Wetlands as Carbon Sources

3.3.1 Wetlands as Carbon Sources

The important greenhouse gases carbon dioxide, methane, and nitrous oxide can

be released from natural and constructed wetlands (Le Mer and Roger 2001; Whit- ing and Chanton 2001; Malmer et al. 2005; Liikanen et al. 2006; Mander et al. 2008). Processes such as denitrification and methane production are dependent on the oxygen status of soil and sediment. Anoxic soils and sediments produce meth- ane, while well-drained soils act as a sink for atmospheric methane due to methane oxidation (Hanson and Hanson 1996).

The water table level of wetlands influences not only the amount of methane emitted to the atmosphere but also the removal of methane from the atmosphere. For example, Harris et al. (1982) determined that peat from the Great Dismal Swamp contributes to the removal of atmospheric methane when the water table level is below the surface of peat during dry periods. In contrast, when peat is well saturated with water, it becomes an important methane source. Furthermore, Au- gustin et al. (1998) concluded that lowering the groundwater table of minerotro- phic fens in the northeast of Germany increased the release of nitrous oxide and the reduction of methane. Relatively high methane emissions could be observed when the groundwater table was high and soil temperatures were higher than 12°C. They

also point out that these fens release approx. between 0.6 and 9.0 mg CH 4 –C/m 2 /h. Natural wetlands emit approx. 1.45 × 10 11 kg CH 4 –C/a to the atmosphere. This equates to about 25% of the total emissions from all anthropogenic and natural sources. Wetland methane flux rates are commonly 10 –6 kg/m 2 /d and represent the net effects of microbial production and consumption (Whalen 2005).

The studies carried out so far have demonstrated that constructed and restored wetlands also have high nitrous oxide (Fey et al. 1999; Xue et al. 1999; Johansson et al. 2003; Mander et al. 2005; Stadmark and Leonardson 2005; Teiter and Man-

3.3 Are Wetlands Carbon Sources or Sinks? 133

der 2005; Liikanen et al. 2006; Picek et al. 2007) and methane (Schipper and Reddy 1994; Tanner et al. 1997; Cao et al. 1998; Tuittila et al. 2000; Johansson et al. 2004; Mander et al. 2005; Teiter and Mander 2005; Altor and Mitsch 2006; Liikanen et al. 2006; Picek et al. 2007) emissions. However, data are highly vari- able due to different designs and operations, and more importantly as a result of system locations in different climates. The methane emissions from vegetated constructed treatment wetlands can be similar to those from productive natural wetlands. Methanogenesis can be an important decomposition process in con- structed wetlands treating organic wastewaters. For example, a flux of approx.

between 28 and 278 mg CH –C/m 2 4 /h has been reported for pilot-scale constructed wetlands (Tanner et al. 1997).

Brix et al. (2001) reported that in wetlands dominated by Phragmites australis, the primary productivity is high, and approx. 50% of the net primary production is respired to carbon dioxide and methane in the sediment. In the growing season, the process of methanogenesis is primarily limited by organic matter availability, while at other times temperature is the most important factor.

Like natural wetlands, rice paddies have been identified as one of the important sources of atmospheric methane. The majority of studies reporting on methane emissions from wetlands have been conducted in natural ecosystems or heavily managed rice paddies (Cao et al. 1998; Crutzen 1995). Cao et al. (1998) estimated

that the global annual methane emission from wetlands is 1.45 × 10 11 kg, of which

0.92 × 10 11 kg comes from natural wetlands and 0.53 × 10 11 kg from rice paddies.

A limited number of studies consider carbon dioxide fluxes from constructed wetlands (Mander et al. 2005; Liikanen et al. 2006). Many authors report that carbon dioxide emissions increase with increasing temperature and are higher under drained than flooded conditions (Bridgham and Richardson 1992; Moore and Dalva 1993, 1997; Price and Waddington 2000; Scanlon and Moore 2000; Waddington et al. 2001).

Deep wetlands generally capture carbon dioxide from and release methane into the atmosphere (Whiting and Chanton 2001). The combination of these two fluxes determines whether these countervailing processes make a wetland system an overall contributor to the greenhouse effect. The ratio of methane release to carbon dioxide consumption determines the carbon exchange balance with the atmosphere for any wetland ecosystem (Kayranli et al. 2010).

A better understanding of the critical processes regulating greenhouse gases as- sociated with wetlands such as freezing–thawing cycles and pulsing hydrological regimes are important for assessing carbon dioxide, methane, and nitrous oxide fluxes. The production and consumption of greenhouse gases are partly regulated by microbial processes, which in turn are influenced by soil moisture and tempera- ture. Nitrification and denitrification are the key processes that produce nitrous oxide. However, nitrous oxide production in frozen soils is more likely to be regu- lated by denitrification (Mørkved et al. 2006; Öquist et al. 2007). Van Bochove et al. (2001) highlighted that nitrous oxide fluxes are high in winter because of the sudden release of stored nitrous oxide. Maljanen et al. (2007) reported that nitrous oxide and carbon dioxide accumulated in the soil during winter and were released

134 3 Carbon Storage and Fluxes Within Wetland Systems

swiftly during thawing in spring. During winter, methane concentrations in the soil remained lower than in the atmosphere and subsequently increased as tempera- tures increased after thawing.

Zhang et al. (2005) observed that during thawing, methane and carbon dioxide emissions increased rapidly (4.5 to 6 times the winter emissions) for continuously flooded and seasonally flooded marshes. They estimated that a continuously flooded and a seasonally flooded wetland in Sanjiang (northeast China) released

0.5 ± 0.19 and 0.18 ± 0.15 mg CH 4 –C/m 2 /h methane, respectively. In comparison, naturally flooded forests and floating grass mats in Brazil (Amazon floodplain) emitted between 8 and 92 mg CH 4 –C/m 2 /h into the atmosphere (Bartlett et al. 1988). Environmental parameters such as temperature, pH, depth of water table, plant- ing regime (Waddington et al. 1996; Schlesinger 1997; Trettin and Jurgensen 2003; Whalen 2005; Inamori et al. 2007; Picek et al. 2007; Knoblauch et al. 2008), substrate type and quality (Bellisario et al. 1999; Joabsson et al. 1999; Ström et al. 2003), and specialized microbes (Fischer and Pusch 1999; Whalen 2005; Buesing and Gessner 2006; Picek et al. 2007; Sleytr et al. 2007; Ström and Christensen 2007; Tietz et al. 2008) impact on gas production and, ultimately, net methane emission rates. Furthermore, the methane exchange between wetland ecosystems and the atmosphere can be affected by the presence of plants because the convective flow process in plants facilitates a faster diffusion of gases through water, and particularly by the species composition of vascular plants. These plants affect important aspects of methane dynamics such as production, consumption, and transport; for example, the root exudates are decomposed by microbes and transformed into methane and carbon dioxide (Zhu and Sikora 1995; Joabsson and Christensen 2001; Tanner 2001; Picek et al. 2007; Ström and Christensen 2007).

Walter and Heimann (2000) emphasize that most wetland plants root below the water table and that methane flux from the soil to the atmosphere takes place via aerenchyma in the vascular tissue of the plants. Additional transport pathways are diffusion and bubble generation. Above the water table, methane is oxidized in the soil pores by methanotrophic bacteria.

Landry et al. (2009) claim that constructed wetlands emit between 2 and 10 times more greenhouse gases than natural wetlands. This is likely due to high loading rates. They observed that methane was the most important greenhouse gas in unplanted wetland systems and that the presence of plants decreased methane fluxes but favored carbon dioxide production.

Alford et al. (1997) estimated fluxes of between approx. 6 and 38 mg

CH4−C/m 2 /h for swamp forests and marshes near New Orleans, LA, USA. Barlett and Harris (1993) reported fluxes of approx. 4 mg CH4–C/m 2 /h for forested swamps and marshes. Kang and Freeman (2002) reported that bog and forested swamps in North Wales (UK) emit up to approx. 3 mg CH4–C/m 2 /h into the atmo- sphere. The relatively high data variability is likely due to different climatic regions. Hou et al. (2000) pointed out that the reduction of various oxidants in homoge- neous soil suspensions occurs sequentially at corresponding soil redox potentials. The availability of soil oxidants such as oxygen and carbon dioxide used as elec-

3.3 Are Wetlands Carbon Sources or Sinks? 135

tron acceptors for organic matter degradation contributes significantly to soil microbiological processes. They found that emissions of methane were strongly correlated with changes in the soil redox potential. Significant methane emissions occurred only at soil redox potentials, which were lower than approx. −100 mV.

It is the water table level that largely determines the presence of aerobic and anaerobic conditions occurring at different depths of wetlands. These conditions control the methanogenic and methanotrophic processes (Kelley et al. 1995). Me- thanogenesis is a rigid anaerobic process and is evoked during flooding periods, when the water table level rises. In contrast, with a decrease in flooding periods, methane production decreases. An inverse relation is observed for methane oxida- tion (Kayranli et al. 2010).

Grünfeld and Brix (1999) compared vegetated organic sediments at different water table depths below the surface with vegetated inundated sediments. They found that due to the high water-holding capacity of organic sediments, rates of methanogenesis and methane emission in organic sediments with a water table of

8 cm below the sediment surface were only slightly, but not statistically signifi- cantly, different from rates in inundated sediments. Sandy sediments with water tables of 8 cm below the sediment surface had very low methanogenic activity as compared with organic sediments.

Methane can be transported to the atmosphere through pathways such as mo- lecular diffusion, gas bubbling up (ebullition) from the sediments, and vascular plant stems (Walter and Heimann 2000). King (1996) pointed out that the amount of methane oxidized did not correlate with the total potential methane fluxes from

a wetland. Oxygen distribution and availability controls the rates of methane oxi- dation within wetlands. Moreover, oxygen penetration within peat varies from 1 to

7 mm, with some diurnal variation coupled to benthic photosynthesis.

Moore and Dalva (1993) and Moore and Roulet (1993) reported that the mean position of the water table level is the best indicator of methane emissions. Appar- ently, a critical depth exists at which maximal emissions occur. It has been deter- mined that a water table depth greater than 18 cm does not produce high emis- sions, since methane production (methanogenesis) decreases and its consumption increases (methanotrophy). However, when the depth of the water table is 12 cm below the surface of peat, or exceeds it, methane fluxes are high. Bubier et al. (1993) and Daulat and Clymo (1998) estimated that methane emitted into the atmosphere from experimental digs was between 5 and 60 times higher than that produced in hillocks (small hills or mounds) due to the digs’ having a lower water table depth than the hillocks. Roulet et al. (1993) discovered that peatlands are converted from a source into a sink of methane when the water table drops to

25 cm below the peat surface due to increased methane oxidation. Kelley et al. (1995) studied methane emissions across a tidally flooded riverbank in North Carolina, USA. Their study showed the highest methane fluxes when the water level was close to the surface and the lowest fluxes at both high and low water table levels. Similarly, Smith et al. (2000) estimated that methane emissions stopped when the soil moisture content fell below approx. 25%, as floodwaters receded in Venezuela’s Orinoco River floodplain.

136 3 Carbon Storage and Fluxes Within Wetland Systems

Glatze et al. (2004) pointed out that the highest rates of anaerobic methane pro- duction can be measured for samples close to the soil surface with fresh peat ac- cumulation and a high water table. In contrast, the lowest rates were observed for samples from the sub-surface of sites with a low water table. Anaerobic methane production was significantly positively correlated with aerobic and anaerobic carbon dioxide production. These production potentials show that drainage (Salm et al. 2009), harvesting, and restoration change the ability of the peat profile to produce and emit carbon dioxide and methane.