Keywords

Introduction

The United Nations Framework Convention on Climate Change defines climate change as “a change of climate which is attributed directly or indirectly to human activity that alters the composition of the global atmosphere and which is in addition to natural climate variability observed over comparable time periods.” Human alteration of atmospheric composition is expected to lead to an increase of global surface temperatures for the period 2090–2099 as compared to the period 1980–1999 by 1.8 °C to 4.0 °C (uncertainty range 1.1–6.4 °C) depending on the underlying scenario. The temperature increase is expected to be most pronounced over many land areas and be accompanied by an increase in global mean precipitation, indicating an intensification of the hydrological cycle. The spatial variability of precipitation will generally increase, with decreased rainfall in the subtropics and an increase of rainfall at higher latitudes and in parts of the tropics. These predicted changes in climate are likely to affect plant and microbial processes driving carbon (C) and nitrogen (N)turnover and cross-compartment exchange (hydrosphere-atmosphere-biosphere) of nutrients and trace gases at terrestrial ecosystem and landscape levels. In addition, the functioning of C and N cycling at ecosystem, landscape, and global scales is directly and indirectly affected by changes in concentrations of atmospheric substances such as CO2 and O3 and the human perturbation of the regional and global N cycling, which is resulting in increased inputs of reactive N to natural and semi-natural ecosystems (Erisman et al. 2008).

This chapter evaluates how climate change may influence microbial production and consumption of the atmospheric trace gases nitrous oxide (N2O) and methane (CH4) in soils. Methane contributes 18.1 %, and N2O contributes 6.24 % to the overall global radiative forcing (Forster et al. 2007). About 2/3 of all sources for both greenhouse gases (GHGs) are associated with microbial production processes in soils (Conrad 2009), with fluxes at the soil surface being the net product of simultaneous occurring production and consumption processes. Therefore, soils and microbial processes in soils are of outstanding importance for regulating atmospheric concentration of CH4 and N2O. As all biological processes are affected by climate conditions, climate change is likely to affect soil microbial processes and associated GHG production and consumption processes in soils too. Hence, the explicit understanding of the feedbacks between temperature and precipitation and soil microbial processes involved in N2O and CH4 exchange is an indispensable prerequisite to better predict how climate change will alter the source (and sink) strength of soils for atmospheric N2O and CH4. Consequently, such mechanistic understanding between environmental drivers and soil microbial processes is required to develop mitigation and adaption strategies targeted at a minimization of the source strength or a maximization of the sink strength of soils for CH4 and N2O.

Climate Change and CH4 Exchange

Methane is predominantly produced in anaerobic sites of soils and sediments by methanogenic bacteria as a final step of the anaerobic decomposition of organic matter. CH4 produced in soils may be emitted to the atmosphere by several pathways (diffusion, bubble, or plant-mediated transport) with CH4 getting potentially oxidized while passing by sites dominated by CH4-oxidizing microbial communities. These communities are mostly using O2 – but under certain circumstances also sulfate or nitrate – as electron acceptors (Conrad 2009). Oxidation of CH4 in soil, sediments, and even in the aerenchyma cavities of vascular plants is a significant process. In rice paddies, 10–30 % of the produced CH4 may get oxidized before emission (Conrad 2009). High-affinity methanotrophic bacteria are capable to gain energy from CH4 concentrations <1.7 ppmv in air. Methanotrophic bacteria inhabits predominantly oxic upland soils. Though net CH4 uptake rates are rather low (approx. 0.1–5 kg CH4-C ha−1 year−1) (Dutour and Verchot 2008), at a global scale upland soils are representing a significant sink (approx. 10 %) within the global budget of atmospheric CH4.

Climate change interacts in several ways with CH4 production and consumption processes in soils (Figs. 38.1 and 38.2). On the one hand, climate changes will directly affect soil environmental conditions, namely, moisture and temperature, and by this the balance of oxidative to reductive processes. For example, temperature increases will – as far as water availability is not limiting – likely result in an increase in respiration, thus decreasing O2 availability and the CH4 oxidizing capacity of upland soils. On the other hand, global change and thus also changes in atmospheric [CO2] will affect plant biomass production, the ratio of aboveground to belowground biomass production, root exudation, and litter quality. All these changes will finally affect ecosystem CH4 exchange, with results being different across different ecosystem types and climate zones. Finally, climate change will also affect regional water balances and thus landscape groundwater levels. This will ultimately control the expansion of wetlands and emission magnitudes of CH4 at landscape scales (Figs. 38.1 and 38.2). For example, Liu et al. (2009) demonstrated for landscapes in Inner Mongolia that CH4 emissions from riparian areas outweigh CH4 uptake by adjacent upland steppe soils though the riparian areas contributed only 1.5 % to the total area. Comparable landscape effects of climate change are also expected with regard to the thawing of permafrost areas in northern latitudes, which may not only go along with the release of CH4 captured in so far permanently frozen soils but also with the generation of new wetlands and resulting net emissions of CH4 from these areas following the anaerobic decomposition of inundated organic matter.

Fig. 38.1
figure 1

Conceptual scheme of global change-induced effects on CH4 production by methanogens in soils and sediments. The figure is composed by (a) an inner ring divided into two sectors representing two major parameters controlling methanogenesis (anaerobic conditions and C availability) which may be affected by climate change factors (temperature increase, atmospheric [CO2] increase, precipitation increase, and increase in precipitation variability (droughts/flooding)). Here, the sector size represents an approximate of the importance of the relevant climate change factor for overall methanogenesis. (b) The second ring shows the effects of changes in environmental conditions on ecosystem processes (e.g., root exudation, evapotranspiration) while (c) the outer ring represents alterations driven by landscape level and microbial community changes, which will likely affect the drivers for methanogenesis, too. Dotted circumferences are representing a response function of 1. Therefore, a sector of the circle above the dash-lined inner circumference indicates a positive effect of each expected change on the respective controlling process; a sector below the circumference indicates a negative effect on the respective controlling process. The overall size of each circle sector is calculated by summing up the contribution of the single effects driven by each global change-expected variation. For example, temperature increase will on the one hand enhance soil anaerobiosis due to the stimulation of respiratory processes but on the other hand also increase evapotranspiration and, thus, decrease soil moisture and anaerobiosis. This is indicated by expanding the sector for the soil respiration effect beyond the dotted ring, while it is opposite for the evapotranspiration effect. Shaded areas are indicating uncertainties

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Fig. 38.2
figure 2

Conceptual scheme of global change-induced effects on CH4 consumptions by methanotrophic bacteria in soils. For a further explanation of interactions between climate change factors (inner circle) and controllers of methanotrophy in soils (diffusion of O2/CH4, NH4 competition effects) which can be depicted from the graph, see explanations for Fig. 38.1

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Emission of N2O from Soils and Effects of Climate Change

Nitrous oxide production and consumption in soils is mainly performed by a series of microbial N turnover processes, which can either be oxidative or reductive. Oxidative processes involved in N2O production, i.e., autotrophic and heterotrophic nitrification, use reduced N-forms such as ammonia (NH3) or organic N-forms. Both autotrophic and heterotrophic nitrification are carried out by various groups of bacteria and archaea (Hayatsu et al. 2008). Nitrifiers oxidize NH3 or organic N compounds to nitrite and nitrate, thereby obligatorily involving molecular O2 and releasing N2O as a trace by-product. Oxidized inorganic N substances such as nitrate, nitrite but also nitric oxide, and even N2O can be used, e.g., by denitrifying microorganisms (bacteria, archaea, fungi) as electron acceptors if anaerobiosis occurs. Furthermore, nitrate and/or nitrite may be used by fermentative bacteria in the process of dissimilatory nitrate reduction to ammonium (DNRA) to regain ammonium (NH4 +) in strictly anaerobic environments. DNRA has been shown to occur predominantly in anaerobic sludge and sediments and is considered as a potential important process for securing ecosystem N retention (Silver et al. 2001). Since both DNRA and denitrification appear to be favored by similar soil conditions (low redox potential, high nitrate (NO3 −), and labile C availability) (Butterbach-Bahl et al. 2011), global change may affect both processes in the same way. As for CH4, N2O produced in the soil is not necessarily emitted but can undergo microbial consumption before reaching the soil-atmosphere interface.

There is an increasing number of studies showing net uptake of atmospheric N2O by soils or consumption of N2O in soil layers during its diffusion from production sites to the soil surface. Nitrous oxide consumption is likely associated to the activity of the N2O reductase, the final enzyme in the denitrification chain. N2O removal can be rather effective: up to 2/3 or even 90 % (Vieten et al. 2009) of the produced N2O may get further reduced to N2. N2O reduction does not only occur in anaerobic but also in predominantly aerobic soils. Goldberg and Gebauer (2009) determined isotopic (15N/14N) signatures of N2O along a soil profile of a temperate forest and simultaneous N2O flux measurements at the soil-atmosphere interface. Their data indicate that soil N2O consumption may be less affected by prolonged drought as compared to subsoil N2O production. This may be explained by reduced nitrate but improved N2O diffusion to anaerobic microsites, thus, providing a concept why in drought affected soils atmospheric N2O consumption may prevail.

The mentioned oxidative and reductive processes involved in N2O production and consumption in soils are complex and closely interwoven with ecosystem and landscape-scale N cycling (Figs. 38.3 and 38.4). For example, plant-microbe competition for mineral and organic N in soils, volatilization, and redeposition of NH3 at landscape scales or lateral transport of nutrients by leaching or erosion/deposition processes need to be considered for predicting how changes in climate will finally feedback on regional soil N2O emissions (Butterbach-Bahl and Dannenmann 2011) (Figs. 38.3 and 38.4). These indirect effects may override direct climate change effects such as the immediate response of nitrification and denitrification to changes in soil temperature and moisture. However, even the understanding of direct temperature effects on microbial production and consumption processes in soils is limited due to our still hampered ability to unravel which specific processes or microbial communities are driving N2O exchange at the soil-atmosphere interface. For example, reported Q10 values for denitrification – a measure of the rate change upon temperature increases – widely vary from 2 to 10 (Abdalla et al. 2009). Such Q10 values of high variability and magnitude indicate that the temperature response of denitrification is not solely enzymatic but that temperature affects the delicate balance of oxidative and reductive processes in soils and, thus, also the relative importance of nitrification versus denitrification. Therefore, the temperature response of soil N2O emissions has been attributed to increases in the anaerobic volume fraction, brought about by an increased respiratory sink for O2. Also spring-thaw N2O pulse emissions, which in some ecosystems (e.g., steppe, temperate forest, and temperate agricultural land) may dominate annual N2O fluxes, are largely decoupled from long-term temperature changes but strongly depending on the severeness and frequency of frost period and the availability of soil water in the winter and spring period (Wolf et al. 2010). The situation is even further complicated by likely shifts in the product ratios of denitrification (N2O:N2) due to changes in soil environmental conditions. Predicting the response of European forest soil N2O and N2 emissions to future climate conditions (2031–2039) compared to present-day climate (1991–2000) by use of the biogeochemical model PnET-N-DNDC revealed a decrease in N2O emissions by 6 % (Kesik et al. 2006). This decrease in N2O emissions was mainly due to a shift in the N2O:N2 ratio driven by enhanced denitrification. The favoring of N2 as end product of denitrification at elevated temperatures can be explained by the decreasing availability of electron acceptors with increasing anaerobiosis forcing denitrifiers to be more resource efficient and, thus, to express the full chain of denitrification enzymes (Butterbach-Bahl and Dannenmann 2011). It is worthwhile to notice that also further increases in atmospheric [CO2] will feedback on soil N2O emissions (Figs. 38.3 and 38.4) due to both increased soil moisture following improved water use efficiency of plants at higher [CO2] and increased root exudation alleviating potential C substrate limitation of denitrification (Butterbach-Bahl and Dannenmann 2011).

Fig. 38.3
figure 3

Conceptual scheme of global change-induced effects on N2O production via nitrification. See Figs. 38.1 and 38.2 for explanations on how to interpret the graph

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Fig. 38.4
figure 4

Conceptual scheme of global change-induced effects on N2O production via denitrification. See Figs. 38.1 and 38.2 for explanations on how to interpret the graph

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These few examples, and the complexity of interacting processes as shown in Figs. 38.3 and 38.4, demonstrate that a better prediction of how climate change will affect soil N2O emissions will require both a detailed process understanding and a specific assessment at regional scale. The latter is of outstanding importance, since at the landscape level not only N use is affecting N availability in soils but also changes in hydrological fluxes, land management, and land use will finally drive the response of the soil microbial community to climate changes and offer possibility for intervention to mitigate soil N2O emissions in a changing climate.

Conclusions

It is obvious that we are still not capable to predict how climate change will finally affect the emission and the deposition of the greenhouse gases CH4 and N2O from/to soils. The reason is not only a partially missing process understanding but even more our still restricted ability to relate surface fluxes at the soil-atmosphere interface to underlying microbial, physicochemical, and plant processes in the soil. To unravel complexity, controlled studies under laboratory conditions have helped to improve our understanding. Nevertheless, the complex C, N, and water interactions at ecosystem and landscape levels are still not sufficiently understood to allow for generalization and to predict climate change effects on CH4 and N2O exchange. Furthermore, regional drivers of greenhouse gas exchange such as regional hydrology or the dispersion of nutrients due to volatilization and redeposition, lateral transport of nutrients by soil water, and surface erosion have been neglected in most studies, though the mentioned processes will as well be affected by climate change. Therefore, targeted regional climate change studies are needed to overcome uncertainties, to improve process understanding, and to better predict climate change feedbacks of ecosystem processes at site, regional, and global levels.