Selenium Contamination in Water. Группа авторов
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by assimilation into solid phases (e.g. pyrite) (Velinsky and Cutter 1991). FeSe2 on oxidation can be converted to selenate according to the reaction (Eq. 2.1) (Bailey et al. 2012).
Gustafsson and Johnson (1992) have reported that considerable amount of Se fractionation is present in OM rich Bs horizon of Swedish podzols due to the high adsorption capacity of anions by OM. Wang and Gao (2001) have reported that despite having sufficient Se concentration (0.1–0.3 μg/g) in soils, Keshan and Kashin –Beck diseases (Se deficiency) are widespread in the northeastern regions of China. This is because the soil is rich in OM. This result in immobilization of Se and makes it less bioavailable. Se concentration in wetland soils (Cooke and Bruland 1987; Thompson‐Eagle and Frankenburger 1992) and evaporation ponds (Fan and Higashi 1998) also depends on biogeochemical volatilization of Se.
In the environment, reduction of dissolved Se oxyanions to particulate Se in oxygenated water occurs via plant uptake and by microorganisms and affects Se bioavailability. Through a dissimilatory reduction (energy metabolism) process microorganisms can use soluble selenate (SeO42−) and selenite (SeO32−) as electron acceptors and reduce to insoluble and less bioavailable Se(0) or Se(‐II) species. In this process SeO32− acts as an intermediate. These insoluble Se(0) and Se(‐II) are further transformed to volatile organic‐Se species and escape to the atmosphere. Hamdy and Gissel‐Nielsen (1976) have studied under laboratory conditions and reported that the volatilization of reduced Se is higher in sandy soil than a humic soil which is about 8 and 5% of added Se, respectively. The selenate owing to its kinetically inert nature does not get precipitated and remains in equilibrium with other selenium species formed in soil or ground water that are rich in humic acid. The following reactions (Eqs. 2.2 and 2.3) for the reduction of Se oxianions to elemental Se have been reported by Bailey et al. (2012) with the generic carbon species:
(where CH2O represents generic organic carbon compound).
This process provides an efficient way for detoxification of environment as insoluble Se and organic‐Se compounds are less toxic than soluble Se oxyanions (Bailey et al. 2012). Appropriate electron (e−) donor substances such as sugars, alcohols, humic substances, organic acids and hydrogen (Bailey et al. 2012; Korom 1992) have been used by Se‐reducing microorganisms. Se oxyanions are behaved as electron (e−) acceptor in dissimilatory reduction reactions. Dissimilatory reduction process contributes significantly to detoxification of a seleniferous environment as compared to the assimilatory reduction process. Dissimilatory reduction offers a cost‐effective way for environmental remediation and therefore attracts the attention of researchers. Eswayah et al. (2016) have reviewed in detail the microbial transformations of unlike forms of Se, assimilation, dissimilation reduction process, anaerobic respiration of several microorganisms supported by dissimilatory reduction, formation of nanoparticulate elemental selenium, etc. Retention of Se by cattle manure has also been reported by Ogaard et al. (2006). In the batch studies the group observed that on addition of cattle manure there is a decrease in adsorption of both selenite and selenate to a loam soil, since manure contains organic acids of low molecular weight such as acetic, formic, and oxalic acids which contend with Se for available sorption sites (Ogaard 1996).
2.5 Interaction of Selenium with Clay Mineral
In soil the presence of clay and minerals plays a vital role in sorption and desorption process of Se species and thus affect their bioavailability. The content of Se is better in soil with elevated clay content than in sandy soil (Gissel‐Nielsen 1975). Gissel‐Nielsen (1971) has studied the effect of texture and pH value on Se uptake from sandy and loamy soils. Caolinite (Al2Si2O5(OH)4), a clay mineral, is formed by octahedral sheets of alumina having a positive charge at their edges, that serve as binding sites for Se (Bar‐Yosef and Meek 1987). Several other minerals such as Fe/Al oxyhydroxides, manganese oxide, and hydroxyapatite (Monteil‐Rivera et al. 2000) are also present in the soils (Eich‐Greatorex et al. 2010); the presence of such minerals varies from place to place. The available binding sites for these minerals, i.e. positively or negatively charged, are dependent on reaction conditions, e.g. in alkaline and acidic reaction conditions the charge on aluminum hydroxide minerals is negative and positive, respectively. Over the surfaces of these minerals selenate and selenite get sorbed under acidic soil conditions (Balistrieri and Chao 1990). However, selenate has lesser adsorption capacity and more bioavailable than selenite. The presence of citric and oxalic acid affect the sorption of selenate and selenite. The presence of acetic and formic acids facilitates the adsorption of selenate but does not affect the adsorption of selenite (Dynes and Huang 1997; Wijnja and Schulthess 2000). This kind of behavior of selenate and selenite can be elucidated by the formation of an outer‐sphere complex of selenate which bounds with non‐specific anion exchange sites with minerals, whereas selenite forms an inner‐sphere complex by ligand exchange method (i.e. form covalent bonds) that is irreversible in nature (Peak and Sparks 2002). Similar to selenate and selenite, acetic and formic acids form outer‐sphere complexes, whereas citric and oxalic acids form inner‐sphere complexes at the mineral‐water interface (Dynes and Huang 1997; Wijnja and Schulthess 2000). Besides this, sorption of selenite in the presence of fulvic acids has no effect on aluminum oxides (Zuyi et al. 2000), but promotes the adsorption on amorphous iron oxides (Tam et al. 1995).
The adsorption process is affected by the existence of contending ions such as phosphate (PO43−) and sulfate (SO42−), and the nature of bonding formed between the competing ions has also governed the mobility of Se species by affecting their sorption in soil and water bodies. In soil, these species are mainly added through agricultural products including gypsum‐ and phosphate‐rich fertilizers. For Japanese soils it has been reported that sorption of Se decreases in the presence of PO43− ion; for selenite this effect is more pronounced than for selenate. Therefore, in phosphate‐rich areas selenite is more bioavailable (Nakamaru et al. 2006; Nakamaru and Sekine 2008) whereas for sulfate‐rich soil selenate is more available for biota. This kind of behavior of selenite and selenate is due to their chemical similarities to phosphate (Eich‐Greatorex et al. 2010) and sulfate, respectively. This supports the study of Nakamaru and Sekine (2008), i.e. an increase in concentration of sulfate ions does not change selenite sorption. To reduce the selanate sorption for plants it has been reported that BaCl2 is more effective than sulfate ions as it forms an insoluble BaSeO4 compound (Mikkelsen et al. 1989). Further, in neutral medium Se uptake decreased in the presence of sulfate (Williams and Thoronton 1972), whereas increased Se uptake has been observed in an alkaline medium (Carter et al. 1972).
Singh (1979) has studied the consequences of supplementary Se and P on arid matters and also examined the chemical composition of Brassicajuncea Cos, in a greenhouse test. The study reveals that when Se is added without and with P there is a manifold improvement in the concentration of Se in plants with respect to control. Se uptake by Trifolium alexandrinum has increased effectively for all variants of Se on addition of 50 and 100 ppm P. The effect is also dependent on the presence of auxiliary ions in fertilizers (Khattak et al. 1991) such as superphosphate (Ca(H2PO4)2) (Fleming 1962), which reduces the Se uptake that may be due to the existence