and/or highly reactive organic matter from enhanced root growth, inclusive of root exudates (Bernal et al., 2017; A. A. Wolf et al., 2007). In some peatlands exposed to elevated CO2, the activity of the extracellular enzymes β‐glucosidase and phenol oxidase decreased (Fenner et al., 2007) or did not change (Kang et al., 2005), perhaps because reactive carbon was not limiting at those sites. Using elevation change as a proxy, elevated CO2 enhanced belowground productivity and increased soil carbon storage in a brackish tidal marsh (Langley et al., 2009).
The intrusion of saline water into freshwater systems can affect wetland–atmosphere CO2 exchanges. Net ecosystem production is often depressed by saltwater intrusion (Herbert et al., 2018; Neubauer, 2013) but can be unchanged in some years or in response to transient salinity increases (Herbert et al., 2018). The changes in net ecosystem production reflect salinity‐related declines in plant CO2 fixation (Neubauer, 2013; Sutter et al., 2014) and variable heterotrophic respiration responses to increased salinity (Herbert et al., 2015). Changes in heterotrophic respiration could reflect a shift from methanogenesis to energetically favorable SO42– reduction (Weston et al., 2011), reduced activity of extracellular enzymes (Jackson & Vallaire, 2009; Neubauer et al., 2013), or indirect effects that are mediated through soil organic matter availability and composition, microbial community structure, soil O2 availability, and/or nutrient availability (Herbert et al., 2015; Tully et al., 2019).
Fire is an increasingly common feature in many wetlands, especially during drought or periods of seasonal water drawdown (Hope et al., 2005; Turetsky, Kane et al., 2011) and intentional land clearing activities (Marlier et al., 2015). Fire represents a pathway for the abiotic oxidation of wetland biomass and soil organic matter, generating emissions of CO2 (and much smaller amounts of CH4; Kuwata et al., 2016). Surface fires cause a short‐term burst of CO2 emissions as surface vegetation and litter are burned but may promote a decrease in long‐term CO2 emissions if thermally altered organic matter becomes more resistant to microbial decomposition (Flanagan et al., 2020). Smoldering fires can burn tens of centimeters of soil organic matter, converting hundreds to thousands of years of accumulated carbon back to CO2 and significantly increasing global CO2 emissions (Page et al., 2002; Turetsky et al., 2015; Turetsky, Donahue et al., 2011).
Methane (CH4)
Under anaerobic conditions, the final step of the mineralization of organic carbon results in the production of CH4, which is carried out by a subset of the Archaea called methanogens (Bridgham et al., 2013; Megonigal et al., 2004). Methane emissions to the atmosphere reflect the balance between rates of CH4 production (methanogenesis) and CH4 oxidation (methanotrophy). Methane also can be produced abiotically by the burning of vegetation and peat, which can be especially important in years when large peatland fires occur (Kuwata et al., 2016). The last fifteen years have seen reports of aerobic CH4 production by plants (Bruhn et al., 2012; Keppler et al., 2006), fungi (Lenhart et al., 2012), soil macrofauna (Kammann et al., 2009), and in the water column (Damm et al., 2010; Grossart et al., 2011); the importance of these pathways in wetlands is unknown. Globally, wetlands are the largest source of CH4 to the atmosphere, with natural wetlands accounting for 30% of all CH4 emissions (natural + anthropogenic) and paddies associated with rice cultivation adding another 5% to the total (Saunois et al., 2016). Although CH4 is a powerful greenhouse gas, wetland CH4 emissions are not contributing to recent climate change, except to the extent that these emissions have changed in the last ~250 years (Section 3.2).
Methanogenesis has the lowest yield of the terminal metabolic pathways so it tends to be most important when other terminal metabolic pathways are limited by low rates of electron acceptor resupply/regeneration and/or when supply rates of acetate, H2, and other suitable electron donors are high enough to relieve competition with other anaerobic decomposers. The production of CH4 typically requires anaerobic conditions, such that rates of CH4 emissions are inversely related to soil O2 levels (Smyth et al., 2019) and rates of methanogenesis drop sharply in response to decreases in wetland water levels (MacDonald et al., 1998; T. R. Moore & Knowles, 1989). Oxygen inputs can stimulate aerobic respiration (Mueller, Jensen et al., 2016; A. A. Wolf et al., 2007) and/or reoxidize alternate terminal electron acceptors (Laanbroek, 2010; Neubauer, Givler et al., 2005), such that methanogens may be unable to successfully compete for electron donors. Rates of methanogenesis can also be suppressed by the delivery of alternate terminal electron acceptors – largely NO3–, Fe(III), and/or SO42– – from saltwater intrusion (L. G. Chambers et al., 2013; Kroeger et al., 2017; Neubauer et al., 2013), fertilizer runoff (Bodelier, 2011; Kim et al., 2015), atmospheric deposition (Gauci et al., 2004; Watson & Nedwell, 1998), and river flooding (Luo et al., 2020).
There is tight coupling between plant activity and CH4 emissions (Whiting & Chanton, 1993), in part because plants produce low molecular weight organic molecules that can be used by methanogens (Dorodnikov et al., 2011; Megonigal et al., 1999). Plants can also prime the decomposition of soil organic matter (Basiliko et al., 2012; Bernal et al., 2017), thus providing substrates that fuel methanogenesis. Plant species composition affects CH4 cycling (Kao‐Kniffin et al., 2010) due to differences in the reactivity of carbon supplied by each vegetation type (e.g., Chanton et al., 2008). Humic substances inhibit the production of CH4, either through direct competition between microbial humic reducers and methanogens or, alternately, by abiotically reoxidizing reduced sulfur compounds and therefore supporting sulfate reducers that outcompete the methanogens (Heitmann et al., 2007; Keller, Weisenhorn et al., 2009). The polyphenol sphagnum acid and the polysaccharide sphagnan, both of which are produced by Sphagnum mosses, can interfere with methanogenic activity (van Breemen, 1995; Bridgham et al., 2013) and help explain why some peatlands have low rates of methanogenesis despite low concentrations of inorganic terminal electron acceptors such as Fe(III) and SO42– (Galand et al., 2010; Keller & Bridgham, 2007; Vile et al., 2003).
Methanotrophy, which oxidizes CH4 to CO2, can proceed aerobically using O2 as the electron acceptor or anaerobically using the entire suite of alternate terminal electron acceptors (Bridgham et al., 2013). Whether a wetland emits gas as CH4 or CO2 is unimportant in the context of a wetland’s carbon budget but has large implications for the radiative balance of the wetland. On a global basis, the aerobic oxidation of CH4 can prevent 40–70% of the CH4 produced in wetlands from reaching the atmosphere (Megonigal et al., 2004), but it is rare that annual wetland CH4 oxidation exceeds methanogenesis (that is, very few wetlands are net sinks for CH4; Bridgham et al., 2006; Harriss et al., 1982; Petrescu et al., 2015). Beyond the first‐order control that the aerobic oxidation of CH4 requires O2, the availability of O2 can regulate methanotrophy when there is a narrow aerobic zone, when CH4 spends little time in the aerobic zone before being emitted to the atmosphere (as would happen when most CH4 emissions are via ebullition and/or transport through plants), and/or when rates of CH4 production are high (Megonigal et al., 2004). Conversely, methanotrophy can be limited by the availability of CH4 when rates of CH4 production are low and/or there is a large diffusive aerobic zone (Megonigal & Schlesinger, 2002). High concentrations of ammonium (NH4+) inhibit methane oxidation because both CH4 and NH4+ compete for the same sites on the enzyme methane monooxygenase (Bodelier & Frenzel, 1999; Crill et al., 1994). However, it is also possible that methanotrophs can be nitrogen limited, such that fertilization increases rates of CH4 oxidation (Bodelier et al., 2000). Like all biological processes, rates of aerobic methanotrophy increase with increasing temperatures, although methanotrophy is less sensitive to temperature than is methanogenesis (Segers, 1998).
Rates of the anaerobic oxidation of CH4 can be of the same magnitude as aerobic oxidation (Smemo &
