SCOPE 48 - Sulphur Cycling on the Continents


The Interactions of Sulphur with other Element Cycles in Ecosystems

Section of Ecology and Systematics, Cornell University, Ithaca, NY, USA
Department of Soil Science, University of Saskatchewan, Saskatoon,Canada


The sulphur cycle is intimately connected with the cycles of all other major elements and many minor elements. The interactions of sulphur with these other element cycles has been discussed extensively in other SCOPE volumes (Likens, 1981; Ivanov and Freney, 1983; Bolin and Cook, 1983; Brimble-combe and Lein, 1989; Degens et al., in press) and is an underlying theme in most of the chapters in this book. It is not our intention here to thoroughly review the interactions of sulphur with other element cycles or to repeat the information found in the other chapters. Rather, we wish to lay out a framework for understanding the diversity of sulphur-element interactions.

The ecological interest in nitrogen and phosphorus stems primarily from their role as plant nutrients. Sulphur also is a plant nutrient, although it is less often limiting than either nitrogen or phosphorus. But sulphur has numerous interactions with other elements beyond its role in plant nutrition. Below we first discuss the role of sulphur as an essential nutrient and then summarize these other types of element interactions.


Sulphur is an essential element for all living plants, animals, and microorganisms. Three amino acids found in almost all proteins (cysteine, cystine, and methionine) contain sulphur. Sulphur is also found in sulpholipids, some vitamins, sulphate esters, and a variety of other compounds. Biogeochemists often distinguish between sulphate-ester sulphur, where the sulphur atom is bonded to oxygen, and carbon-bonded sulphur (which includes the sulphur in amino acids). Usually, most of the sulphur in vascular plants is in amino acids, resulting in a tight and predictable coupling between sulphur, carbon, nitrogen, and phosphorus in plant tissues (Johnson, 1984; Maynard, Stewart and Bettany, 1984; Mitchell, David and Harrison, this volume; Schoenau and Germida, this volume).


The sulphur content of vascular plants is fairly plastic, and sulphur can accumulate in terrestrial plant tissues when soil sulphur is in excess compared to other nutrients such as nitrogen and phosphorus. Most of this sulphur which accumulates in plant tissues is in the form of sulphate esters and/or inorganic sulphate, and carbon-bonded sulphur is relatively non-plastic (Johnson, 1984; Maynard, Stewart and Bettany, 1984; Mitchell, David and Harrison, this volume). Although sulphate esters have not been specifically identified in trees (Mitchell, David and Harrison, this volume), sulphate esters definitely accumulate in crop plants subject to excess sulphate; currently available data do not distinguish between sulphate esters and inorganic sulphate in tree tissues (Johnson, 1984; Maynard, Stewart and Bettany, 1984).

If one considers only carbon-bonded sulphur in plants (that is, excluding inorganic sulphate and sulphate esters) , then the molar N : S ratio varies little (N : S = 31 to 39) in a wide variety of vascular plants and plant tissues, including conifer needles (Kelly and Lambert, 1972), deciduous tree leaves (Johnson, 1984), graminous plants (Dijkshoorn and Van Wijk, 1967), and legumes (Dijkshoorn and Van Wijk, 1967). Again considering only carbon-bonded sulphur, the P : S ratio in vascular plants is typically near 1 : 1 (Maynard, Stewart and Bettany, 1984). Thus, we can define sulphur excesses in soils in terms of the ratios of available N : S and P : S. That is, we would expect that sulphur accumulates in vascular plants in the form of sulphate or sulphate esters when the ratio of available nitrogen to available sulphur in the soil is less than 30 or so (for a nitrogen limited system) or when the ratio of available phosphorus to available sulphur is less than 1 (for a phosphorus limited system).


In the absence of sulphur fertilization, the removal of sulphur in crop harvests can sometimes lead to sulphur limitation, both for rice crops (Lefroy et al., this volume) and for upland agriculture (Maynard, Stewart and Bettany, 1984; Tabatabai, 1984; Schoenau and Germida, this volume). In forests, sulphur is not usually limiting, although sulphur deficiencies are found in some regions of low atmospheric deposition of sulphur and in some soils rich in nitrogen (Mitchell, David and Harrison, this volume). In addition to the potential to limit primary production, sulphur deficiencies in forests can increase the susceptibility of trees to fungal infections by increasing the concentration of non-sulphur-containing amino acids in leaves (Johnson, 1984; Mitchell, David and Harrison, this volume).


Mineralization occurs when microbes catabolize organic molecules, respiring CO2 and releasing sulphur. Carbon, N, S, and P cycles are interdependent because the elements are combined in organic matter in stoichiometric relationships. Although the N and S cycles are similar in many respects, mechanistic differences and imbalanced supplies can cause wide variations among the element ratios of nutrient pools in soils. Diverse communities of decomposers convert heterogeneous mixtures of organic N and S to simple inorganic forms. The chemical compositions of soil organic N and S, however , are largely undefined, and the degradation pathways are poorly understood.

The microbes responsible for mineralization are not confined to a particular group (Schoenau and Germida, this volume). Successional groups of organisms, rather than single species, frequently are implicated in multistep decomposition schemes. Soils are generally nutrient-poor environments, teeming with a diverse heterotrophic community consisting of bacteria, actinomycetes, fungi and protozoa that are in a state of starvation (Williams, 1985; Lynch, 1982). Comminution and mineralization of organic matter is also fostered by soil fauna (Swift, Heal and Anderson, 1979). Since the saprophytic organisms responsible for mineralization are non-specific, decomposer microbes are often studied collectively as 'functional groups' (Coleman, Reid and Cole, 1983). Faunal populations sometimes are better indicators of decomposition than is total microbial biomass, because fauna consume the microbes (Andren, Paustian and Rosswall, 1988).

During mineralization the molecular size of organic matter progressively decreases (depolymerization) before inorganic N and S are released. Organic matter must be broken down into its oligomeric or monomeric constituents before crossing the cell membranes of decomposers, although fauna may feed on particles (Swift, Heal and Anderson, 1979). Extracellular enzymes depolymerize polysaccharides, proteins, peptides, lipids, lignin, and other organic constituents. Aminopolysaccharides, such as the repeating units of N-acetylglucosamine in the chitin of microbial cell wall, typically account for 3 to 10% of soil organic N and likely contribute to N mineralization (Sowden, Chen and Schnitzer, 1977) .Algae and bacteria may produce sulphated polysaccharides in soils, but these ester sulphates do not accumulate (Freney, 1986).

Proteins account for major proportions of soil organic N and S, and the amino acids released by proteases are important sources of mineralizable N and S. Most amino acids are bound in organic polymers or organo-mineral complexes, because free amino acids are rapidly taken up by decomposers and autotrophs (Loll and Bollag, 1983). Soil amino acids usually are determined after acid hydrolysis, but free amino acids are barely detectable in mild extracts (Freney, Stevenson and Beavers, 1982). Free amino acids, however, are the most likely source of mineralizable N and S, and production of NH4+ and H2S during amino acid degradation are viewed as the terminal steps of mineralization.

Nitrogen mineralization is measured as NH4+ plus NO3- production, because NH4+ is often oxidized to NO3-. In contrast, H2S is rarely detected during the degradation of amino-acid-S, and mineralization of soil organic S is measured as sulphate production. Cysteine-S may be oxidized while attached to the organic moiety, and organic S gases may be intermediaries in methionine degradation (Trudinger and Loughlin, 1981; Bremner, 1977). The S in ester sulphates is hydrolysed by sulphatase enzymes and released directly as inorganic sulphate (Fitzgerald, 1976). Sulpholipids in plant leaves account for major fluxes of S in terrestrial ecosystems, but the catabolic pathways of sulphonic acids (-C-SO2-OH) in lipids are uncertain (Harwood and Nicholls, 1979).


Chemical fractionation of organic S identifies the major linkages between organic C and S. Organic sulphate is commonly defined as the fraction of organic S which is reduced by a mixture of hydriodic, formic, and hypophosphorous acids (Freney, 1961). The reducing mixture releases H2S from organic S that is not directly bonded to C; therefore, 'organic sulphate' includes ester sulphate (-C-O-SO2-O-), sulphamates (-C-N-SO2-O-), and the second S in sulphocysteine (-C-S-S-). Organic sulphate typically accounts for 30 to 70% of organic S in soils (Freney, Melville and Williams, 1970). Carbon-bonded sulphur is reducible by Raney nickel. Some workers contend that Raney nickel reduces a well- defined fraction of organic S corresponding to amino-acid-S, and that the non-reducible C-bonded S is probably aliphatic sulphonates or heterocyclic S (Scott and Anderson, 1976; Scott, Bick and Anderson, 1981).

Chemical and physical separations expose general relationships between organic matter and soil environment, management history, and nutrient availability (Bettany, Stewart and Halstead, 1973; Bettany, Saggar and Stewart, 1980). Intensive extractions of organic matter with strong alkali and sonication suggested that organic sulphates were part of the aliphatic side chains of humic molecules, and fertility decreased when the side chains were eliminated (Bettany, Stewart and Saggar, 1979; Bettany, Saggar and Stewart, 1980). Anderson and co-workers (1981) observed that organic N and S were preferentially associated (narrow C-N-S) with the fine clay fractions, and suggested the clay-associated organic matter was a medium-term reservoir of available S.

Chemical separations distinguish between the bond classes of organic S, and some workers consider organic sulphate to be more 'susceptible to mineralization' or 'readily available' than C-bonded S (McLaren and Swift, 1977; Biderbeck, 1978). The possibility that organic sulphate may be more readily available, at least in the short term, is based on its chemical nature and physical accessibility: C-O-S linkages are easily hydrolysed by exocytoplasmic enzymes and chemical reagents (Houghton and Rose, 1976; Fitz-gerald, 1978), and R-C-O-S is situated on aliphatic side chains rather than the condensed aromatic core of humus (Bettany, Stewart and Saggar , 1979). Chemical separations of organic matter, however, are too broad to generalize about the sources of mineralized N and S (McLachlan and De Marco, 1975). Incubation studies, aided by isotopic tracers, indicate that chemical fractions lack 'biological significance', because mineralizable N and S are derived from several different fractions (Chichester , Legg and Stanford, 1975; Freney, Melville and Williams, 1975). Care must therefore be taken with overall generalization of the processes (McGill and Cole, 1981), as mineralization of sulphate to obtain energy and enzymatic hydrolysation of sulphate can occur concurrently.


In contrast to studies in terrestrial ecosystems, the role of sulphur in plant nutrition and the processes of organic sulphur decomposition have received fairly little study in aquatic ecosystems. Because of the high concentration of sulphate in seawater (28 mM; 0.9 g (S) l-1), sulphur is probably never limiting to phytoplankton in marine ecosystems. In freshwater lakes, sulphur has not been found to limit primary production in any lake where it has been studied; however, the potential for sulphur limitation in lakes having extremely low sulphate concentrations has not been examined (Cook and Kelly, this volume). The ratio of C: P: S (molar) in freshwater phytoplankton is approximately 100 : 1: 1 (Cook and Kelly, this volume). An appreciable amount of the sulphur in phytoplankton is probably present as sulphate esters (David and Mitchell, 1985). None the less, in contrast to plants in terrestrial ecosystems, the sulphur content of freshwater phytoplankton seems fairly constant and unresponsive to changes in ambient sulphate concentrations (Cook and Kelly, this volume) . Similarly, even though sulphate concentrations are much higher in seawater than in freshwaters, Cuhel and Waterbury (1984) have reported a C: S ratio for a marine phytoplankter which is very similar to that found for phytoplankton in freshwater lakes.


One aspect of the sulphur metabolism of plants deserves special mention because of its atmospheric consequences: many species of plants, both algae and vascular plants, produce large amount of dimethylsulphonium propionate (DMSP) from sulphur-containing amino acids (Andreae and Jaeschke, this volume). This DMSP is converted to dimethyl sulphide (DMS), either within living plants or upon the death of plant tissues; the resulting production of DMS in the oceans is the largest source of biogenic sulphur to the atmosphere globally (Andreae and Jaeschke, this volume). Although the biochemical pathways of DMSP and DMS production are known, the physiological and ecological significance of this process remain unclear. For marine phytoplankton, DMSP and DMS production have been hypothesized to play a role in osmoregulation (Vairavamuthry, Andreae and Iverson, 1985; Andreae, 1986; Turner et al., 1988) or in buoyancy control (Andreae, 1980; Barnard, Andreae and Iverson, 1984). It has also been postulated that DMSP production is indirectly related to the nitrogen nutrition of algae, with DMSP being a store for excess sulphate taken up while assimilating the molybdenum necessary to synthesize nitrate reductase or nitrogenase (Rudnick and Howarth, 1990; see also Section 4.6 below). This could also be true for vascular plants with high nitrate-reductase activities or with symbiotic nitrogen-fixing bacteria.



Sulphate can serve as a terminal electron acceptor for bacterial respiration in the absence of oxygen, and so dissimilatory sulphate reduction is coupled to carbon oxidation and nutrient mineralization in anoxic environments. Sulphate-reducing bacteria themselves metabolize only a limited range of relatively low-molecular-weight organic substrates (Sørensen, Christensen and Jørgensen, 1981; Ivanov et al., 1989). Other bacteria ferment the complex organic matter in sediments, and their fermentation products become the substrates for the sulphate-reducing bacteria. Thus, it is a bacterial assemblage of fermenters and sulphate reducers which decompose organic carbon. Overall, two moles of carbon are oxidized to carbon dioxide for every mole of sulphate reduced (Jørgensen, 1977, 1982a; Howarth and Teal, 1979; Berner, 1984). The mineralization of other elements in organic matter is stoichiometrically related to the mineralization of carbon.

In coastal marine sediments and in marine wetlands, sulphate reduction is the major form of respiration; this process and the associated fermentation reactions which feed substrates to sulphate-reducing bacteria are responsible for most carbon and nutrient mineralization in these sediments (Jørgensen, 1982a; Howarth, 1984; Ivanov et al., 1989; Giblin and Wieder, this volume). In freshwater environments, sulphate concentrations are much lower than in seawater, and sulphate often limits the rate of sulphate reduction in lake and freshwater-wetland sediments (Lovely and Klug, 1986; Cook and Kelly, this volume; Giblin and Wieder, this volume). None the less, sulphate reduction and associated fermentations frequently account for a significant percentage of the carbon and nutrient mineralization in freshwater sediments (Cook and Kelly, this volume; Giblin and Wieder, this volume). Dissimilatory sulphate reduction is not generally a major process in non-flooded, terrestrial ecosystems since soils tend to be oxic.


The formation of methane and fluxes of methane from sediments are intimately tied to the sulphur cycle. In the absence of sulphate, methanogenesis is the dominant process of organic matter degradation in anoxic sediments (Cook and Kelly, this volume; Giblin and Wieder, this volume). It was once thought that sulphate-reducing bacteria could completely out-compete methanogenic bacteria for substrates, and so methane should not be produced in sulphate-rich marine sediments. There is now abundant evidence that some methane is formed even in the presence of reasonably high concentrations of sulphate (see review by Ivanov et al., 1989). However , methane is also rapidly oxidized in the presence of sulphate. Although the mechanism of this anoxic consumption of methane remains somewhat unclear, it appears to be coupled to sulphate reduction, either directly or indirectly (Ivanov et al., 1989; Cook and Kelly, this volume; Giblin and Wieder, this volume). The net result is that methane fluxes from sediments are inversely correlated with sulphate concentrations in freshwater and brackish environments (King and Wiebe, 1980; Bartlett et al., 1987; Yavitt, Lang and Wieder, 1987; Giblin and Wieder, this volume). Flooded rice paddies are thought to be major sources of methane to the world's atmosphere (Crutzen, 1987; Aselmann and Crutzen, 1989; Schutz, Seiler and Rennenberg, 1990). The addition of sulphur fertilizer to rice paddies may reduce the emission of methane from these systems.


Energy is conserved in the reduced sulphur compounds formed during sulphate reduction (Fenchel and Jørgensen, 1977; Howarth, 1984). Relative to aerobic respiration, approximately 75% of the potential energy in the organic matter consumed is transferred and stored in reduced sulphur during sulphate reduction (Howarth and Teal, 1980). When these reduced sulphur compounds are re-oxidized to sulphate, the energy is available to support bacterial growth. A variety of chemosynthetic bacteria can oxidize reduced sulphur compounds using oxygen or nitrate as electron acceptors and fixing carbon dioxide (Kelly, 1982; Jørgensen, 1982b). These chemosynthetic baceria also assimilate other major and minor elements in stoichiometric proportion to carbon fixation, just as with primary production by plants. Chemosynthetic sulphur oxidizers are commonly found at the oxic/anoxic interface in sediments where both reduced sulphur and oxygen are abundant (Lein and Ivanov, this volume). Calculations based on energy flow suggest that chemosynthetic bacteria may fix enough carbon dioxide into organic carbon to support between 3 and 18% of the total rate of respiration in various near- shore marine sediments (Howarth, 1984).

The chemosynthetic bacteria require oxygen or nitrate to oxidize reduced sulphur compounds. However, sulphide can also be oxidized in the absence of oxygen and nitrate if light is present; a variety of photosynthetic bacteria, including purple and green sulphur bacteria and some cyanobacteria, use sulphide as an electron donor in anoxygenic photosynthesis (Pfennig, 1975; Gorlenko, 1979; Cohen, Gorlenko and Bonch-Osmolovskaya, 1989). Anoxygenic, photosynthetic bacteria occur in large numbers in the upper layers of anoxic water columns when enough light is present to support their photosynthesis. They are the dominant primary producers in some meromictic lakes, particularly those with relatively little iron to precipitate sulphides (Gorlenko, in press). Photosynthetic sulphur bacteria are also very active in bacterial mats on the surface of sediments receiving light (Cohen, Gorlenko and Bonch-Osmolovskaya, 1989). As with plants and chemosynthetic bacteria, carbon fixation by photosynthetic sulphur bacteria is stoichiometrically coupled to nutrient assimilation.


Most oxidation reactions of sulphur produce protons (acidity), while most reduction reactions consume protons. The process of sulphate reduction generates alkalinity and consumes protons in stoichiometric proportion as follows:

SO42- + 2(CH2O) + 2H+ = H2S + 2CO2 + 2H2O

Sulphate reduction in lakes helps mitigate the effects of acid deposition by generating alkalinity and consuming some of the acidity associated with sulphuric acid inputs (Schindler et al., 1986; Cook et al., 1986; Cook and Kelly, this volume). However, the mitigating effect persists only as long as the reduced sulphur formed from sulphate reduction is stored within the lake; if the reduced sulphur is re-oxidized to sulphate, acidity is produced and alkalinity consumed. For example:

H2S + 2O2 = SO42- + 2H+

Reduced sulphur formed during sulphate reduction can be stored in sediments either as metal sulphide minerals (principally iron sulphides) or as organic sulphur compounds formed by nucleophilic attack of sulphide on organic matter (Nriagu and Soon, 1985; Rudd, Kelly and Furitani, 1986; Cook and Kelly, this volume; Luther and Church, this volume). In some lakes, the availability of iron may limit the formation of iron sulphides, thereby limiting the storage of reduced sulphur and the extent of permanent alkalinity generation from sulphate reduction (Schindler, 1985; Carignan and Tessier, 1988; Giblin et al., 1990).

Many wetland soils contain large amounts of pyrite (FeS2) as a result of sulphate reduction and sulphide precipitation. Marine wetlands are particularly rich in pyrite (Howarth, 1984; Giblin and Wieder, this volume; Luther and Church, this volume). If these soils are drained, oxygen penetration increases, and pyrite is oxidized. The pyrite oxidation stoichiometrically creates protons:

FeS2 + 15/4O2 + 5/2H2O = 2SO42- + FeOOH + 4H+

When marine and brackish wetland are drained, the soils can become so acidic as to be unable to support agricultural use or other plant growth (Van Breemen, 1982).

Coal seams also often contain large amounts of pyrite, a result of sulphate reduction in the ancient sediments from which the coal evolved. When the coal is mined, the pyrite is exposed to oxygen in air and oxygenated drainage waters. The pyrite is oxidized, and the resulting acid-mine drainage is a major environmental problem in many mining regions.

Elemental sulphur is often added to agricultural soils as a sulphur fertilizer (Schoenau and Germida, this volume). This sulphur is oxidized to sulphate, creating acidity:

Sº + 3/2O2 + H2O = SO42- + 2H+

The addition of sulphur as sulphate in fertilizers, such as in CaSO4 or in superphosphate, will not acidify the soil.


The inorganic chemistry of sulphur can affect the movement and availability of a variety of other elements. In many soils, sulphate is the most abundant divalent anion and is often the most abundant anion of any charge. As such, its abundance can influence the movement and adsorption of basic and acidic cations. In reducing environments, sulphides and polysulphides precipitate and complex with a variety of metals, often controlling the chemistry of the metals. Indirectly, precipitation of metal sulphides can control the availability of other elements, such as phosphorus. Below, we briefly present examples of some of these interactions.


In soils where sulphate is mobile, it is commonly the major anion serving as a 'carrier' of protons and other cations, determining their mobility (Johnson, 1984; Mitchell and Fuller, 1988; Mitchell, David and Harrison, this volume). The cation exchange capacity and base saturation of a soil determine which cations leach in association with the mobile sulphate; protons and the Al3+are leached in those soils which have very low concentrations of basic cations and which have low cation exchange capacities (Mitchell, David and Harrison, this volume). In such soils, sulphate anions thus partially regulate the down- stream export of acid and toxic aluminium.

The mobility of sulphate anions in soils is a function of interactions with other elements. The mechanisms of sulphate retention in soils are complex and are not completely understood, but adsorption on to aluminium and iron hydroxides appears to be an important mechanism in many soils (Rajan, 1978; Johnson, 1984; Mitchell, David and Harrison, this volume). Precipitation of aluminium sulphate minerals may also occur when the pH is low and aluminium concentrations are high (Matzner and Ulrich, 1981; Johnson, 1984; Mitchell, David and Harrison, this volume). Iron sulphate minerals can be important in retaining sulphate in drained (oxic) wetland soils (Van Breemen, 1982).


In reducing sediments and flooded wetland soils, soluble sulphides (H2S, HS-, S2-) are formed from sulphate reduction. Most of the sulphides are rapidly precipitated as metal sulphides, predominantly iron monosulphides and pyrite (FeS2) (Berner, 1984; Howarth, 1984; Howarth and Jorgensen, 1984; Luther and Church, this volume; Giblin and Wieder, this volume). The precipitation keeps concentrations of soluble sulphides relatively low, which can be important since soluble sulphides are toxic to a variety of organisms. In many salt marshes, the rapid precipitation of sulphides in pyrite may minimize the toxic effects of sulphide and allow for the high plant productivity typically observed (Howarth and Teal, 1979; Chalmers, 1982). In some wetlands, however, a low availability of iron may limit pyrite formation, favouring higher concentrations of soluble sulphides and organic sulphur (Howarth, 1984).


Most of the iron for precipitation of iron sulphides comes from reduction of iron(III) oxides and hydroxides within the sediment (Lord and Church, 1983; Berner, 1984; Canfield, 1989; Luther and Church, this volume). Significant amounts of oxidized iron persist even at depth in reducing sediments, but the most reactive oxidized iron is reduced and converted to iron sulphides (Berner, 1984). Since phosphate is highly sorptive on reactive iron(III) minerals, phosphate is released as the iron is reduced. This has long been recognized as a mechanism for returning phosphorus from sediments to the water column in lakes (Mortimer, 1941), although non-iron phases are also efficient adsorbers of phosphate in many lake sediments (Schindler, Hesslein and Kipphut, 1977; Bostrom, Jansson and Forsber, 1982; Wodka, Effler and Driscoll, 1985) and in calcaraeous marine sediments (Morse et al., 1987). Caraco, Cole and Likens (1989) have proposed that the release of phosphorus from both freshwater and marine sediments is correlated with the sulphate concentration of the overlying water. This may explain why there tends to be less phosphorus adsorption in non-calcaraeous marine sediments than in freshwater lake sediments (Caraco, Cole and Likens, 1990).


Although iron sulphides are the most abundant metal sulphides in most sediments, sulphides form highly insoluble minerals with a variety of other metals, including lead, copper, cadmium, zinc, silver, and mercury (Framson and Leckie, 1978). These trace metals can also co-precipitate in iron monosulphides and pyrite (Luther et al., 1980). However, many of these metals also form stable, soluble complexes with sulphides and polysulphides (Boulegue et al.,1982; Luther and Church, this volume). Thus, the mobility of many trace metals in anoxic sediments, soils, and waters is controlled by the balance between precipitation as metal sulphides and complexation with polysulphides (Boulegue et al., 1982; Emerson, Jacobs and Tebo, 1982).


Many trace elements have biogeochemical cycles which closely parallel those of more abundant elements. For instance, the following analogues and substitutions are well known: strontium for calcium, arsenic for phosphorus, and germanium for silica. The cycles of two trace elements, selenium and molybdenum, are intimately related to the cycle of sulphur.


Selenium is an essential trace element which is toxic if present in excessive concentrations (Fergusson, 1990). In natural systems, it exists as selenate (SeO42-), selenite (SeO32-), and organic and inorganic forms of selenide (Se2-) (Cutter and Bruland, 1984; Fergusson, 1990). Chemically, these are all similar to forms of sulphur: sulphate, sulphite, and sulphides. In the selenide form, selenium substitutes for sulphur in proteins (Stadtman, 1974). Sulphate can inhibit the assimilation of selenate by some bacteria (Brown and Shrift, 1980; Bryant and Laishley, 1988).

Sulphate-reducing bateria are capable of reducing selenate to selenide (Zehr and Oremland, 1987). However, even low levels of sulphate inhibit selenate reduction by sulphate-reducing bacteria (Zehr and Oremland, 1987). Selenate reduction in sulphate-rich environments is probably mediated by other types of bacteria which are not inhibited by sulphate (Oremland et al., 1989; Steinberg and Oremland, 1990).


Molybdenum is toxic to mammals if present in excessive concentrations (Fergusson, 1990) but is essential for the processes of nitrate reduction in plants and bacteria and of nitrogen fixation in bacteria (Fogg and Wolfe, 1954). Reduced forms of molybdenum bear little resemblance to sulphur compounds (Manheim and Landergren, 1978) .However, the most oxidized form of molybdenum, molybdate (MoO42-), is very similar in stereochemistry to sulphate (Howarth and Cole, 1985). As a result, sulphate can inhibit the assimilation of molybdate. This inhibition has been demonstrated in tomato plants (Stout and Meagher, 1948), bacteria (Elliott and Mortenson, 1975), mammalian intestines (Huising and Matrone, 1975; Cardin and Mason, 1976), and planktonic algae and cyanobacteria (Howarth and Cole, 1985; Cole et al., 1986; Howarth, Marino and Cole, 1988). Sulphate is sometimes fed to sheep to reduce the toxic effects of excessive molybdenum in their food (Huising and Matrone, 1975). In oxic natural waters, molybdate is the thermodynamically stable form of molybdenum (Manheim and Landergren, 1978). Since the environmental abundance of sulphate in natural waters is thousands- to millions-fold higher than that of molybdenum, the inhibitory effect of sulphate can make molybdenum availability quite low. This may contribute to low rates of nitrogen fixation and nitrogen limitation in coastal marine ecosystems and in some saline lakes (Howarth and Cole, 1985; Howarth, Marino and Cole, 1988; Marino et al., 1990).


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