Wetland Carbon and Environmental Management. Группа авторов

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 SO₄^2– – 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 SO₄^2– (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 & Yavitt, 2007) and, globally, may be comparable to the total CH₄ emissions from freshwater wetlands (Segarra et al., 2015). Rates of anaerobic oxidation of CH₄ in wetlands and wet soils are correlated with rates of CH₄ production (Blazewicz et al., 2012; Segarra et al., 2015). The anaerobic oxidation of CH₄ can be coupled with the reduction of NO₃^– or nitrite (NO₂^–) (Hu et al., 2014; Raghoebarsing et al., 2006), Mn(III, IV) and Fe(III) (Beal et al., 2009; Egger et al., 2015), humic acids (Smemo & Yavitt, 2011; Valenzuela et al., 2017), or SO₄^2– (Egger et al., 2015; Knittel & Boetius, 2009). It is not always straightforward to identify which electron acceptors drive the oxidation of CH₄ (Gupta et al., 2013; Segarra et al., 2013), but the electron acceptor likely varies in freshwater vs. saline wetlands, organic vs. mineral soils, and oligotrophic vs. eutrophic sites, as is the case for terminal metabolism (see Anaerobic metabolism in Section 3.3.2).

       Nitrous Oxide (N2O)

      Although this chapter is focused on carbon, we will briefly discuss the emissions of N₂O. Recent global wetland emissions of N₂O are “negligible” (Anderson et al., 2010), but management activities and environmental changes have the potential to increase emissions of this powerful greenhouse gas. The production of N₂O, which is a byproduct of both denitrification and nitrification, is largely controlled by nitrogen availability and soil redox status (Davidson et al., 2000). Nitrous oxide emissions are greatly enhanced in wetlands exposed to high nutrient loading (Hefting et al., 2003; Moseman‐Valtierra et al., 2011) and inversely related to soil C:N ratios (Klemedtsson et al., 2005). Further, peatlands that experience drought or anthropogenic lowering of the water table have higher N₂O emissions than those with a high water table (Pärn et al., 2018; Prananto et al., 2020). The production of N₂O is also affected by the availability of electron acceptors and electron donors, concentrations of hydrogen sulfide, temperature, and pH (Cornwell et al., 1999; Joye & Hollibaugh, 1995; Megonigal et al., 2004; Pärn et al., 2018; Parton et al., 1996).

       Emission Pathways

      There are three major pathways by which gases produced in wetland soils can be emitted to the atmosphere: diffusion, transport through plants, and ebullition. The rate of diffusion of gases out of a wetland soil is a function of the concentration gradient between soil pore spaces and the overlying water column or atmosphere, the wetness of the soil, and the amount of atmospheric/water column turbulence (Lai, 2009; Le Mer & Roger, 2001). Because molecular diffusion is a relatively slow process, rates of CH₄ oxidation can be more important when diffusion is the major route of export from the wetland (Bridgham et al., 2013). However, while a low water table increases the distance CH₄ has to diffuse through oxidized soils and therefore provides more opportunities for the aerobic oxidation of CH₄ (Roslev & King, 1996), this can occur at the radiative expense of higher rates of N₂O production (Pärn et al., 2018).

The aerenchyma tissues that allow vascular wetland plants to transport O₂ to their roots permit gases produced in soils to be efficiently vented through plants by passive diffusion or (faster) convective gas flows (Colmer, 2003). Gas transport through both herbaceous and woody plants can account for a substantial portion of total wetland CH₄ emissions (Covey & Megonigal, 2019; Gauci et al., 2010; Neubauer et al., 2000; Pangala et al., 2017; Whiting & Chanton, 1992). Methane that is transported through plants spends less time in oxidized surface soils and therefore is less susceptible to being oxidized to CO₂ (Joabsson et al., 1999), although CH₄ oxidation can be enhanced in the rhizosphere due to root O₂ loss (van Bodegom et al., 2001). There is a temporal coupling between CH₄ production and emission in vegetated wetlands, but this relationship breaks down in unvegetated sediments because the lack of vegetation reduces CH₄ emissions and promotes transient CH₄ storage (Reid et al., 2013) that leads to enhanced ebullition.

      Ebullition (bubbling) occurs when the local hydrostatic pressure decreases due to changes in temperature, air pressure, and water levels (Chanton et al., 1989; Männistö et al., 2019; Tokida et al., 2007), allowing gas bubbles to rise. As with plant‐mediated gas transport, the rapid vertical movement of gas bubbles allows CH₄ to quickly transit active CH₄ oxidation regions (Lai, 2009). Rates of ebullition are spatially patchy and temporally variable but can be the major