Common section

Fundamentals of the Sulfur, Phosphorus, and Mineral Cycles

Biogeochemical cycles can easily be grouped into two major categories based on the significance of soil minerals to nutrient availability. As previously noted, in soils not receiving anthropogenically fixed organic carbon and nitrogen, photosynthesis and nitrogen fixation rates are major controllers of biological productivity of the soils. Once carbon dioxide and dinitrogen are incorporated into living biomass, the bulk of the cycling of these nutrients involves their movement into and out of the soil organic matter pools. Biomass levels are controlled, at least in part, by dissolution of soil minerals.

The main objective of this chapter is to introduce the primary processes associated with solubilization of sulfur, phosphate, and metal nutrients and the incorporation of the mineral forms of these essential nutrients. Biological inputs into these nutrient cycles are emphasized.

15.1 Sulfur in the Soil Ecosystem

Sulfur atoms occur in soil in a wide variety of organic and inorganic compounds. Sulfur is found in humus, animal, plant and microbial biomass, sulfur‐bearing minerals and a variety of water‐soluble compounds. Volatile sulfur‐containing compounds are found among the gases making up the soil atmosphere and ionic forms contribute to the salt content of interstitial water. The primary water‐soluble ionic sulfur form in soils is sulfate. As a gas, sulfur atoms are detected in hydrogen sulfide, which may occur at significant concentrations in anoxic, waterlogged soils. Sulfur contributes to the soil particulate fraction in both soil organic matter and soil minerals. Examples of sulfur‐bearing minerals include pyrite (FeS₂), occurring in igneous rocks, and gypsum (CaSO₄.2H₂O) and epsomite (MgSO₄.7H₂O).

Quantities of organic and mineral sulfur in soil are extremely variable, ranging from as low as 0.002% in highly leached, weathered soils of humid regions to concentrations as high as 5% in calcareous and saline soils of arid and semiarid regions (Stevenson 1986). Much of the soil sulfur is derived from internal sources but external inputs, such as fertilizers, air pollutants (e.g. sulfur dioxide), acid mine drainage (sulfuric acid), and a variety of xenobiotic compounds (e.g. pesticides), may also be significant sulfur sources. Basically, essential pools of sulfur in soil include not only major components of living biomass (e.g. biotin, coenzyme A, thiamine, glutathione, cysteine, cystine, methionine) (Figure 15.1), but also inorganic sulfur‐based energy sources (e.g. sulfide, S²⁻, elemental sulfur S⁰), as well as terminal electron acceptors supporting microbial metabolism (e.g. sulfate, SO₄ ²⁻).

Structures of organic sulfur compounds found in soil, such as thiamine, biotin, glutathione, and coenzyme A.

Figure 15.1 Examples of organic sulfur compounds found in soil.

A primary property of the sulfur atom contributing to its utility and versatility in biological systems is its wide range of commonly occurring oxidation states. Common valence states of sulfur relating to biological active molecules are −2 (e.g. in organic compounds as sulfhydrals and in mineral sulfides), 0 (elemental sulfur), and + 6 (e.g. organosulfates and mineral sulfates). See Table 15.1 for further demonstration of the variation in valence of sulfur atoms in commonly occurring soil sulfur compounds. Microbes are the primary mediators for cycling the sulfur atom among these various oxidation states, although spontaneous chemical processes contribute to these transformations to a limited degree.

15.2 Biogeochemical Cycling of Sulfur in Soil

A generalized schematic of the sulfur cycle is depicted in Figure 15.2. As with those biogeochemical cycles previously evaluated, sulfur atoms are transferred between a variety of organic and inorganic reservoirs. Primary soil‐resident sulfur pools are organic residues (plant, animal, and microbial biomass plus their decomposition products and the humic substances derived from them) and soil minerals. In contrast to the nitrogen cycle, except in limited situations (e.g. anthropogenically generated sulfur dioxide produced primarily by the burning of fossil fuels), atmospheric contributions of sulfur compounds to this biogeochemical cycle are limited. This diversity of organic and inorganic sulfur reservoirs suggests a complexity in sulfur cycling in soil not previously encountered in our study of carbon and nitrogen cycles. Sulfur cycling is controlled to a large degree by levels of interaction between soil mineral pools and living biomass in a manner unique to the microbial community (that is, inorganic sulfur serves as both a terminal electron accepter, e.g. metabolic reduction of sulfate to sulfide, and a microbial energy source, e.g. oxidation of sulfide to sulfate).

Table 15.1 Oxidation state of major sulfur compounds in soil

Compound	Oxidation state
Organic S (R‐SH)	−2
Sulfide (S²⁻)	−2
Elemental S (S⁰)	0
Thiosulfate (S₂O₃ ²⁻)	2+ (average/S)
Tetrathionate (S₄O₆ ²⁻)	+2 (average/S)
Sulfur dioxide (SO₂)	+4
Sulfite (SO₃ ²⁻)	+4
Sulfate (SO₄ ²⁻)	+6

Diagram depicting the major features of the sulfur cycle, with boxes labeled “soil minerals,” “microbial biomass,” “humic substances,” “organic residues,” “plants,” “animals,” etc.

Figure 15.2 Major features of the sulfur cycle.

· Cycling of sulfur between soil biomass reservoirs and water‐soluble sulfate (assimilation and mineralization): Sulfate incorporated into biomass may be derived from mineralization of native soil organic matter as well as microbial, plant, and animal residues. Organic sulfur compounds contained in the decaying biomass are mineralized and incorporated into newly synthesized cells. Additionally, the microbial cells may also directly incorporate components of partially decayed biomass, including sulfur‐bearing amino acids and vitamins, into their cellular structure, as exemplified by the mineralization of organic wastes (e.g. farmyard manure) in soil, where portions of the sulfur‐bearing decay products of the organic matter are incorporated directly into growing plant and microbial biomass.

· Complexing of sulfide and sulfate into soil minerals and their solubilization to biological available substances: Some mineral sulfur may be lost from the above mineralization‐assimilation cycle through transfer to water‐insoluble reservoirs. Sulfide accumulation in soil minerals has been defined as sulfidization whereas the exposure of sulfide‐bearing minerals to oxidizing conditions resulting in production of sulfate is sulfurization. These processes may involve dissolution or precipitation of the minerals through chemical processes (e.g. precipitation of sulfide by iron or solubilization of minerals under acidic conditions) as well as biological oxidation or reduction of the primary soil mineral components. This portion of the overall sulfur cycle is highlighted because it represents the formation of an inorganic sulfur pool with long‐term implications for the geological portion of the soil biogeochemical cycle.

This model of the sulfur cycle emphasizes the primary sulfur resources in soil ecosystems. Examples of other sulfur‐containing reservoirs that may be of significant environmental importance, at least on a localized basis, include (i) societal waste products incorporated into soil (e.g. sewage biosolids, composts, organic matter in household wastes, and fly ash from electric power production), (ii) agricultural chemicals (including pesticides), and (iii) air pollutants (e.g. products of fossil fuel combustion in electrical power generation and in automobile exhausts). These substances enter the soil sulfur cycle through reactions similar to those already depicted. For example, sulfur dioxide produced from burning fossil fuels enters the interstitial water where it is oxidized to sulfate. The anthropogenically produced organic sulfur is mineralized as described above for native biomass of the ecosystem.

The most obvious atmospheric contribution to the sulfur cycle is the previously mentioned sulfur dioxide, but it must be understood that particulate (dry deposition) sulfur inputs may also be significant on a localized basis. As with the sulfur dioxide, particulate sulfur inputs (primarily as elemental sulfur but also as various sulfates) contribute to soil acidity by increasing sulfate concentrations. An extreme example of the impact of deposition of elemental sulfur on an ecosystem is provided by studies of forest ecosystems in Alberta, Canada. These sites received large quantities of industrially produced sulfur. Kennedy et al. (1988) and Maynard et al. (1986) both report dramatic declines in aboveground vegetation quantity and diversity due to increasing soil sulfur loadings. Sulfur inputs reported by Maynard et al. (1986) ranging from 4100 to 51 400 μg g⁻¹ soil resulted in decreasing soil pH (4.4–2.4) and augmentation of the Thiobacillus thiooxidans populations. Visser and Parkinson (1989), in their study of forest soils receiving comparable elemental sulfur loadings, found a decline in microbial biomass that correlated with soil pH and a parallel reduction in the ability of the soil community to mineralize glucose. The environmental impact of the perturbation of the sulfur cycle resulting from deposition of unusually large quantities of elemental sulfur is a logical outgrowth of the capacity of the native soil microbial population to use this sulfur substance as an energy source.

15.3 Biological Sulfur Oxidation

Chemoorganotrophic and chemolithotrophic bacteria derive their energy from oxidation of organic and mineral substrates. The large range of oxidation states of the sulfur atom (−2 to +6) make it a nearly ideal candidate as a primary energy source for resourceful soil microbes. In soil ecosystems, a variety of soil microbes have the capacity to derive their energy from oxidation of inorganic sulfur compounds. This exploitation of sulfur oxidation is a minor biogeochemical process in that terrestrial ecosystems are predominantly driven by heterotrophic processes. Sulfur oxidation can be an important energy source for some portions of the terrestrial ecosystem, such as exotic deep sea hydrothermal vents. Both sulfide and elemental sulfur are substrates in this energy‐yielding process.

The diversity of the oxidation status of the sulfur atom allows for the concurrent oxidation and reduction of the same compound by the same microbe. In this disproportionation process, the sulfur atom is serving as both an electron donor and electron acceptor for the microorganism. Examples of the process are demonstrated by the disproportionation of thiosulfate and sulfite by Desulfovibrio sulfodismutans. Sulfate and sulfide are formed from the sulfur compounds by the following reactions:

15.3.1 Microbiology of Sulfur Oxidation

Oxidation of sulfur‐containing compounds, especially elemental sulfur, has a major impact on overall ecosystem function, primarily due to the fact that the terminal product of the reaction, sulfate, results in acidification of the soil. Elemental sulfur has long been used as a soil amendment when lowering of soil pH is required. Environmental degradation has been linked with excessive oxidation of sulfur, as can result from mining of sulfide‐bearing ores (acid mine drainage) or atmospheric deposition of sulfur dioxide due to acid rain produced during energy recovery from fossil fuels. Thus, an understanding of the biological interactions involved with the oxidation of sulfur compounds and the nuances of environmental management of the process is essential for proper stewardship of soil resources as well as for reclamation of mismanaged systems.

15.3.1.1 The Thiobacilli

Although a diverse group of soil microbes are capable of oxidizing sulfur and sulfide for energy, members of the bacterial genus Thiobacillus are the primary mediators of this process in soil ecosystems. Some studies suggest that heterotrophic bacteria and fungi may also be of some importance in this biogeochemical process, but the significance of their contribution has yet to be elucidated.

The thiobacilli are chemoautotrophic or facultatively chemoautotrophic bacteria that are widely distributed in soil ecosystems. These generally motile, gram‐negative bacteria are distinguished from other members of the soil community by their capacity to catalyze at least some portion of the oxidative continuum of sulfide, elemental sulfur, thiosulfate, and/or tetrathionate to sulfate. Among the energy substrates of the five most studied members of this genus (T. thiooxidans, T. ferrooxidans, T. thioparus, T. novellus, and T. denitrificans ), elemental sulfur, thiosulfate, tetrathionate, and sulfide are the most commonly listed. It must be noted that each species of the genus Thiobacillus is capable of oxidizing a specific portion of the spectrum of oxidizable sulfur compounds occurring in soil. For example, not all species or strains within a given species are capable of oxidizing elemental sulfur. Furthermore, not all Thiobacillus species are limited to deriving their energy from oxidation of sulfur compounds; fixed carbon and ferrous iron may also be oxidized. The described species can be divided into three groups based on their sensitivity to soil pH and their capability to oxidize fixed carbon: neutrophilic chemoautotrophs, neutrophilic facultative autotrophs, and acidophiles (Kuenen et al. 1992). Of the most studied members of this genus, T. novellus is a facultative chemoautotroph and T. ferrooxidans recovers some energy from ferrous iron oxidation (Table 15.2).

These sulfur‐oxidizing bacteria are further distinguished by the nature of their terminal electron acceptor and their resistance to acidic conditions. Resistance to extreme acidity by thiobacilli is particularly valuable in that they may produce sufficient sulfate to lower the soil pH to values between 1 and 3. Most members of the genus are obligate aerobes, although T. denitrificans can use oxygen or nitrate as its terminal electron acceptor and T. ferrooxidans transfers electrons to ferric ion under anoxic conditions. A wide range of tolerance to acidity is characteristic of the genus. For example, T. thiooxidans and T. ferrooxidans have pH optima between 2.0 and 3.5. In contrast, T. thioparus, T. denitrificans , and T. novellus grow optimally at near neutral pH values (Table 15.2). For a more general discussion of the properties and function of Thiobacillus species, see Alexander (1977), Harrison (1984), Kuenen et al. (1992), Stevenson (1986) or Visniac and Santer (1957).

Table 15.2 Metabolic properties of commonly studied Thiobacillus species

Thiobacillus species	Optimum pH	Energy substrates	Electron acceptors
T. thiooxidans	2.0–3.5	S⁰, S₂O₃ ²⁻, S₄O₆ ²⁻	O₂
T. ferrooxidans	2.0–3.5	S⁰, S₂O₃ ²⁻, Fe²⁺	O₂
T. thioparus	Approx. 7.0	S⁰, S²⁻, S₂O₃ ²⁻, S₄O₆ ²⁻	O₂
T. novellus	Approx. 7.0	S₂O₃ ²⁻, organic compounds	O₂
T. denitrificans	Approx. 7.0	S⁰, S²⁻, S₂O₃ ²⁻, S₄O₆ ²⁻	O₂, NO₃ ⁻

The kinetics of sulfate formation by these microbes can be exemplified by a list of the reactions involved in elemental sulfur, thiosulfate, and tetrathionate by T. thiooxidans:

T. ferrooxidans oxidizes elemental sulfur and thiosulfate by reactions similar to those listed above for T. thiooxidans, but this oxidation is combined with the oxidation of ferrous to ferric iron as follows:

This species also contains enzymes capable of reducing ferric iron (Fe³⁺) (Sugio et al. 1987), tetravalent manganese (Mn⁴⁺) (Sugio et al. 1988a), cupric ion (Cu²⁺) (Sugio et al. 1990), and molybdic ion (Mo⁶⁺) (Sugio et al. 1988b). The environmental significance of the latter processes has not been elucidated.

The oxidative process for T. denitrificans is coupled to the reduction of nitrate as follows:

The environmental significance of this process is that it provides a mechanism for sulfur oxidation in anoxic environments. Alternatively, ferric ion may also serve as a terminal electron acceptor for sulfur oxidizing T. ferrooxidans growing in the absence of molecular oxygen. Pronk et al. (1992) have shown a linear relationship between ferrous iron accumulation and cell density in T. ferrooxidans reducing ferric iron and oxidizing elemental sulfur. Suzuki et al. (1990) have also shown that several strains of T. ferrooxidans are capable of oxidizing elemental sulfur under anoxic conditions when using ferric ion as a terminal electron acceptor. Some of the strains oxidized the sulfur with ferric ion as the electron acceptor at rates comparable to those of the aerobic process, but with others, sulfur oxidation occurred at a rate 50% or more lower than that occurring in the presence of oxygen.

A variety of other metal‐mobilizing, sulfur‐oxidizing Thiobacillus species have been evaluated, e.g. T. acidophilus (Meulenbert et al. 1992), T. prosperus (Huber and Stetter 1989), T. versutus (Beffa et al. 1991) and T. cuprinus (Huber and Stetter 1990). Other autotrophic bacteria that catalyze sulfur oxidation under aquatic or more specialized conditions include phototrophic bacteria (Brune 1989) and Sulfolobus species that function in hot springs (Alexander 1977).

15.3.1.2 Heterotrophic Sulfur Oxidizers

Several heterotrophic microbes also catalyze sulfur oxidation. Aspergillus niger, Mucor flavus, and Trichoderma harzianum, for example, oxidize elemental sulfur (Grayston et al. 1986). These fungi produce substantial amounts of sulfate from sulfur oxidation with thiosulfate (A. niger) or thiosulfate and tetrathionate (M. flavus) accumulating as intermediates. Aureobasidium pullulans (de Bary) Arnaud, isolated from sycamore phylloplane previously exposed to atmospheric pollution, oxidized elemental sulfur to dithionite, tetrathionate, and sulfate (Killham et al. 1981). Similar products are produced by yeast isolates of the genus Rhodotorula (Kurek 1979, 1985) and bacteria (Pepper and Miller 1978) and actinomycete (Yagi et al. 1971) isolates. These heterotrophic microorganisms utilize carbonaceous substrates for their carbon and energy source and generally accumulate oxidized sulfur compounds after the period of active growth. Thus, their role in sulfur oxidation in native soil ecosystems is questioned.

Indirect evidence suggests that there is at least some role for the heterotrophs in elemental sulfur oxidation in soil. Lawrence et al. (1988) evaluated the effect of elemental sulfur fertilization on a variety of measures of sulfur oxidation in two Gray luvisolic soils from Canada. In one soil, increases in Thiobacillus populations were induced by elemental sulfur amendment whereas in the other soil no change in these populations was detected. Heterotrophic sulfur oxidizers were the most abundant sulfur oxidizers in both soils. In a related study (Lawrence and Germida 1988), elemental sulfur oxidation in 28 Canadian soils correlated linearly with soil microbial biomass carbon (r = 0.68, p < 0.01) and soil respiration (r = 0.88, p < 0.01). Further evidence implicating a heterotrophic contribution to the sulfur oxidation included (i) augmentation of heterotrophic biomass through glucose amendment increased sulfur oxidation rates and (ii) inhibition of general heterotrophic microbial activity reduced sulfur oxidation rates.

In contrast, Lee et al. (1987), in their study of 48 New Zealand soils, found that elemental sulfur amendment resulted in increases of thiobacilli populations with no change or a decline in heterotrophic sulfur‐oxidizing populations. These reports suggest that environmental selection is likely of great significance in determining the dominant populations involved in sulfur oxidation in soil ecosystems. Of primary importance in considering the impact of autotrophic vs heterotrophic populations in modified sulfur cement oxidation is the requirement of oxidizable carbon substrates for function of the heterotrophs. These substrates are concentrated in surface horizons of the soil profile and are limiting in subsoils (Tate 1987).

15.3.2 Environmental Conditions Affecting Sulfur Oxidation

Although sulfur‐oxidizing microbes are considered to be ubiquitous, populations of autotrophic thiobacilli can be essentially undetectable in some soils. Increases in populations can result from sulfur amendment to the soil. For example, thiobacilli population augmentation by sulfur amendment has been demonstrated with several soils from Kansas (Attoe and Olson 1966), Canada (Janzen and Bettany 1987a), and New Zealand (Lee et al. 1988a). The increased sulfur‐oxidizing population induced by sulfur amendments can result in increased sulfur oxidation rates in subsequent growing seasons (Lee et al. 1988b). These observations relate specifically to Thiobacillus populations. Since, as indicated above, a variety of heterotrophic microbes are also capable of oxidizing sulfur in soil, it is reasonable to conclude that sulfur oxidation is not precluded in soils due to nonexistence of requisite microbes. Significant autotrophic or heterotrophic populations can be anticipated to occur in any biologically active soil. The rate may be limited due to initially low population densities but the numbers of autotrophic microbes should increase following amendment with sulfur substrates. Therefore, it can be concluded that since the requisite sulfur oxidizers can be assumed to be present in any given soil ecosystem, oxidation of sulfur is dependent primarily upon occurrence of physical and chemical conditions conducive to the function of these microbes.

Basically, for a functional soil microbial community to develop there must be both an energy source (an oxidizable substrate) and an electron acceptor. For most living organisms, these are a fixed carbon supply, such as simple sugars, carbohydrates or amino acids, and oxygen. With elemental sulfur oxidation, the sulfur can serve as the electron source, i.e. the oxidizable substrate, and oxygen is the electron acceptor. As noted above, one species of elemental sulfur oxidizer, T. denitrificans , that is capable of oxidizing sulfur under anoxic conditions has been isolated. Cytochromes are used in the electron transport system and the energy yield approaches that gained when oxygen serves as the electron acceptor. Although some strains of T. denitrificans have been shown to be unable to oxidize extracellularly available elemental sulfur anaerobically (Schedel and Trueper 1980), others function quite efficiently with nitrate as an electron acceptor and elemental sulfur as the energy source. For example, a T. denitrificans ‐based system is sufficiently efficient that it has been proposed as a means of removing excess nitrate from waste water. Elemental sulfur serves as a cheap source of energy for the denitrification process (Batchelor and Lawrence 1978). High percentages of denitrification (over 95%) have been demonstrated using this process (Hasimoto et al. 1987).

The use of elemental sulfur as an energy source in soil is limited by the capacity of the microorganism to interact with the substrate. Physical attachment of thiobacilli to sulfur particles is required for enzymatic oxidation of sulfur. This interaction of the microbes with sulfur particles can be the rate‐limiting factor determining the sulfur oxidation rate. This was shown clearly 80 years ago in a study of T. oxidans conducted by Vogler and Umbreit (1941). Additionally, Espejo and Romero (1987) demonstrated that for T. ferrooxidans only bacteria attached to the sulfur particle were capable of growth. No replication occurred once the bacteria were detached from the sulfur particle. Growth of the microbes was not uniform on the sulfur particle surface but rather it occurred in distortions present on the particle surface. Staining of sulfur prills with crystal violet revealed that even after more than four months incubation, intact regions remained on the substrate surface whereas destruction of the surface had occurred quite deeply at other locations on the prill.

Strains of T. ferrooxidans also have a more versatile energy recovery capability. When attached to sulfur particles, sulfur oxidation is the primary energy‐yielding mechanism. Microbes that become detached from the sulfur particles can survive through the oxidation of ferrous ion to ferric ion. The sulfur oxidation rate was not affected by the presence of ferrous ion in the growth medium (Espejo et al. 1988). For T. albertis specific cellular adhesions, bacterial glycocalyx, are produced to facilitate the interaction of the thiobacilli cells with the sulfur particle (Bryant et al. 1984). Elemental sulfur oxidation by the latter Thiobacillus species is proportional to particle size, surface area per unit weight and the crystal microstructure of the elemental sulfur (Laishley et al. 1986).

Since the elemental sulfur particles do not dissolve in water and direct contact of the microbial cell with the sulfur particle is necessary for oxidation to proceed, it can be concluded that surface area of the sulfur particles controls their oxidation rate in soils. This has been demonstrated repeatedly in a wide variety of soils over the last several decades (e.g. see Bertramson et al. 1950; Frederick and Starkey 1948; Janzen and Bettany 1987b; Laishley et al. 1986; Lee et al. 1988a, b; McCaskill and Blair 1987; Nor and Tabatabai 1977; Watkinson 1989). For each of these studies, a size distribution of sulfur particles was added to soil and sulfate concentration or pH changes due to sulfur oxidation were quantified as a function of particle size. In each case, rapid oxidation was detected with small particles and limited reaction of the larger particles occurred. For example, Lee et al. (1988a) found that after 340 days, 80–90% of elemental sulfur particles less than 0.15 mm diameter were oxidized as opposed to only 24–55% of the particles greater than 0.15 mm diameter. McCaskill and Blair (1987) found that with mixtures of particle sizes, a biphasic oxidation curve could be detected. A reduction in conversion rate occurred once the 0.1 and 0.2 mm diameter particles were exhausted. Frederick and Starkey (1948) evaluated particles of sulfur in pipe sealing mixtures. They found that over a three‐week incubation period, sulfur oxidation was essentially nondetectable with particle sizes greater than about 3 mm mean diameter.

Along with availability of sulfur‐oxidizing microbes in soil and their growth substrates, sulfur oxidation is controlled by soil moisture, pH, and temperature. Because of the physical stability of elemental sulfur particles in soil, the impact of these factors on sulfur oxidation relates to their effect on the microbes catalyzing the reactions.

In soil, the rates of oxidation of sulfur‐containing compounds and elemental sulfur tend to increase with increasing soil moisture, reach an optimum level, and then rapidly decline (e.g. Attoe and Olson 1966; Janzen and Bettany 1987b; Lettl et al. 1981; Moser and Olson 1953). With slight variations dependent upon soil texture, optimum soil moisture levels are near field capacity.

Complex interactions between soil moisture, soil texture, and microbial populations alter elemental sulfur oxidation rates (Janzen and Bettany 1987a, b; Watkinson 1989). In one study of soils with limiting oxidizer population densities (as indicted by responses to inoculation), elemental S conversion rate was not affected by clay contents ranging from 9% to 52% (McCaskill and Blair 1987). This contrasts with data from Janzen and Bettany (1987b) who found that with 40 diverse Canadian soils, the elemental sulfur oxidation rate was negatively correlated with soil clay contents. Watkinson (1989) found that the improved soil porosity resulting from amendment of phosphate rock to mineral soils augmented elemental sulfur oxidation. It can be concluded that the prevailing controller of elemental sulfur oxidation in soils of varying moisture and texture is the availability of the obligatory electron acceptor, molecular oxygen.

Although elemental sulfur is used to reduce soil pH, its oxidation rate actually generally increases when pH is raised. Both Attoe and Olson (1966) and Janzen and Bettany (1987b) found a positive correlation of elemental sulfur oxidation rate and soil pH in a wide variety of soils. Further, Lettl et al. (1981) noted that oxidation of elemental sulfur was stimulated by amendment of soil with calcium carbonate.

Elemental sulfur oxidation occurs between 4% and 55 °C (Stevenson 1986), a range well within that normally observed in soil systems. Therefore, the sulfur oxidation rate is generally directly proportional to the temperature in soil systems (Ahonen and Tuovinen 1990; Attoe and Olson 1966; Chapman 1989; Janzen and Bettany 1987c; Nor and Tabatabai 1977). A wide range of Q₁₀ values was reported in these studies: 1.9–3.1 (Chapman 1989), 2.1 (Ahonen and Touvinen 1990), 3.2–4.3 (Janzen and Bettany 1987c).

15.4 Biological Sulfur Reduction

As is the situation with reduction of nitrate in the nitrogen cycle, sulfate can be reduced via both assimilatory and dissimilatory pathways. Assimilatory sulfate reduction is the source of sulfur for biomass synthesis. Therefore, this sulfate reduction process is catalyzed by a variety of organisms, including higher plants, algae, fungi, and bacteria. In contrast, dissimilatory sulfate reduction, sulfidogenesis, is catalyzed by a specialized group of obligately anaerobic bacteria. In soil, the organisms involved with the latter reaction are generally found in the genera Desulfovibrio and Desulfotomaculum. As with nitrate in denitrification, sulfate is the terminal electron acceptor for dissimilatory sulfate reducers. In both dissimilatory processes, the electrons are transferred through a cytochrome‐based electron transport process to the terminal electron acceptor.

With dissimilatory sulfate reduction, sulfate is reduced through the transfer of eight electrons produced by the oxidation of a fixed carbon substrate, as described by the following reaction:

The sulfate atom is activated in this process through the consumption of adenosine triphosphate (ATP). A variety of fixed carbon compounds may be oxidized to provide the electrons for this process. The most commonly studied energy sources are organic acids (e.g. lactate, acetate, fatty acids), but an increasing number of more complex substances, such as various aromatic compounds, are being added to the list of electron donors. Coupling of this process with mineralization of organic soil pollutants is being explored as a means of bioremediation of soils.

Dissimilatory sulfate reduction in soil systems is distinguished by its occurrence only in specialized ecosystems. Aside from the requirement for an oxidizable substrate and sulfate, the primary environmental property delimiting function of these anaerobic bacteria is anoxic conditions. Sulfidogenesis is only detected in soil ecosystems exhibiting extremely reducing conditions. A reducing potential of about −0.25 V is required before dissimilatory sulfate reduction occurs. Thus, this sulfate reduction process is limited to sites similar to those where methanogenesis occurs, i.e. swampy or chronically flooded soils.

Although sulfate is a common component of soil interstitial water, its presence in concentrations supportive of extensive dissimilatory sulfate reduction is limited. A typical example of a land‐based ecosystem that would allow extensive sulfide generation is a marsh site inundated with sea or brackish water. Anoxic conditions prevail in the marsh soils and sulfate is a major anion in seawater. The best indication of current or previously existent sulfidogenic conditions is detection of metal sulfide deposits in the soil profile or detection of hydrogen sulfide fluxes from the site.

It should be noted that although the primary substrate considered for sulfidogenesis is sulfate, elemental sulfur can also be reduced by a variety of microorganisms. For example, Desulfurolobus ambivalens has the capacity to oxidize or to reduce elemental sulfur (Zillig et al. 1986). Desulfotomaculum termoacetoxidans (Min and Zinder 1990), Desulfuromonas acetoxidans (Gebhardt et al. 1985), and Spirillum 5175 N (Zoephel et al. 1988) all reduce elemental sulfur to hydrogen sulfide.

Sulfidogenesis may be managed to ameliorate the impact of a variety of environmental pollutants, both organic and inorganic. For example, this process may be explored as a means of reducing organic contamination of chronically flooded soils, reducing drainage water acidity, and ameliorating the difficulties associated with heavy metal contamination of soils and waters.

15.4.1 Anaerobic Biodegradation

A variety of xenobiotic compounds containing halogenated aromatic rings may enter soil ecosystems in undesirable concentrations. Under aerobic conditions, the halogenated substances are reasonably stable but under anoxic conditions, bacteria have been shown to be capable of removing the halogen atom. Since at least some of these substances are mineralized by the sulfidogenic bacteria, the combination of dehalogenating and sulfate‐reducing populations may be used to remove this troublesome class of compounds from sites that can be easily managed under anoxic conditions. These techniques may be particularly operative in sediments polluted with xenobiotic compounds.

15.4.2 Reducing Acidity of Acid Mine Drainage

Acid mine drainage results from exposure of previously buried sulfide‐bearing soil minerals to oxidative conditions in the presence of an active sulfur‐oxidizing population. The environmental impact of the highly acidified waters is further accentuated by their high metal loadings resulting from the dissolution of the soil and mine slag minerals. Thus, soils and waters receiving these drainage waters are adversely impacted not only by the acidic conditions but also by toxic levels of heavy metals. A means of reducing the environmental impact in these situations is to pass the drainage waters through a swampy area where dissimilatory sulfate reducers can convert the sulfate to sulfide. The metal concentrations of the waters are reduced by their precipitation as metal sulfides. See Mills (1985) for a more detailed discussion of this process.

15.4.3 Reduction of Complications of Metal Contamination in Soil

Sulfidogenesis can also be exploited in situations where soils have been contaminated by heavy metals. As with the acid mine drainage problem outlined above, metal loadings of the leachate waters from the contaminated site may be removed by precipitation as metal sulfides in a sulfate‐reducing system. Mobility of the metals may also be reduced in the site through creation of appropriate conditions for sulfide genesis within the soil profile. The primary difficulty with both the utilization of sulfidogenesis to remove metal contaminants from acid mine drainage and to reclaim metal contaminated sites is that the anoxic conditions of the purification site must be maintained. Should the soil become aerobic, sulfur‐oxidizing bacteria will utilize the sulfide for an energy source, thereby mobilizing the previously stabilized heavy metals.

15.5 Mineralization and Assimilation of Sulfurous Substances

Mineralization of organic sulfur compounds is a microbially catalyzed process analogous to carbon and nitrogen mineralization. Although the relative rates of nitrogen, carbon, and sulfur mineralization in a specific soil system may differ (see Bettany et al. 1980 for an example of a field study and Tabatabai and Al‐Khafaji 1980 for a laboratory‐based study demonstrating this difference), each is dependent upon activity of a variety of heterotrophic bacteria and fungi. This commonalty of mediators results in a similar impact of variation of environmental conditions on process rates. That is, sulfur mineralization is anticipated to occur in any ecosystem wherein heterotrophic microbes can function. Under aerobic conditions, complete mineralization of an organic substrate yields carbon dioxide, ammonium, and sulfate. Under anoxic conditions carbon dioxide, methane, sulfide, and ammonium are primary products. Optimal conditions for mineralization of sulfur‐containing organic compounds include a near neutral pH, mesophilic temperatures, and approximately field capacity soil moisture. It is easily concluded a priori that since sulfur mineralization occurs over essentially the totality of the physical and chemical conditions supportive of microbial life, the process is catalyzed by an extremely diverse group of common soil bacteria and fungal species.

The mineral sulfur produced through the catalytic activity of the soil decomposer community may be assimilated by the plant community or immobilized (assimilated) by the soil microbes. Since ecosystem conditions generally favor assimilation of sulfate into plant or microbial biomass, the bulk of the sulfur contained in soil systems tends to be in the organic fractions (e.g. David et al. 1982; Swank et al. 1984). Since these assimilatory processes are catalyzed by the plant community and the heterotrophic soil bacteria and fungi, the physical and chemical limitations to the activities would be similar to those previously described for overall soil microbial activity.

As has been observed with carbon and nitrogen mineralization, the rate of mineralization of sulfur compounds is proportional to the biodegradation susceptibility of the organic compound itself as well as its physical availability to the degrader population. That is, the decomposition rate is directly dependent upon the capability of the microbial population to synthesize the requisite enzymes necessary for the catalysis and the proximity of the organic substance to the active microbial population. Since both biodegradation‐sensitive (common cellular substituents) and more biodegradation‐resistant sulfurous compounds (humified substances) are found in soil (e.g. Anderson et al. 1981 or Bettany et al. 1979), the conversion of organic sulfur to sulfate in an aerobic soil system may or may not be proportional to the total soil organic matter content. This relationship is totally dependent upon the relative contribution of biodegradable and accessible organic sulfur to biologically stabilized (e.g. humic acid) sulfur.

15.6 The Phosphorus Cycle

Processes associated with transformation of phosphorus occurring in soil occur by the same principles as the other soil‐based biogeochemical cycles. Mineralization and assimilation are catalyzed by the general soil biological community. The phosphorus cycle, like the sulfur cycle, involves the interactions of mineral and organic nutrient forms, with the quantity of water‐soluble phosphate being controlled not only by biological mineralization processes but also by dissolution rates from soil mineral fractions. Unique aspects of the phosphorus cycle include the fact that phosphorus does not undergo any valence changes in the cycle and that there is no gaseous component to the cycle. Based on the similarities of the biological and chemical properties of the phosphorus cycle to those of other nutrient cycles, only the salient features of the phosphorus cycle will be described herein. For a more detailed analysis of the nuances of the phosphorus cycle, see Stevenson (1986).

In a depiction of the phosphorus cycle with biomass accentuated, the pivotal process is mineralization of organic matter into plant and microbially available phosphate (Figure 15.3). As with the mineralization processes central to the carbon, nitrogen, and sulfur cycles, phosphorus mineralization is catalyzed by the general soil heterotrophic microbial population (e.g. Molla et al. 1984). Thus, as in the previously described process, the mineralization rate is controlled by the physical and chemical factors delimiting overall microbial heterotrophic activity. Phosphate mineralization and its assimilation into biomass are generally optimal under those soil conditions suitable for aboveground plant community development. Mineralization occurs under both aerobic and anaerobic conditions, but maximal activity is detected in the presence of molecular oxygen.

A variety of relatively easily decomposed organic phosphorus compounds are found in biomass. The most abundant phosphatic substances are inositol phosphates, which make up 10–50% of the organic phosphate fraction. Other organic phosphates are phospholipids (1–5%) and nucleic acids (0.2–2.5%) (Stevenson 1986). Lesser concentrations of phosphopyridines (e.g. NAD and NADH) and nucleotides (e.g. ATP) are detectable in soil organic phosphorus fractions.

Diagram depicting the major soil phosphorus cycle reservoirs, with interconnecting boxes labeled “soluble mineral P,” “microbial biomass,” “plants,” “animals,” “organic residues,” “humified P,” etc.

Figure 15.3 Major soil phosphorus cycle reservoirs.

Native phosphate in soils is derived from rock phosphate, apatite, which has the empirical formula 3(Ca₃(PO₄)₂)^.CaX₂ where X equals Cl⁻, F⁻, OH⁻ or CO₃ ²⁻. Secondary minerals that are formed from phosphate in soil are wavellite (Al₃(PO₄)₂(OH)₃ ^.5H₂O), variscite (Al(PO₄)^.2H₂O), dufrenite (FePO₄ ^·Fe(OH)₃), strengite (Fe(PO₄)^.H₂O), and vivianite (Fe₃(PO₄)₂ ^.8H₂O). Along with the effect of phosphate mineralization rate, the quantities of plant‐available phosphate present are determined by the solubilization rate of phosphatic minerals. Organophosphate levels in soil range from 15% to 80% of the total phosphates (Stevenson 1986). Therefore, in soils with high soil organic matter pools, the phosphorus needs of the plant community can be readily met via turnover of living biomass and mineralization of available soil organic matter pools whereas in the contrasting situation, with high mineral phosphate and low organic matter levels, the rate of solubilization of the minerals would be anticipated to be a major controller of plant growth.

Soil microbes are not only mediators of phosphate biological mineralization and immobilization processes, but they can also be instrumental in solubilization of rock phosphates via the synthesis and release of organic acids and carbon dioxide into the soil system. The organic acids liberate phosphate through chelation of the associated metal ions in the phosphatic minerals. Carbon dioxide contributes to the acidity of soil water by formation of carbonic acid. As the pH of an acidic soil increases, organic acid production becomes increasingly important in determining dissolution rate of rock phosphates (e.g. Jayashree et al. 2011; Traina et al. 1986).

Were this the end of the story, understanding the ecosystem processes affecting phosphate transformations in the soil environment would be easy. Differential levels of water‐soluble and insoluble phosphate arise from the fact that a variety of ecosystem properties control soil phosphate distribution within and between soil phosphate pools. Examples of additional ecosystem properties affecting phosphate availability include soil water contents (drought or wet conditions) (Dijkstra et al. 2015), the nature of the plant community (including intercropping) and the relationship of “hot spots” of phosphatase activity in the vicinity of the rhizosphere (Bettencourt et al. 2012; Spohn et al. 2015; Tang et al. 2014). Increased solubilization of phosphate minerals has also been proposed to result from inoculation of the soil with organic acid‐synthesizing microbes (Giles et al. 2014).

15.7 Microbially Catalyzed Soil Metal Cycling

Soil‐resident reservoirs of biomass‐nutrient cations (as well as toxic cations) are easily grouped into three categories: contained in biomass, dissolved in interstitial water, and retained in soil minerals (i.e. water‐insoluble forms) (Figure 15.4). Biomass productivity may be controlled by existence of excessive quantities (toxic metals) and occurrence of minimal levels (required nutrients). The soil biological community is instrumental in catalyzing a variety of processes that directly or indirectly result in the transfer of cations between these three soil reservoirs.

Diagram depicting the primary metal reservoirs in soil, with boxes labeled “biomass associated” and “organic residues” (top), “water soluble” (middle), and “soil minerals” and “precipitates” (bottom).

Figure 15.4 Primary metal reservoirs in soil.

15.7.1 Interactions of Soil Metals with Living Systems

Benefits from maintenance of appropriate levels of cations in interstitial waters result from their essential roles in cellular biochemistry. Plant essential trace metals include iron, zinc, manganese, calcium, boron, molybdenum, and, most likely, nickel. Animals also require cobalt, chromium, and tin. In contrast, cations of major interest in relation to negative interactions with biomass production include cadmium, mercury, and lead.

In most ecosystems, the balance between aboveground productivity and trace metal availability is maintained by the movement of soil‐resident cations between the various pools (Figure 15.4). Limitations of these plant nutrients to overall biomass production are generally restricted to ecosystems where augmentation of biomass production above that essentially dictated by native soil mineral cycling rates is necessary (i.e. intensive agricultural production) or in cases where metals are removed from the system through harvesting of aboveground biomass. In these situations, aboveground biomass needs are met through external amendments and/or modification of soil pH to optimize availability of soil resources.

In contrast to the situation exemplified by intensive agricultural systems, metal limitations for microbial productivity in soil are generally a rarity. Soil microbes possess highly efficient mechanisms for transport of cations into the cell. Furthermore, as discussed below, many members of the soil microbial community can produce chelators that increase water‐soluble cation resources.

An increasing array of environmental problems are arising where management of the soil for stimulation of ecosystem sustenance and/or development results in alteration of metal availability to the microbial community. Historically, the most commonly cited toxicity problem involving soil cations is associated with aluminum concentrations in acidic soils. Aluminum is toxic to biological processes. As the soil pH declines, the solubility of aluminum increases. Thus, best management practices of acidic soils for crop production have frequently included addition of lime to the soil to raise the pH as well as amendment with organic matter to complex the water‐soluble ions (see Stevenson 1982 for a more complete analysis of aluminum solubility in soil). More recent considerations involving toxic metals in soil have been associated with environmental contamination (e.g. smelter and foundry operations) and from disposal of societal wastes (e.g. land disposal of metal‐contaminated biosolids and composts).

The implications of increased soil loadings of toxic cations on ecosystem function are best exemplified by the succession of aboveground and belowground communities associated with metal contamination of soils by smelter operations. The availability of the toxic cation may become the controlling factor in community development, as was observed by Nordgren et al. (1983) in soils near a brass mill in South Sweden. In these soils sampled along a copper and zinc gradient (with maximum levels of 20 000 μg copper and zinc g⁻¹ soil), the influence of soil organic matter content and moisture on fungal community composition was minimal compared to the effect of heavy metal pollution on these populations.

A more general impact on ecosystem communities has been noted for heavy metal‐contaminated soils near a zinc smelter at Palmerton, Pennsylvania. Soil heavy metal loadings (litter horizon within 1 km of the smelter) in the 1970s included about 26 000 μg zinc g⁻¹ soil, 10 000 μg iron g⁻¹ soil, 2300 μg cadmium g⁻¹ soil, and 340 μg copper g⁻¹ soil (Strojan 1978) with as much as 90% of the metals being retained within the top 15 cm of the soil profile (Baucher 1973). These heavy metal loadings declined to approximately one‐tenth these values at a control area about 40 km east of the smelter. In this 1978 study, consideration of the average amount of organic matter in the more highly polluted site and comparison of decomposition rates of leaf litter in litter bags led to the conclusion that organic matter decomposition rates were reduced by the heavy metal contamination. In a 1975 report (Jordan 1975), reduced germination of tree seeds due to high zinc concentrations was proposed as a mechanism preventing the establishment of invader tree species in regions where indigenous trees, shrubs, and herb populations were in decline from the metal contamination. Additionally, in this site, the prevalent metal contaminant concentrations exceed those necessary to result in major reductions of total bacterial, actinomycete, and fungal populations (Jordan and Lechevalier 1975). Major alterations of the soil microbial as well as aboveground populations resulted from the metal contamination as demonstrated by changes in the soil metabolic diversity (Kelly and Tate 1998) and phospholipid fatty acid profiles (Kelly et al. 2003).

Currently, the aboveground status of the site at Palmerton can be defined as an “establishing” grass community due to overlaying of the soil with a sludge‐fly ash‐limestone‐grass seed mixture or a essentially barren system in untreated sites (Figure 15.5). The once flourishing forest ecosystem has been reduced to a bare mountainside. Reclamation of the site requires redevelopment of a functional soil microbial community accompanied with establishment of aboveground plant community consisting of a grass community.

With both of these examples, extremely high metal loadings were detected in the environmental samples. It must be remembered that the toxicity of the metals is proportional to their availability to the microbial community. Thus, inhibitory concentrations of metals derived from pure culture study where all the toxicant is water soluble frequently do not correspond to field observations where total soil metal concentration can be distributed between a variety of chemical forms that vary in availability to the soil microbial community (e.g. Lighthart et al. 1983). In a related study (Sullivan et al. 2013), variations in soil chemical properties and soil bacterial and archaeal community composition were strongly related in a muck soil with elevated sulfur and metal content. Therefore, when evaluating metal loading toxicity, soil properties that alter metal distribution between water‐soluble and water‐insoluble species must be considered. For example, in a study of planktonic, sediment and epilithic bacterial community adjustment to heavy metal concentrations (Dean‐Ross and Mills 1989), changes in microbial populations did not reflect the known toxicity of the contaminating metals. It was concluded that the occurrence of high pH in the samples that reduced the solubility of the toxic metals apparently resulted in lack of correlation between metal concentrations and response of the bacterial communities to the toxicants.

Image described by caption.

Figure 15.5 Metal‐contaminated site in Palmerton, PA (USA), showing effect of organic matter, lime, and fly ash amendments on grass growth.

The meaningful impact on microbial community structure in soils containing high loadings of toxic metals is reasonably anticipated a priori, but long‐term effects on microbial community structure in soils receiving moderate to low levels of such metals have also been reported. For example, MacDonald et al. (2011) found that changes in bacterial and fungal community structure resulting from zinc‐ and copper‐containing sewage biosolides were detectable more than 11 years post amendment,

15.7.2 Microbial Response to Elevated Metal Loading

The most obvious effect of high metal loadings on the soil microbial community is toxicity. Death of sensitive microbes and plants may result directly from dissolved cations as well as indirectly from the release of cations following catabolism of the organic matter to which they are bound. As implied by these data, alteration of soil microbial community dynamics and functions can be anticipated to result from soil contamination with heavy metals. A primary impact of metal contamination is the inhibition of catabolism of organic substances by the toxic metals.

Microbial populations may be inhibited by release of organic matter‐associated cations through microbial catabolism of the organic matter. The microbial recovery of energy from the fixed carbon substances results in accumulation of toxic concentrations of the metal. This phenomenon is exemplified by studies of the decomposition of iron, aluminum, zinc, and copper salts complexed to polysaccharides (Martin et al. 1966), metal‐nitrilotriacetate (calcium, manganese, magnesium, copper, zinc, cadmium, iron, and sodium chelates were examined) complexes (Firestone and Tiedje 1975) as well as copper, mercury, zinc, and cadmium salts of 2,4‐dichloro‐phenoxyacetic acid methyl ester (Said and Lewis 1991). As the concentration of the metal liberated from the organic complex increased, microbial activity decreased. The degree of inhibition varied with toxicity of the metal to the decomposer populations.

Overall reduction of microbial respiration by increasing concentrations of a variety of heavy metals is exemplified by the study of Lighthart et al. (1983). Additionally, examples of the capacity of heavy metal contamination to reduce or inhibit soil biological processes include a variety of enzymatic activities (e.g. Cole 1977; Doelman and Haanstra 1979, 1989; Stott et al. 1985) as well as various soil nitrogen transformations (immobilization, mineralization and nitrification (Chang and Broadbent 1982), anaerobic nitrogen transformations (Blais et al. 1988), and nitrogen fixation (Wickliff et al. 1980).

A variety of organisms must exist in soil with augmented resistance to heavy metal pollutants. For example, soil bacteria have been isolated with augmented resistance to metal contaminants, including cadmium (Bopp et al. 1983), copper (Dressler et al. 1991; Yang et al. 1993), mercury (Kelley and Reanney 1984), nickel (Schmidt et al. 1991), and silver (Haefeli et al. 1984). Selective enrichment of metal‐tolerant bacterial populations in polluted soils has been demonstrated with classical microbiological procedures (e.g. Duxbury and Bicknell 1983) as well as DNA probe‐mediated methods (e.g. Diels and Mergeay 1990).

15.7.3 Microbial Modifications of Metal Mobility in Soils

Soil properties controlling the availability of both essential nutrient cations and toxic metals are primarily soil pH, redox potential and organic matter concentrations. Soil‐buffering capacity, chlorinity, inorganic cations and anions, clay mineralogy, and moisture (e.g. Babich and Stotzky 1983) also contribute to metal solubility. The microbial community is instrumental in modifying each of these soil factors in a manner that can enhance or reduce soil interstitial water metal loadings. Furthermore, microbial interactions affecting metal mobility in soil can be grouped as indirect or direct. Indirect interactions include pH alteration, siderophore production, and modification of the physical environment (e.g. changing the redox potential). Direct interactions include valance change, substitution, methylation, transalkylation, incorporation into cell substituents, and release from organic association through mineralization.

15.7.3.1 Acidification and Chelation

Soil microbes produce a variety of inorganic and organic acids that contribute to solubilization of metals. The inorganic acids could be considered to be by‐products of energy metabolism. Two of the more commonly encountered products are carbonic and sulfuric acid. Aerobic microbes yield carbon dioxide as their terminal oxidation product. This gaseous substance exists in equilibrium in soil between carbonic acid, bicarbonate, and carbonate. Carbonic acid is a weak acid capable of dissolving soil minerals.

A potentially more dramatic example of mineral acid production is the sulfuric acid produced by oxidation of sulfides and elemental sulfur by soil thiobacilli. This process has been exploited by recovery of mineral from ores (e.g. Carlson et al. 1992) and has been proposed as a means of removing mineral sulfides from municipal sewage biosolids (e.g. Jain and Tyagi 1992). An example of ecosystem destruction resulting from sulfur oxidation by soil microbes and alteration of metal solubility production is the orange‐colored leachate associated with acid mine drainage (e.g. Mills 1985).

Organic acids that may solubilize soil minerals either directly through acidification of the microsite or through chelation reactions are produced by aerobic and anaerobic soil microbial processes. Under anaerobic conditions, a variety of organic acids accumulate due to incomplete oxidation of fixed carbon substrates. Commonly produced substances under aerobic conditions are siderophores. These chelators are important in iron nutrition of bacteria and higher plants. Bacterially synthesized siderophores may be instrumental in solubilizing significant quantities of iron for higher plant biomass synthesis. Such substances produced by rhizosphere bacteria have been shown to stimulate associated plant productivity (e.g. Kloepper et al. 1980). Similarly, rhizobial siderophores have been shown to stimulate clover productivity (Derylo and Skorupska 1992). A large number of members of the general soil community also produce these metal chelators, for example Pseudomonas species (e.g. Ankenbauer et al. 1988; Bar‐Ness et al. 1991; Cody and Gross 1987; Hoefte et al. 1991). The fact that these substances are synthesized within the soil ecosystem is supported by the capacity to isolate a variety of such substances from soil samples (e.g. Akers 1983a, b; Powell et al. 1982).

15.7.3.2 Direct and Indirect Oxidation and Chemical Reduction of Soil Minerals

Oxidation of metals by soil bacteria is exemplified by the oxidation of ferrous ion by T. ferrooxidans and mercury oxidation by a variety of bacteria (e.g. Holm and Cox 1975) as well as by the oxidation of metal oxides (e.g. conversion of arsenite to arsenate by Alcaligenes faecalis [Phillips and Taylor 1976]). Similarly, a number of bacterial species are capable of chemically reducing a number of metallic cations and metallic oxides, including iron (Lovley and Phillips 1986), manganese and iron (Nealson and Myers 1992), uranium (Lovley et al. 1991; Lovley and Phillips 1992), and chromate (Eary and Rai 1991).

Indirect modification of metal mobility by changing its valence state can result from microbial modification of the redox potential of the soil microsite. This process is envisioned in soil ecosystems where microbial energy metabolism results in exhaustion of molecular oxygen, nitrate and other electron acceptors, creating the highly reducing condition necessary for spontaneous chemical reduction of metals (such as ferric ion reduction to ferrous iron). The result of these reactions frequently results in changes in soil physical structure. For example, iron reduction in waterlogged soils can result in gley formation (see Tate 1987). In this situation, the precipitation of the ferrous ion by the concurrently formed sulfide causes a gray‐to‐black coloration of the soil. This alteration of the soil color within the profile can typically be used to determine transient occurrence of a high water table and the resulting imposition of anoxic conditions.

An interesting indirect modification of oxidation state of metal compounds may occur in the rhizosphere of plants that actively oxygenate their root zone. These plants, such as rice, transport molecular oxygen from the aboveground tissue into the root zone. Oxidation of technetium in rice ( Oryza sativa L.) was shown to result from this mechanism (Sheppard and Evenden 1991).

15.7.3.3 Bioaccumulation and Solubilization of Metal Ions through Mineralization

As is true of all living cells, soil microbial biomass incorporates essential metals into cell substituents. Also, accumulation of metallic cations on or within living cells can greatly exceed these metabolic requirements. This bioaccumulation is delineated by the achievement of cellular loadings of metals far beyond any levels that might be associated with cellular metabolism. This process is exemplified among the bacterial populations by strontium accumulation by Micrococcus luteus (Faison et al. 1990) and precipitation of cadmium by Clostridium thermoaceticum. In the latter example, cadmium is accumulated at the cell surface as well as in the surrounding medium (Cunningham and Lundie 1993). Similarly, cesium (Tomioka et al. 1992) and zinc (Sakurai et al. 1990) can be accumulated by bacteria. It must be noted that the cells need not be living for augmentation of cell‐associated metals to occur. Kurek et al. (1982) found that dead microorganisms were capable of absorbing more cadmium than did the live cells used in their studies.

The return of biomass‐associated metals to soil water‐soluble or mineral forms occurs through mineralization of the microbial biomass following death of the cell. As opposed to the cycling of carbon, nitrogen, and sulfur discussed above, the return of the metal to a mineral state likely provides little benefit to the decomposer population and is generally not directly enzymatically mediated. This metal cycling could be classified as a gratuitous product of microbial cycling of carbon.

15.7.3.4 Other Soil Metal Transformations

As was indicated above, soil minerals may also be methylated and transalkylated by soil microbial populations. The attachment of alkyl groups to the metal may alter the metal mobility within the ecosystem through enhanced volatility and water solubility. For example, monomethyl and dimethylmercury are volatile. Therefore, their formation in soils can result in reduction in the level of mercury contamination a the specific soil site (see Tate 1987 for further discussion of the behavior of metals in soil ecosystems).

15.7.4 Managing Soils Contaminated with Toxic Metals

Metal‐contaminated soils present a unique challenge to the soil microbiologist charged with the development of soil renovation plans. Conceptually, most microbial reclamation procedures involve management of the availability of the contaminant to the soil community of degraders and encouragement of mineralization of the offending substance. In contrast, reduction of elevated metal loadings involves either removal of the contaminant from the site (leaching or excavation of the soil itself) or modification of the properties of the polluted soil in a manner that encourages sequestering of the toxic metal into biologically unavailable forms. Since the three major soil properties controlling metal mobility are pH, redox potential, and organic matter content, traditional soil reclamation involving metal management has included manipulation of these parameters.

Some shifting of soil metals from water‐soluble to particulate fractions can be achieved through soil amendment with lime and/or organic matter. Increase in soil pH to approximate neutrality through lime amendment encourages a change in equilibrium between soluble and insoluble metal forms. With the addition of organic matter, the soluble metal loadings are reduced through complexing of the metal with the organic material itself. (See Tate 1987 for a more detailed discussion of control of soil metals through these mechanisms.) These nonbiological management procedures only provide a temporary solution to the metal toxicity. For example, once the organic matter retaining the metal is decomposed, any associated cations are released into the interstitial soil water. Similarly, from the view of management of soil pH, the natural tendency of soil systems is to become increasingly acidic. Thus, continued monitoring and adjustment of soil acidity may be necessary in particularly sensitive ecosystems if on‐site retention of the metal contaminant is essential.

Neither of these procedures, pH adjustment and organic matter amendment, presents any difficulty to the development or continued function of the soil microbial community. Organic matter encourages the stability of soil decomposer communities through the provision of carbon and energy. Furthermore, site management may require raising the soil pH to approximating neutrality (i.e. within the range of optimal function of the majority of the members of the soil biological community), and negative effects on biological parameters are not anticipated.

A more intensive stewardship of metal‐contaminated sites may actually involve alteration of the biological properties of the system. Activity of both the aerobic and the anaerobic microbial populations can be manipulated to encourage sequestering of metal contaminants. A logical extension of an understanding of the dynamics of sulfidogenesis and chemical interactions of the sulfide produced with soil metals is management of the cycle to purify soil interstitial waters. Provision of reducing conditions in the presence of sulfate ions may allow production of sufficient sulfide to reduce soluble metal loadings to acceptable levels. Metal management through control of the sulfur cycle is commonly accomplished by creation of artificial wetland. The difficulty with this procedure is that the wetland created must be maintained in an anoxic condition. If oxygen is introduced into the anoxic wetland, sulfur oxidizers utilize the sulfide as an energy source, thereby liberating the associated metal ion.

15.8 Conclusion

The foregoing analysis of the sulfur, phosphorus, and mineral cycles in soils underscores not only their essentiality to total ecosystem biomass productivity but also a primary role in reversal of adverse situations resulting from anthropogenic activities. That is, excessively acidic soils may result from elevated influxes of reduced sulfur compounds, heavy metal contamination may preclude sustenance of soil microbial communities, and mobility of soil phosphate may accelerate eutrophication of regional surface waters. Ameliorating these unfavorable soil or water conditions, while optimizing biomass productivity in soil ecosystems containing restrictive concentrations of the essential components of these cycles, requires a firm appreciation for not only the biological aspects of the biogeochemical cycles but also the chemistry of formation and dissolution of soil minerals. These biogeochemical cycles are highlighted by the major role of soil primary and secondary minerals in maintaining concentrations of bioavailable water‐soluble nutrients. This intersection of the mineral with the biological world and the implications of societal mismanagement of soil ecosystems involving soil phosphate, sulfur, and metal loadings can be managed to encourage development of sustainable ecosystems.

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