initiative Technical Support Document for the Procedure to Determine Bioaccumulation Factors. Office of Water. Washington, DC. EPA/ 820/ B 95/ 005.

USEPA. 1997a. Guidance for Assessing Chemical Contaminant Data for Use in Fish Advisories. Volume II: Risk Assessment and Fish Consumption Limits. Second Edition. Office of Water. Washington DC. EPA/ 823/ B 97/ 009.

USEPA. 1997b. Exposure Factors Handbook. National Center for Environmental Assessment, Office of Research and Development. Washington, DC. EPA/ 600/ P 95/ 002Fa. August.

USEPA. 1998. Guidance for Conducting Fish and Wildlife Consumption Surveys. Office of Science and Technology, Office of Water. Washington, DC. EPA 823 B 98 007. November.

USEPA. 2000a. Estimated Per Capita Water Ingestion in the United States: Based on Data Collected by the United States Department of Agriculture's 1994 96 Continuing Survey of Food Intakes by Individuals. Office of Science and Technology, Office of Water. Washington, DC. EPA 822 00 008. April.

USEPA. 2000b. Estimated Per Capita Fish Consumption in the United States: Based on Data Collected by the United States Department of Agriculture's 1994 1996 Continuing Survey of Food Intake by Individuals. Office of Science and Technology, Office of Water, Washington, DC. March.

WESTAT. 2000. Memorandum on Body Weight Estimates Based on NHANES III data, Including Data Tables and Graphs. Analysis conducted and prepared by WESTAT, under EPA Contract No. 68 C 99 242. March 3, 2000. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health ( 2000)

Chapter 5 5. BIOACCUMULATION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 1 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 1 5.2 Definitions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 6 5.3 Framework for Determining National Bioaccumulation Factors . . . . . . . . . . . . . . . 5 10 5.4 National Bioaccumulation Factors for Nonionic Organic Chemicals . . . . . . . . . . . . . 5 16 5.5 National Bioaccumulation Factors for Ionic Organic Chemicals . . . . . . . . . . . . . . . . 5 55 5.6 National Bioaccumulation Factors for Inorganic and Organometallic Chemicals . . . . 5 57 5.7 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 63 TABLES AND FIGURES Figure 5 1 Framework for Deriving a National BAF . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 13 Figure 5 2 BAF Derivation for Nonionic Organic Chemicals . . . . . . . . . . . . . . . . . . . . . 5 17 Table 5 1 Food Chain Multipliers for Trophic Levels 2, 3 and 4 . . . . . . . . . . . . . . . . . 5 36 5 1 ( Equation 5 1) 5. BIOACCUMULATION 5.1 INTRODUCTION Aquatic organisms can accumulate certain chemicals in their bodies when exposed to these chemicals through water, their diet, and other sources. This process is called bioaccumulation. The magnitude of bioaccumulation by aquatic organisms varies widely depending on the chemical but can be extremely high for some highly persistent and hydrophobic chemicals. For such highly bioaccumulative chemicals, concentrations in aquatic organisms may pose unacceptable human health risks from fish and shellfish consumption even when concentrations in water are too low to cause unacceptable health risks from drinking water consumption alone. These chemicals may also biomagnify in aquatic food webs, a process whereby chemical concentrations increase in aquatic organisms of each successive trophic level due to increasing dietary exposures ( e. g., increasing concentrations from algae, to zooplankton, to forage fish, to predatory fish).

In order to prevent harmful exposures to waterborne chemicals through the consumption of contaminated fish and shellfish, national 304( a) water quality criteria for the protection of human health must address the process of chemical bioaccumulation in aquatic organisms. For deriving national 304( a) criteria to protect human health, EPA accounts for potential bioaccumulation of chemicals in fish and shellfish through the use of national bioaccumulation factors ( BAFs). A national BAF is a ratio ( in L/ kg) that relates the concentration of a chemical in water to its expected concentration in commonly consumed aquatic organisms in a specified trophic level. An illustration of how national BAFs are used in the derivation of 304( a) criteria for carcinogens using linear low dose extrapolation is shown in the following equation:

where:

RSD = Risk specific dose ( mg/ kg day) BW = Human body weight ( kg) DI = Drinking water intake ( L/ day) FI i = Fish intake at trophic level I, where I= 2, 3, and 4; BAF i = National bioaccumulation factor at trophic level I, where I= 2, 3, and 4 The purpose of this chapter is to present EPA's recommended methodology for deriving national bioaccumulation factors for setting national 304( a) water quality criteria to protect human health. A detailed scientific basis of the recommended national BAF methodology is provided in the Bioaccumulation TSD. While the methodology detailed in this chapter is 5 2 intended to be used by EPA for deriving national BAFs, EPA encourages States and authorized Tribes to derive BAFs that are specific to certain regions or waterbodies, where appropriate. Guidance to States and authorized Tribes for deriving site specific BAFs is provided in the Biaccumulation TSD.

5.1.1 Important Bioaccumulation and Bioconcentration Concepts Several attributes of the bioaccumulation process are important to understand when deriving national BAFs for use in setting national 304( a) criteria. First, the term " bioaccumulation" refers to the uptake and retention of a chemical by an aquatic organism from all surrounding media ( e. g., water, food, sediment). The term " bioconcentration" refers to the uptake and retention of a chemical by an aquatic organism from water only. For some chemicals ( particularly those that are highly persistent and hydrophobic), the magnitude of bioaccumulation by aquatic organisms can be substantially greater than the magnitude of bioconcentration. Thus, an assessment of bioconcentration alone would underestimate the extent of accumulation in aquatic biota for these chemicals. Accordingly, EPA's guidelines presented in this chapter emphasize the measurement of chemical bioaccumulation by aquatic organisms, whereas EPA's 1980 Methodology emphasized the measurement of bioconcentration.

Another noteworthy aspect of bioaccumulation process is the issue of steady state conditions. Specifically, both bioaccumulation and bioconcentration can be viewed simply as the result of competing rates of chemical uptake and depuration ( chemical loss) by an aquatic organism. The rates of chemical uptake and depuration can be affected by various factors including the properties of the chemical, the physiology of the organism in question, water quality and other environmental conditions, ecological characteristics of the waterbody ( e. g., food web structure), and the concentration and loadings history of the chemical. When the rates of chemical uptake and depuration are equal, tissue concentrations remain constant over time and the distribution of the chemical between the organism and its source( s) is said to be at steady state. For constant chemical exposures and other conditions, the steady state concentration in the organism represents the highest accumulation potential of the chemical in that organism under those conditions. The time required for a chemical to achieve steady state has been shown to vary according to the properties of the chemical and other factors. For example, some highly hydrophobic chemicals can require long periods of time to reach steady state between environmental compartments ( e. g., many months), while highly hydrophilic chemicals usually reach steady state relatively quickly ( e. g., hours to days).

Since national 304( a) criteria for the protection of human health are typically designed to protect humans from harmful lifetime or long term exposures to waterborne contaminants, the assessment of bioaccumulation that equals or approximates steady state accumulation is one of the principles underlying the derivation of national BAFs. For some chemicals that require relatively long periods of time to reach steady state in tissues of aquatic organisms, changes in water column concentrations may occur on a much more rapid time scale compared to the corresponding changes in tissue concentrations. Thus, if the system departs substantially from steady state conditions and water concentrations are not averaged over a sufficient time period, 5 3 the ratio of the tissue concentration to a water concentration may have little resemblance to the steady state ratio and have little predictive value of long term bioaccumulation potential. Therefore, BAF measurements should be based on water column concentrations which are averaged over a sufficient period of time ( e. g., a duration comparable to the time required for the chemical to reach steady state). In addition, BAF measurements should be based on adequate spatial averaging of both tissue and water column concentrations for use in deriving 304( a) criteria for the protection of human health.

For this reason, a BAF is defined in this Methodology as representing the ratio ( in L/ kgtissue of a concentration of a chemical in tissue to its concentration in the surrounding water in situations where the organism and its food are exposed and the ratio does not change substantially over time ( i. e., the ratio which reflects bioaccumulation at or near steady state). A bioconcentration factor ( BCF) is the ratio ( in L/ kg tissue) of the concentration of a substance in tissue of an aquatic organism to its concentration in the ambient water, in situations where the organism is exposed through the water only and the ratio does not change substantially over time.

5.1.2 Goal of the National BAF The goal of EPA's national BAF is to represent the long term, average bioaccumulation potential of a chemical in edible tissues of aquatic organisms that are commonly consumed by humans throughout the United States. National BAFs are not intended to reflect fluctuations in bioaccumulation over short time periods ( e. g., a few days) because 304( a) human health criteria are generally designed to protect humans from long term exposures to waterborne chemicals. National BAFs are also intended to account for some major chemical, biological, and ecological attributes that can affect bioaccumulation in bodies of water across the United States. For example, separate procedures are provided for deriving national BAFs depending on the type of chemical ( i. e., nonionic organic, ionic organic, inorganic and organometallic). In addition, EPA's national BAFs are derived separately for each trophic level to account for potential biomagnification of some chemicals in aquatic food webs and broad physiological differences between trophic levels that may influence bioaccumulation. Because lipid content of aquatic organisms and the amount of organic carbon in the water column have been shown to affect bioaccumulation of nonionic organic chemicals, EPA's national BAFs are adjusted to reflect the lipid content of commonly consumed fish and shellfish and the freely dissolved fraction of the chemical in ambient water for these chemicals.

5.1.3 Changes to the 1980 Methodology Numerous scientific advances have occurred in the area of bioaccumulation since the publication of the 1980 Methodology for deriving AWQC for the protection of human health ( USEPA, 1980). These advances have significantly increased our ability to assess and predict the bioaccumulation of chemicals in aquatic biota. As a result, EPA has revised the bioaccumulation portion of the 1980 Methodology to reflect the current state of the science and to improve accuracy in assessing bioaccumulation for setting 304( a) criteria for the protection of 5 4 human health. The changes contained in the bioaccumulation portion of the 2000 Human Health Methodology are mostly designed to:

C Improve the ability to incorporate chemical exposure from sediments and aquatic food webs in assessing bioaccumulation potential,

C Expand the ability to account for site specific factors which affect bioaccumulation, and C Incorporate new data and assessment tools into the bioaccumulation assessment process.

A summary of the key changes that have been incorporated into the bioaccumulation portion of the 2000 Human Health Methodology and appropriate comparisons to the1980 Methodology are provided below.

<IP_ADDRESS> Overall Approach The 1980 Methodology for deriving 304( a) criteria for the protection of human health emphasized the assessment of bioconcentration ( uptake from water only) through the use of the BCF. Based on the 1980 Methodology, measured BCFs were usually determined from laboratory data unless field data demonstrated consistently higher or lower accumulation compared with laboratory data. In these cases, " field BCFs" ( currently termed field measured BAFs) were recommended for use. For lipophilic chemicals where lab or field measured data were unavailable, EPA recommended predicting BCFs from the octanol water partition coefficient and the following equation from Veith et al. ( 1979): " log BCF = ( 0.85 log K ow)  0.70".

The 2000 Human Health Methodology revisions contained in this chapter emphasize the measurement of bioaccumulation ( uptake from water, sediment, and diet) through the use of the BAF. Consistent with the 1980 Methodology, measured data are preferred over predictive approaches for determining the BAF ( i. e., field measured BAFs are generally preferred over predicted BAFs). However, the 2000 Human Health Methodology contains additional methods for deriving a national BAF that were not available in 1980. The preference for using the BAF methods also differs depending on the type and properties of the chemical. For example, the BAF derivation procedure differs for each of three broadly defined chemical categories: ( 1) nonionic organic, ( 2) ionic organic, and ( 3) inorganic and organometallic chemicals. Furthermore, within the category of nonionic organic chemicals, different procedures are used to derive the BAF depending on a chemicals' hydrophobicity and extent of chemical metabolism that would be expected to occur in aquatic biota.

<IP_ADDRESS> Lipid Normalization In the 1980 Methodology, BCFs for lipophilic chemicals were normalized by the lipid fraction in the tissue of fish and shellfish used to determine the BCF. Lipid normalization enabled BCFs to be averaged across tissues and organisms. Once the average lipid normalized 5 5 BCF was determined, it was adjusted by the consumption weighted lipid content of commonly consumed aquatic organisms in the United States to obtain an overall consumption weighted BCF. A similar procedure has been retained in the 2000 Human Health Methodology, whereby BAFs for nonionic organic chemicals are lipid normalized and adjusted by the consumptionweighted lipid content of commonly consumed organisms to obtain a BAF for criteria calculations. However, the 2000 Human Health Methodology uses more up to date lipid data and consumption data for deriving the consumption weighted BAFs.

<IP_ADDRESS> Bioavailability Bioconcentration factors derived according to the 1980 Methodology were based on the total concentration of the chemical in water, for both lipophilic and nonlipophilic chemicals. In the 2000 Human Health Methodology, BAFs for nonionic organic chemicals are derived using the most bioavailable fraction ( i. e., the freely dissolved fraction) to account for the influence of particulate and dissolved organic carbon on a chemical's bioavailability. Such BAFs are then adjusted to reflect the expected bioavailability at the sites of interest ( i. e., by adjusting for organic carbon concentrations at the sites of interest). Procedures for accounting for the effect of organic carbon on bioaccumulation were published previously by EPA under the Great Lakes Water Quality Initiative ( GLWQI or GLI) rulemaking ( USEPA, 1995a, b). Bioavailability is also considered in developing BAFs for the other chemical classes defined in the 2000 Human Health Methodology ( e. g., ionic organics, inorganics/ organometallics) but is done so on a chemical bychemical basis.

<IP_ADDRESS> Trophic Level Considerations In the 1980 Methodology, BCFs were determined and used for criteria derivation without explicit regard to the trophic level of the aquatic organism ( e. g., benthic filter feeder, forage fish, predatory fish). Over the past two decades, much information has been assembled which demonstrates that an organism's trophic position in the aquatic food web can have an important effect on the magnitude of bioaccumulation of certain chemicals. In order to account for the variation in bioaccumulation that is due to trophic position of the organism, the 2000 Human Health Methodology recommends that BAFs be determined and applied on a trophic level specific basis.

<IP_ADDRESS> Site Specific Adjustments The 1980 Methodology contained little guidance for making adjustments to the national BCFs to reflect site or region specific conditions. The 2000 Human Health Methodology has greatly expanded the guidance to States and authorized Tribes for making adjustments to national BAFs to reflect local conditions. This guidance is contained in the Bioaccumulation TSD. In the Bioaccumulation TSD, guidance and data are provided for adjusting national BAFs to reflect the lipid content in locally consumed aquatic biota and the organic carbon content in the waterbodies of concern. This guidance also allows the use of appropriate bioaccumulation models for deriving site specific BAFs. EPA also plans to publish detailed guidance on designing and conducting field 5 6 ( Equation 5 2) bioaccumulation studies for measuring BAFs and biota sediment accumulation factors ( BSAFs). In general, EPA encourages States and authorized Tribes to make site specific modifications to EPA's national BAFs provided such adjustments are scientifically defensible and adequately protect the designated use of the waterbody.

While the aforementioned revisions are new to EPA's Methodology for deriving national 304( a) criteria for the protection of human health, many of these refinements have been incorporated in prior Agency guidance and regulations. For example, the use of food chain multipliers to account for the biomagnification of nonionic organic chemicals in aquatic food webs when measured data are unavailable was introduced by EPA in three documents: Technical Support Document for Water Quality Based Toxics Control ( USEPA, 1991), a draft document entitled Assessment and Control of Bioconcentratable Contaminants in Surface Waters ( USEPA, 1993), and in the Great Lakes Water Quality Initiative ( GLI) ( USEPA, 1995b). Similarly, procedures for predicting BAFs using BSAFsand incorporating the effect of organic carbon on bioavailability were used to derive water quality criteria under the GLI.

5.1.4 Organization of This Section The methodology for deriving national BAFs for use in deriving National 304( a) Human Health AWQC is provided in the following sections. Important terms used throughout this chapter are defined in Section 5.2. Section 5.3 provides an overview of the BAF derivation guidelines. Detailed procedures for deriving national BAFs are provided in Section 5.4 for nonionic organic chemicals, in Section 5.5 for ionic organic chemicals, and in Section 5.6 for inorganics and organometallic chemicals. Literature cited is provided in Section 5.7.

5.2 DEFINITIONS The following terms and definitions are used throughout this chapter.

Bioaccumulation. The net accumulation of a substance by an organism as a result of uptake from all environmental sources.

Bioconcentration. The net accumulation of a substance by an aquatic organism as a result of uptake directly from the ambient water, through gill membranes or other external body surfaces.

Bioaccumulation Factor ( BAF). The ratio ( in L/ kg tissue) of the concentration of a substance in tissue to its concentration in the ambient water, in situations where both the organism and its food are exposed and the ratio does not change substantially over time. The BAF is calculated as: 5 7 ( Equation 5 3)

( Equation 5 4) where:

C t = Concentration of the chemical in the specified wet tissue C w = Concentration of chemical in water Bioconcentration Factor ( BCF). The ratio ( in L/ kg tissue) of the concentration of a substance in tissue of an aquatic organism to its concentration in the ambient water, in situations where the organism is exposed through the water only and the ratio does not change substantially over time. The BCF is calculated as:

where:

C t = Concentration of the chemical in the specified wet tissue C w = Concentration of chemical in water Baseline BAF ( BAF R f d). For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), a BAF ( in L/ kg lipid) that is based on the concentration of freely dissolved chemical in the ambient water and the lipid normalized concentration in tissue.

Baseline BCF ( BCF R f d). For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), a BCF ( in L/ kg lipid) that is based on the concentration of freely dissolved chemical in the ambient water and the lipid normalized concentration in tissue.

Biomagnification. The increase in tissue concentration of a chemical in organisms at successive trophic levels through a series of predator prey associations, primarily through the mechanism of dietary accumulation.

Biomagnification Factor ( BMF). The ratio ( unitless) of the tissue concentration of a chemical in a predator at a particular trophic level to the tissue concentration in its prey at the next lower trophic level for a given waterbody and chemical exposure. For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), a BMF can be calculated using lipid normalized concentrations in the tissue of organisms at two successive trophic levels as: 5 8 ( Equation 5 5)

( Equation 5 6) where:

C R ( TL, n) = Lipid normalized concentration in appropriate tissue of predator organism at a given trophic level ( TL " n") C R ( TL, n 1) = Lipid normalized concentration in appropriate tissue of prey organism at the next lower trophic level from the predator ( TL " n 1")

For inorganic, organometallic, and certain ionic organic chemicals where lipid and organic carbon partitioning does not apply, a BMF can be calculated using chemical concentrations in the tissue of organisms at two successive trophic levels as:

where:

C t ( TL, n) = Concentration in appropriate tissue of predator organism at trophic level " n" ( may be either wet weight or dry weight concentration so long as both the predator and prey concentrations are expressed in the same manner) C t ( TL, n 1) = Concentration in appropriate tissue of prey organism at the next lower trophic level from the predator ( may be either wet weight or dry weight concentration so long as both the predator and prey concentrations are expressed in the same manner)

Biota Sediment Accumulation Factor ( BSAF). For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), the ratio of the lipid normalized concentration of a substance in tissue of an aquatic organism to its organic carbon normalized concentration in surface sediment ( expressed as kg of sediment organic carbon per kg of lipid), in situations where the ratio does not change substantially over time, both the organism and its food are exposed, and the surface sediment is representative of average surface sediment in the vicinity of the organism. The BSAF is defined as:

where:

C R = The lipid normalized concentration of the chemical in tissues of the biota ( µ g/ g lipid) 5 9 ( Equation 5 7)

( Equation 5 8) C soc = The organic carbon normalized concentration of the chemical in the surface sediment ( µ g/ g sediment organic carbon)

Depuration. The loss of a substance from an organism as a result of any active or passive process.

Food Chain Multiplier ( FCM). For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), the ratio of a baseline BAF R f d for an organism of a particular trophic level to the baseline BCF R f d ( usually determined for organisms in trophic level one). For inorganic, organometallic, and certain ionic organic chemicals where lipid and organic carbon partitioning does not apply, a FCM is based on total ( wet or dry weight) concentrations of the chemical in tissue.

Freely Dissolved Concentration. For nonionic organic chemicals, the concentration of the chemical that is dissolved in ambient water, excluding the portion sorbed onto particulate or dissolved organic carbon. The freely dissolved concentration is considered to represent the most bioavailable form of an organic chemical in water and, thus, is the form that best predicts bioaccumulation. The freely dissolved concentration can be determined as:

where:

C w f d = Freely dissolved concentration of the organic chemical in ambient water C w t = Total concentration of the organic chemical in ambient water f fd = Fraction of the total chemical in ambient water that is freely dissolved Hydrophilic. A term that refers to the extent to which a chemical is attracted to partitioning into the water phase. Hydrophilic organic chemicals have a greater tendency to partition into polar phases ( e. g., water) compared to chemicals of hydrophobic chemicals.

Hydrophobic. A term that refers to the extent to which a chemical avoids partitioning into the water phase. Highly hydrophobic organic chemicals have a greater tendency to partition into nonpolar phases ( e. g., lipid, organic carbon) compared with chemicals of lower hydrophobicity.

Lipid normalized Concentration ( C R ) . The total concentration of a contaminant in a tissue or whole organism divided by the lipid fraction in that tissue or whole organism. The lipidnormalized concentration can be calculated as: 5 10 ( Equation 5 9) where:

C t = Concentration of the chemical in the wet tissue ( either whole organism or specified tissue) f R = Fraction lipid content in the organism or specified tissue Octanol water Partition Coefficient ( Kow). The ratio of the concentration of a substance in the n octanol phase to its concentration in the aqueous phase in an equilibrated two phase octanolwater system. For log K ow, the log of the octanol water partition coefficient is a base 10 logarithm.

Organic Carbon normalized Concentration ( Csoc). For sediments, the total concentration of a contaminant in sediment divided by the fraction of organic carbon in sediment. The organic carbon normalized concentration can be calculated as:

where:

C s = Concentration of chemical in sediment f oc = Fraction organic carbon in sediment Uptake. Acquisition by an organism of a substance from the environment as a result of any active or passive process.

5.3 FRAMEWORK FOR DETERMINING NATIONAL BIOACCUMULATION FACTORS 5.3.1 Four Different Methods Bioaccumulation factors used to derive national BAFs can be measured or predicted using some or all of the following four methods, depending on the type of chemical and its properties. These methods are:

( 1) a measured BAF obtained from a field study ( i. e., a field measured BAF);

( 2) a BAF predicted from a field measured BSAF;

( 3) a BAF predicted from a laboratory measured BCF ( with or without adjustment by an FCM); and 5 11 ( 4) a BAF predicted from a chemical's octanol water partition coefficient ( K ow ), with or without adjustment using an FCM.

A brief summary of each of the four methods is provided below. Additional details on the use of these four methods is provided in Section 5.4 ( for nonionic organics), Section 5.5 ( for ionic organics) and Section 5.6 ( for inorganics and organometallics).

1. Field Measured BAF. Use of a field measured BAF, which is the most direct measure of bioaccumulation, is the only method that can be used to derive a national BAF for all types of chemicals ( i. e., nonionic organic, ionic organic, and inorganic and organometallic chemicals). A field measured BAF is determined from a field study using measured chemical concentrations in the aquatic organism and its surrounding water. Because field studies are conducted in natural aquatic ecosystems, a field measured BAF reflects an organism's exposure to a chemical through all relevant exposure pathways ( i. e., water, sediment, and diet). A field measured BAF also reflects any metabolism of a chemical that might occur in the aquatic organism or its food web. Therefore, field measured BAFs are appropriate for all chemicals, regardless of the extent of chemical metabolism in biota.

2. Field measured BSAF. For nonionic organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies), a BAF can also be predicted from BSAFs. A BSAF is similar to a field measured BAF in that the concentration of a chemical in biota is measured in the field and reflects an organism's exposure to all relevant exposure routes. A BSAF also reflects any chemical metabolism that might occur in the aquatic organism or its food web. However, unlike a fieldmeasured BAF which references the biota concentration to the water concentration, a BSAF references the biota concentration to the sediment concentration. Use of the BSAF procedure is restricted to organic chemicals which are classified as being moderately to highly hydrophobic.

3. Lab measured BCF. A laboratory measured BCF can also be used to estimate a BAF for organic and inorganic chemicals. However, unlike a field measured BAF or a BAF predicted from a field measured BSAF, a laboratory measured BCF only reflects the accumulation of chemical through the water exposure route. Laboratory measured BCFs may therefore under estimate BAFs for chemicals where accumulation from sediment or dietary sources is important. In these cases, laboratory measured BCFs can be multiplied by a FCM to reflect accumulation from non aqueous ( i. e., food chain) pathways of exposure. Since a laboratory measured BCF is determined using the measured concentration of a chemical in an aquatic organism and its surrounding water, a laboratory measured BCF reflects any metabolism of the chemical that occurs in the organism, but not in the food web.

4. Kow. A chemical's octanol water partition coefficient, or K ow, can also be used to predict a BAF for nonionic organic chemicals. This procedure is appropriate only for nonionic 5 12 organic chemicals ( and certain ionic organic chemicals where similar lipid and organic carbon partitioning behavior applies). The K ow has been extensively correlated with the BCF for nonionic organic chemicals that are poorly metabolized by aquatic organisms. Therefore, where substantial metabolism is known to occur in biota, the K ow is not used to predict the BAF. For nonionic organic chemicals where chemical exposure through the food web is important, use of the K ow alone will under predict the BAF. In such cases, the K ow is adjusted with a FCM similar to the BCF procedure above.

5.3.2 Overview of BAF Derivation Framework Although up to four methods can be used to derive a BAF as described in the previous section, it is evident that these methods do not apply equally to all types of chemicals. In addition, experience demonstrates that the required data will usually not be available to derive a BAF value using all of the applicable methods. As a result, EPA has developed the following guidelines to direct users in selecting the most appropriate method( s) for deriving a national BAF.

Figure 5 1 shows the overall framework of EPA's national BAF methodology. This framework illustrates the major steps and decisions that will ultimately lead to calculating a national BAF using one of six hierarchical procedures shown at the bottom of Figure 5 1. Each procedure contains a hierarchy of the BAF derivation methods discussed above, the composition of which depends on the chemical type and certain chemical properties ( e. g., its degree of hydrophobicity and expected degree of metabolism and biomagnification). The number assigned to each BAF method within a procedure indicates its general order of preference for deriving a national BAF value. The goal of the framework and accompanying guidelines is to enable full use of available data and methods for deriving a national BAF value while appropriately restricting the use of certain methods to reflect their inherent limitations.

The first step in the framework is to define the chemical of concern. As described in Section 5.3.3, the chemical used to derive the national BAF should be consistent with the chemical used to derive the critical health assessment value. The second step is to collect and review all relevant data on bioconcentration and bioaccumulation of the chemical of concern ( see Section 5.3.4). Once pertinent data are reviewed, the third step is to classify the chemical of concern into one of three broadly defined chemical categories: ( 1) nonionic organic chemicals, ( 2) ionic organic chemicals, and ( 3) and inorganic and organometallic chemicals. Guidance for classifying chemicals into these three categories is provided in Section 5.3.5.

After a chemical has been classified into one of the three categories, other information is used to select one of six hierarchical procedures to derive the national BAF. The specific procedures for deriving a BAF for each chemical group are discussed in Section 5.4 for nonionic organics, Section 5.5 for ionic organics, and Section 5.6 for inorganics and organometallics. 5 13 CLASSIFY CHEMICAL OF CONCERN Nonionic Organic Figure 5 1. Framework for Deriving a National BAF Ionic Organic Inorganic & Organometallic Yes ( Log KOW > 4) ( Log KOW < 4)

PROCEDURE # 4 HYDROPHOBICITY?

METABOLISM?

Low High Low High METABOLISM?

1. Field BAF or PROCEDURE # 3 Lab BCF 2. KOW 1. Field BAF PROCEDURE # 2 3. Lab BCF 1. Field BAF PROCEDURE # 1 2. BSAF 4. KOW* FCM BIOMAGNIFICATION?

PROCEDURE # 6 3. Lab BCF* FCM Low COLLECT & REVIEW DATA DEFINE CHEMICAL OF CONCERN Moderate High PROCEDURE # 5 No IONIZATION NEGLIGIBLE?

No Yes 2. BSAF 1. Field BAF or Lab BCF 1. Field BAF or Lab BCF 1. Field BAF 2. Lab BCF* FCM 5 14 Detailed guidance concerning the first three steps of the derivation process ( i. e, defining the chemical of concern, collecting and reviewing data, and classifying the chemical of concern) is provided in the following three sections.

5.3.3 Defining the Chemical of Concern Defining the chemical of concern is the first step in deriving a national BAF. This step involves precisely defining the form( s) of the chemical upon which the national BAF value will be derived. Although this step is usually straightforward for single chemicals, complications can arise when the chemical of concern occurs as a mixture. The following guidelines should be followed for defining the chemical of concern.

1. Information for defining the chemical of concern should be obtained from the health and exposure assessment portions of the criteria derivation effort. The chemical( s) used to derive the national BAF should be consistent with the chemical( s) used to derive the reference dose ( RfD), point of departure/ uncertainty factor ( POD/ UF), or cancer potency factor.

2. In most cases, the RfD, POD/ UF, or cancer potency factor will be based on a single chemical. In some cases, the RfD, POD/ UF, or cancer potency factor will be based on a mixture of compounds, typically within the same chemical class ( e. g., toxaphene, chlordane). In these situations, the national BAF should be derived in a manner that is consistent with the mixture used to express the health assessment.

a. If sufficient data are available to reliably assess the bioaccumulation of each relevant compound contained in the mixture, then the national BAF( s) should be derived using the BAFs for the individual compounds of the mixture and appropriately weighted to reflect the mixture composition used to establish the RfD, POD/ UF, or cancer potency factor. An example of this approach is shown in the derivation of BAFs for PCBs in the GLI Rulemaking ( USEPA, 1997).

b. If sufficient data are not available to reliably assess the bioaccumulation of individual compounds of the mixture, then the national BAF( s) should be derived using BAFs for the same or appropriately similar chemical mixture as that used to establish the RfD, POD/ UF, or cancer potency value.

5.3.4 Collecting and Reviewing Data The second step in deriving a national BAF is to collect and review all relevant bioaccumulation data for the chemical of concern. The following guidance should be followed for collecting and reviewing bioaccumulation data for deriving national BAFs.

1. All data on the occurrence and accumulation of the chemical of concern in aquatic animals and plants should be collected and reviewed for adequacy. 5 15 2. A comprehensive literature search strategy should be used for gathering bioaccumulationrelated data. An example of a comprehensive literature search strategy is provided in the Bioaccumulation TSD.

3. All data that are used should contain sufficient supporting information to indicate that acceptable measurement procedures were used and that the results are probably reliable. In some cases it may be appropriate to obtain additional written information from the investigator.

4. Questionable data, whether published or unpublished, should not be used. Guidance for assessing the acceptability of bioaccumulation and bioconcentration studies is found in Sections 5.4, 5.5, and 5.6.

5.3.5 Classifying the Chemical of Concern The next step in deriving a national BAF consists of classifying the chemical of concern into one of three categories: nonionic organic, ionic organic, and inorganic and organometallic ( Figure 5 1). This step helps to determine which of the four methods described in Section 5.3.1 are appropriate for deriving BAFs. The following guidance applies for classifying the chemical of concern.

1. Nonionic Organic Chemicals. For the purposes of the 2000 Human Health Methodology, nonionic organic chemicals are those organic compounds that do not ionize substantially in natural bodies of water. These chemicals are also referred to as neutral or nonpolar organics in the scientific literature. Due to their neutrality, nonionic organic chemicals tend to associate with other neutral ( or near neutral) compartments in aquatic ecosystems ( e. g., lipid, organic carbon). Examples of nonionic organic chemicals which have been widely studied in terms of their bioaccumulation include polychlorinated biphenyls ( PCBs), polychlorinated dibenzo p dioxins and furans, many chlorinated pesticides, and polynuclear aromatic hydrocarbons ( PAHs). Procedures for deriving a national BAF for nonionic organic chemicals are provided in Section 5.4.

2. Ionic Organic Chemicals. For the purposes of the 2000 Human Health Methodology, ionic organic chemicals are considered to include those chemicals that contain functional groups with exchangeable protons such as hydroxyl, carboxylic, and sulfonic groups and functional groups that readily accept protons such as amino and aromatic heterocyclic nitrogen ( pyridine) groups. Ionic organic chemicals undergo ionization in water, the extent of which depends on pH and the pKa of the chemical. Because the ionized species of these chemicals behave differently from the neutral species, separate guidance is provided for deriving BAFs for ionic organic chemicals. Procedures for deriving national BAFs for ionic organic chemicals are provided in Section 5.5.

3. Inorganic and Organometallic Chemicals. The inorganic and organometallic category is considered to include inorganic minerals, other inorganic compounds and elements, 5 16 metals ( e. g., copper, cadmium, chromium, zinc), metalloids ( selenium, arsenic) and organometallic compounds ( e. g., methylmercury, tributyltin, tetraalkyllead). Procedures for deriving BAFs for inorganic and organometallic chemicals are provided in Section 5.6.

5.4 NATIONAL BIOACCUMULATION FACTORS FOR NONIONIC ORGANIC CHEMICALS 5.4.1 Overview This section contains the methodology for deriving national BAFs for nonionic organic chemicals as defined in Section 5.3.5. The four general steps of this methodology are:

1. Selecting the BAF derivation procedure, 2. Calculating individual baseline BAF R f ds, 3. Selecting the final baseline BAF R f ds, and 4. Calculating the national BAFs from the final baseline BAF R f ds.

A schematic of this four step process is shown in Figure 5 2.

Step 1 of the methodology ( selecting the BAF derivation procedure) determines which of the four BAF procedures summarized in Figure 5 1 will be appropriate for deriving the national BAF. Step 2 involves calculating individual, species specific BAF R f ds using all of the methods available within the selected BAF derivation procedure. Calculating the individual baseline BAF R f ds involves using data from the field site or laboratory where the original data were collected to account for site specific factors which affect the bioavailability of the chemical to aquatic organisms ( e. g., lipid content of study organisms and freely dissolved concentration in study water). Step 3 of the methodology consists of selecting the final baseline BAF R f ds from the individual baseline BAF R f ds by taking into account the uncertainty in the individual BAFs and the data preference hierarchy selected in Step 1. The final step is to calculate a BAF ( or BAFs) that will be used in the derivation of 304( a) criteria ( i. e., referred to as the national BAF). This step involves adjusting the final baseline BAF R f d( s) to reflect certain factors that affect bioavailablity of the chemical to aquatic organisms in waters to which the national 304( a) criteria will apply ( e. g., the freely dissolved fraction expected in U. S. waters and the lipid content of consumed aquatic organisms). Baseline BAF R f ds are not used directly in the derivation of the 304( a) criteria because they do not reflect the conditions that affect bioavailability in U. S. waters.

Section 5.4.2 below provides detailed guidance for selecting the appropriate BAF derivation procedure ( Step 1 of the process). Guidance on calculating individual baseline BAF R f ds, selecting the final baseline BAF, and calculating the national BAF ( Steps 2 through 4 of the process) is provided in separate sections under each of the four BAF derivation procedures. 5 17 5 18 5.4.2 Selecting the BAF Derivation Procedure This section describes the decisions that should be made to select one of the four available hierarchical procedures for deriving a national BAF for nonionic organic chemicals ( Procedures # 1 through # 4 of Figure 5 1). As shown in Figure 5 1, two decision points exist in selecting the BAF derivation procedure. The first decision point requires knowledge of the chemical's hydrophobicity ( i. e., the K ow of the chemical). Guidance for selecting the K ow for a chemical is provided in the Bioaccumulation TSD. The K ow provides an initial basis for assessing whether biomagnification may be a concern for nonionic organic chemicals. The second decision point is based on the rate of metabolism for the chemical in the target organism. Guidance for assessing whether a high or low rate of metabolism is likely for a chemical of concern is provided below in Section <IP_ADDRESS>. With the appropriate information for these two decision points, the BAF derivation procedure should be selected using the following guidelines.

<IP_ADDRESS> Chemicals with Moderate to High Hydrophobicity 1. For the purposes of the 2000 Human Health Methodology, nonionic organic chemicals with log K ow values equal to or greater than 4.0 should be classified as moderately to highly hydrophobic. For moderately to highly hydrophobic nonionic organic chemicals, available data indicate that exposure through the diet and other non aqueous routes can become important in determining chemical residues in aquatic organisms ( e. g., Russell et al., 1999; Fisk et al., 1998; Oliver and Niimi, 1983; Oliver and Niimi, 1988; Niimi, 1985; Swackhammer and Hites, 1988). Dietary and other non aqueous exposure can become extremely important for those nonionic organic chemicals that are poorly metabolized by aquatic biota ( e. g., certain PCB congeners, chlorinated pesticides, and polychlorinated dibenzo p dioxins and furans).

2. Procedure # 1 should be used to derive national BAFs for moderately to highly hydrophobic nonionic organic chemicals in cases where:

( a) the rate of chemical metabolism by target aquatic organisms is expected to be sufficiently low such that biomagnification is of concern, or ( b) the rate of chemical metabolism by target aquatic organisms is not sufficiently known.

Procedure # 1 accounts for non aqueous exposure and the potential for biomagnification in aquatic food webs through the use of field measured values for bioaccumulation ( i. e., field measured BAF or BSAF) and FCMs when appropriate field data are unavailable. Guidance on deriving national BAFs using Procedure # 1 is found below in Section 5.4.3.

3. Procedure # 2 should be used to derive the national BAFs for moderately to highly hydrophobic nonionic organic chemicals in cases where: 5 19 ( a) the rate of chemical metabolism by target aquatic organisms is expected to be sufficiently high such that biomagnification is not of concern.

Procedure # 2 relaxes the requirement of using FCMs and eliminates the use of K ow based estimates of the BAF, two procedures that are most appropriate for poorly metabolized nonionic organic chemicals. Guidance on deriving national BAFs using Procedure # 2 is found below in Section 5.4.4.

<IP_ADDRESS> Chemicals with Low Hydrophobicity 1. For the purposes of these guidelines, nonionic organic chemicals with log K ow values less than 4.0 should be classified as exhibiting low hydrophobicity. For nonionic organic chemicals that exhibit low hydrophobicity ( i. e., log K ow < 4.0), available information indicates that non aqueous exposure to these chemicals is not likely to be important in determining chemical residues in aquatic organisms ( e. g., Fisk et al., 1998; Gobas et al., 1993; Connolly and Pedersen, 1988; Thomann, 1989). For this group of chemicals, laboratory measured BCFs and K ow predicted BCFs do not require adjustment with FCMs for determining the national BAF ( Procedures # 3 and # 4), unless other appropriate data indicate differently.

Other appropriate data include studies clearly indicating that non aqueous exposure is important such that use of a BCF would substantially underestimate residues in aquatic organisms. In these cases, Procedure # 1 should be used to derive the BAF for nonionic organic chemicals with log K ow < 4.0. Furthermore, the data supporting the K ow determination should be carefully reviewed for accuracy and appropriate interpretation, since the apparent discrepancy may be due to errors in determining K ow.

2. Procedure # 3 should be used to derive national BAFs for nonionic organic chemicals of low hydrophobicity in cases where:

( a) the rate of chemical metabolism by target aquatic organisms is expected to be negligible, such that tissue residues of the chemical of concern are not substantially reduced compared to an assumption of no metabolism, or ( b) the rate of chemical metabolism by target aquatic organisms is not sufficiently known.

Procedure # 3 includes the use of K ow based estimates of the BCF to be used when lab or field data are absent. Guidance on deriving national BAFs using Procedure # 3 is found below in Section 5.4.5.

3. Procedure # 4 should be used to derive national BAFs for nonionic organic chemicals of low hydrophobicity in cases where: 5 20 ( a) the rate of chemical metabolism by target aquatic organisms is expected to be sufficiently high, such that tissue residues of the chemical of concern are substantially reduced compared with an assumption of no metabolism.

Procedure # 4 eliminates the option of using K ow based estimates of the BAF because the K ow may over predict accumulation when a chemical is metabolized substantially by an aquatic organism. Guidance on deriving national BAFs using Procedure # 4 is found below in Section 5.4.6.

<IP_ADDRESS> Assessing Metabolism Currently, assessing the degree to which a chemical is metabolized by aquatic organisms is confounded by a variety of factors. First, conclusive data on chemical metabolism in aquatic biota are largely lacking. Such data include whole organism studies where the metabolic rates and breakdown products are quantified in fish and other aquatic organisms relevant to human consumption. However, the majority of information on metabolism is derived from in vitro liver microsomal preparations in which primary and secondary metabolites may be identified and their rates of formation may or may not be quantified. Extrapolating results from in vitro studies to the whole organism involves considerable uncertainty. Second, there are no generally accepted procedures for reliably predicting chemical metabolism by aquatic organisms in the absence of measured data. Third, the rate at which a chemical is metabolized by aquatic organisms can be species and temperature dependent. For example, PAHs are known to be metabolized readily by vertebrate aquatic species ( primarily fish), although at rates much less than those observed for mammals. However, the degree of metabolism in invertebrate species is generally much less than the degree in vertebrate species ( James, 1989). One hypothesis for this difference is that the invertebrate species lack the detoxifying enzymes and pathways that are present in many vertebrate species.

Given the current limitations on assessing the degree of chemical metabolism by aquatic organisms, the assessment of metabolism should be made on a case by case basis using a weightof evidence approach. When assessing a chemical's likelihood to undergo substantial metabolism in a target aquatic organism, the following data should be carefully evaluated:

( 1) in vivo chemical metabolism data, ( 2) bioconcentration and bioaccumulation data, ( 3) data on chemical occurrence in target aquatic biota, and ( 4) in vitro chemical metabolism data.

1. In vivo Data. In vivo data on metabolism in aquatic organisms are from studies of chemical metabolism using whole organisms. These studies are usually conducted using large fish from which blood, bile, urine, and individual tissues can be collected for the identification and quantification of metabolites formed over time. In vivo studies are considered the most useful for evaluating a chemical's degree of metabolism in an organism because both oxidative ( Phase I) and conjugative ( Phase II) metabolism can be 5 21 assessed in these studies. Mass balance studies, in which parent compound elimination is quantified separately from biotransformation and elimination of metabolites, allow calculation of conversion rate of parent to metabolite as well as metabolite elimination. This information might be used to estimate loss due to metabolism separately from that due to elimination of the parent compound for adjustment of K ow predicted BAFs. However, due to the analytical and experimental challenges these studies pose, data of this type are limited. Less rigorous in vivo metabolism studies might include the use of metabolic blockers to demonstrate the influence of metabolism on parent compound kinetics. However, caution should be used in interpretation of absolute rates from these data due to the lack of specificity of mammalian derived blockers in aquatic species ( Miranda et al., 1998).

2. Bioconcentration or Bioaccumulation Data. Data on chemical bioconcentration or bioaccumulation in aquatic organisms can be used indirectly for assessing metabolism. This assessment involves comparing acceptable lab measured BCFs or field measured BAFs ( after converting to baseline values using procedures below) with the chemical's predicted value based on K ow. The theoretical basis of bioconcentration and bioaccumulation for nonionic organic chemicals indicates that a chemical's baseline BCF should be similar to its K ow predicted value if metabolism is not occurring or is minimal ( see the Bioaccumulation TSD). This theory also indicates that baseline BAFs should be similar to or higher than the K ow for poorly metabolized organic chemicals, with highly hydrophobic chemicals often exhibiting higher baseline BAFs than K ow values. Thus, if a chemical's baseline BCF or BAF is substantially lower than its K ow, this may be an indication that the chemical is being metabolized by the aquatic organism of concern. Note, however, that this difference may also indicate problems in the experimental design or analytical chemistry, and that it may be difficult to discern the difference.

3. Chemical Occurrence Data. Although by no means definitive, data on the occurrence of chemicals in aquatic biota ( i. e., residue studies) may offer another useful line of evidence for evaluating a chemical's likelihood to undergo substantial metabolism. Such studies are most useful if they have been conducted repeatedly over time and over wide geographical areas. Such studies might indicate a chemical is poorly metabolized if data show that the chemical is being biomagnified in the aquatic food web ( i. e., higher lipidnormalized residues in successive trophic levels). Conversely, such studies might indicate a chemical is being metabolized substantially if residue data show a decline in residues with increasing trophic level. Again, other reasons for increases or decreases in concentrations with increasing trophic level might exist and should be carefully evaluated ( e. g., incorrect food web assumptions, differences in exposure concentrations).

4. In vitro Data. In vitro metabolism data include data from studies where specific subcellular fractions ( e. g., microsomal, cytosolic), cells, or tissues from an organism are tested outside the body ( i. e., in test tubes, cell or tissue culture). Compared with in vivo studies of chemical metabolism in aquatic organisms, in vitro studies are much more plentiful in the literature, with the majority of studies characterizing oxidative ( Phase I) 5 22 reactions de coupled from conjugative ( Phase II) metabolism. Cell, tissue, or organ level in vitro studies are less common but provide a more complete assessment of metabolism. While such studies are particularly useful for identifying the pathways, rates of formation, and metabolites formed, as well as the enzymes involved and differences in the temperature dependence of metabolism across aquatic species, they suffer from uncertainty when results are extrapolated to the whole organism. This uncertainty results from the fact that dosimetry ( i. e., delivery of the toxicant to, and removal of metabolite from, the target tissue) cannot currently be adequately reproduced in the laboratory or easily modeled.

When assessing chemical metabolism using the above information, the following guidelines apply.

a. A finding of substantial metabolism should be supported by two or more lines of evidence identified using the data described above.

b. At least one of the lines of evidence should be supported by either in vivo metabolism data or acceptable bioconcentration or bioaccumulation data.

c. A finding of substantial metabolism in one organism should not be extrapolated to another organism or another group of organisms unless data indicate similar metabolic pathways exist ( or are very likely to exist) in both organisms. In vitro data may be particularly useful in cross species extrapolations.

d. Finally, in situations where sufficient data are not available to properly assess the likelihood of significant metabolism in aquatic biota of concern, the chemical should be assumed to undergo little or no metabolism. This assumptions reflects a policy decision by EPA to err on the side of public health protection when sufficient information on metabolism is lacking.

5.4.3 Deriving National BAFs Using Procedure # 1 This section contains guidance for calculating national BAFs for nonionic organic chemicals using Procedure # 1 shown in Figure 5 1. The types of nonionic organic chemicals for which Procedure # 1 is most appropriate are those that are classified as moderately to highly hydrophobic and subject to low ( or unknown) rates of metabolism by aquatic biota ( see Section 5.4.2 above). Non aqueous contaminant exposure and subsequent biomagnification in aquatic food webs are of concern for chemicals that are classified in this category. Some examples of nonionic organic chemicals for which Procedure # 1 is considered appropriate include:

C tetra, penta & hexachlorobenzenes; C PCBs; C octachlorostyrene; C hexachlorobutadiene; 5 23 C endrin, dieldrin, aldrin; C mirex, photomirex; C DDT, DDE, DDD; and C heptachlor, chlordane, nonachlor.

Under Procedure # 1, the following four methods may be used in deriving a national BAF:

C using a BAF from an acceptable field study ( i. e., a field measured BAF); C predicting a BAF from an acceptable field measured BSAF; C predicting a BAF from an acceptable laboratory measured BCF and FCM; and C predicting a BAF from an acceptable K ow and FCM.

As shown in Figure 5 2, once the derivation procedure has been selected, the next steps in deriving a national BAF for a given trophic level include: calculating individual baseline BAF R f ds ( step 2), selecting the final baseline BAF R f d ( step 3), and calculating the national BAF from the final baseline BAF R f d ( step 4). Each of these three steps is discussed separately below.

<IP_ADDRESS> Calculating Individual Baseline BAF R f ds Calculating an individual baseline BAF R f d involves normalizing the field measured BAF tT ( or laboratory measured BCF tT ) which are based on total concentrations in tissue and water by the lipid content of the study organisms and the freely dissolved concentration in the study water. Both the lipid content in the organism and the freely dissolved concentration ( as influenced by organic carbon in water) have been shown to be important factors that influence the bioaccumulation of nonionic organic chemicals ( e. g., Mackay, 1982; Connolly and Pederson, 1988; Thomann, 1989, Suffet et al., 1994). Therefore, baseline BAF R f ds ( which are expressed on a freely dissolved and lipid normalized basis) are considered more amenable to extrapolating between different species and bodies of water compared to BAFs expressed using the total concentration in the tissue and water. Because bioaccumulation can be strongly influenced by the trophic position of aquatic organisms ( either due to biomagnification or physiological differences), extrapolation of baseline BAF R f ds should not be performed between species of different trophic levels.

1. For each species for which acceptable data are available, calculate all possible baseline BAF R f ds using each of the four methods shown above for Procedure # 1.

2. Individual baseline BAF R f ds should be calculated from field measured BAF tT s, fieldmeasured BSAFs, laboratory BCF tT s, and the K ow according to the following procedures.

A. Baseline BAF R f ds from Field Measured BAFs A baseline BAF R f d should be calculated from each field measured BAF tT using information on the lipid fraction in the tissue of concern for the study organism and the fraction of the total chemical that is freely dissolved in the study water. 5 24 ( Equation 5 10)

( Equation 5 11) 1. Baseline BAF R f d Equation. For each acceptable field measured BAF tT , calculate a baseline BAF R f d using the following equation:

where:

Baseline BAF R f d = BAF expressed on a freely dissolved and lipid normalized basis Measured BAFtT = BAF based on total concentration in tissue and water f R = Fraction of the tissue that is lipid f fd = Fraction of the total chemical that is freely dissolved in the ambient water The technical basis of Equation 5 10 is provided in the Bioaccumulation TSD. Guidance for determining each component of Equation 5 10 is provided below.

2. Determining the Measured BAFt T. The field measured BAF tT shown in Equation 5 10 should be calculated based on the total concentration of the chemical in the appropriate tissue of the aquatic organism and the total concentration of the chemical in ambient water at the site of sampling. The equation to derive a measured BAF tT is:

where:

C t = Total concentration of the chemical in the specified wet tissue C w = Total concentration of chemical in water The data used to calculate a field measured BAF tT should be reviewed thoroughly to assess the quality of the data and the overall uncertainty in the BAF value. The following general criteria apply in determining the acceptability of field measured BAFs that are being considered for deriving national BAFs using Procedure # 1.

a. Aquatic organisms used to calculate a field measured BAF tT should be representative of aquatic organisms that are commonly consumed in the United States. An aquatic organism that is not commonly consumed in the United States can be used to calculate an acceptable field measured BAF tT provided that the 5 25 organism is considered to be a reasonable surrogate for a commonly consumed organism. Information on the ecology, physiology, and biology of the organism should be reviewed when assessing whether an organism is a reasonable surrogate of a commonly consumed organism.

b. The trophic level of the study organism should be determined by taking into account its life stage, diet, size, and the food web structure at the study location. Information from the study site ( or similar sites) is preferred when evaluating trophic status. If such information is lacking, general information for assessing trophic status of aquatic organisms can be found in USEPA ( 2000a, b, c).

c. The percent lipid of the tissue used to determine the field measured BAF tT should be either measured or reliably estimated to permit lipid normalization of the chemical's tissue concentration.

d. The study from which the field measured BAF tT is derived should contain sufficient supporting information from which to determine that tissue and water samples were collected and analyzed using appropriate, sensitive, accurate, and precise analytical methods.

e. The site of the field study should not be so unique that the BAF cannot be reasonably extrapolated to other locations where the BAF and resulting criteria will apply.

f. The water concentration( s) used to derive the BAF should reflect the average exposure of the aquatic organism that corresponds to the concentration measured in its tissue of concern. For nonionic organic chemicals, greater temporal and spatial averaging of chemical concentrations is required as the K ow increases. In addition, as variability in water concentrations increase, greater temporal and spatial averaging is also generally required. Greater spatial averaging is also generally required for more mobile organisms.

g. The concentrations of particulate organic carbon and dissolved organic carbon in the study water should be measured or reliably estimated.

EPA is currently developing guidance for designing and conducting field studies for determining field measured BAFtT s, including recommendations for minimum data requirements. A more detailed discussion of factors that should be considered when determining field measured BAFtT s is provided in the Bioaccumulation TSD.