SCOPE 48 - Sulphur Cycling on the Continents

5.1 INTRODUCTION

Wetland ecosystems are distributed world-wide from boreal regions to the tropics. Total wetland area has been estimated as 5.8 x 10⁸ ha, although this estimate should be regarded as quite rough (Table 5.1). Saline and/or brackish wetlands comprise approximately 9% of the total global wetland area, while freshwater wetlands occupy the remaining area.

Primary production is quite variable among wetland types. Salt marshes and mangroves are considered to be among the most productive ecosystems of the world, with estimates of above ground net primary production ranging from 125 to greater than 1500 g (C) m^-2 a^-1 and total net primary production (above ground + below ground) reaching more than 4000 g (C) m^-2 a^-1 (Table 5.1; reviews by Odum, McIvor and Smith, 1982; Schubauer and Hopkinson, 1984). Production in freshwater wetlands also spans a wide range (see review by Brinson, Lugo and Brown, 1981). Among inland Sphagnum- dominated wetlands, above ground net primary production ranges from less than 40 to 440 g (C) m^-2 a^-1 (Wieder et al., 1989). While net primary production in tidal freshwater wetlands may exceed that of saline wetlands (Whigham et al., 1978), in general net primary production in freshwater wetlands is less than saline wetlands.

Wetland ecosystems, whether salt marshes, brackish water marshes, freshwater bogs, fens, marshes or swamps, are all characterized as having soils that are frequently or continuously waterlogged and that are often high in organic matter content (Cowardin et al., 1979). The duration and frequency of waterlogging, however, are quite varied. For example, many salt marshes and some freshwater marshes are regularly flooded with each tidal cycle. Bottom-land hardwood swamps may be dry during most of the year, but inundated during periods of high rainfall and/or runoff as often occurs during the spring months. Freshwater bogs may be waterlogged year-round, although water table fluctuations influence the depth of the waterlogged soil. Under saturated conditions, wetland soils are typically oxic at the surface and anoxic at depth (Teal and Kanwisher, 1961; Ingram, 1978; Howest et al., 1981), yet the distinctness of the transition between oxic and anoxic zones is quite variable depending on such factors as soil drainage, horizontal rates of water flow, and the presence or absence of vascular plants that may deliver oxidants to otherwise anoxic soil layers (Sparling, 1966; Armstrong and Boatman, 1967; Armstrong, 1979; Clymo and Hayward, 1982; King, 1983; Dacey and Howes, 1984).

Table 5.1. Characteristics of the world's wetlands 

There has been intense interest in understanding the sulphur cycle in wetland ecosystems because high inputs of organic matter into wetland soils, along with oxic surface and anoxic subsurface zones, potentially allow sulphur to play a critical role in the biogeochemistry of wetlands. Interest in understanding the sulphur cycle in wetland ecosystems has focused on a variety of issues:

1. Microbial dissimilatory sulphate reduction may account for much of the total organic matter decomposition (i.e. terminal carbon mineralization) in some wetlands.

2. Hydrogen sulphide may reach concentrations in soils that are inhibitory or toxic to wetland plants, thereby affecting net primary production.

3. Reduced sulphur may be an important energy source for chemoautotrophic bacteria, leading to new carbon fixation and thus affecting net ecosystem production and carbon balance.

4. Wetlands may be a significant source of sulphur gases to the atmosphere, some of which may be oxidized and redeposited to the earth as acid precipitation.

5. Sulphate reduction in wetlands may control methanogenesis and therefore methane emissions from wetlands, of potential importance because methane is a greenhouse gas.

6. The oxidation of reduced sulphur compounds in wetland soils produces acidity and therefore may affect the use of such drained soils for agriculture.

7. Studying sulphur storage in contemporary wetland soils may provide insights into the origins of sulphur in coal.

This chapter reviews what is known about the sulphur cycle in saline and freshwater wetlands.

Conceptually, the major features of the sulphur cycle are similar in saline and freshwater wetlands. However, dissolved sulphate concentrations in wetlands span more than three orders of magnitude, from greater than 45 mM (1.44 g (S) l^-1 in some salt marsh soils (Luther and Church, 1988), to less than 15 µM (0.48 mg (S) l^-1) in freshwater bogs remote from sources of atmospheric sulphur deposition (Gorham and Detenbeck, 1986). For this, and other less well understood reasons, the quantitative importance of specific parts of the sulphur cycle differ considerably between saline and freshwater wetlands. Most of the information presented in this chapter is derived from studies of salt marshes and freshwater Sphagnum-dominated wetlands, since the largest data base exists for these types of systems. Whenever possible, generalizations to other wetland types are made.

5.2 FORMS OF SULPHUR IN WETLAND SOILS

Sulphur exists in wetland soils in a variety of oxidation states and may be present in gaseous, soluble, and/or solid forms. A complete understanding of sulphur cycling in wetland soils has been hampered by the analytical difficulty in the unambiguous determination of specific sulphur forms. Although considerable progress has been made regarding methodology, assessment of published data on sulphur forms is made with caution, in light of the various approaches to sulphur fractionation used by different authors.

The total sulphur concentration in wetland soils ranges from 0.2 to 16% of the dry weight (0.06 to 5.0 mmol g^-l dry weight) (Howarth and Teal, 1979; Casagrande, Gronli and Sutton, 1980; Postma, 1982; Altschuler et al., 1983; Lowe and Bustin, 1985; Giblin, 1986; Wieder and Lang, 1986). In general, saline wetland soils contain considerably higher sulphur concentrations than freshwater wetlands. Sulphur concentrations in soils from low-salinity, brackish wetlands (Lowe and Bustin, 1986; Feijtel, 1986; Giblin, 1988) are as high as, or higher than, those reported for salt marshes with higher salinity porewaters, indicating that at salinities above a few parts per thousand, sulphate concentration is not a major factor in determining sulphur content. In freshwater wetlands, sulphur concentrations in peat may be related to sulphate inputs (Wieder and Lang, 1986; Ferguson, Robinson and Press, 1984; Brown, 1985).

The majority of sulphur in wetland soils is present as reduced inorganic sulphur minerals (pyrite, iron monosulphides, and elemental sulphur), organic forms, or sulphate. In general, sulphur gases (Wieder, 1985; Behr, 1985) and dissolved reduced sulphur compounds (Howarth et al., 1983) contribute little to the total sulphur in wetland soils. Soluble sulphate concentrations are sometimes low but often account for 10 to 20% or more of the total sulphur even at depth in waterlogged soils (Casagrande et al., 1977; Altschuler et al., 1983; Howarth, 1984; Behr, 1985; Neue and Mamaril, 1985; Spratt, Morgan and Good, 1987; Urban, Eisenreich and Grigal, 1989; Wieder and Lang, 1988). When wetland soils are drained for agriculture, the large store of reduced sulphur in the sediments may be oxidized, producing acid and sulphate. High concentrations of sulphate minerals (ferrous sulphate, jarosite, and aluminium sulphates) may be present in these soils (see review by Van Breemen, 1982).

In salt marshes and brackish swamps, a majority of the total sulphur is often present as reduced inorganic sulphur minerals (Howarth, 1984; Postma, 1982; Giblin, 1988), although organic sulphur has been found to be dominant in some brackish water marshes and swamps (Casagrande et al., 1977; Lowe and Bustin, 1985) and in the surface peat in some salt marshes (Giblin, 1988). In salt marshes the majority of reduced inorganic sulphur is present as pyrite (FeS₂), with elemental sulphur and FeS typically making up less than a few per cent of the reduced inorganic sulphur (Kaplan, Emery and Rittenberg, 1963; Howarth and Teal, 1979; Lord and Church, 1983; Howarth, 1984; Giblin, 1988). Pyrite is also the overwhelmingly dominant form of sulphur in many brackish swamps (Postma, 1982). In contrast to marine and brackish wetland soils, most of the sulphur in freshwater wetland soils is typically present as organic rather than inorganic sulphur, with carbon bonded sulphur (reduced organic sulphur) much more abundant than ester-sulphate sulphur (oxidized organic sulphur) (Casagrande et al., 1987; Altschuler et al., 1983; Behr, 1985; Lowe, 1986; Spratt, Morgan and Good, 1987; Urban, Eisenreich and Grigal, 1989; Wieder and Lang, 1988). Of the reduced inorganic sulphur in freshwater peats, pyrite is often the dominant form, although concentrations are much lower than in salt marshes (Behr, 1985; Spratt, Morgan and Good, 1987; Urban, Eisenreich and Grigal, 1989; Wieder and Lang, 1988).

5.3 DISSIMILATORY SULPHATE REDUCTION AND ITS MEASUREMENT

Of the many possible transformations of sulphur in wetland soils (see Luther and Church, this volume), the process of dissimilatory sulphate reduction has received the most attention. Dissimilatory sulphate reduction involves a number of steps in which fermentative microorganisms break down organic matter into small organic molecules (e.g. lactate, acetate, alcohols, etc.) that then are oxidized to CO₂ by sulphate-reducing bacteria using sulphate as the terminal electron acceptor. Overall, two moles of organic carbon are oxidized for every mole of sulphate reduced (Howarth and Teal, 1979; Jørgensen, 1982; Grinenko and Ivanov, 1983; Howarth and Stewart, this volume). Sulphate reduction has received attention in wetland soils because in anoxic waterlogged soils it is an important, and sometimes dominant, pathway of carbon mineralization. In addition, sulphate reduction generates alkalinity, which buffers porewater pH, and produces hydrogen sulphide, which may be deleterious to plants.

Experimental studies in subtidal marine sediments (Westrich and Berner, 1984) have shown that where high levels (> 5 mM) of sulphate typical of the marine environment occur, organic carbon availability most often limits sulphate reduction. Sulphate may be limiting in freshwater environments where sulphate concentrations are lower (Cook, 1981), or at depth in some highly organic marine and brackish sediments where sulphate is exhausted by sulphate-reducing bacteria.

Sulphate reduction rates in wetlands have been estimated using a variety of techniques, including mathematical modelling of sulphate and bicarbonate profiles (Lord and Church, 1983; Casey and Lasaga, 1987), measurement of sulphate disappearance in sealed cores (Jørgensen, 1977; Howarth and Teal, 1979; King and Wiebe, 1980), calculation of sulphur retention from whole watershed sulphur budgets (Hemond, 1980; Bayley, Behr and Kelly, 1986; Bayley et al.,1987; Ogden, 1982; Wieder, Yavitt and Lang, this volume), and direct measurement using ³⁵SO₄^2- (Howarth, 1979; Howarth and Teal, 1979; Howarth and Giblin, 1983; King, 1988; Hines, Knollmeyer and Tagel, 1989). Skyring (1987) and Ivanov et al. (1989) provide comprehensive discussions of the advantages, disadvantages, and assumptions associated with each of these approaches to measuring sulphate reduction.

The most commonly used method of measuring sulphate reduction is a radiotracer method in which tracer quantities of ³⁵SO₄^2-, are injected into sediments and the formation of reduced ³⁵S is measured. The basic method was first used in subtidal marine sediments by Ivanov (1956); Jørgensen (1978) described the approach in some detail and provided a theoretical framework for interpreting the method. The method as used by most early investigators working in subtidal marine systems assayed only for acid-volatile sulphides as products of sulphate reduction. These products include both the free sulphides in the porewater and sulphide which has precipitated with iron to form iron monosulphides.

In his original work, Ivanov (1956) pointed out that some of the ³⁵S reduced during sulphate reduction in marine sediments was incorporated into a non-acid-volatile pool, pyrite (FeS₂). Unfortunately, this early report received little attention in the Western literature. More than 20 years later, Howarth (1979) reported that pyrite was the major end-product of short-term ³⁵SO₄ ^2- reduction measurements in salt-marsh sediments in Massachusetts, and that measuring the incorporation of ³⁵S only into acid volatile sulphur substantially underestimated sulphate reduction rates because a considerable quantity of reduced ³⁵S was recovered from forms which were not acid volatile. Howarth and colleagues extracted these forms using aqua regia, finding 5 to 10 times more ³⁵S in pyrite and elemental sulphur (also a non-acid-volatile form) than in soluble sulphide and iron monosulphides (the acid-volatile forms) (Howarth, 1979; Howarth and Teal, 1979; Howarth and Giblin, 1983; Howarth and Marino, 1984; Howarth and Merkel, 1984).

The use of aqua regia to analyse for ³⁵S in pyrite and elemental sulphur during ³⁵SO₄^2- reduction measurements can overestimate rates of reduction if care is not taken to fully rinse away unreduced ³⁵SO₄^2-. Although this has caused some concern in salt-marsh studies (King, 1983; Howes, Dacey and Teal, 1983), other investigators find little or no unreduced ³⁵SO₄^2- remains after careful rinsing of marsh sediments (Howarth and Merkel, 1984) or lake sediments (Rudd, Kelly and Furutani, 1986; Cook and Kelly, this volume). 

Any ambiguity regarding the determination of reduced ³⁵S during the measurement of sulphate reduction rates has largely been resolved by the application of the chromium reduction technique, which uses an acid solution of Cr(II) to reduce pyrite and elemental sulphur to H₂S (Zhabina and Volkov, 1978). The H₂S that is produced can be distilled and trapped, assayed for radioactivity, and quantified. Because chromium does not reduce sulphate to sulphide (Zhabina and Volkov, 1978; Howarth and Jørgensen, 1984; Howarth and Merkel, 1974; Wieder, Lang and Granus, 1985; Canfield et al., 1986), the possible presence of any remaining unreduced sulphate in the soil sample is not of concern. Moreover, tests with yeast and specific organic compounds (amino acids, ester sulphates) have indicated that organic sulphur also is not liberated by chromium reduction (Howarth and Jørgensen, 1984; Wieder, Lang and Granus, 1985; Canfield et al., 1986).

Howarth and Merkel (1984) compared the use of the chromium reduction assay with the aqua regia extraction in measuring sulphate reduction rates in salt marshes in Massachusetts and Georgia. They found excellent agreement between the two methods and confirmed that pyrite was the major short-term product of sulphate reduction. Using the chromium reduction assay, other workers have confirmed that pyrite is often a major product of sulphate reduction in marshes (Howes, Dacey and King, 1984; King et al., 1985; Hines, Knollmeyer and Tugel, 1989), although the extent of pyrite formation can vary in space and time even within a marsh (Hines, Knollmeyer and Tugel, 1989). The formation of ³⁵S-labelled pyrite during sulphate reduction measurements has also been noted in subtidal marine sediments (Ivanov, Lein and Kashparova, 1976; Ivanov et al., 1980; Lein et al., 1982; Howarth and Jørgensen, 1984). The incorporation of the radiolabel into pyrite is probably a result of rapid formation of pyrite, as originally proposed by Howarth (1979), since isotopic exchange between soluble sulphides and pyrite is not observed within the time scale of a day (Fossing and Jørgensen, 1990). See Luther and Church (this volume) for further discussion of the mechanisms involved.

Chromium reduction is now the most commonly used method of determining the reduced inorganic ³⁵S end-product of sulphate reduction in studies in salt marshes, but rates of sulphate reduction still can be underestimated if a substantial quantity of ³⁵S becomes incorporated into organic, carbon-bonded sulphur during incubation. Carbon-bonded sulphur can be formed either biotically through sulphate uptake and assimilatory sulphate reduction, or abiotically as H₂S reacts with organic matter via nucleophilic attack (Casagrande, Gronli and Sutton, 1980; Altschuler et al., 1983; Brown, 1986; Francois, 1987; Luther and Church, this volume). There is no evidence for significant rates of formation of organic sulphur during short-term incubations with saline or brackish wetland soils with ³⁵SO₄^2- (Howarth and Merkel, 1984). However, in freshwater wetland soils, failure to account for the incorporation of ³⁵S into carbon-bonded S can lead to substantial underestimates of sulphate reduction (Wieder and Lang, 1988; Spratt and Morgan, submitted). The rapid incorporation of radiolabel into carbon-bonded sulphur has also been observed in lake sediments (Rudd, Kelly and Furutani, 1986).

5.3.1 SULPHATE REDUCTION IN SALTWATER WETLANDS

Much of the primary production in salt marshes takes place below ground (Valiela, Teal and Persson, 1976; Schubauer and Hopkinson, 1984; Good and Frasco, 1982). In studies where total below ground carbon budgets are available, sulphate reduction and associated fermentation reactions are the major anaerobic decomposition processes in soils; moreover, anaerobic decomposition equals or exceeds oxic respiration (Howarth and Teal, 1979; Howes, Dacey and King, 1984; King, 1988). In spite of the importance of sulphate reduction to carbon cycling in saltwater wetlands, there have been relatively few studies over annual cycles, and these largely have been confined to temperate salt marshes dominated by Spartina alterniflora (Table 5.2). To our knowledge, no seasonal studies have been carried out in tropical mangrove wetlands.

Table 5.2. Published annual sulphate reduction rates for saline marshes. The method used to make the measurement is also shown 

Sulphate reduction rates in salt marshes are generally higher than rates in subtidal marine sediments (Howarth, 1984; Skyring, 1987). Variation in sulphate reduction rates over time, and between sites can be attributed to at least three factors: temperature, substrate availability, and partitioning between oxic and anoxic respirations. Howarth and Teal (1979) found a strong temperature response, but they also found a higher rate of sulphate reduction in the fall than in the spring at similar temperatures in a marsh in Massachusetts. They speculated that the higher autumn rates might be due to an increase in the amount of carbon available below ground when plants senesce. A similar increase during autumn was noted in soil CO₂ production in the same marsh (Howes, Dacey and Teal, 1983). King (1988) found that within a South Carolina salt marsh, temperature accounted for most of the variation in sulphate reduction rates over the season. However, Morris and Whitting (1986) observed a large difference between spring and fall rates of CO₂ production in South Carolina even though the temperatures were similar. Hines, Knollmeyer and Tugel (1989) found the seasonal pattern of sulphate reduction followed changes in temperature in a high marsh Spartina patens zone but not in a tall Spartina alterniflora zone; rates of sulphate reduction in the tall S. alterniflora zone were highest when plant growth was most rapid. Hines, Knollmeyer and Tugel (1989) suggested that the release of dissolved organic matter from plant roots during active growth may fuel sulphate reduction.

Patterns of primary production and biomass allocation seem to be responsible for many of the differences in the annual sulphate reduction rate reported for salt marshes. When a comparison is made among short S. alterniflora sites (Table 5.2), there does appear to be a relationship between sulphate reduction rate and below ground production. For example, higher sulphate reduction rates have been reported for soils from a Massachusetts marsh (Great Sippewissett) than more southern marshes where the growing season is longer and temperatures are higher, but where below ground production is lower (Table 5.2). Differences in sulphate reduction rates within sites in marshes both in Georgia and in South Carolina appear to be related to differences in below ground biomass and production (Howarth and Giblin, 1983; King, 1988). Sulphate reduction rates reported from salt marsh pans and creeks are generally much lower than those reported for vegetated areas (Table 5.2), suggesting that the roots, rhizomes, and perhaps exudates from the dominant macrophytes are fueling these high rates. Hartman (1984) compared sulphate reduction rates in areas of the same marsh where there were different amounts of above ground biomass, including an area which had been bare for more than three years, and a pan. Over the three-month study, sulphate reduction rates in undisturbed vegetated areas were considerably higher than in pans or areas of bare peat.

The degree of oxidation of marsh soils also affects sulphate reduction rates. Skyring, Oshrain and Wiebe (1979) initially reported that sulphate reduction rates were higher in more oxidized creekbank zones than in areas of short S. alterniflora. However, they did not include non-acid-volatile end-products of sulphate reduction in their measurements. Both Howarth and Giblin (1983) and King (1988) have reported that sulphate reduction rates are lower in the more oxidized and more productive creekbank stands of S. alterniflora than in interior marsh sites. Greater porewater drainage and air entry may allow more total decomposition via aerobic pathways in creekbank areas when compared to interior marsh soils. OeDema (1988) found very low sulphate reduction rates in a highly oxidized marsh in the Netherlands. Hines, Knollmeyer and Tugel (1989) found that sulphate reduction rates in the high marsh S. patens zone, where desiccation occasionally caused the soils to become oxidized to 11 cm, were lower than in the nearby S. alterniflora zone where redox potentials were more negative.

5.3.2 SULPHATE REDUCTION IN FRESHWATER WETLANDS 

Sulphate reduction rates have been directly measured in only a few freshwater systems (Table 5.3). These rates are somewhat lower than reported for salt marshes, although the values overlap considerably. It should be emphasized that Cedar Swamp, Big Run Bog, and Buckle's Bog receive considerable inputs of SO₄^2- from sources other than directly incident precipitation. Wieder, Yavitt and Lang (1990) have argued that the advent of high SO₄^2- deposition via acid precipitation, coupled with minerotrophic drainage into Big Run Bog from the 276 ha of surrounding upland forest may have created conditions conducive to the dynamic sulphur cycling. A similar argument could be made for Cedar Swamp, with its substantial groundwater input. In the absence of further study, it presently is questionable whether or not the high rates of sulphate reduction reported in these studies are typical of all Sphagnum-dominated wetlands.

Very little is known about the controls on sulphate reduction rates in freshwater wetlands. Spratt and Morgan (1990) reported that sulphate reduction rates under mats of Sphagnum were up to five-fold lower than nearby areas where Sphagnum was absent in a New Jersey cedar swamp, possibly because photosynthesis from the moss oxidized the soil. The depth of the water table has been identified as another major factor controlling sulphate reduction rates (Spratt, Morgan and Good, 1987).

Traditionally, anaerobic carbon mineralization in Sphagnum wetlands was thought to be dominated by fermentation reactions, including methanogenesis. However, at Big Run Bog and Buckle's Bog, methane production accounted for only 3 to 12% of total anaerobic carbon mineralization of approximately 660 g (C) m^-2 a^-l (55 mol m^-2 a^-l) (Wieder, Yavitt and  Lang, 1990; Yavitt, Lang and Wieder, 1987). Carbon dioxide production, resulting substantially from sulphate reduction, dominated anaerobic carbon mineralization (Wieder, Yavitt and Lang, 1990). While oxic respiration and methanogenesis are probably more important than sulphate reduction in decomposition in most freshwater wetlands, sulphate reduction rates may be very important in wetlands receiving sulphate from sources such as acid deposition or mine drainage.

Table 5.3. Sulphate reduction rates, sulphur inputs and sulphur outputs for selected freshwater wetlands

Some studies have reported sulphate reduction rates which are considerably higher than would be indicated by the net retention of sulphur (Table 5.3). Therefore, the importance of sulphur reduction to carbon cycling may not be apparent from sulphur input-output budgets. The dynamic cycling of sulphur within the peat, discussed below, allows for more carbon to be mineralized via sulphate reduction than would be calculated from net sulphur retention. Such recycling is also important in salt marshes (Howarth and Teal, 1979; Hines, et al., 1988; Gardner, 1990; Luther and Church, this volume).

5.4 THE CYCLING OF SULPHUR WITHIN THE SOIL

5.4.1 SALINE WETLANDS

Rates of sulphur burial (permanent accretion in the sediment, or precipitation in excess of re-oxidation) in saltwater wetlands are usually much less than 1 mol m^-2 a^-l (32 g (S) m^-2 a^-l) and are small compared with sulphate reduction rates (Howarth, 1984; King, 1988). Therefore, nearly all of the sulphide produced by sulphate reduction must be lost from the soils by some other mechanism. Although some sulphide is lost by volatilization and advection (below), the majority is oxidized in situ.

Sulphide may be directly oxidized in the porewater or converted to sulphide minerals which are later oxidized. Howarth and Teal (1979) and Giblin and Howarth (1984) proposed that pyrite formation and oxidation occurred on a seasonal basis. They suggested that a significant portion of the sedimentary pyrite was oxidized in the spring when sulphate reduction rates were low, but grass growth was rapid. It now appears that the importance of sulphur cycling through solid phases of sulphur may vary considerably from marsh to marsh, and that weather and tides may help determine sediment oxidation rates (Giblin, 1988). King (1988) found no seasonal pattern in concentrations of reduced inorganic sulphur, although concentrations did fluctuate over the course of the year. In a Louisiana marsh, Feitjel (1986) found that maximal pyrite oxidation occurred during periods of low water typical of the winter and early summer months. Luther and Church (1988) observed a very strong seasonal cycle of pyrite oxidation in a Delaware marsh. Hines, Knollmeyer and Tugel (1989) found that tidal and rainfall events produced large changes in soil redox in Spartina patens soils which resulted in rapid changes in the precipitation and dissolution of iron and in the magnitude and spatial distributions of sulphate reduction.

The mechanism by which pyrite is oxidized is not fully understood. Drainage and evapotranspiration allow air entry into the sediments (Dacey and Howes, 1984), but it is unlikely that oxidation proceeds by purely chemical mechanisms. In coal surface mines, more than 75% of the pyrite oxidation is carried out by chemoautotrophic bacteria of the genus Thiobacillus (Taylor , Wheeler and Nordstrom, 1984). The very rapid oxidation observed by Luther and Church (1988) suggests that microbial processes are driving the reaction. Luther et al., (1986) have also observed large concentrations of thiols in porewater during pyrite oxidation events, which they suggested are products of pyrite oxidation. Thiols would not be products of the chemical oxidation of sulphur (Luther et al., 1986; Luther and Church, this volume). Previously, the cycling of organic sulphur in saltwater wetlands had largely been ignored because the cycling of organic sulphur through the vegetation is much smaller than the amount of sulphur used for respiration. The production of thiols during pyrite oxidation suggests that there may be a link between organic and inorganic sulphur pools (Luther et al., 1986) which should be explored further .

5.4.1.1 The Significance of the Sulphur Cycle to Energy Flow in Saline Wetlands

In oxic systems the energy trapped during the formation of organic matter is liberated during decay so the flow of energy follows the flow of carbon. However, when decomposition occurs anoxically, part of the energy present in the organic carbon is not released, but remains conserved in high energy reduced compounds such as hydrogen sulphide (see discussion in Howarth and Teal, 1980; Howarth, 1984). During dissimilatory sulphate reduction, only 25% of the energy present in the organic matter is available for microbial growth and respiration, while nearly 75% of the energy remains trapped as sulphides. Hence carbon and energy flow are decoupled.

The energy retained in reduced compounds is released when they are oxidized. If this oxidation is carried out chemically, the energy originally present as photosynthetically fixed carbon is lost from the ecosystem. However, a variety of chemoautotrophic bacteria are capable of using the energy released in the oxidation of sulphide and other reduced sulphur compounds to support new carbon fixation. In saline wetlands, where much of the decomposition is carried out by sulphate-reducing bacteria, the amount of energy carried in reduced sulphur compounds is very large (Howarth, 1984). If we assume that 75% of the total decomposition proceeds by sulphate reduction, then more than half of the energy from photosynthesis is trapped, at least temporarily, as reduced sulphur.

A portion of the reduced sulphur produced each year is exported to creeks before being oxidized (Figure 5.1). Porewater fluxes to creeks may exceed 101 m^-2 day^-1 (Howarth et al., 1983), although these high rates may be restricted to areas within a few metres of creeks (Hemond and Fifield, 1982). Howarth et al., (1983) calculated that at most, 6 mol m^-2 a^-1 (192 g (S) m^-2 a^-1) of reduced sulphur are exported to creeks in Great Sippewissett Marsh in such an advection of porewaters. It is not known how much of this energy is actually captured as opposed to being lost by chemical oxidation. At least a portion of the sulphide reaching creeks is used biologically, as evidenced by the presence of large mats of Beggiatoa, a chemoautotrophic bacterium, wherever porewater seeps occur. The majority of oxidation of reduced sulphur occurs within the soil, or at the soil surface. As discussed above, a large portion of this oxidation may be biologically catalysed. The fate of the large amount of energy released in this process is unknown but could potentially support CO₂ fixation by chemoautotrophic bacteria in the soil at rates as high as 275 to 500 g (C) m^-2 a^-1 (23 to 42 mol m^-2 a^-1) (Howarth, 1984). However, a number of heterotrophic bacteria are also able to oxidize reduced sulphur compounds under oxic conditions, probably with little or no energy gain from the oxidation (Schoenau and Germida, this volume).

Figure 5.1. A model of energy flow in a salt marsh ecosystem. Adapted from Howarth and Teal (1980)

5.4.2 SULPHUR CYCLING IN FRESHWATER WETLANDS

In contrast to the extensive studies of sulphur cycling in saline and brackish wetlands, sulphur cycling in freshwater wetland systems has only recently generated much interest. One reason for the minimal effort in freshwater wetlands was the traditional belief that in freshwater wetlands dissimilatory sulphate reduction should be negligible, the process being limited by low sulphate concentrations (see review by Nedwell, 1984). However, the modern phenomenon of elevated atmospheric sulphur deposition in acid precipitation focused new efforts towards understanding sulphur cycling in freshwater wetland ecosystems (Gorham, Bayley and Schindler, 1984).

Most of the available information regarding sulphur cycling in freshwater wetlands comes from studies of Sphagnum-dominated wetland systems. In contrast to the high sulphate inputs to salt marshes, inputs of sulphur to freshwater wetlands via precipitation are quite low (Table 5.3). Sphagnum-dominated wetlands have consistently been identified as net sinks for atmospheric sulphur, rather than net sources. For truly ombrotrophic wetlands, receiving all of their water and nutrients from precipitation alone, the quantity of sulphur retained annually is less than input of sulphur from the atmosphere. However, in Sphagnum wetlands, with additional inputs from either groundwater or runoff from surrounding upland areas (i.e. minerotrophic sites), annual sulphur retention can exceed inputs from precipitation
(Table 5.3).

Most of the sulphur in freshwater wetland soils is present as organic sulphur, but sulphur retention is not simply the result of slow incorporation of sulphur inputs directly into organic sulphur forms. Rather, both inorganic and organic sulphur fractions are involved in what appears to be a surprisingly rapid and dynamic cycling of sulphur, at least in Sphagnum wetlands. Radiotracer experiments have shown that in short-term incubations ³⁵SO₄^2- becomes incorporated into reduced inorganic sulphur, carbon-bonded sulphur, and ester-sulphate sulphur, with the majority of the reduced label being recovered from inorganic, rather than organic forms (Wieder and Lang, 1988; Spratt and Morgan, 1990). Apparently, the reduced inorganic sulphur pool turns over rapidly; rates of reduced sulphur oxidation are similar to rates of sulphur reduction (Wieder and Lang, 1988). In longer term studies following the fate of field-applied ³⁵SO₄^2-, the majority of the reduced ³⁵S is recovered as carbon-bonded sulphur rather than reduced inorganic S (Behr, 1985; Brown and MacQueen, 1985; Brown, 1985). It appears that newly deposited SO₄^2- is initially converted to reduced inorganic sulphur, which turns over rapidly, on the time scale of days. A portion of this sulphide is incorporated into organic forms which turn over more slowly.

Although the number of sites examined thus far are few, in freshwater Sphagnum wetlands the process of sulphate reduction seems to be far more important than was previously imagined. Surprisingly, annual rates of sulphate reduction reported for freshwater Sphagnum wetlands overlap considerably with rates reported for saline wetlands, even though concentrations of dissolved sulphate in the Sphagnum wetland soils are typically less than 300 µM (9.6 mg (S) l^-1 (Spratt, Morgan and Good, 1987; Wieder and Lang, 1988). Annual sulphate reduction far exceeds estimates of sulphur retention in Cedar Swamp and Big Run Bog (Table 5.3), attesting to the importance of reduced sulphur oxidation and rapid turnover of the reduced inorganic sulphur pool. This observation brings into question whether or not sulphate reduction in these Sphagnum wetlands is indeed limited by low sulphate concentrations. Radiotracer determination of sulphate reduction in short-term incubations of Cedar Swamp peat (Spratt, pers. comm.) and Big Run Bog peat (Wieder, Yavitt and Larig, 1990), in which background SO₄^2- concentrations were augmented up to 100-fold, did not result in a stimulation of sulphate reduction. In over 600 determinations of sulphate reduction in Big Run Bog and Buckle's Bog peat, rates of sulphate reduction were not positively correlated with the ambient sulphate pool size (Wieder, Yavitt and Lang, 1990).

Clearly, for sulphate reduction to continue, there must be an adequate supply of available SO₄^2-. However, in Sphagnum wetlands like Cedar Swamp, Big Run Bog, and Buckle's Bog, the size of the dissolved SO₄^2- pool apparently is not a good indicator of SO₄^2- availability. Rather, SO₄^2- availability is the result of dynamic cycling; the SO₄^2- pool turns over rapidly, being depleted by sulphate reduction and also being replenished by the oxidation of reduced inorganic sulphur.

The dynamic cycling of sulphur in these Sphagnum-dominated wetlands has implications for carbon balance and peat accumulation. Although most of the carbon mineralization in Sphagnum-dominated wetlands occurs aerobically, rates of peat accumulation are more affected by inputs of carbon into the anaerobic zone and by rates of anaerobic carbon mineralization (Clymo, 1984). At Buckle's Bog and Big Run Bog, net primary production is approximately balanced by carbon mineralized via oxic respiration and methanogenesis. However, with the addition of sulphate reduction as an anaerobic carbon mineralization pathway, these wetlands are presently in a state of negative carbon balance (Wieder, Yavitt and Lang, 1990).

The sulphur requirement for plant growth in freshwater wetlands has been estimated to be of the order of 0.1 to 1.0 mol m^-2 a^-l (3 to 32 g m^-2 a^-l) (calculated from Urban et al., 1988 and production data in Table 5.1), which exceeds sulphur inputs in many locations (Table 5.3; Urban, Eisenreich and Grigal, 1989). Therefore, the recycling of sulphur within the bog is important for maintaining new primary production in addition to the role that sulphate plays in carbon decomposition. Sulphur appears to be mineralized more rapidly than C or N in aerobic portions of wetland soils (Urban, Eisenreich and Grigal, 1989).

5.5 FEEDBACKS BETWEEN THE SULPHUR CYCLE AND PRIMARY PRODUCTION

Sulphate reduction benefits plants by regenerating nutrients and buffering soil pH. However, the sulphides produced are potentially toxic and may inhibit nutrient uptake (Vamos and Koves, 1972; Goodman and Williams, 1961; Mitsui, 1965); It has been proposed that the concentration of sulphide in the sediment strongly influences both primary production (Howarth and Teal, 1979; Mendelssohn and Seneca, 1980; Howes et al., 1981; King et al., 1982) and plant species distribution (Nickerson and Thibodeau, 1985) in wetlands.

Wetland plants possess structural and metabolic adaptations to deal with the presence of sulphides (Crawford, 1978; Teal and Kanwisher, 1961; Thibodeau and Nickerson, 1986; Burdick and Mendelssohn, 1987). Some species of wetland plants are capable of lowering sulphide concentrations in sediments by allowing air to diffuse out of roots (Teal and Kanwisher, 1961) or by allowing sulphides to diffuse into roots where they can be oxidized or incorporated into organic compounds (Carlton and Forrest, 1982; Fry et al., 1982). In spite of these adaptations roots and rhizomes may experience anoxia (Mendelssohn, McKee and Patrick, 1981). It has frequently been observed that wetland plants are more productive in areas of higher water movement (Sparling, 1966; references in Chalmers, 1982). A number of explanations for this phenomenon have been advanced including greater nutrient availability where water flow is rapid, removal of sulphides or other toxic metabolic end-products by flushing, and in saline wetlands, lower salinity stress. Research in salt marshes now suggests that primary production in these systems is regulated by a complex interaction between water flow and the sulphur cycle due to the dual role of sulphate reduction in generating both needed nutrients and metabolic poisons.

We have summarized the possible reactions between the sulphur cycle and marsh production in Figure 5.2. Some of these interactions are well known and the reader is referred to the summary by Chalmers (1982) as well as recent work by Wiegert, Chalmers and Randerson (1983), DeLaune, Smith and Tolley (1984), and Howes, Dacey and Goehringer (1986) for more details. The negative interactions on production postulated here under conditions of high drainage seem to be borne out by the long-term experiments of Chalmers, Wiegert and Wiebe (submitted), who observed that prolonged drainage ultimately resulted in declines in productivity in some sites and no changes in others.

Figure 5.2 A model of the interaction of the sulphur cycle and the primary production in a salt marsh ecosystem

Briefly, sulphate reduction generates sulphides which decrease plant nutrient uptake. Lowering dissolved sulphide concentration by oxidation or porewater exchange increases the ability of the plants to take up nutrients, and may reduce the excretion of root fermentation products, leading to lower rates of sulphate reduction and sulphide production, in a positive feedback loop. Water removal from evapotranspiration or porewater exchange also may lead to a negative feedback on production by increasing sediment salinity, or removing nutrients. Extensive oxidation will completely eliminate sulphide production and decomposition will proceed aerobically. This will further increase the plant's ability to take up nutrients by eliminating dissolved sulphides, but will promote nutrient immobilization by aerobic heterotrophs and phosphorus precipitation with iron oxides resulting in decreased nutrient concentrations.

5.6 SULPHUR GAS FLUXES TO THE ATMOSPHERE

A variety of volatile sulphur compounds are produced in wetlands including hydrogen sulphide (H₂S), methylmercaptan (MeSH), dimethyl sulphide (DMS), dimethyl disulphide (DMDS), carbonyl sulphide (COS), and carbon disulphide ( CS₂) .A portion of these biogenic sulphur gases are emitted to the atmosphere where they are eventually oxidized to SO₂ and SO₄^2-, contributing to the acidification of rainfall (Andreae and Jaeschke, this volume). Much of the research on sulphur gas emission in wetlands has focused on the relative importance of biogenic sulphur vs anthropogenically released sulphur as potential contributors to the acidification of rainfall. Measurements of sulphur gas fluxes have been complicated by analytical, methodological, and logistical difficulties (see Andreae and Jaeschke, this volume).

Most studies of sulphur gas fluxes from saline wetlands have been in marshes dominated by Spartina alterniflora, although several mangrove and Juncus sites have been examined. These studies have all shown that H₂S and DMS are the dominant gases being released to the atmosphere (Table 5.4). Some studies have demonstrated that emissions change on a diurnal basis (Cooper et al., 1987a,b; DeMello et al., 1987), or during the tidal cycle (Cooper et al., 1987a; Steudler and Peterson, 1985). Estimates of emissions of both individual gases and the total gaseous sulphur flux span more than two orders of magnitude (Table 5.4). As discussed in Steudler and Peterson (1985) and Lamb et al. (1987) the cause for this variability has not been established. Methodological problems certainly contribute (Andreae and Jaeschke, this volume), but high spatial and temporal variability in fluxes should be expected in these ecosystems which have such dynamic cycles of sulphur.

Table 5.4. Sulphur gas fluxes from wetlands

The very high DMS fluxes which have been reported in areas colonized by S. alterniflora may not be representative of other saline wetlands. Steudler and Peterson (1984) found that DMS fluxes from unvegetated tidal creeks were only about 5% of those measured from sites vegetated by S. alterniflora, and DMS emission has been related more closely to S. alterniflora biomass than soil parameters (DeMello et al., 1987). Dacey, King and Wakeham (1987) report finding high concentrations of dimethylsulphoniopropionate (DMSP) in the leaves of S. alterniflora; they speculated that DMSP may function in regulating osmotic pressure. DMSP concentrations were below detection levels in the leaves of seven other wetland species examined, including mangrove, although it has been reported in high concentrations in S. anglia leaves (Larher, Hamelin and Stewart, 1977). Dacey, King and Wakeham (1987) suggest that high fluxes of DMS occur when DMSP is degraded to DMS enzymatically in the leaves. It is not known whether DMS emissions are influenced by the chamber techniques used to measure gas fluxes. Most studies on sulphur gas emission have only been carried out for short time periods (Table 5.4; see also Andreae and Jaeschke, this volume), usually in the summer. If we extrapolate these rates to yearly fluxes (Figure 5.3) we find the loss of volatile sulphur gases represents only a small percentage of the total sulphur reduced by sulphate reduction (Steudler and Peterson, 1985) and is a negligible term in the total sulphur budget.

Sulphur gas fluxes from freshwater wetlands are generally lower than fluxes from saline wetlands (Table 5.4). These fluxes are discussed in detail else-where (Andreae and Jaeschke, this volume). Low rates of sulphur gas emission, and in particular H₂S, from freshwater wetlands may be a result of low rates of sulphate reduction. However, at Big Run Bog where 550 g (S) m^-2 a^-l (17 mol m^-2 a^-l) of sulphate is reduced (Table 5.3), H₂S emission is still minimal at 9.7 mg (S) (0.3 mmol m^-2 a^-l) (Wieder, Yavitt and Lang, this volume). At Big Run Bog, sulphur gas flux does not substantially affect the wetland or watershed sulphur budget. In contrast, however, Nriagu, Holdway and Coker (1987) calculated dimethyl sulphide fluxes from bogs and swamps in northern Ontario to range from 25 to 184 mg (S) m^-2 a^-l (0.78 to 5.8 mmol m^-2 a^-l), based on porewater concentrations and application of a simple diffusion model. Although DMS emission was not directly measured, with sulphur inputs via precipitation of only 1 to 3 g (S) m^-2 a^-l (30 to 100 mmol m^-2 a^-l), a substantial fraction of the sulphur input could be returned to the atmosphere by DMS emission if the estimates of Nriagu, Holdway and Coker (1987) are correct (for a discussion of the assumptions used in making these calculations see Andreae and Jaeschke, this volume).

Figure 5.3. Sulphur gas emissions from fresh and saltwater wetlands. The values given in Table 5.4 were extrapolated to annual values. Values are reported separately for vegetated and bare areas in Spartina marshes

On a global basis, freshwater wetlands occupy about 10 times the area of saline wetlands (Table 5.1), but reported sulphur gas emissions from saline wetlands average about 1000 times greater than emissions from freshwater wetlands (Adams et al., 1981a; Table 5.4; Figure 5.3). It seems likely that freshwater wetlands contribute substantially less sulphur to the global atmosphere than do saline wetlands.

5.7 EFFECT OF THE SULPHUR CYCLE ON METHANE FLUX

Wetlands make a significant contribution to the global methane budget (Matthews and Fung, 1987). It is well recognized that methane fluxes from freshwater wetlands are considerably greater than those from saline wetlands (Bartlett et al., 1987 and references therein). This difference in methane emissions is due at least in part to changes in both the production and consumption of methane in the presence of sulphate-reducing bacteria. Sulphate reduction liberates more energy than methanogenesis, so sulphate reducers out-compete methanogens and dominate anaerobic processes when sulphate is available. At very low concentrations of sulphate, perhaps in the range of 10 to 30 µM SO₄^2- (0.3 to 1 µg (S) l^-1), methanogenesis becomes the dominant pathway for carbon decomposition. Methane is also anoxically oxidized to carbon dioxide in the presence of sulphate; this seems coupled to sulphate-reducing bacteria, although the mechanism remains unclear (see Ivanov, 1989; see also Howarth and Stewart, this volume). The anoxic consumption of methane obviously reduces the flux of methane to the atmosphere.

In saline and brackish water wetlands, a tight coupling between the sulphur cycle and methane release has been demonstrated in both field and laboratory experiments. Methane release from saltwater wetlands is correlated with sulphate depletion (King and Wiebe, 1980), and there is an inverse correlation of methane release and soil salinity across a broad range of saline and brackish wetland types (Bartlett et al., 1987). The situation in freshwater Sphagnum wetlands appears to be more complex. The addition of high levels of sulphate to freshwater wetland soils decreases methane production (Yavitt, Lang and Wieder, 1987). However, at very low sulphate concentrations (<300 µM; 10 mg (S) l^-1), methane production does not appear to be closely related to sulphate concentration (Yavitt and Wieder, unpublished data).

5.8 SUMMARY

The major aspects of the sulphur cycle discussed above have been summarized for saline (Figure 5.4) and freshwater wetlands (Figure 5.5). The fluxes represent the range of values in literature and should only be taken as order of magnitude estimates. The fluxes and standing stocks are most representative of temperate salt marshes (Figure 5.4) and temperate bogs (Figure 5.5).

The tidal fluxes of sulphur through saline wetlands are usually very large in comparison to the rates of sulphur processing within the ecosystem (see Peterson et al., 1983). Most of the sulphur reduced in the sediment is subsequently re-oxidized or exported to creeks, so only a small per cent of the reduced sulphur produced each year is buried. This rapid cycling occurs primarily through inorganic reduced forms of sulphur. Gaseous fluxes are high in comparison to many other ecosystems but represent only a tiny fraction of the total cycled within the wetland.

Figure 5.4. The sulphur cycle of saline wetlands. Fluxes and standing stocks, taken from the literature, are representative of temperate salt marshes. (See text for details)

 Figure 5.5. The sulphur cycle of freshwater wetlands. Fluxes and standing stocks, taken from the literature, are representative of temperate bogs and swamps. (See text for details)

Sulphur inputs to freshwater wetlands are much lower than saline wetlands. In contrast to saline wetlands, the amount of sulphur cycled within freshwater ecosystems equals or exceeds new inputs, perhaps by orders of magnitude. Another difference between fresh and saline marshes is the importance of sulphur uptake by the vegetation in the total sulphur budget. Gaseous losses are apparently lower than saline wetlands, but because of the low sulphur inputs, these losses may represent a significant fraction of the sulphur budget. The cycling of organic sulphur is much more important in freshwater wetlands than saline wetlands (at least as a percentage of total rate of sulphur cycling). Sulphur burial is lower in freshwater wetlands than in saline wetlands, but in general represents a greater percentage of the sulphur inputs.

REFERENCES 

Adams, D. F., Farwell, S. O., Pack, M. R. and Robinson, E. (1981a). Biogenic sulfur gas emissions from soils in Eastern and Southeastern United States. J. Air Poll. Control Assoc., 31, 1083-9.

Adams, D. F., Farwell, S. O., Robinson, E., Pack, M. R. and Bamesberger, W. (1981b). Biogenic sulfur source strengths. Environ. Sci. Technol., 15,1493-8.

Altschuler, Z. S., Schnepfe, M. M., Silber, C. C. and Simon, F. O. (1983). Sulfur diagenesis in Everglades peat and origin of pyrite in coal. Science, 221, 221- 7.

Andreae, M. O. and Jaeschke, W. This volume.

Aneja, V. P. (1986). Characterization of emissions of biogenic atmospheric hydrogen sulfide. Tellus, 38B, 81-6.

Aneja, V. P., Overton, J. H. and Aneja, A. P. (1981). Emission survey of biogenic sulfur flux from terrestrial surfaces. J. Air Poll. Control Assoc., 31, 256-8.

Aneja, V. P., Overton, J. H., Cupitt, L. T., Durham, J. L. and Wilson, W. E. (1979). Direct measurements of emission rates of some atmospheric biogenic sulfur compounds. Tellus, 31, 174-8.

Armstrong, W, (1979). Aeration in higher plants. Adv. Bot. Res., 7, 225-32. 

Armstrong, W. and Boatman, D. J. (1967). Some field observations relating the growth of bog plants to conditions of soil aeration. J. Ecol., 55, 101-10.

Bartlett, K. B., Bartlett, D. S., Harriss, R. C. and Sebacher, D. I. (1987). Methane emissions along a salt marsh salinity gradient. Biogeochemistry, 4, 183-202.

Bayley, S. E., Behr, R. S. and Kelly, C. A. (1986). Retention and release of S from a freshwater wetland. Water Air Soil Pollut.,31, 101-14.

Bayley, S. E., Vitt, D. H., Newberry,R. W., Beatty, K. G., Behr, R. and Miller, C. (1987). Experimental acidification of a Sphagnum-dominated peatland: First year results. Can. J. Fish. Aquat. Sci., 44, 194-205.

Behr, R. S. (1985). Sulfur dynamics in an experimentally acidified mire in northwestern Ontario. M.S. Thesis, University of Manitoba.

Brinson, M. M., Lugo, A. E. and Brown, S. (1981). Primary productivity, decomposition and consumer activity in freshwater wetlands. Annu. Rev. Ecol. Syst., 12, 123- 61.

Brown, K. A. (1985). Sulphur distribution and metabolism in waterlogged peat. Soil Biol. Biochem., 17, 39-45.

Brown, K. A. (1986). Formation of organic sulphur in anaerobic peat. Soil Biol. Biochem., 18,131-40.

Brown, K. A. and MacQueen, J. F. (1985). Sulphate uptake from surface water by peat. Soil Biol. Biochem., 17, 411-20.

Burdick, D. M. and Mendelssohn, I. A. (1987). Waterlogging responses in dune, swale and marsh populations of Spartina patens under field conditions. Oecologia, 74, 321-9.

Canfield, D. E., Raiswell, R., Westrich, J. T., Reaves, C. M. and Berner, R. A. (1986). The use of chromium reduction in the analysis of reduced inorganic sulfur in sediments and shales. Chem. Geol., 54, 149-55.

Carlson, P. R. and Forrest, J. (1982). Uptake of dissolved sulfide by Spartina alterniflora: Evidence from natural sulfur isotope abundance ratios. Science, 216, 633-5.

Casagrande, D. J., Gronli, K. and Sutton, N. (1980). The distribution of sulfur and organic matter in various fractions of peat: Origins of sulfur in coal. Geochim. Cosmochim. Acta, 44, 25-32.

Casagrande, D. J., Siefert, K., Berschinski, C. and Sutton, N. (1977). Sulfur in peatforming systems of the Okefenokee Swamp and Florida Everglades: Origins of sulfur in coal. Geochim. Cosmochim. Acta, 41,161-7.

Casey, W. H. and Lasaga, A. C. (1987). Modeling solute transport and sulfate reduction in marsh sediments. Geochim. Cosmochim. Acta, 51, 1109-20.

Castro, M. S. and Dierberg, F. E. (1987). Biogenic hydrogen sulfide emissions from selected Florida wetlands. Water Air Soil Pollut., 33, 1-13.

Chalmers, A. G. (1982). Soil dynamics and the productivity of Spartina alterniflora. In: Kennedy, V. S. (Ed.). Estuarine Comparisons. Academic Press, New York, pp. 231-42.

Chalmers, A. G., Wiegert, R. G. and Wiebe, W. J. (submitted). Effects of increased drainage in a short Spartina alterniflora marsh.

Clymo, R. S. (1984). The limits to peat bog growth. Phil. Trans. R. Soc. Lond. B Biol. Sci., 303, 605-54.

Clymo, R. S. and Hayward, P. M. (1982). The ecology of Sphagnum. In: Smith, A. J . E. (Ed.). Ecology of Byrophytes. Chapman Hall, London, pp. 229-89.

Cook, R. B. (1981). The biogeochemistry of sulfur in two small lakes. Ph.D. dissertation, Columbia University, New York, 234pp.

Cook, R. B. and Kelly, C. A. This volume.

Cooper, D. J., De Mello, W. Z., Cooper, W. J., Zika, R. G., Saltzman, E. S., Prospero, J. M. and Savoie, D. L. (1987a). Short-term variability in biogenic sulphur emissions from a Florida Spartina alterniflora marsh. Atmos. Environ., 21, 7-12.

Cooper, W. J., Cooper, D. J., Saltzman, E. S., De Mello, W. Z., Savoie, D.L., Zika, R. G. and Prospero, J. M. (1987b). Emissions of biogenic sulphur compounds from several wetland soils in Florida. Atmos. Environ., 21, 1491-5.

Cowdardin, L. M., Cater, V., Gollet, F. C. and Laroe, E. T. (1979). Classification of Wetlands and Deepwater Habitats of the United States. FWS/OBS-79/31, US Fish and Wildlife Service.

Crawford, R. M. M. (1978). Metabolic adaptations to anoxia. In: Hook, D. D. and Crawford, R. M. M. (Eds). Plant Life in Anaerobic Environments. Ann Arbor, Mich, Ann Arbor Science, pp. 119-36.

Dacey, J. W. and Howes, B. L. (1984). Water uptake by roots controls water table movement and sediment oxidation in short Spartina marsh. Science, 224, 487-9.

Dacey, J. W. H., King, G. M. and Wakeham, S. G. (1987). Factors controlling emission of dimethylsulphide from salt marshes. Nature, 330, 643-5.

DeLaune, R. D., Smith, C. J. and Tolley, M. D. (1984). The effect of sediment redox potential on nitrogen uptake, anaerobic root respiration and growth of Spartina alterniflora Loisel. Aquat. Bot., 18, 223-30.

Delmas, R., Baudet, J., Servant, J. and Baziard, Y. (1980). Emissions and concentrations of hydrogen sulfide in the air of the tropical forest of the Ivory Coast and of temperate regions in France. J. Geophys. Res., 85, 4468-74.

De Mello, W. Z., Cooper, D. J., Cooper, W. J., Saltzman, E. S., Zika, R. G., Savoie, D. L. and Prospero, J. M. (1987). Spatial and diel variability in the emissions of some biogenic sulfur compounds from a Florida Spartina alterniflora coastal zone. Atmos. Environ., 21, 987-90.

Feijtel, T. C. J. (1986). Biogeochemical cycling of metals in Barataria Basin. Ph.D dissertation. Louisiana State University.

Ferguson, P., Robinson, R. N. and Press, M. C. (1984). Element concentrations in five Sphagnum species in relation to atmospheric pollution. J. Byrol., 13, 107.

Fossing, H. and Jorgensen, B. B. (1990). Measurement of bacterial sulfate reduction in sediments: Evaluation of a single-step chromium reduction method. Biogeochemistry, 9, 223-46.

Francois, R. (1987). A study of sulphur enrichment in the humic fraction of marine sediments during early diagenesis. Geochim. Cosmochim. Acta, 51, 17-27.

Fry, B., Scalan, R. S., Winters, J. K. and Parker, P. L. (1982). Sulphur uptake by salt grasses, mangroves, and seagrasses in anaerobic sediments. Geochim. Cosmochim. Acta, 46, 1121-4.

Gardner, L. R. (1990). Simulation of the diagenesis of carbon, sulfur, and dissolved oxygen in salt marsh sediments. Ecol. Monogr., 60 (1), 91-111.

Giblin, A. E. (1988). Pyrite formation during early diagenesis. Geomicrobiol. J., 6, 77-97.

Giblin, A. E. and Howarth, R. W. (1984). Porewater evidence for a dynamic sedimentary iron cycle in salt marshes. Limnol. Oceanogr., 29, 47-63.

Goldan, P. D., Kuster, W. C., Albritton, D. L. and Fehsenfeld, F. C. (1987). The measurement of natural sulfur emissions from soils and vegetation: Three sites in the Eastern United States revisted. J. Atmos. Chem., 5, 439-67.

Goldberg, A. B., Maroulis, P. J., Wilner, L. A. and Bandy, A. R. (1981). Study of H₂S emissions from a salt water marsh. Atmos. Environ., 15, 11-18.

Good, R. E., Good, N. F. and Frasco, B. R. (1982). A review of primary production and decomposition dynamics of the belowground marsh component. In: Kennedy, V. S. (Ed.). Estuarine Comparisons. Academic Press, New York, pp. 139-57.

Goodman, P. J. and Williams, W. T. (1961). Investigations into 'dieback' in Spartina townsendii. J. Ecol., 49, 391-8.

Gorham, E., Bayley, S. E. and Schindler, D. W. (1984). Ecological effects of acid deposition upon peatlands: A neglected field in 'acid-rain' research. Can. J. Fish. Aquat. Sci., 41, 1256-68.

Gorham, E. and Detenbeck, N. E. (1986). Sulfate in bog waters: A comparison of ion chromatography with Mackereth's cation-exchange technique and a revision of earlier views on cause of bog acidity. J. Ecol., 74, 899-903.

Grinenko, V. A. and Ivanov, M. V. (1983). Principal reactions of the global biogeochemical cycle of sulfur. In: Ivanov, M. V. and Freney, J. R. (Eds). The Global Biogeochemical Sulfur Cycle, SCOPE 19. Wiley, New York, pp. 1-23.

Hartman, J. M. (1984). The role of wrack in disturbance in the vegetation of a New England salt marsh. Ph.D dissertation, University of Connecticut, 130pp.

Hemond, H. F. (1980). Biogeochemistry of Thoreau's Bog, Concord, Massachusetts. Ecol. Monogr., 50, 507-26.

Hemond, H. F. and Fifield, J. L. (1982). Subsurface flow in salt marsh peat: A model and field study. Limnol. Oceanogr., 27, 126-36.

Hines, M. E. (in press). Emissions of sulfur gases from wetlands. In: Adams, D. D., Seitzinger, S. P. and Crill, P. M. (Eds). Cycling of Reduced Gases in the Hydrosphere.

Hines, M. E., Knollmeyer, S. L. and Tugel, J. B. (1989). Sulfate reduction and other sedimentary biogeochemistry in a northern New England salt marsh. Limnol. Oceanogr., 34, 578-90.

Howarth, R. W. (1979). Pyrite: Its rapid formation in a salt marsh and its importance in ecosystem metabolism. Science, 203, 49-51.

Howarth, R. W. (1984). The ecological significance of sulfur in the energy dynamics of salt marsh and coastal marine sediments. Biogeochemistry, 1, 5-27.

Howarth, R. W. and Giblin, A. E. (1983). Sulfate reduction in the salt marshes at Sapelo Island, Georgia. Limnol. Oceanogr., 28, 70-82.

Howarth, R. W. and Jørgensen, B. B. (1984). Formation of ³⁵S-labelled elemental sulfur and pyrite in coastal marine sediments (Limfjorden and Kysing Fjord, Denmark) during short-term ³⁵SO₄ reduction measurements. Geochim. Cosmochim. Acta, 48, 1807-18.

Howarth, R. W. and Marino, R. (1984). Sulfate reduction in salt marshes, with some comparisons to sulfate reduction in microbial mats. In: Cohen, Y., Castenholz, R. W. and Halvorson, H. O. (Eds). Microbial Mats: Stromatolites. Alan R. Liss, New York, pp. 245-63.

Howarth, R. W. and Merkel, S. (1984). Pyrite formation and the measurement of sulfate reduction in salt marsh sediments. Limnol. Oceanogr., 29, 598-608.

Howarth, R. W. and Stewart, J. W. B. This volume.

Howarth, R. W. and Teal, J. M. (1979). Sulfate reduction in a New England salt marsh. Limnol. Oceanogr., 24, 999-1013.

Howarth, R. W. and Teal, J. M. (1980). Energy flow in a salt marsh ecosystem: The role of reduced inorganic sulfur compounds. Am. Nat., 116, 862-72.

Howarth, R. W., Giblin, A., Gale, J., Peterson, B. J. and Luther, G. W. (1983). Reduced sulfur compounds in the pore waters of a New England salt marsh. In: Hallberg R. O. (Ed.). Environmental Biogeochemistry. Ecol. Bull. (Stockholm), 35, 135-52.

Howes, B. L., Dacey, J. W. H. and Goehringer, D. D. (1986). Factors controlling the growth form of Spartina alterniflora: Feedbacks between above-ground production, sediment oxidation, nitrogen and salinity. J. Ecol., 74, 881-98.

Howes, B. L., Dacey, J. W. H. and King, G. M. (1984). Carbon flow through oxygen and sulfate reduction pathways in salt marsh sediments. Limnol. Oceanogr., 29, 1037-51.

Howes, B. L., Dacey, J. W. H. and Teal, J. M. (1983). Annual carbon mineralization and below ground production of Spartina alterniflora in a New England salt marsh. Ecology, 66, 595-605.

Howes, B. L., Howarth, R. W., Valiela, I. and Teal, J. M. (1981). Oxidation- reduction potentials in a salt marsh: Spatial patterns and interactions with primary production. Limnol. Oceanogr., 26, 350-60.

Ingram, H. A. P. (1978). Soil layers in mires: Function and terminology. J. Soil Sci., 29, 224-7.

Ivanov, M. V. (1956). Isotopes in the determination of the sulfate-reduction rate in Lake Belovod. Microbiologia, 25, 305-9.

Ivanov, M. V., Lein, A. Yu and Kashparova, E. V. (1976). Intensity of formation and diagenetic transformation of reduced sulfur compounds in sediments of the Pacific Ocean. In: The Biogeochemistry of Diagenesis of Ocean Sediments. Mauka, Moscow, pp. 171-8 (in Russian).

Ivanov, M. V., Lein, A. Yu., Beylaev, S. S., Nesterov, A. I., Bonder, V. A. and Zhabina, N. N. (1980). Geochemical activity of sulfate reducing bacteria in benthal sediments of the North-West Indian Ocean. Geokhim., 8, 1238-49 (in Russian).

Ivanov, M. V., Lein, A. Yu., Reeburgh, M. S. and Skyring, G. W. (1989). Interaction of sulphur and carbon cycles in marine sediments. In: Brimblecombe, P. and Lein, A. Yu. (Eds). Evolution of the Global Biogeochemical Sulphur Cycle. Wiley, Chichester, pp. 125-79.

Jørgensen, B. B. (1977). The sulfur cycle of a coastal marine sediment (Limfjorden, Denmark). Limnol. Oceanogr., 22, 814-32.

Jørgensen, B. B. (1978). A comparison of methods for quantification of bacterial sulfate reduction in coastal marine sediments. I. Measurement with radiotracer techniques. Geomicrobiol. J., 1,11-27.

Jørgensen, B. B. (1982). Mineralization of organic matter in the sea bed: The role of sulfate reduction. Nature, 296, 643-5.

Kaplan, I. R., Emery, K. O. and Rittenberg, S. C. (1963). The distribution and isotopic abundance of sulfur in recent marine sediments off southern California. Geochim. Cosmochim. Acta, 27, 297-331.

King, G. M. (1983). Sulfate reduction in Georgia salt marsh soils: An evaluation of pyrite formation using ³⁵S and ⁵⁵Fe tracers. Limnol. Oceanogr., 28, 987-95.

King, G. M. (1988). Patterns of sulfate reduction and the sulfur cycle in a South Carolina salt marsh. Limnol. Oceanogr., 33, 376-90.

King, G. M. and Wiebe, W. J. (1980). Regulation of sulfate concentrations and methanogenesis in salt marsh soils. Estuarine Coastal Mar. Sci., 10, 215-23.

King, G. M., Klug, M. J., Wiegert, R. G. and Chalmers, A. G. (1982). Relation of soil water movement and sulfide concentration to Spartina alterniflora production in a Georgia salt marsh. Science, 218, 61-3.

Lamb, B., Westberg, H., Allwine, G., Bamesberger, L. and Guenther, A. (1987). Measurement of biogenic sulfur emissions from soils and vegetation: Application of dynamic enclosure methods with Natusch Filter and GC/FPD analysis. J. Atmos. Chem., 5, 469-91.

Larcher, F., Hamelin, J. and Stewart, G. R. (1977). L'acid dimethylsulfonium-3 propanique-3 de Spartina alglica. Phytochemistry, 16, 2019-20.

Lein, A. Yu., Vaynshteyn, M. B., Namsarayev, B. B., Kashparova, E. V., Matrosov, A. G., Bondar, V. A. and Ivanov, M. V. (1982). Biogeochemistry of anaerobic diagenesis of recent Baltic Sea sediments. Geochem. Int., 19, 90-103.

Lord, C. J. and Church, T. M. (1983). The geochemistry of salt marshes: Sedimentary ion diffusion, sulfate reduction, and pyritization. Geochim. Cosmochim. Acta, 47, 1381-91.

Lowe, L. E. (1986). Application of sequential extraction procedure to the determination of the distribution of sulfur forms in selected peat materials. Can. J. Soil Sci., 66, 337-45.

Lowe, L. E. and Bustin, R. M. (1985). Distribution of sulfur forms in six facies of peats of the Fraser River delta. Can. J. Soil Sci., 65, 531-41.

Luther, G. W. and Church, T. M. (1988). Seasonal cycling of sulfur and iron in porewaters of a Delaware salt marsh. Mar. Chem., 23, 295-309.

Luther, G. W. and Church, T. M. This volume.

Luther, G. W., Church, T. M., Scudlark, J. R. and Cosman, M. (1986). Inorganic and organic sulfur cycling in salt-marsh pore waters. Science, 232, 746-9.

Malt by, E. (1988). Global wetlands-History, current status and future. In. Hook, D. D. (Ed.). The Ecology and Management of Wetlands. Timber Press, Portland, Oregon, pp. 3-14.

Matthews, E. and Fung, I. (1987). Methane emission from natural wetlands: Global distribution, area, and environmental characteristics of sources. Global Biogeochemical Cycles, 1, 61-86.

Mendelssohn, I. A., McKee, K. L. and Patrick, W. H. (1981). Oxygen deficiency in Spartina alterniflora roots: Metabolic adaptation to anoxia. Science, 214, 439-41.

Mendelssohn, I. A. and Seneca, E. D. (1980). The influence of soil drainage on the growth of salt marsh cordgrass Spartina alterniflora in North Carolina. Estuarine Coastal Shelf Sci., 11, 27-40.

Mitsui, S. (1965). Dynamic aspects of nutrient uptake. In: The Mineral Nutrition of the Rice Plant. Symp. Proc. Int. Rice Res Inst. John Hopkins Press, pp. 53-62.

Morris, J. T. and Whiting, G. J. (1986). Emission of gaseous CO₂ from salt marsh sediments and its relationship to other carbon losses. Estuaries, 9, 9-19.

Nedwell, D. B. (1984). The input and mineralization of organic carbon in anaerobic aquatic sediments. In: Marshall, K. C. (Ed.). Advances in Microbial Ecology. Plenum Press, New York, pp. 93-131.

Nedwell, D. B. and Abram, J. W. (1978). Bacterial sulphate reduction in relation to sulphur geochemistry in two contrasting areas of saltmarsh sediment. Estuarine Coastal Shelf Sci., 6, 341-51.

Nedwell, D. B. and Takii, S. (1988). Bacterial sulphate reduction in sediments of a European salt marsh: Acid-volatile and tin-reducible products. Estuarine Coastal Shelf Sci., 26, 599-606.

Neue, H. W. and Mamaril, C. P. (1985). Zinc, sulfur, and micronutrients in wetland soils. In: Proc. Wetland Soils Characterization, Classification and Utilization. International Rice Research Institute, pp. 307-19.

Nickerson, N. H. and Thibodeau, F. R. (1985). Association between pore water sulfide concentrations and the distribution of mangroves. Biogeochemistry, 1, 183- 92.

Nriagu, J. O., Holdway, D. A. and Coker, R. D. (1987). Biogenic sulfur and the acidity of rainfall in remote areas of Canada. Science, 237, 1189-92.

Odum, W. E., McIvor, C. C. and Smith, T. J. (1982). The Ecology of the Mangroves of South Florida: A Community Profile. FWS/OBS-81/24, Fish and Wildlife Service.

Oenema, O. (1988). Early diagenesis in recent fine-grained sediments in the Eastern Scheldt. Ph.D dissertation, University of Utrecht, Netherlands, 223pp.

Ogden, J. G. (1982). Seasonal mass balance of major ions in three small watersheds in a maritime environment. Water Air Soil Pollut., 17, 119-30.

Peterson, B. J., Steudler, P. A., Howarth, R. W., Friedlander, A. I., Juers, D. and Bowles, F. P. (1982). Tidal export of reduced sulfur from a salt marsh ecosystem. Environ. Biogeochem. Ecol. Bull. (Stockholm), 35, 153-65.

Postma, D. (1982). Pyrite and siderite formation in brackish and freshwater swamp sediments. Am. J. Sci., 282, 1151-83.

Rudd, J. W. M., Kelly, C. A. and Furutani, A. (1986). The role of sulfate reduction in  long term accumulation of organic and inorganic sulfur in lake sediments. Limnol. Oceanogr., 31, 1281-91.

Schoenau, J. J. and Germida, J. J. This volume.

Schubauer, J. P. and Hopkinson, C. S. (1984). Above- and below ground emergent macrophyte production and turnover in a coastal marsh ecosystem, Georgia. Limnol. Oceanogr., 29, 1052-65.

Skyring, G. W. (1987). Sulfate reduction in coastal ecosystems. Geomicrobiol. J., 5, 295-374.

Skyring, G. W., Oshrain, R. L. and Wiebe, W. J. (1979). As assessment of sulfate reduction rates in Georgia marshland soils. Geomicrobiol. J., 1, 398-400.

Sparling, J. H. (1966). Studies on the relationship between water movement and water chemistry in mires. Can. J. Bot., 44, 747-58.

Spratt Jr, H. G. and Morgan, M. D. (1990). Sulfur cycling in a cedar dominated freshwater wetland. Limnol. Oceanogr., 35, 1586-93.

Spratt, Jr H. G., Morgan, M. D. and Good, R. E. (1987). Sulfate reduction in peat from a New Jersey pinelands cedar swamp. Appl. Environ. Microbiol., 53, 1406-11.

Steudler, P. A. and Peterson, B. J. (1984). Contribution of gaseous sulphur from salt marshes to the global sulphur cycle. Nature, 311, 455-7.

Steudler, P. A. and Peterson, B. J. (1985). Annual cycle of gaseous sulfur emissions from a New England Spartina alterniflora marsh. Atmos. Environ., 19, 1411-16.

Swider, K. T. (1988). Transformations of sulfur compounds in salt marsh sediments. M.S. Thesis, SUNY at Stony Brook.

Tabatabai, M. A. This volume.

Taylor, B. E., Wheeler, M. C. and Nordstrom, D. K. (1984). Stable isotope geochemistry of acid mine drainage: Experimental oxidation of pyrite. Geochim. Cosmochim. Acta, 48, 2669-78.

Teal, J. M. and Kanwisher, J. (1961). Gas exchange in a Georgia salt marsh. Limnol. Oceanogr., 6, 388-99.

Thibodeau, F. R. and Nickerson, N. H. (1986). Differential oxidation of mangrove substrate by Avicennia germinans and Rhizophora mangle. Am. J. Bot., 73, 512-16.

Urban, N. R. and Bayley, S. E. (1986). The acid-base balance of peatlands: A short-term perspective. Water Air Soil Polut., 30, 791-800.

Urban, N. R., Eisenreich, S. J. and Grigal, D. F. (1989). Sulfur cycling in a forested Sphagnum bog in northern Minnesota. Biogeochemistry, 7, 81-109.

Valiela, I., Teal, J. M. and Persson, N. Y. (1976). Production and dynamics of experimentally enriched salt marsh vegetation: Below ground biomass. Limnol. Oceanogr., 21, 245-52.

Vamos, R. and Koves, E. (1972). Role of light in the prevention of the poisoning action of hydrogen sulfide in the rice plant. Ecology, 53, 519-25.

Van Breemen, V. (1982). Genesis, morphology and classification of acid sulfate soils in coastal plains. In: Kit trick, J. A., Fanning, D. S. and Hossner, L. R. (Eds). Acid Sulfate Weathering. Soil Science Society of America, Madison, WI, pp. 95-108.

Westrich, J. T. and Berner, R. A. (1984). The role of sedimentary organic matter in bacterial sulfate reduction: The G model tested. Limnol. Oceanogr., 29, 236-49.

Whigham, D. F., McCormick, J., Good, R. E. and Simpson, R. L. (1978). Biomass and primary production in freshwater tidal marshes of the Middle Atlantic Coast. In: Good, R. E., Whigham, D. F. and Simpson, R. L. (Eds). Freshwater Wetlands: Ecological Process and Management Potential. Academic Press, New York, pp. 3- 20. .

Wieder, R. K. and Lang, G. E. (1986). Fe, Al, Mn, and S chemistry of Sphagnum peat in four peatlands with different metal and sulfur inputs. Water Air Soil Pollut., 29, 309-20.

Wieder, R. K. and Lang, G. E. (1988). Cycling of inorganic and organic sulfur in peat from Big Run Bog, West Virginia. Biogeochemistry, 5, 221-42.

Wieder, R. K., Lang, G. E. and Granus, V. A. (1985). An evaluation of wet chemical methods for quantifying sulfur fractions in freshwater wetland peat. Limnol. Oceanogr., 30, 1109-15.

Wieder, R. K., Yavitt, J. B. and Lang, G. E. (1990). Methane production and sulfate reduction in two Appalachian peatlands. Biogeochemistry, 10 (2), 81-104.

Wieder, R. K., Yavitt, J. B. and Lang, G. E. This volume.

Wieder, R. K., Yavitt, J. B., Lang, G. E. and Bennett, C. A. (1989). Annual aboveground net primary production at Big Run Bog, West Virginia. Castanea (in press).

Wiegert, R. G., Chalmers, A. G, and Randerson, P. F. (1983). Productivity gradients in salt marshes: The response of Spartina alterniflora to experimentally manipulated soil water movement. Oikos, 41, 1-6.

Yavitt, J. B., Lang, G. E. and Wieder, R. K. (1987). Control of carbon mineralization to CH₄ and CO₂ in anaerobic, Sphagnum-derived peat from Big Run Bog, West Virginia. Biogeochemistry, 4, 141-57.

Yuretich, R. F., Crerar, D. A., Kinsman, D. J. J., Means, J. T. and Borsik, M. P. B. (1982). Hydrogeochemistry of the New Jersey Coastal Plain. I. Major element cycles in precipitation and rainwater. Chem. Geol., 33, 1-21.

Zhabina, N. N. and Volkov, I. I. (1978). A method of determination of various sulfur compounds in sea sediments and rocks. In: Krumbein W. E. (Ed.). Environmental Biogeochemistry and Geomicrobiology: Methods, Metals and Assessment. Ann Arbor, Ann Arbor Science Publishers, pp. 735-45.

Freshwater peatlands cover some 450 million ha of the earth's land surface (Kivinen and Pakarinen, 1981), mainly north of 49° N latitude. Large areas of peatland, especially in northeastern North America and in northern Europe, receive large inputs of sulphur in acid deposition. The effect of this deposition on peatland ecosystems has received much less study than have effects on forests and lakes (Gorham, Bayley and Schindler, 1984). We have hypothesized that one effect of sulphate deposition to peatlands may be to increase rates of anoxic decomposition and decrease the storage of organic carbon in the peat (Wieder, Yavitt and Lang, 1990). This hypothesis is based on our research on sulphur biogeochemistry in Big Run Bog, West Virginia. Big Run Bog receives a considerable loading of atmospheric sulphur, some 1.8 g (S) m^-2 a^-1 (0.056 mol m^-2 a^-1 in wet deposition alone).

Total sulphur concentration in Big Run Bog peat averages 3.9 mg g^-l dry mass (123 µmol g^-l dry mass), with 81% as carbon-bonded S, 10% as ester-sulphate S, and less than 10% as inorganic S, including pyrite, iron monosulphides, elemental sulphur, and sulphate (Wieder and Lang, 1988). Given that organic sulphur pools rather than inorganic sulphur pools dominate in Big Run Bog peat, one might suppose that sulphur fluxes would be mainly through the organic pools (via mineralization and immobilization) rather than through the inorganic pools (via oxidation and reduction). Also, traditional views of anaerobic carbon mineralization in freshwater environments held that sulphate reduction should be of minimal importance, being limited by a low sulphate concentration (see review by Nedwell, 1984). Thus, we were surprised to find that the major fate of ³⁵SO₄^2- in short-term incubations was the formation of reduced inorganic sulphur via the process of dissimilatory sulphate reduction (Table 5A.1; Wieder and Lang, 1988). Furthermore, rates of sulphate reduction were comparable to rates reported for coastal marine sediments, high sulphate environments (Giblin and Wieder, this volume). High rates of sulphate reduction in Big Run Bog apparently can persist despite a low instantaneous dissolved sulphate concentration because there is a continuous replenishment of the dissolved sulphate pool via the oxidation of reduced inorganic sulphur (Wieder and Lang, 1988; Table 5A.1).

High sulphur deposition at Big Run Bog, along with the unexpectedly high rates of sulphate reduction in Big Run Bog peat, may have important implications for carbon balance. Big Run Bog is in many ways structurally and functionally similar to more northern peatland ecosystems. However, one difference is that at Big Run Bog, a maximum of only 2.25 m of peat has accumulated over a relatively long period of time (approximately 13 000 years; Wieder, 1985). Such slow rates of peat accumulation are typical of Appalachian Sphagnum-dominated peatlands (Maxwell and Davis, 1982; Watts, 1979; Cotter, 1983). In contrast, in many northern peatlands considerably greater quantities of peat have accumulated over much shorter periods of time. The slow rate of peat accumulation at Big Run Bog is probably not because of low carbon inputs by net primary production. Present day net primary production at Big Run Bog is quite high (approximately 450 g (C) m^-2 a^-l), and in general net primary production in peatlands increases with decreasing latitude (Wieder and Lang, 1983; Wieder et al., 1989). Moreover , it does not appear to be the case that high atmospheric sulphur deposition has inhibited the growth of Sphagnum species at Big Run Bog, as has been suggested for peatlands in the southern Pennines of Great Britain (Austin and Wieder,1987).

Table 5A.1 Summary of sulphur pool sizes and transformation rates in Sphagnum-derived peat from Big Run Bog. Values are means of 36 determinations on peat collected in March 1986, and are based on direct measurements on the short-term fate of ³⁵SO₄^2- injected into peat samples. Details are in Wieder and Lang (1988). Reduced inorganic S is equivalent to chromium-reducible S, and therefore includes H₂S, FeS, S^o, and FeS₂. The formation rate listed for sulphate is calculated as the negative value of the sum of the formation of reduced inorganic S, ester-sulphate S, and carbon-bonded S. Because formalin-killed controls exhibited essentially no conversion of added label into either organic or inorganic products, the transformations listed below are biologically mediated. Formation rates and pool sizes are one gram wet weight basis

Historically, longer growing seasons and generally higher mean annual temperatures associated with decreasing latitude probably have led to net primary production being greater in southern peatlands than in northern peatlands (Damman, 1979). Southern peatlands like Big Run Bog would seem likely sites for high rates of carbon accumulation because of the high rate of net primary production. That this is not true and that rates of peat accumulation are low implies not only that carbon mineralization to CO₂ and CH₄ must be relatively high, but also that annual fluxes of carbon to and from the atmosphere must be high.

At Big Run Bog, anaerobic carbon mineralization in the top 35 cm of peat is dominated by sulphate reduction (Table 5A.2). Of the total annual anaerobic carbon mineralization to CO₂ and CH₄ in the top 35 cm of peat of 53 mol m^-2 (640 g (C) m^-2 a^l), 64% is attributable to sulphate reduction, 23% to methanogenesis, and 12% to other processes (Wieder, Yavitt and Lang, 1990). The dominance of sulphate reduction is apparently not at the expense of methanogenesis. Rates of sulphate reduction and methane production in replicate peat subsamples are not negatively correlated (Wieder, Yavitt and Lang, 1990). Moreover, sulphate additions to peat do not consistently inhibit methane production. Although sulphate additions to concentrations of 1 mM (32 mg l^-1) or greater sometimes inhibit methane production, this inhibition is not consistently accompanied by an increase in anaerobic CO₂ production, as would be expected if sulphate-reducing bacteria were out-competing methane producers at these enhanced sulphate levels (Yavitt, Wieder and Lang, 1988).

Table 5A.2 Summary of anaerobic carbon mineralization in the top 35 cm of Big Run Bog peat. Methane production and anaerobic carbon dioxide production were obtained by measurement of gas accumulation in the headspace of flasks containing peat during incubation in the laboratory .Sulphate reduction data were determined radiometrically. Values are based on 191 determinations for CH₄ production and CO₂ production and on 333 determinations of sulphate reduction in peat samples collected on eight dates throughout the course of an entire year. Data are from Wieder, Yavitt and Lang, 1990.

Most of the world's peatlands receive considerably less atmospheric sulphur input than that received by Big Run Bog. The high sulphate input in acid deposition to Big Run Bog, and the spiralling of added sulphur between oxidized and reduced inorganic states during its residence in the bog, have created conditions where sulphate reduction contributes much more to anaerobic carbon mineralization than the sulphate deposition rate alone might suggest. Sulphate reduction in the top 35 cm of Big Run Bog peat is 17 mol m^-2 a^-1 (540 g (S) m^-2 a^-1), whereas annual sulphate input from wet and dry deposition is only 0.11 mol m^-2 (3.6 g (S) m^-2 a^-1) (Wieder, Yavitt and Lang, 1990; see Tables 5A.1 and 5A.2). Carbon-balance calculations for Big Run Bog indicate that in the absence of sulphate reduction, the Big Run Bog peat deposit would more than likely be either at steady state or in a condition of net peat accumulation. However, under extant conditions aerobic and anaerobic carbon mineralization to CO₂ and CH₄ exceeds carbon fixation by net primary production. The carbon balance is negative and the peat deposit is apparently being degraded (Wieder, Yavitt and Lang, 1990).

Our work at Big Run Bog has considerably broader implications. Although estimates vary, peatland ecosystems (including tundra, boreal forests, and wetlands) may contain as much as 34% of the global soil carbon pool (Schlesinger, 1977; Houghton et al., 1985; Trabalka, 1985). Most of the carbon in peatland ecosystems is stored as peat that has accumulated since the most recent glaciation. As such, peatlands have served as global net sinks for atmospheric carbon over the past several thousand years, and continue to represent significant carbon reservoirs as well as potential sources and sinks for atmospheric carbon (Miller, 1981; Trabalka, 1985; Sebacher et al., 1986; Matthews and Fung, 1987; Moore and Knowles, 1987). It is quite possible that in regions that are not exposed to high atmospheric sulphur deposition, anoxic carbon mineralization may be dominated by methanogenesis. However, our findings from Big Run Bog suggest that if such peatlands were to receive increased atmospheric sulphur deposition; the effect may be to enhance sulphate reduction without inhibiting methane production. As a result, CH₄ emissions to the atmosphere may remain unaffected, while CO₂ emissions may be stimulated, thereby augmenting the role of peatlands as sources of atmospheric carbon. According to such a scenario, peatland ecosystems may provide an important link between anthropogenically produced acid precipitation and global carbon cycling.

REFERENCES

Austin, K. A. and Wieder, R. K. (1987). Effects of elevated H⁺, SO₄^2-, NO₃^-, and NH₄⁺ in simulated acid precipitation on the growth and chlorophyll content of 3 North American Sphagnum species. Bryologist, 90, 221-9.

Cotter, I. F. P. (1983). The minimum age of the Woodfordian deglaciation of northeastern Pennsylvania and northwestern New Jersey. Ph D dissertation, Lenigh University, Betheleham, PA, USA.

Damman, A. W. H. (1979). Geographic patterns in peatland development in eastern North America. Proceedings of the International Symposium on Classification of Peat and Peatlands. Hyytiala, Finland, pp. 42-57.

Giblin, A. E. and Wieder, R. K. (This volume).

Gorham, E., Bayley, S. E. and Schindler, D. W. (1984). Ecological effects of acid deposition upon peatlands: a neglected field in 'acid rain' research. Can. J. Fish. Aquat. Sci., 41, 1256-68.

Houghton, R. A., Schlesinger, W. H., Brown, S. and Richards, J. F. (1985). Carbon dioxide exchange between the atmosphere and terrestrial ecosystems. In: Trabalka, J. R. (Ed.). Atmospheric Carbon Dioxide and the Global Carbon Cycle. DOE/ER- 0239 US Dept. of Energy, Off. Energy Research, Carbon Dioxide Research Div., Washington DC, pp. 113-40.

Kivinen, E. and Pakarinen, P. (1981). Geographical distribution of peat resources and major peatland complex types in the world. Annales Academiae Scientiarum Fennicae, Series A, III. Geologica-Geographica, 132.

Matthews, E. and Fung, I. (1987). Methane emission from natural wetlands: global distribution, area, and environmental characteristics of sources. Global Biogeochem. Cycles, 1, 61-87.

Maxwell, J. A. and Davis, M. B. (1972). Pollen evidence of Pleistocene and Holocene vegetation of the Allegheny Plateau, Maryland. Quaternary Research, 2, 506-30.

Miller, P. C. (1981). Carbon balance in northern ecosystems and the potential effect of carbon dioxide induced climate change, CONF-8003118, US Dept. of Energy , Off. Health Environ. Research, Washington DC.

Moore, T. R. and Knowles, R. (1987). Methane and carbon dioxide evolution from subarctic fens. Can. J. Soil Sci., 67, 321-43.

Nedwell, D. B. (1984). The input and mineralization of organic carbon in anaerobic aquatic sediments. Adv. Microb. Ecol., 7, 93-131.

Schlesinger, W. H. (1977). Carbon balance in terrestrial detritus. Annu. Rev. Ecol. Syst., 8, 51-81.

Sebacher, D. I., Harriss, R. C., Bartlett, K. B., Sebacher, S. M. and Grice, S. S. (1986). Atmospheric methane sources: Alaska tundra bogs, an alpine fen, and a subarctic boreal marsh. Tellus, 38B, 1-10.

Trabalka, J. R. (1985). Atmospheric carbon dioxide and the global carbon cycle, DOE/ER-0239. US Dept. of Energy, Off. Energy Research, Carbon Dioxide Research Div., Washington DC.

Watts, W. A. (1979). Late quaternary vegetation of central Appalacia and the New Jersey coastal plain. Ecol. Monog., 49, 427-69.

Wieder, R. K. (1985). Peat and water chemistry at Big Run Bog, a peatland in the Appalachian Mountains of West Virginia. Biogeochemistry, 1, 277-302.

Wieder, R. K. and Lang, G. E. (1983). Net primary production of the dominant bryophytes in a Sphagnum-dominated wetland in West Virginia. Bryologist, 86, 280-6.

Wieder, R. K. and Lang, G. E. (1988). Cycling of inorganic and organic sulphur in peat from Big Run Bog, West Virginia. Biogeochemistry, 5, 221-42.

Wieder, R. K., Yavitt, J. B. and Lang, G. E. (1990). Methane production and sulphate reduction in two Appalachian peatlands. Biogeochemistry, 10, 81-104.

Wieder, R. K., Yavitt, J. B., Lang, G. E. and Bennett, C. A. (1989). Aboveground net primary production at Big Run Bog, West Virginia. Castanea (in press).

Yavitt, J. B., Wieder, R. K. and Lang, G. E. (1988). Control of carbon mineralization to CH₄ and CO₂ in anaerobic Sphagnum-derived peat from Big Run Bog, West Virginia. Biogeochemistry, 4, 141-57.