Development of aquatic life criteria for selenium: a regulatory perspective on critical issues and research needs

Aquatic Toxicology 57 (2002) 101– 113

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Development of aquatic life criteria for selenium: a regulatory perspective on critical issues and research needs Keith G. Sappington * US En6ironmental Protection Agency, Office of Research and De6elopment, National Center for En6ironmental Assessment (8623D), 1200 Pennsyl6ania A6enue, N.W., Washington, DC 20460, USA Received 10 October 2000; received in revised form 3 February 2001; accepted 17 March 2001

Abstract The US is currently in the process of revising its freshwater, chronic aquatic life criterion for selenium. The fundamental issues being addressed include which environmental compartment(s) support the most reliable expression of the criterion, which form(s) of selenium should be measured in the medium (media) of choice, and which site-specific water quality (or other factors) should be linked to the expression of the criterion. Literature reviews and a recent workshop were conducted to assess the state of the science on various issues related to water-, tissue- and sediment-based criteria for selenium. Evaluation of many of these issues is ongoing. In terms of water column criteria issues, data limitations will likely restrict the expression of a criterion to operationally defined forms (e.g. total recoverable, dissolved). The specific identity of organoselenium in natural systems is lacking and may not be appropriately represented by free seleno-amino acids (e.g. selenomethionine). The available data do not appear to support quantitative relationships between chronic toxicity and water quality characteristics. In terms of a tissuebased criterion, reproductive tissue (ovary, egg) has been recommended as the tissue of choice, but practical concerns and data availability require consideration of other tissues (e.g. whole-body). Organoselenium (bound to peptides or proteins) is thought to be the form of greatest toxicological importance in fish, however, direct measurements of organoselenium compounds in tissues are very limited. Route of exposure (food vs. water uptake) may prove important for establishing diagnostic tissue residues for selenium based on laboratory data. Data on toxicological aspects of selenium in sediments appear sparse, particularly in relation to different sedimentary forms. Reliable assessments of bioaccumulation will likely be critical for making site-specific modifications to chronic selenium criteria, however, many technical issues for assessing bioaccumulation remain. The need for improved analytical methods for directly speciating organoselenium in various environmental media underpins many of the current data gaps. Improving analytical methodologies to enable affordable and reliable measurement of organoselenium compounds holds significant promise for advancing selenium ecotoxicological research. © 2002 Elsevier Science B.V. All rights reserved. Keywords: Selenium; Aquatic life criteria; Research needs

* Tel./fax: +1-202-564-1834. E-mail address: [email protected] (K.G. Sappington). 0166-445X/02/$ - see front matter © 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 6 - 4 4 5 X ( 0 1 ) 0 0 2 6 7 - 3

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1. Introduction Selenium, a metalloid, is toxic to aquatic life at relatively low concentrations (US EPA, 1987). The occurrence of selenium in surface waters is widespread resulting from a variety of natural and anthropogenic sources. These sources include natural weathering and irrigation- induced leaching of selenium containing rocks and soils, mobilization and discharge from mining and smelting activities, flue gas emission from fuel oil and coal combustion, and fly-ash disposal practices (i.e. pond leachate and runoff from land disposal areas). Selenium contamination in aquatic ecosystems has been linked to adverse ecological effects in several field settings (for review, see Skorupa, 1998) that include reproductive and developmental impairment of aquatic birds and fish (Ohlendorf et al., 1986a,b; Cumbie and Van Horn, 1978; Lemly, 1985). A hypothesis has also been extended that selenium contamination may have contributed to the historical decline of some endangered fish species in the Colorado River Basin (Hamilton, 1999). Selenium is also an essential trace element to many aquatic and terrestrial species, with required levels approaching just one order of magnitude below those causing toxicological effects in fish (Hilton et al., 1980; Hodson and Hilton, 1983; Lemly, 1998a; US EPA, 1998). Therefore, when setting aquatic toxicological benchmarks for selenium, consideration must be given to ensure that both sufficiency and toxicity concerns are appropriately balanced. Besides sufficiency versus toxicity concerns, several other factors complicate the establishment of aquatic toxicological thresholds for selenium. These factors are associated with the complex biogeochemistry of selenium in aquatic ecosystems. First, selenium occurs in several different oxidation states in the aquatic environment that include oxidized selenates (Se + 6) and selenites (Se + 4), elemental selenium (Se0) and reduced selenides (Se − 2). Each form is known to differ in bioavailability and toxicity to aquatic organisms. Second, selenium can undergo biotransformations between inorganic and organic forms as a result of biotic and abiotic processes, which are not well characterized. Third, selenium also has been

shown to bioaccumulate in aquatic food webs to the extent that dietary exposure to selenium becomes a critical exposure pathway for top predatory aquatic and aquatic-dependent organisms. Since traditional laboratory chronic toxicity tests rarely include realistic exposures through the diet, they are less relevant for directly assessing the toxicity of selenium in natural settings. The US Environmental Protection Agency (EPA) establishes aquatic toxicological benchmarks for various pollutants in the form of aquatic life criteria (US EPA, 1999). Aquatic life criteria are expressed in the form of an acute criterion (or criterion maximum concentration) and a chronic criterion (or criterion continuous concentration). Acute criteria are designed to protect against unacceptable adverse effects resulting from short-term exposures to chemical concentrations, while chronic criteria are designed to protect against unacceptable adverse effects resulting from long-term (continuous) exposures (Stephan et al., 1985). The EPA most recently revised its chronic freshwater aquatic life criterion for selenium in 1987 (US EPA, 1987). This criterion (5 mg/l, as total recoverable selenium) is based on field investigations of the effects of selenium on fish populations in Belews Lake, North Carolina, as well as corroborating laboratory tests indicating selenium effects in the field can occur below those observed in the laboratory. Over the past decade, a substantial body of literature has accumulated on the toxicity of selenium to aquatic life. In response to this and other new information, EPA initiated various activities to support revision of its national freshwater chronic criterion. In addition, EPA is also exploring the development of guidelines to enable site-specific adjustments of the national criterion where appropriately justified. Such guidance is viewed as extremely important for selenium because of potential variability in selenium cycling for different sites. Likewise, the importance of dietary exposure renders conventional site-specific criteria methods (e.g. water effect ratio, US EPA, 1994) inappropriate for selenium. In 1998, EPA conducted a peer consultation workshop which brought together leading experts to assess the state of the science on selenium toxicity and

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bioaccumulation (US EPA, 1998). Results from this workshop are being used by EPA to complement the traditional process of criteria derivation which involves reviewing the literature, conducting additional testing and analysis, and writing criteria documents. This manuscript is intended to provide a regulatory perspective on the most pressing issues facing the Agency during its revision of aquatic life criteria for selenium, primarily those associated with the chronic aquatic life criterion. In addition, important data gaps and research needs are highlighted where further analyses of existing data or the generation of new data are warranted. While efforts to address a number of the research needs identified in this manuscript will undoubtedly extend beyond the time allotted for EPA’s expected revision of its selenium aquatic life criteria, conducting further research in these areas would likely lead to significant improvements in future criteria development efforts and aquatic life risk assessments involving selenium. The issues and research needs identified below are motivated by several fundamental decisions the Agency is addressing in revising the chronic, freshwater criterion for selenium. These decisions include: (1) choosing the environmental compartment(s) for expressing the chronic criterion (e.g. water, sediment, tissue); (2) determining the form(s) of selenium represented by the criterion in one or more environmental media; and (3) deciding whether ample data exist for expressing the criterion as a function of various water quality or hydrological factors. Accordingly, the following discussion is organized by the environmental medium for which toxicological benchmarks could conceivably be derived (e.g. water, tissue, sediment). In addition, issues which transcend the various environmental compartments are also discussed (e.g. analytical chemistry, bioaccumulation). Research needs in the following discussion are identified through recent literature reviews on selenium aquatic toxicology, but in a number of instances, are informed by the work and opinions of experts from the peer consultation workshop (US EPA, 1998). The views and opinions expressed in this manuscript are those of the author and do not necessarily reflect those of the EPA.

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2. Water column criterion issues The water column traditionally has been the medium of choice for expressing EPA’s aquatic life criteria. It has been the preferred medium primarily because toxicity data supporting aquatic life criteria are usually expressed as water concentrations and because a direct linkage can be established between a water-based criterion and chemical loads. For selenium, there is significant conceptual appeal for expressing criteria in terms of tissue residues. However, a mechanism for translating selenium residues in tissues to concentrations in other environmental media will likely be needed for regulatory purposes (e.g. to translate from tissue residues to water column concentrations).

2.1. Form(s) in water The EPA is currently evaluating which form(s) of selenium in water provide(s) the most reliable means of expressing (or translating to) a waterbased chronic criterion. This decision includes consideration of the chemical specie (e.g. selenite, selenate, organic selenides), in addition to operationally-defined forms such as total recoverable or dissolved selenium. The most comprehensive measure of selenium (i.e. total selenium) may not be the most reliable form for predicting chronic risks, because some forms may have reduced bioavailability to aquatic organisms. In concept, the form(s) of selenium that most closely correlate(s) with bioaccumulation and chronic toxicity in natural settings is most desirable for expressing, or translating to, water column-based criteria. Selenite and selenate usually dominate the inorganic fraction of selenium dissolved in aerobic water of aquatic ecosystems (Cutter, 1991; US EPA, 1998), although colloidal selenium which includes reduced organoselenides (Se − 2) can be a significant fraction of total selenium in some reservoirs (Cutter, 1989). However, knowledge of the exact nature of organoselenium forms in water is lacking. Importantly, some forms of organoselenium such as selenomethionine have been shown to bioaccumulate to a much greater extent than selenite and selenate (Besser et al., 1993). Experts

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contend the most toxicologically important forms of organoselenium dissolved in water are likely protein or peptide-bound forms, since the occurrence of freely-dissolved seleno-amino acids (e.g. selenomethionine) is thought to be very low (US EPA, 1998). Although selenomethionine appears to be a reasonable surrogate for naturally occurring organoselenium in the diet (US EPA, 1998), it might not be representative of organoselenium in the water. This has important ramifications for use and extrapolation of bioaccumulation factors (BAFs) and water-based toxicity data from seleno-amino acid data. Clearly, additional data on selenium speciation in aquatic ecosystems is needed to identify the most toxicologically-relevant forms of organoselenium dissolved in water. Once the relevant organoselenium specie(s) in water are identified, appropriate tests can be conducted to compare their bioavailability and toxicological importance to aquatic organisms. Obtaining this information is directly dependent upon the availability of accurate and affordable analytical methods (as discussed later in this manuscript). Until these methods are developed, practical limitations of available data (i.e. most data are expressed as total selenium) and analytical methods will greatly restrict the options for expressing selenium criteria in water, most likely to expressions of dissolved or total recoverable selenium.

2.2. Factors affecting toxicity and bioaccumulation Traditionally, EPA aquatic life criteria have been expressed as a function of various water quality parameters (e.g. hardness, pH, temperature) to the extent that quantitative relationships between the parameter and toxicity allow. Such relationships enable national aquatic life criteria to better reflect bioavailability differences among appropriate sites. Significant concern and debate have occurred over the effect that differences in water quality characteristics (e.g. sulfate, hardness, pH, etc.) and hydrogeology (e.g. lentic vs. lotic habitats) have on selenium toxicity in aquatic ecosystems (Chapman, 1999; Canton and Van Derveer, 1997; Van Derveer and Canton, 1997;

US EPA, 1998; Skorupa, 1998; Hamilton and Lemly, 1999). Although some studies have evaluated the effect of hardness, temperature, pH and other parameters on selenium toxicity, sulfate has perhaps been most widely studied in relation to selenium uptake and toxicity in aquatic organisms. Antagonistic effects from sulfate on either uptake or acute toxicity of selenate to aquatic organisms have been reported for algae (Sarma and Jayaraman, 1984; Williams et al., 1994; Riedel and Sanders, 1996), aquatic macrophytes (Bailey et al., 1995), zooplankton and midges (Ogle and Knight, 1996; Hansen et al., 1993) and two amphipods and fathead minnow (EPRI, 1999; Brix et al., 2001). On the contrary, sulfate did not affect the distribution and bioaccumulation of a mixture of selenite, selenate and selenomethionine in aquatic microcosms and associated biota; however, this may have resulted from the low proportion of selenate relative to the other selenium forms present (Besser et al., 1989). With respect to chronic toxicity, experts at the 1998 peer consultation workshop considered available data too sparse to support quantitative relationships with water quality characteristics (US EPA, 1998). Furthermore, extrapolation of acute toxicity relationships such as sulfate dependency to chronic toxicity was judged to be highly uncertain. Such uncertainty is due in part because sulfate dependency usually applies to selenate (and not selenite), the routes of acute and chronic exposure differ, and because of the apparent absence of sulfate dependency with respect to selenium toxicity and bioaccumulation in real world situations (US EPA, 1998; Skorupa, 1998). Additional data are needed to evaluate the effect of environmentally realistic ranges of water quality characteristics on selenium bioaccumulation and chronic toxicity.

3. Tissue residue criterion issues The implicit or explicit use of tissue residues to characterize contaminant toxicity is fairly widespread (e.g. Mancini, 1983; Abernethy et al., 1988; McCarty et al., 1992; McCarty and Mackay,

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1993; Hickie et al., 1995; Meyer et al., 1999). Expressing aquatic life criteria for selenium in terms of tissue residues in aquatic organisms1 is conceptually appealing due to its complex biogeochemistry in the aquatic environment and the potential ‘normalizing effect’ tissue residues might have on expressing risk across different ecosystems. Specifically, tissue residues are closer to the site(s) of toxic action than concentrations in water or sediment and, therefore, inherently account for species-specific toxicokinetics and site-specific differences in bioavailability. Notably, however, diagnostic tissue residues would not address toxicodynamic differences across species (i.e. the response elicited by a given residue at the site of toxic action). Thus, variability in diagnostic tissue residues would still be expected across species, although it may be reduced relative to toxicity values expressed in the water column or other external media. For selenium, diagnostic tissue residues have been developed for aquatic life (Lemly, 1993a) and have been recently critiqued (DeForest et al., 1999). EPA is currently exploring the use of a tissue-residue based approach for setting chronic aquatic life criteria for selenium, and in particular, for setting site-specific criteria (US EPA, 1998). One such approach would rely on diagnostic tissue residues, above which the potential for adverse effects on aquatic life would be considered unacceptable. In its simplest form, site-specific criteria (expressed in terms of concentrations in water) could, in theory, be determined by assessing bioaccumulation at a site in appropriate target organisms (e.g. by using a model or a BAF) and dividing the diagnostic tissue residue value by the site-specific BAF per Eq. (1): SSC =

1

DTR SSBAF

(1)

For the sake of brevity, this discussion is focused on diagnostic tissue residues in target aquatic species rather than their diet. This does not imply that dietary tissue residues are not a viable approach for evaluating selenium risks or expressing criteria.

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where SSC=Site-specific criterion (mg/l); DTR = Diagnostic tissue residue (mg/kg); SSBAF= Site-specific bioaccumulation factor (l/kg). Lemly (1998b) recently described a similar approach for making site-specific adjustments to EPA’s chronic aquatic life criterion for selenium using a combination of monitoring data and diagnostic tissue residues. However, adjustments to the criterion using Lemly’s approach are made based on fixed percentages rather than an assessment of bioaccumulation relative to water concentrations. Regardless of how diagnostic tissue residues might ultimately be used in setting aquatic life criteria, issues such as adequacy of data base, choice of tissue, and route of exposure require further evaluation. Some of these issues may be resolved with further analysis of existing data, but others will require additional research.

3.1. Adequacy of tissue residue database Given the expected variation in diagnostic tissue residues due to toxicodynamic differences among aquatic species, the species sensitivity range must be adequately captured by the available tissue residue data. This evaluation is ongoing, but cursory inspection of the residue data base (Jarvinen and Ankley, 1999; DeForest et al., 1999) reveals that whole body diagnostic residue values are available for five species of fish (rainbow trout, chinook salmon, fathead minnow, bluegill, and largemouth bass) representing three families. However, data summarized by Jarvinen and Ankley (1999) for largemouth bass represent a no observed effect concentration (NOEC) for survival without an accompanying lowest observed adverse effect concentration (LOEC). Additional diagnostic tissue residue values might be derived for other fish species based on data from Lemly (1993b), although residue-response relationships require further evaluation on a speciesspecific basis. Diagnostic tissue residue values have been summarized for three species of invertebrates, representing a rotifer, cladoceran, and a chironomid (Jarvinen and Ankley, 1999). Diag-

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nostic tissue residues in ovaries (a desirable tissue based on the reproductive effects of selenium) appear to be limited to two species representing two families (fathead minnow, bluegill). Diagnostic residues expressed in the diet appear to be available for at least five species (rainbow trout, chinook salmon, fathead minnow, bluegill and striped bass). Considering the whole-body residue data from Jarvinen and Ankley (1999), DeForest et al. (1999, 1999) reasonably diverse assemblage of aquatic species appears to be available. When compared with the minimum database requirements set forth in EPA’s aquatic life criteria guidelines (Stephan et al., 1985), whole-body residue data appear to be lacking for a benthic crustacean and possibly one other family. Diagnostic ovarian tissue residues are obviously very limited, but the bluegill, generally considered to be among the more sensitive fish species, is represented. Perhaps more importantly may be the need for reproductive data for coldwater fish (e.g. Salmonids). It appears that whole-body diagnostic tissue residues for chinook salmon (3– 7.6 mg/kg d.w.) and rainbow trout (3.3– 4 mg/kg d.w.) are among the lowest values for the fish species reported by DeForest et al. (1999) table 3; original data from Hamilton et al., 1990; Hunn et al., 1987). While these data represent exposure to early life stages of fish, they do not reflect reproductive effects of selenium which may be more sensitive than the survival and growth endpoints quantified here (i.e. parental exposure and subsequent maternal transfer of selenium from the ovary to the egg). Thus, a reproductive study involving a salmonid and both maternal and F1 generation exposure is desirable for selenium.

3.2. Choice of tissue Another issue facing the use of diagnostic tissue residues for criteria derivation is the choice of tissue(s) upon which to express the residue value. This question was posed to a panel of experts at a workshop on selenium toxicity and bioaccumulation (US EPA, 1998), which overwhelmingly recommended reproductive tissue (ovary, egg) as the tissue of choice. Reproductive tissue is recom-

mended based on knowledge that critical effects on several fish species involves maternal transfer of selenium from the ovary to the egg, with subsequent effects on larval survival and development (Woock et al., 1987; Schultz and Hermanutz, 1990; Hermanutz et al., 1992). Thus, the ovary or egg concentration would be closest to the site of toxic action. However, practical considerations of implementing selenium diagnostic residues based solely on ovary or egg concentrations (i.e. the limited availability of these tissues for monitoring purposes) may require consideration of other tissues such as whole body, muscle, or liver. Furthermore, the data on diagnostic residues in ovaries is limited (as noted above) with apparently only one study (Coyle et al., 1993) that has bracketed the residue-response curve for reproductive effects (i.e. containing both a NOEC and a LOEC). When ovary residue-response data for larval survival are combined across available studies for fathead minnow and bluegill, the characterization of the residue-response relationship is more complete (DeForest et al., 1999 figure 3). Graphical data presented by Coyle et al. (1993) lend support for the use of whole-body residue data as a surrogate for ovary residues. Specifically, carcass, ovary and egg selenium concentrations in bluegill appear to be well correlated over a range of treatments containing combined water and dietary exposures, with ovary concentrations typically averaging twice those in the carcass. Data from Hamilton et al. (1990) also demonstrate a well-defined dose-response relationship for whole-body selenium and larval mortality of chinook salmon. When data were combined across years and species for Centrarchids, the frequency of abnormalities in fish exposed to selenium in Belews Lake is reasonably well correlated with total selenium in whole- body (Lemly, 1993b). However, it is interesting that whole-body selenium residues in samples of normal and abnormal fish (displaying teratogenic effects) from Belews lake are not substantially different (Lemly, 1993b). Since the frequency of larval abnormalities has been shown to vary widely with spawning event (Woock et al., 1987; Hermanutz et al., 1992), it seems plausible that a combination of high variability in larval abnormality responses

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among spawns (i.e. producing both normal and abnormal larvae) and subsequent selenium uptake by ‘normal’ F1 juveniles and adults (after the critical exposure period during oogenesis) could explain some of the similarity in whole-body residues in abnormal and normal fish. However, this explanation is purely speculative. Suffice it to say that as additional selenium toxicological studies are conducted, measurement of selenium in a variety of tissues should become a prerequisite so that the reliability of diagnostic residues expressed in different tissues can be more fully evaluated.

3.3. Form and route of selenium exposure The existing tissue residue data base of selenium effects on aquatic organisms consists of studies employing multiple selenium forms (selenite, selenate, selenomethionine) and exposure routes (water, diet and combinations thereof) (Lemly, 1993a; Jarvinen and Ankley, 1999; DeForest et al., 1999). The derivation of diagnostic tissue residue values from studies with heterogenous dosing strategies raises the question of whether form of selenium or route of exposure influences the diagnostic residue value. In other words, is a given response of an organism to a specified tissue concentration of selenium independent of its initial form or its exposure route to the organism? Aspects of this issue were raised in both reviews by Jarvinen and Ankley (1999), DeForest et al. (1999). There appears to be some difference of opinion on the answer to this question. Hamilton (1998) summarized data which suggest comparability among tissue residue thresholds based on dietary and waterborne only exposures. For example, similar diagnostic tissue residues (whole body) were reported for rainbow trout exposed to selenite via the water column (i.e. 5.2 mg/g d.w. for 60 day survival and growth effects; Hunn et al., 1987) compared with selenite exposure through the diet (i.e. 4–4.5 mg/g d.w. based on 16 week growth effects; Hilton and Hodson, 1983). Results from Cleveland et al. (1993) on the effects of dietary and waterborne selenium on bluegills are open to interpretation with respect to similarity in tissue residue thresholds derived from

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waterborne and dietary exposures. The waterborne exposure study by Cleveland et al. (1993) consisted of exposing 5-month old bluegill sunfish to measured concentrations ranging from 0.16 to 2.8 mg/l total selenium in water (based on a 6:1 mixture of selenate to selenite) for 60 days. The dietary exposure study by Cleveland et al. (1993) consisted of exposure of 3-month-old bluegills to measured dietary concentrations ranging from 2.3 to 25.0 mg/kg total selenium in the diet (based on seleno-L-methionine) for 90 days. High mortality was observed in the waterborne study. After 30 days exposure to waterborne selenium, 87.5% mortality was observed for fish in the highest exposure treatment, which based on graphical results, corresponded to a whole-body selenium residue of approximately 14.4 mg Se/kg d.w. After 60 days, mortality due to lower waterborne selenium exposures significantly increased (52.5% and 70.0%) for fish containing whole-body residues of approximately 5.2 and 9.5 mg Se/kg d.w., respectively, compared with control fish (10%). Relatively little mortality was observed in the dietary study. After 90 days of exposure, a statistically significant increase in mortality (22.5%) was observed for fish containing mean wholebody residues of approximately 4.8 mg Se/kg d.w. compared with control fish (5%). However, mortality (15 and 17.5%) from dietary exposure did not significantly increase relative to controls for fish containing whole-body residues of approximately 7.8 and 13.4 mg/kg d.w., respectively. Regardless of whether one interprets the mortality results at 4.8 mg Se/kg from the dietary study as anomalous (per DeForest et al., 1999) or not, it is apparent that mortality was higher for fish containing comparable whole-body residues in the waterborne study compared with the dietary study (e.g. 87.5% mortality at 14.4 mg Se/kg d.w. from water vs. 17.5% mortality at 13.4 mg/kg d.w. from the diet). Different responses of growth condition factor also occurred for waterborne and dietary exposed fish containing similar wholebody residues. Based on kinetic differences, Kleinow and Brooks (1986), Bertram and Brooks (1986) have suggested that more than one metabolic pool of selenium may exist in fish (an unbound inorganic pool and an organically-

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bound pool) which may reflect differences in form and route of exposure. Although the results from Cleveland et al. (1993) are suggestive that route of exposure may affect the diagnostic tissue residue (at least in bluegill), two important caveats apply. First, waterborne exposure concentrations in this study are extremely high (0.16– 2.7 mg/l) relative to concentrations of concern for long-term survival of fish populations (0.005 mg/l). Second, the results from this study are based on direct exposure to juvenile bluegills whereas in field settings, mortality and abnormalities in F1 generation larvae resulting from parental exposure are generally considered to be the principal effects of concern from chronic low-level selenium exposure. Thus, the relevancy of these results to common selenium exposure scenarios encountered in field situations can be questioned. Additional analyses of studies relating selenium residues in various tissues to effects at environmentally realistic concentrations are needed to determine whether waterborne-based diagnostic tissues residues are appropriate for application to natural settings.

nearly all measurements being total selenium (US EPA, 1998). Although good correlation between toxicological effects and total selenium in tissues has been reported for laboratory and field exposures (e.g. Hamilton et al., 1990; Lemly, 1993b), it is nevertheless desirable to obtain direct evidence confirming the fraction and identity of organoselenium species in various tissues. Fan et al. (this issue) provide evidence that proteinaceous selenium can constitute a substantial fraction of total selenium in tissues of fish inhabiting irrigation drainage evaporation ponds, but can be variable (i.e. from 15–82% of total selenium). Furthermore, the fraction of proteinaceous selenomethionine, preliminarily correlated with histopathological effects, varied from less than 1% to slightly more than 20% in fish tissue (Fan et al., 2002 this issue). Therefore, if different forms of organoselenium in tissues are shown to exert differential toxicity to aquatic organisms, the importance of speciating selenium in tissues of aquatic organisms is clear.

4. Sediment criterion issues

3.4. Form of selenium in tissue The present understanding is that organoselenium (possibly as peptide or protein-bound forms) is likely to be the most toxicologically relevant fraction of selenium in tissues of top predatory aquatic organisms (US EPA, 1998). This hypothesis is supported by observations of the biotransformation of selenium into selenoamino acids (e.g. selenomethionine, selenocysteine) by primary producers (e.g. Wrench, 1978; Cutter and Bruland, 1984; Cutter, 1985; Fan et al., 1997; Fan and Higashi, 1998) and subsequent transfer to higher food chain organisms (e.g. Reinfelder and Fisher, 1991; for review, see [Cutter, 1991]). As reviewed by Lemly (1998a) excessive amounts of selenium may alter critical protein or enzyme function perhaps through competition with sulfur during protein synthesis. However, with the exception of data presented by Fan et al., 2002 (this issue), little data are available that have speciated total or organoselenium in tissues of higher trophic level aquatic organisms, with

Sediments are a dominant sink for selenium in aquatic ecosystems and represent an important link and exposure source to the benthic-driven food web. Elemental and organic selenide (in detritus) tend to dominate the fraction of total selenium in aquatic sediments, although most measurements have been made on total selenium (Cutter, 1991; US EPA, 1998). Sediment has been proposed as a medium for expressing aquatic life criteria for selenium in western streams (Canton and Van Derveer, 1997; Van Derveer and Canton, 1997), but not without controversy (Hamilton and Lemly, 1999). Issues associated with deriving sediment-based criteria are similar to those discussed previously for the other environmental media. These issues include identifying and relating the most bioavailable form(s) of selenium in sediments to toxicological effects, relating bioavailability and toxicity to various sediment-quality or water- body characteristics, and quantifying appropriate sediment concentration-response relationships.

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Currently, it appears that the data base relating sedimentary selenium to toxicological effects is relatively sparse (US EPA, 1998). In addition, the data do not appear to support broad-based quantitative relationships between toxicity and sediment quality characteristics, although a relationship between dissolved selenium and total organic carbon and bulk sediment concentrations has been derived with data from western streams (Van Derveer and Canton, 1997). Based on the selenium toxicity and bioaccumulation workshop (US EPA, 1998), some important areas of research relating sedimentary selenium to chronic toxicity include: 1. Assessing the relationship between detrital selenium and food web accumulation. 2. Performing laboratory studies of sedimentary selenium and its accumulation and subsequent effects on aquatic invertebrates. 3. Understanding factors that may cause variation in sedimentary selenium accumulation in biota, such as speciation, interspecies differences, assimilation rates, and the relationship between preferred feeding sites and sediment concentrations. 4. Better defining the mechanisms of selenium accumulation from sediments. 5. Evaluating the merit of depurating specimens prior to correlation with selenium in other environmental compartments.

5. Cross cutting issues

5.1. Analytical chemistry A common link to most of the aforementioned criteria issues is the availability of analytical methodologies for measuring selenium in water, sediment and tissue. One fundamental reason for the limited selenium speciation data for organoselenium compounds is the lack of standardized and/or affordable analytical methodologies. Operationally defined fractions of organoselenium can be measured (e.g. Cutter, 1978, 1986), but speciating organoselenium requires knowledge of specific compounds and the use of techniques such as high performance liquid chromatography (HPLC) and gas chromatography– mass spectrometry (GC–

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MS). Furthermore, some methods rely on the indirect determination of selenium specie through subtraction of different operationally-defined selenium forms. Such methods have the potential to propagate errors in selenium measurements. Continued development and standardization of affordable analytical methods for selenium speciation in water, sediment and tissues is likely the most pressing need for advancing selenium ecotoxicological research.

5.2. Bioaccumulation From a regulatory perspective, understanding and quantifying selenium bioaccumulation in aquatic ecosystems is critical, particularly for setting site-specific criteria. Biogeochemical cycling of selenium in water bodies is expected to vary as a function of hydrology (e.g. retention time), rates of biotic and abiotic selenium transformations, redox conditions in sediments, food web structure, species-specific differences in uptake and depuration kinetics, and other factors. In practice, a global model attempting to relate selenium in the water column to selenium in food chain organisms (mostly corixids and chironomids) reveals high variability in bioaccumulation relative to water concentrations across different sites (up to two orders of magnitude) (Adams et al., 1998). Lower variability (about a factor of 10) was observed for taxon-specific relationships between water and tissue residues. Fan et al. (2002) also report a large range in BAFs calculated for microphytes and macroinvertebrates inhabiting irrigation drainage evaporation ponds. Variability in BAFs may also be introduced due to temporal asynchrony between water and biota that result from kinetic limitations on selenium uptake and depuration by higher trophic level organisms. To address kinetic and biogeochemical concerns, a mechanistic model of bioaccumulation has also been developed (Bowie et al., 1996) which has been used to predict concentrations in water, sediment, and various biological compartments including fish. Inputs to the model require a significant amount of site-specific data. Currently, validation of this model appears to be limited to Hyco reservoir (Carlton, 2002 personal communication).

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Given the regulatory need to relate concentrations in biota back to chemical loads to water bodies, the demand for mechanisms to accomplish this task will remain. Additional analysis and data are required to evaluate the reliability of both empirical (e.g. BAF) or mechanistic (e.g. biogeochemical models) approaches to assessing bioaccumulation. The model by Bowie et al. (1996) appears promising, but its application and validation need to be extended to other water bodies. Use of a BAF approach needs further analysis to determine the most important factors affecting variability in selenium BAFs and the extent to which BAFs can be extrapolated spatially and temporally.

6. Conclusions Evaluation of many of the selenium issues discussed above is ongoing. Preliminarily, however, the following summary statements can be made based on literature reviews and a workshop on selenium bioaccumulation and aquatic toxicology (US EPA, 1998). In terms of expressing a selenium chronic criterion as a water column concentration, selenite, selenate, and organoselenium are generally the forms of most concern. However, direct confirmation of the most toxicologically important form(s) of organoselenium is lacking. It is likely that consideration of how to express a chronic criterion in water will be limited to operationally defined forms (e.g. total recoverable or dissolved selenium). Furthermore, the available data do not appear to support quantitative relationships between chronic toxicity and water quality characteristics. In regards to a tissue-based criterion, reproductive tissue (ovary, egg) are preferred, but practical concerns and data availability require consideration of other tissues (e.g. whole-body). Several studies have demonstrated good correlation between whole-body residues and chronic toxicity, and the assemblage of species represented by whole-body diagnostic tissue residues appears reasonably diverse. The importance of the route of exposure (food vs. water uptake) in setting diagnostic tissue residues for selenium is not clear.

The available data base may not enable a definitive answer to this question, but additional evaluation is planned. Organoselenium, likely bound to proteins or peptides, is thought to be most toxicologically important in tissues of fish. However, direct confirmation of the toxicologically important forms of organoselenium is lacking. Recent data suggest the fraction of total selenium in the proteinaceous form can vary substantially in fish tissue (e.g. Fan et al., 2002 this issue). These data also suggest that proteinaceous selenomethionine may play a role in the onset of histopathological effects. Elemental and organoselenium (in detritus) are dominant selenium forms in aquatic sediments, although most measurements have been made on total selenium. The sediment toxicity data base is very limited and does not support broad-based quantitative relationships between toxicity and sediment quality characteristics, although the establishment of some relationships has been attempted with limited data. Additional research is needed for assessing the relationship between detrital selenium and food web accumulation, understanding factors that may cause variation in sedimentary selenium accumulation in biota, and better defining the mechanisms of selenium accumulation from sediments. Reliable assessments of bioaccumulation will likely be critical for making site-specific modifications to chronic selenium criteria. Relating water concentrations to tissue concentrations is complicated by the complex biogeochemistry of selenium in aquatic environments and data limitations. Global correlations between water and biota concentrations demonstrate high variability, although some improvement is seen for taxon-specific relationships. However, wider validation of empirical (e.g. BAF) and mechanistic approaches for assessing bioaccumulation is needed. The lack of analytical methods for directly speciating organoselenium in various environmental media is linked to many of the current data gaps and research needs. Improving analytical methodologies to enable affordable and reliable measurement of organoselenium compounds and assessing their relationship to toxicological effects will likely be important for furthering selenium ecotoxicological research.

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Acknowledgements This manuscript was written while the author worked for EPA’s Office of Water, Health and Ecological Criteria Division, which is responsible for developing water quality criteria for various pollutants including selenium. The author is grateful to the hard work and useful recommendations made by experts participating in EPA’s 1998 workshop on selenium bioaccumulation and aquatic toxicity. These experts include: Dr W. Adams, Dr G. Bowie, Dr G. Chapman, Dr G. Cutter, Dr A. Fairbrother (chair), Dr T. Fan, Dr A. Lemly, Dr S. Hamilton, Dr G. Riedel, Dr J. Skorupa and W. Van Derveer. This author also thanks H. Bell, C. Delos, J. Mitchell, E. Sappington, and E. Winchester for their helpful technical and editorial reviews of this manuscript. The views and opinions expressed in this manuscript are those of the author and do not necessarily reflect those of the EPA. References Abernethy, S.G., Mackay, D., McCarty, L.S., 1988. ‘Volume fraction’ correlation for narcosis in aquatic organisms: the key role of partitioning. Environ. Toxicol. Chem. 7, 469 – 481. Adams, W.J., Brix, K.V., Cothern, K.A., Ter, L.M., Cardwell, R.D., Fairbrother, A., Toll, J., 1998. Assessment of selenium food chain transfer and critical exposure factors for avian wildlife species: need for site-specific data. In: Little, E.E., DeLonay, A.J., Greenberg, B.M. (Eds.), Environmental Toxicology and Risk Assessment, vol. 7. American Society for Testing and Materials, Philadelphia, PA ASTM STP 1333. Bailey, F.C., Knight, A.W., Ogle, R.S., Klaine, S.J., 1995. Effect of sulfate level on selenium uptake by Ruppia maritima. Chemosphere 30, 579 –591. Bertram, P.E., Brooks, A.S., 1986. Kinetics of accumulation of selenium from food and water by fathead minnows. Water Res. 20, 877 – 884. Besser, J.M., Huckins, J.N., Little, E.E., La Point, T.W., 1989. Distribution and bioaccumulation of selenium in aquatic microcosms. Environ. Pollut. 62, 1 – 12. Besser, J.M., Canfield, T.J., La Point, T.W., 1993. Bioaccumulation of organic and inorganic selenium in a laboratory food chain. Environ. Toxicol. Chem. 12, 57 –72. Bowie, G.L., Sanders, J.G., Riedel, G.F., Gilmour, C.C., Breitburg, D.L., Cutter, G.A., Porcella, D.B., 1996. Assessing selenium cycling and accumulation in aquatic ecosystems. Water Air Soil Pollut. 90, 93 –104.

111

Brix, K.V., Volosin, J.S., Adams, W.J., Reash, R.J., Carlton, R.G., McIntyre, D.O., 2001. Effects of sulfate on the acute toxicity of selenate to freshwater organisms. Environ. Toxicol. Chem. 20, 1037 – 1045. Canton, S.P., Van Derveer, W.D., 1997. Selenium toxicity to aquatic life: an argument for sediment-based water quality criteria. Environ. Toxicol. Chem. 16, 1255 – 1259. Carlton, R., Personal communication with Rick Carlton, Palo Alto, CA: Electric Power Research Institute. Chapman, P.M., 1999. Selenium: a potential time bomb or just another contaminant. Human Ecol. Risk Assess. 5, 1123 – 1138. Cleveland, L., Little, E.E., Buckler, D.R., Wiedmeyer, R.H., 1993. Toxicity and bioaccumulation of waterborne and dietary selenium in juvenile bluegill (Lepomis macrochirus). Aquat. Toxicol. 27, 265 – 280. Coyle, J.J., Buckler, D.R., Ingersoll, C.G., Fairchild, J.F., May, T.W., 1993. Effect of dietary selenium on the reproductive success of bluegills (Lepomis macrochirus). Environ. Toxicol. Chem. 12, 551 – 565. Cumbie, P.M., Van Horn, S.L., 1978. Selenium accumulation associated with fish mortality and reproductive failure. Proc. Annu. Conf. Southeast Assoc. Fish Wildlife Agency 32, 612 – 624. Cutter, G.A., 1978. Species determination of selenium in natural waters. Anal. Chim. Acta 98, 59 – 66. Cutter, G.A., 1985. Determination of selenium speciation in biogenic particulate material and sediments. Anal. Chem. 57, 2951 – 2955. Cutter, G.A., 1986. Speciation of selenium and arsenic in natural waters and sediments. In: Selenium Speciation, vol. 1. Electric Power Research Institute, Palo Alto, CA EPRI EA-4641. Cutter, G.A., 1989. Freshwater systems. In: Ihnat, M. (Ed.), Occurrence and Distribution of Selenium. CRC Press, Boca Raton, FL, pp. 243 – 262. Cutter, G.A., 1991. Selenium biogeochemistry in reservoirs. EPRI EN-7281, Vol 1, Project 2020-1, Final report. Electric Power Research Institute, Palo Alto, CA, pp.147. Cutter, G.A., Bruland, K.W., 1984. The marine biogeochemistry of selenium: a re-evaluation. Limnol. Oceanogr. 29, 1179 – 1192. DeForest, D.K., Brix, K.V., Adams, W.J., 1999. Critical review of proposed residue-based selenium toxicity thresholds for freshwater fish. Human Ecol. Risk Assess 5, 1187 – 1228. Electrical Power Research Institute, 1999. Effect of sulfate concentration on acute toxicity of selenite and selenate to invertebrates and fish. Technical Report TR-111878. EPRI, Palo Alto, CA. Fan, T.W.-M., Higashi, R.M., 1998. Biochemical fate of selenium in microphytes: Natural bioremediation by volatilization and sedimentation in aquatic environments. In: Frankenberger, W.T., Engberg, R.A. (Eds.), Environmental Chemistry of Selenium. Marcel Dekker, New York, pp. 545 – 563.

112

K.G. Sappington / Aquatic Toxicology 57 (2002) 101–113

Fan, T.W.M., Lane, A.N., Higashi, R.M., 1997. Selenium biotransformations by a euryhaline microalga isolated from a saline evaporation pond. Environ. Sci. Technol. 31, 569 – 576. Fan, T.W.-M., Teh, S.J., Hinton, D.E., Higashi, R.M. Selenium biotransformations into proteinaceous forms by foodweb organisms of selenium-laden drainage waters in California. Aquat. Toxicol. (this issue). Hamilton, S.J., 1998. Responses to technical charge to experts. In: Report on the Peer Consultation Workshop on Selenium Aquatic Toxicity and Bioaccumulation. EPA-822-R98- 007. US Environmental Protection Agency, Office of Water, Office of Science and Technology, Washington, DC. p. C-51– C-66. http://www.epa.gov/ost/selenium/ Hamilton, S.J., 1999. Hypothesis of historical effects from selenium on endangered fish in the Colorado River basin. Human Ecol. Risk Assess. 5, 1153 – 1180. Hamilton, S.J., Lemly, A.D., 1999. Water-sediment controversy in setting environmental standards for selenium. Ecotoxicol. Environ. Saf. 44, 227 –235. Hamilton, S.J., Buhl, K.J., Faerber, N.L., Wiedmeyer, R.H., Bullard, F.A., 1990. Toxicity of organic selenium in the diet of chinook salmon. Environ. Toxicol. Chem. 9, 347 – 358. Hansen, L.D., Maier, K.J., Knight, A.W., 1993. The effect of sulfate on the bioconcentration of selenate by Chironomus decorus and Daphnia magna. Arch. Environ. Contam. Toxicol. 25, 72 – 78. Hermanutz, R.O., Allen, K.N., Roush, T.H., Hedtke, S.F., 1992. Effects of elevated selenium concentrations on bluegills (Lepomis macrochirus) in outdoor experimental streams. Environ. Toxicol. Chem. 11, 217 – 224. Hickie, B.E., McCarty, L.S., Dixon, D.G., 1995. A residuebased toxicokinetic model for pulse- exposure toxicity in aquatic systems. Environ. Toxicol. Chem. 14, 2187 –2197. Hilton, J.W., Hodson, P.V., 1983. Effect of increased dietary carbohydrate on selenium metabolism and toxicity in Rainbow trout (Salmo gairdneri ). J. Nutr. 113, 1241 – 1248. Hilton, J.W., Hodson, P.V., Slinger, S.J., 1980. The requirement and toxicity of selenium in rainbow trout (Salmo gairdneri ). J. Nutr. 110, 2527 –2535. Hodson, P.V., Hilton, J.W., 1983. The nutritional requirements and toxicity to fish of dietary and waterborne selenium. Ecol. Bull. 35, 335 –340. Hunn, J.B., Hamilton, S.J., Buckler, D.R., 1987. Toxicity of sodium selenite to rainbow trout fry. Water Res. 21, 233– 238. Jarvinen, A.W., Ankley, G.T., 1999. Linkage of effects to tissue residues: development of a comprehensive database for aquatic organisms exposed to inorganic and organic chemicals. SETAC Press, Pensacola, FL, p. 364. Kleinow, K.M., Brooks, A.S., 1986. Selenium compounds in the fathead minnow (Pimephales promelas) I. Uptake, distribution, and elimination of orally administered selenate, selenite, and L-selenomethionine. Comp. Biochem. Physiol. 83C, 61 – 69.

Lemly, A.D., 1985. Toxicology of selenium in a freshwater reservoir: Implications for environmental hazard evaluation and safety. Ecotoxicol. Environ. Saf. 10, 314 – 338. Lemly, A.D., 1993a. Guidelines for evaluating selenium data from aquatic monitoring and assessment studies. Environ. Mon. Assess. 28, 83 – 100. Lemly, A.D., 1993b. Teratogenic effects of selenium in natural populations of freshwater fish. Ecotoxicol. Environ. Saf. 26, 181 – 204. Lemly, A.D., 1998a. Pathology of selenium poisoning in fish. In: Frankenberger, W.T., Engberg, R.A. (Eds.), Environmental Chemistry of Selenium. Marcel Dekker, New York, pp. 281 – 296. Lemly, A.D., 1998b. A position paper on selenium in ecotoxicology: a procedure for deriving site-specific water quality criteria. Ecotoxicol. Environ. Saf. 39, 1 – 9. Mancini, J.L., 1983. A method for calculating effects, on aquatic organisms, of time-varying concentrations. Water Res. 10, 1355 – 1362. McCarty, L.S., Mackay, D., 1993. Enhancing ecotoxicological modeling and assessment; body residues and modes of toxic action. Environ. Sci. Technol. 27, 1719 – 1728. McCarty, L.S., Ozburn, G.W., Smith, A.D., Dixon, D.G., 1992. Toxicokinetic modeling of mixtures of organic chemicals. Environ. Toxicol. Chem. 11, 1037 – 1047. Meyer, J.S., Santore, R.C., Bobbitt, J.P., Debrey, L.D., Boese, C.J., Paquin, P.R., Allen, H.E., Bergman, H.L., Di Toro, D.M., 1999. Binding of nickel and copper to fish gills predicts toxicity when water hardness varies, but free-ion activity does not. Environ. Sci. Technol. 33, 913 – 916. Ogle, R.S., Knight, A.W., 1996. Selenium bioaccumulation in aquatic ecosystems: 1. Effects of sulfate on the uptake and toxicity of selenate in Daphnia magna. Arch. Environ. Contam. Toxicol. 30, 274 – 279. Ohlendorf, H.M., Hoffman, D.J., Saiki, M.K., Aldrich, T.W., 1986a. Embryonic mortality and abnormalities of aquatic birds: apparent impacts of selenium from irrigation drainwater. Sci. Total Environ. 52, 49 – 63. Ohlendorf, H.M., Hothem, R.L., Bunck, C.M., Aldrich, T.W., Moore, J.F., 1986b. Relationships between selenium concentrations and avian reproduction. Trans. North Am. Wildlife Nat. Resourc Conf. 51, 330 – 342. Riedel, G.F., Sanders, J.G., 1996. The influence of pH and media composition on the uptake of inorganic selenium by Chlamydomonas reinhardtii. Environ. Tox. Chem. 15, 1577 – 1583. Reinfelder, J.R., Fisher, N.S., 1991. The assimilation of elements ingested by marine copepods. Science 251, 794 – 796. Sarma, Y.S.R.K., Jayaraman, S., 1984. Observations on sulphur-selenium antagonism on the growth of two Desmids. Acta Botanica Indica 12, 57 – 60. Schultz, R., Hermanutz, R., 1990. Transfer of toxic concentrations of selenium from parent to progeny in the fathead minnow (Pimephales promelas). Bull. Environ. Contam. Toxicol. 45, 568 – 573. Skorupa, J.P., 1998. Selenium poisoning of fish and wildlife in nature: lessons from twelve real- world examples. In:

K.G. Sappington / Aquatic Toxicology 57 (2002) 101–113 Frankenberger, W.T., Engberg, R.A. (Eds.), Environmental Chemistry of Selenium. Marcel Dekker, New York, pp. 315 – 354. Stephan, C.E., Mount, D.I., Hansen, D.J., Gentile, J.H., Chapman, G.A., Brungs, W.A., 1985. Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and their Uses. US Environmental Protection Agency, Office of Research and Development, Duluth, MN, p. 98. US EPA, 1987. Ambient aquatic life criteria for selenium. EPA-440/5-87-006. US Environmental Protection Agency, Office of Water, Washington, DC. US EPA, 1994. Interim guidance on determination and use of water-effect ratios for metals. Appendix L In: Water Quality Standards Handbook, Second Edition. US Environmental Protection Agency, Office of Water, Office of Science and Technology, Washington, DC. US EPA, 1998. Report on the Peer Consultation Workshop on Selenium Aquatic Toxicity and Bioaccumulation. EPA822-R-98-007. US Environmental Protection Agency,

113

Office of Water, Office of Science and Technology, Washington, DC. http://www.epa.gov/ost/selenium/ US EPA, 1999. National recommended water quality criteria – correction. EPA-822-Z-99-001. US Environmental Protection Agency, Office of Water, Office of Science and Technology, Washington, DC. Van Derveer, W.D., Canton, S.P., 1997. Selenium sediment toxicity thresholds and derivation of water quality criteria for freshwater biota of western streams. Environ. Toxicol. Chem. 16, 1260 – 1268. Williams, M.J., Ogle, R.S., Knight, A.W., Burau, R.G., 1994. Effects of sulfate on selenate uptake and toxicity in the green alga Selenastrum capricornutum. Arch. Environ. Contam. Toxicol. 27, 449 – 453. Woock, S.E., Garrett, W.R., Partin, W.E., Bryson, W.T., 1987. Decreased survival and teratogenesis during laboratory selenium exposures to Bluegill, Lepomis macrochirus. Bull. Environ. Contam. Toxicol. 39, 998 – 1005. Wrench, J.J., 1978. Selenium metabolism in the marine phytoplankters Tetraselmis tetrathele and Duniella minuta. Mar. Biol. 49, 231 – 236.