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2.2 Biogeochemistry The particular form of selenium is important in predicting its toxicity and mobility. Selenite, because it is typically taken up more quickly than selenate, is more toxic. Apparently, passive adsorption to cell walls is an important aspect of the initial uptake (Bowie et al. 1996). For example, selenite uptake by phytoplankton was much faster than selenate uptake. Over the long term, however, selenite and selenate uptakes were similar and related to their concentration in the water. Apparently, bacteria and phytoplankton are primarily responsible for selenium's cycling in biota, however, toxicity to lower trophic levels requires selenium concentrations one to two orders of magnitude above what is found in the environment (Bowie et al. 1996). Indeed, algae and bacteria can incorporate selenium in their cells to levels more than 100 times greater than background (Masscheleyn and Patrick 1993). Higher trophic levels, such as fish, do not take up as much selenium from water and obtain most of their exposure from food (Bowie et al. 1996). Several studies have examined selenium mobility in soils. For example, laboratory column studies have demonstrated that selenate moved as fast or faster than sulfate although some was retained in the soil (Alemi et al. 1988a). Indeed, when leaching alkaline soils, the addition of sulfate increases the leaching and availability of selenate (Brown and Carter 1969). Another study examined the application of selenium as a sodium salt to soil columns. Selenite was rapidly sorbed at all pH values. In contrast, selenate was mobile at all pH values and was completely leached with less than three pore volumes. Less than one pore volume was needed at pH 7-9 (Alrichs and Hossner 1987). In general, therefore, above pH 7.5, selenate sorption is virtually nonexistent (Masscheleyn and Patrick 1993). Because of the differences in mobility and toxicity, the rate at which selenite or other reduced forms are converted to selenate is an important issue. Fortunately, much is known regarding selenium's biogeochemical cycling. For example, the Eh/pH diagram shown in Fig. 3 demonstrates that selenate is the dominant form in oxidizing waters. Indeed, most selenates are too soluble to persist in soils (Elrashidi et al. 1987). The Eh/pH diagram also shows that selenite and hydrogen selenite dominate in reducing systems. Selenium-VI is reduced to selenium-IV at +200 to +300mv and to elemental selenium or metal selenides below +50mv. Thus, although not as soluble as selenate, selenite and hydrogen selenite contribute to soluble selenium in soils at moderate redox potentials (Elrashidi et al. 1987). Finally, hydrogen selenide is the dominant form in highly reducing systems. In summary, the Eh/pH diagram suggests that under oxidized and moderately reduced conditions, selenium solubility is governed by adsorption-type mechanisms rather than by precipitation/dissolution reactions. Selenium oxyanions can serve as electron acceptors, much as happens with sulfate reduction and Fe and Mn reduction. In other words, direct bacterial mineralization of organic matter occurs by selenate reducers that use selenium as an energy source (Presser 1994). Because of such biological conversions, the selenium species one would expect based on thermodynamics are not always present (Masscheleyn and Patrick 1993). For example, it has been demonstrated that adding nitrate inhibits selenate reduction because the organisms utilize nitrate as an electron acceptor in preference to selenium (Benson 1998). During microbial assimilation, oxidized selenium species are reduced to various selenium-H compounds, such as seleno-amino acids. Organic forms of selenium are similar to those of sulfur and include seleno-amino acids (e.g.; selenocysteine and selenomethionine), methyl selenides and methyl selenones (Dungan and Frankenberger 1999, also see Sect. 3). Organically complexed selenium is reported to be a significant fraction of total soil selenium in some systems. As shown