Next Article in Journal
A One Health Comparative Assessment of Antimicrobial Resistance in Generic and Extended-Spectrum Cephalosporin-Resistant Escherichia coli from Beef Production, Sewage and Clinical Settings
Next Article in Special Issue
Complete Genome of Isoprene Degrading Nocardioides sp. WS12
Previous Article in Journal
Thermotolerant and Thermophilic Mycobiota in Different Steps of Compost Maturation
Previous Article in Special Issue
Isoprene Oxidation by the Gram-Negative Model bacterium Variovorax sp. WS11
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:

Methane Production in Soil Environments—Anaerobic Biogeochemistry and Microbial Life between Flooding and Desiccation

Max-Planck-Institute for Terrestrial Microbiology, Karl-von-Frisch-Str. 10, 35043 Marburg, Germany
Microorganisms 2020, 8(6), 881;
Original submission received: 27 May 2020 / Revised: 8 June 2020 / Accepted: 9 June 2020 / Published: 11 June 2020
(This article belongs to the Special Issue Microbial Cycling of Atmospheric Trace Gases)


Flooding and desiccation of soil environments mainly affect the availability of water and oxygen. While water is necessary for all life, oxygen is required for aerobic microorganisms. In the absence of O2, anaerobic processes such as CH4 production prevail. There is a substantial theoretical knowledge of the biogeochemistry and microbiology of processes in the absence of O2. Noteworthy are processes involved in the sequential degradation of organic matter coupled with the sequential reduction of electron acceptors, and, finally, the formation of CH4. These processes follow basic thermodynamic and kinetic principles, but also require the presence of microorganisms as catalysts. Meanwhile, there is a lot of empirical data that combines the observation of process function with the structure of microbial communities. While most of these observations confirmed existing theoretical knowledge, some resulted in new information. One important example was the observation that methanogens, which have been believed to be strictly anaerobic, can tolerate O2 to quite some extent and thus survive desiccation of flooded soil environments amazingly well. Another example is the strong indication of the importance of redox-active soil organic carbon compounds, which may affect the rates and pathways of CH4 production. It is noteworthy that drainage and aeration turns flooded soils, not generally, into sinks for atmospheric CH4, probably due to the peculiarities of the resident methanotrophic bacteria.

1. Theoretical Background

Environmental methanogenesis is the degradation of organic matter under anaerobic conditions to the gaseous products CH4 and CO2. The methanogenic pathway is catalyzed by a complex microbial community basically consisting of fermenting bacteria and methanogenic archaea. The entire process is anaerobic, since it only takes place in the absence of O2 and other oxidants (reducible inorganic electron acceptors). Therefore, it is not surprising that most reports of environmental CH4 production refer to anoxic environments, such as animal gut systems, aquatic sediments, flooded soils, peatlands and coastal wetlands [1,2,3,4,5,6,7,8]. More recently, however, it was found that the potential to produce CH4 under anaerobic conditions is widespread among various soil ecosystems, including non-flooded oxic upland soils [9]. Therefore, it is worthwhile to examine the consistency of our theoretical background with the occurrence of methanogenic processes.
Our understanding of environmental methanogenesis in anoxic environments is based on physicochemical and microbiological theories, of which the most relevant are the following. Chemically, complete degradation of organic matter to CO2 is usually an oxidation reaction, which primarily uses O2. The O2 required by the degradation process is supplied by diffusion from the atmosphere into the soil. The diffusion coefficient (D) in air is about four orders of magnitude larger (about 2 × 10−1 cm2 s−1) than that in soil water (about 2 × 10−5 cm2 s−1) [10]. Diffusion through x = 1 cm of water requires about t = 7 h, and through 10 cm of water even requires 70 h (calculated from t = x2/[2 D]). Therefore, it is plausible that oxidation processes in flooded soils and other wetlands become rapidly limited by O2. Once O2 is depleted, other oxidized inorganic compounds, such as nitrate, ferric iron, or sulfate can serve as oxidants. The propensity for using a particular oxidant depends on the redox potential of the electron-accepting reaction; i.e., degradation processes with the most negative Gibbs free energy will dominate [11,12]. The actual Gibbs free energy (ΔG) is the result of the ΔG° under standard conditions and the concentrations of the reactants and products involved. As a consequence, oxidation of organic matter usually results in the sequential reduction of available oxidants. In soils, the most common oxidants are O2, nitrate, ferric iron, and sulfate, which undergo sequential reduction along with the oxidative degradation of organic matter. Once inorganic oxidants are no longer available, however, anaerobic degradation of organic matter can only proceed by disproportionation resulting in the production of CO2 (oxidized compound) and CH4 (reduced compound). According to this theoretical picture, organic matter degradation to CH4 should not start before all the available inorganic electron acceptors have been depleted to such an extent that the ΔG of the oxidation reactions is less negative than the ΔG of the CH4-producing disproportionation reaction. Disproportionation of polymeric organic matter, e.g., cellulose, results in the formation of equal amounts of CO2 and CH4. Stoichiometry of methanogenic degradation of cellulose, or other organic compounds with a carbon oxidation state of zero, also constrains the pathways of CH4 formation resulting in >2/3 of CH4 being produced from acetate and <1/3 being produced from H2/CO2 [13,14,15].
The microbiological theory adds knowledge about the activity of individual microorganisms, which all together make up the environmental microbial community. Of course, the activity of microorganisms cannot overcome the laws of physics or chemistry, but the microbial enzyme machinery provides important catalysts without which most of the chemical processes would proceed very slowly or not at all. Therefore, microorganisms with particular catalytic functions must be present in sufficient numbers to allow a process to proceed. If such microorganisms are absent or inactive, processes may fail to proceed even when thermodynamic conditions are favorable. In soils with the potential for methanogenesis, important microbial functions include hydrolysis of complex organic matter, various fermentation processes, methanogenic processes and processes involved in the sequential reduction of inorganic compounds. Hydrolytic and fermenting bacteria degrade complex organic matter to H2, CO2 and simple organic compounds, which are subsequently further fermented to acetate, H2 and CO2. These compounds are the substrates for methanogenic archaea, which convert them to CH4 and CO2 [3,14,16,17]. Methyl compounds (e.g., methanol, trimethylamine, dimethylsulfide) are also possible substrates for methylotrophic methanogenesis, for example during degradation of pectin-producing methanol or in saline environments where methyl compounds are produced from osmolytes [14].
A sufficiently high water potential is a precondition for all microbial activity, which usually strongly decreases below −6 bar, although potential for growth of particular bacteria may be maintained down to −70 bar [18,19]. The decrease in water potential during the desiccation process can initiate the formation of microbial dormancy, including resting stages such as spores and cysts [20]. Dormancy will persist until the water potential becomes permissive again. Since many microorganisms, e.g., methanogenic archaea, do not form resting stages, it is unclear to which extent they can switch into dormancy and survive detrimental conditions. Microorganisms catalyzing the methanogenic degradation of organic matter are believed to be facultatively or obligately anaerobic, meaning that growth and/or activity of these microorganisms requires the absence of O2 and sometimes even a low (<−330 mV) redox potential [15]. Methanogenic archaea in particular are unable to proliferate in the presence of O2, and their activity is strongly inhibited [3].
Another important theoretical microbiological background concerns the concentrations of the growth substrates required for energy generation. Both the catalytic activities and the growth rates of microbial populations depend on substrate concentrations. The relationship is usually a hyperbolic one, described by the Michaelis–Menten and the Monod equations for catalytic activity and for population growth, respectively [21,22]. Both equations are parametrized with respect to their maximum rates (Vmax, µmax, respectively) and the substrate concentrations (Km, KS, respectively) at which the rates are half-maximal. The lower threshold (T) for utilization of the methanogenic substrate is another important parameter [23,24]. As a consequence, activity and growth of microorganisms depend on the concentrations of their substrates (both electron donors and acceptors), for which they may compete. The microorganisms with the better catalytic parameters (e.g., µmax, KS, T) will win the competition over a limiting substrate and, for example, outgrow a competitor. Please note that the result is comparable (but not identical) to that caused by chemical competition according to the ΔG values that are also dependent on the concentrations of substrates (but also on those of the products) according to Nernst law. For example, different hydrogenotrophic methanogenic species have different characteristic kinetic parameters, in particular thresholds for H2 consumption [25,26,27,28]. There are two groups of methanogens distinguished by their threshold concentrations for H2 that operate with different modifications of the CO2 reduction pathway [29]. There are also two groups of acetotrophic methanogens (Methanothrix spp. versus Methanosarcina spp.) having different enzyme systems for acetate activation and which consequently dominate at different acetate concentrations in the environment [30,31,32].

2. Flooding and Desiccation

Flooding and desiccation are key events for methanogenic processes in wetlands, which include peatlands [33], coastal wetlands [8] and flooded soils. In the following, the review will largely focus on flooded mineral soils, excluding peat soils and marine environments. Methanogenesis can only happen in the phase after flooding and before drainage. A standard model is rice field soils, which are seasonally submerged. Rice fields can undergo this cycle of flooding and desiccation once or twice, in some areas even three times a year [34]. However, there can also be multiple cropping, where cultivation of flooded rice is regularly rotated with cultivation of upland crops, such as maize, wheat or soya [35]. Hence, the time periods between upland conditions, which are more or less desiccated, and flooded conditions may vary between months and years. There are also natural areas, which undergo seasonal or irregular flooding, e.g., the flood plain of the Amazon River [36,37]. Moreover, there are lake sediments that are permanently flooded and, therefore, can represent a source for atmospheric CH4 [38]. On the other hand, there are so-called upland soils that are almost never flooded. At the dry extreme, there are desert soils with a low soil moisture content due to rare precipitation [39]. Theoretically, it is expected that methanogenic degradation of organic matter only operates under flooded conditions, when O2 and potential inorganic electron acceptors are all reduced. From a physicochemical point of view, this expectation is straightforward. However, from a microbiological point of view, it cannot be predicted that all the necessary functional types of microorganisms are actually present in all the different soil environments. This is particularly true for methanogenic microbial communities, which are expected to consist of anaerobic microorganisms that are sensitive to O2 and, thus, should not survive desiccation events.
After 1990, molecular methods for characterization of the composition of microbial communities and the abundance of microbial populations were quite well established and became more and more sophisticated and widespread. It has since then been possible to investigate both the function and the structure of the methanogenic microbial communities in environmental samples. Hence, it has been possible to test how the methanogenic process in various environments functions on the basis of both abiotic and microbial variables. This research resulted largely in the confirmation of our theoretical background knowledge, but also in a number of unexpected observations that required modification of our perspectives, in particular concerning the O2 sensitivity of microorganisms and the role of soil organic carbon (SOC) compounds. These aspects will be covered in the next three sections (Section 3, Section 4 and Section 5).

3. Observations in Accordance with Theory

The physicochemical and microbiological theories outlined in Section 1 are based on background knowledge and have repeatedly been tested and challenged by hypothesis-driven biogeochemical and physiological experiments and observations relevant for environmental methanogenic processes. The most important ones, in particular in the context of flooded rice field soils, are listed without extensive discussion or literature review in the following. (1) Flooding of soil initiates sequential reduction processes, which largely proceed in the sequence of reduction of nitrate, ferric iron, and sulfate before CH4 production starts [40,41]. (2) Vigorous CH4 production is suppressed as long as concentrations of the methanogenic substrates (H2, acetate) are kept low by the presence of reducible inorganic compounds [42]. (3) This suppression is consistent with thermodynamic conditions being unfavorable or favorable for CH4 production [43]. (4) Aeration of methanogenic soil results in regeneration of ferric iron and sulfate, thus suppressing CH4 production until these compounds are reduced again [44,45]. (5) Addition of nitrate, ferric iron or sulfate to methanogenic soils results in suppression of CH4 production, mainly due to creating adverse thermodynamic conditions [46,47]. (6) Addition of nitrate (like the addition of O2) results in the partial regeneration of ferric iron and sulfate, prolonging suppression of CH4 even when nitrate has been completely reduced [47]. (7) Organic matter (e.g., polysaccharides) is degraded sequentially by hydrolysis, fermentation, and hydrogenotrophic plus aceticlastic methanogenesis [48].
Similarly, many observations supported the microbiological theories, e.g., observations made in flooded rice field soils. (1) The sequential methanogenic degradation of organic matter is paralleled by changes in the composition of the microbial community initiated upon depletion of O2 [49,50,51]. (2) The absence or inhibition of functional microbial groups results in suppression of CH4 production. For instance, the inhibition of methanogens by bromoethane sulfonate or chloroform, or the specific inhibition of acetotrophic methanogens by methyl fluoride, results in accumulation of the respective precursor metabolites and allows mass balance calculations for fermentation processes or stable isotope fractionation [52,53,54]. (3) Production of CH4 in flooded soils requires the presence of a sufficient number of the microbes involved. Even more importantly, it is necessary that the microbes are not only present but express sufficient activity [55,56,57,58]. (4) The microorganisms involved in CH4 production are more or less sensitive to O2. This observation, in particular, requires closer inspection.

4. Oxygen Sensitivity and Microbial Populations in the Soil Environment

The most important microbiology hypothesis probably concerns the relation of microorganisms to O2, which is highly relevant for anaerobic microbial life. This hypothesis states that anaerobic microorganisms become active in anoxic flooded soil and inactive upon desiccation. Furthermore, obligate anaerobes might be killed after exposure to O2. The removal and exposure of soil to O2 is probably a more important consequence of flooding and drainage, respectively, than the water potential of the soil. Some obligately anaerobic microorganisms are also able to form resting stages, e.g., the endospore-forming Firmicutes, which survive by entering dormancy. Methanogenic archaea, however, do not form resting stages and, therefore, they were believed to be intolerant to desiccation and O2 exposure [15], albeit this demand was qualified early on [3,59]. Indeed, desiccation and/or aeration of anoxic soil results in cessation of CH4 production, albeit the numbers of methanogenic archaea usually do not decline completely [60]. In general, most soils were found to contain low, but significant populations of methanogenic archaea, which start to proliferate when soil is flooded and O2 and other inorganic electron acceptors have been depleted [61,62]. This picture has repeatedly been reproduced, demonstrating that even desert soils contain low populations of methanogenic archaea [39,63]. Additionally, pure cultures of methanogenic species (e.g., Methanosarcina sp.) were found to be not completely killed by exposure to O2, although CH4 production is inhibited [60]. Later on, it was found that although methanogenic archaea cannot grow in the presence of O2, they are able to survive and even partially utilize O2 as an electron acceptor [64,65]. Methanogenic archaea are apparently able to assimilate carbon even when soils are aerated [66,67]. However, not all methanogenic archaea are able to tolerate O2. The tolerant ones can be distinguished from the intolerant ones based on their functional gene content, and are hierarchically clustered into two classes, II and I, respectively [68]. Hence, the microbiology theory concerning O2 tolerance has to be modified, i.e., that some taxa of obligately anaerobic methanogens can survive (maybe even exploit to some extent) oxic conditions.
In the recent years, our group investigated the structure and function of methanogenic microbial communities in different soil environments that are distinguished according to the frequency of flooding. We hypothesized that these environments have characteristic populations of O2-sensitive microorganisms, in particular fermenting bacteria and methanogenic archaea. Thus, environments with more frequent drainage should contain more O2-tolerant and/or desiccation-tolerant microorganisms, such as class-II methanogens and endospore-forming Firmicutes. We also hypothesized that these populations and their biochemical functions change upon desiccation and reflooding in different ways according to the category of the soil environment. Thus, we expected that more frequently drained soil environments exhibit smaller population changes than permanently flooded environments upon drainage or permanently dry environments upon flooding. We distinguished four categories: (1) permanently flooded lake sediments [69,70]; (2) annually flooded rice field soils or river floodplains [71,72,73]; (3) irregularly flooded soils (rice crop rotation) [74,75,76]; and (4) upland soils (including desert soils) [39,62,72,73,74,77]. We checked the potential of CH4 production, the abundance of bacterial and archaeal microbiota, and the community composition of methanogenic archaea and (fermenting) bacteria. In the following, the observations will briefly be reviewed to demonstrate to what extent they are consistent with our hypotheses (Table 1).
Virtually all the different environments exhibited a potential for CH4 production provided the soil was submerged. Furthermore, all the studies demonstrated that potential CH4 production ceased upon drainage and desiccation of the soil. However, whereas CH4 production started immediately in permanently flooded lakes, the other soil environments exhibited a more or less pronounced lag phase. This lag phase was caused by preferential reduction of alternative electron acceptors (e.g., ferric iron, sulfate) until they were depleted, and by growth of anaerobic microorganisms (e.g., methanogenic archaea). Indeed, numbers of methanogenic archaea were typically low in dry upland soils or desert soil crusts, where they only increased upon flooding. By contrast, all the other soil environments always contained large populations of methanogens, which just required reducing conditions and sufficient substrate (H2, acetate) concentrations to become active (Table 1).
Once sufficiently large populations sizes of methanogens were established, drainage of soil did not result in immediate decrease, showing that the methanogens were relatively insensitive to desiccation and O2 exposure. Only prolonged drainage over several seasons resulted in a decrease in these putatively O2-sensitive populations [75,76,78]. The relative insensitivity of the methanogens to O2 is consistent with the observations that dry and occasionally flooded soil environments were dominated by methanogens belonging to the genera Methanosarcina and Methanocella, which both belong to the O2-tolerant class II [68]. Methanosarcina and Methanocella species also increased in relative abundance after desiccation of permanently or frequently flooded soil environments, which, however, also contained O2-sensitive methanogens of class I (e.g., Methanosaetaceae, Methanobacteriales) [71,73]. Hence, the abundance and composition of methanogen communities was consistent with theoretical expectations concerning O2 sensitivity. However, O2 sensitivity was apparently not the sole criterion. For example, in permanently flooded lake sediments [69], Methanomicrobiales belonging to class II decreased upon desiccation, and, in some paddy soils, [71,76] Methanobacteriales belonging to class I increased. Methanogens that were actively assimilating carbon in aerated soils belonged to methanogens of both class I and class II [66,67]. As a conclusion, O2 sensitivity is apparently not the sole explanation for the occurrence of methanogenic taxa; at least, it does not explain the population dynamics of Methanosaetaceae, Methanomicrobiales and Methanobacteriales, which seem to follow criteria other than O2 sensitivity.
The substrate supply of the methanogenic archaea is accomplished by fermenting bacteria. Endospore-forming Firmicutes were an important group of fermenting bacteria in virtually all anoxic soil environments (Table 1). It is not surprising that the relative abundance of Firmicutes was found to increase upon desiccation in most categories of flooded soil environments. Firmicutes was the major taxon being activated upon hydration of desert soil crusts [77] and increased in relative abundance after desiccation and reflooding of lake sediments [69] and of many other flooded soil environments [73]. Nevertheless, some environments contained Firmicutes as only a minor group of bacteria [76], and, in some environments, their relative abundance was not enhanced after desiccation [72]. Instead, Acidobacteria, Actinobacteria, Chloroflexi and Proteobacteria exhibited larger population dynamics. In summary, the abilities to be O2-tolerant and form resting stages are apparently important but not exclusive criteria for survival and activity of bacterial taxa in flooded and drained soil environments.

5. Role of Soil Organic Carbon

The CH4 production potential always recovered when drained soil was flooded again. However, once recovered, CH4 production rates were not necessarily equal to those before drainage or desiccation (Table 1). In some environmental systems, rates were decreased; in others, they increased. For example, CH4 production in upland soils and soils with crop rotation always decreased upon second flooding, whereas, in wetland rice fields, Amazonian wetlands and lake sediments, potentials increased, decreased or remained the same (Table 1). The reason for such changes cannot so far be traced to changes in the size or composition of the methanogenic microbial communities. Instead, it may be hypothesized that desiccation resulted in a change in the composition of the organic matter (e.g., by chemical reactions) and that such change affected the propensity of being degraded under anaerobic conditions [71,79]. In fact, the composition of SOC may be of similar importance for the functioning of the microbial methanogenic communities as their composition. The composition of SOC, for example, defines the recalcitrance of organic matter. Degradation of SOC to methanogenic substrates (e.g., H2, CO2, acetate) is arguably the rate-limiting step in the production of CH4 [80,81,82].
The apparent pathway of CH4 production in different soil environments was usually higher than 50% hydrogenotrophic methanogenesis, except in flooded rice fields, where it was close to the expected value of <33% (Table 1). The reason why the contribution of aceticlastic and hydrogenotrophic methanogenesis did not follow the expectation may be the activity of syntrophic acetate oxidation coupled to hydrogenotrophic methanogenesis, despite of the presence of populations of aceticlastic methanogens [83,84]. However, acetate can also be oxidized by reduction of organic compounds [85]. In fact, the soil organic matter may be degraded, involving redox reactions of SOC compounds [13,14]. It is known that SOC contains redox-active compounds, which may serve as electron donors, electron acceptors or electron shuttles between bacteria and iron minerals [86,87,88,89,90]. Literature data indicate that changes in the oxidation state of organic matter do happen [89,91,92,93,94,95], although resolution of individual carbon compound formulas does not presently allow identification of the compounds involved. Our knowledge of the composition and dynamics of the thousands of different environmental organic carbon compounds [95,96,97,98] is very limited and requires much more research in order for us to be able to amend our theoretical picture of methanogenic degradation pathways accordingly.

6. Methane Oxidation

Flooded soil environments usually have a strong potential for CH4 production and are generally a source of atmospheric CH4, whereas drained environments lose their potential for CH4 production and stop being a source (see Section 4). Literature indicates that many soil environments actually turn into a net sink of CH4 when drained, indicating that CH4-oxidizing microorganisms become active that are able to oxidize CH4 at low atmospheric concentrations. Such observations were made in soils from peat swamps [99,100], river bank margins [101] and acidic wetlands [102]. However, dried soils do not generally possess activity for aerobic CH4 oxidation, but require induction by incubation at sufficiently high CH4 concentrations [103]. Such soils also are not generally able to utilize the low concentrations of atmospheric CH4 [104]. Thus, drained rice fields usually do not become a net sink for atmospheric CH4 [105,106,107], and the CH4 oxidation potential ceases earlier upon drainage than the CH4 production potential [108]. Only flush feeding with sufficient amounts of CH4 was found to guarantee the development of specific methanotrophic bacteria able to oxidize atmospheric CH4 concentrations [109]. These high-affinity methanotrophs belonged to canonical genera of methanotrophic bacteria (e.g., Methylocystis, Methylosarcina). Desert soil has been found to become a net sink of atmospheric CH4 once water is sufficiently available, apparently because high-affinity methanotrophs are activated [110,111].

7. Conclusions

Flooding and drainage of soil environments control the availability of O2 and the operation of methanogenic microbial processes. Empirical studies of the structure and function of soil microbial communities confirm the validity of pertinent biogeochemical and microbiological theories, but also demonstrate some unexpected processes. In particular, it has become apparent that some groups of anaerobic microbes (e.g., methanogenic archaea) are not thus strictly anaerobic, as formerly believed. Some groups of methanogens were found to tolerate O2 to quite some extent and also to express this characteristic in various soil environments ranging from permanently flooded to dry upland soils. Without such substantial O2 tolerance, many wetlands could probably not act as a source for atmospheric CH4, so that natural and artificial wetlands would not constitute one of the most important sources (accounting for about 250–300 Tg a−1 or 50% in the global CH4 budget [112]. Biogeochemical and microbiological theories concerning environmental methanogenesis have to be amended accordingly. Other interesting observations concern the likely involvement of soil organic carbon compounds in redox processes and methanogenesis, so that for the functioning of environmental methanogenesis the composition of SOC may be of similar importance as that of the microbial community. Finally, it is interesting that drainage and aeration does not always turn a flooded soil environment into a net sink for atmospheric CH4. The reason for this behavior may be differences in the methanotrophic microbial communities among the various soil environments, which deserves further research.


This research received no external funding.

Conflicts of Interest

The authors declare no conflict of interest.


  1. Conrad, R. Microbial ecology of methanogens and methanotrophs. Adv. Agron. 2007, 96, 1–63. [Google Scholar]
  2. Oremland, R.S. Biogeochemistry of Methanogenic Bacteria; Zehnder, A.J.B., Ed.; Wiley: New York, NY, USA, 1988; pp. 641–705. [Google Scholar]
  3. Zinder, S.H. Methanogenesis: Ecology, Physiology, Biochemistry and Genetics; Ferry, J.G., Ed.; Chapman & Hall: New York, NY, USA, 1993; pp. 128–206. [Google Scholar]
  4. Brune, A. Symbiotic digestion of lignocellulose in termite guts. Nat. Rev. Microbiol. 2014, 12, 168–180. [Google Scholar] [CrossRef] [PubMed]
  5. Bridgham, S.D.; Cadillo-Quiroz, H.; Keller, J.K.; Zhuang, Q. Methane emissions from wetlands: Biogeochemical, microbial, and modeling perspectives from local to global scales [review]. Glob. Chang. Biol. 2013, 19, 1325–1346. [Google Scholar] [CrossRef] [PubMed]
  6. Andersen, R.; Chapman, S.; Artz, R.R.E. Microbial communities in natural and disturbed peatlands: A review. Soil Biol. Biochem. 2013, 57, 979–994. [Google Scholar] [CrossRef]
  7. Bartlett, K.B.; Harriss, R.C. Review and assessment of methane emissions from wetlands. Chemosphere 1993, 26, 261–320. [Google Scholar] [CrossRef]
  8. Borges, A.V.; Abril, G.; Bouillon, S. Carbon dynamics and CO2 and CH4 outgassing in the Mekong delta. Biogeosciences 2018, 15, 1093–1114. [Google Scholar] [CrossRef][Green Version]
  9. Kim, D.G.; Vargas, R.; Bond-Lamberty, B.; Turetsky, M.R. Effects of soil rewetting and thawing on soil gas fluxes: A review of current literature and suggestions for future research [review]. Biogeosciences 2012, 9, 2459–2483. [Google Scholar] [CrossRef][Green Version]
  10. Lerman, A. Geochemical Processes. In Water and Sediment Environments; Wiley: New York, NY, USA, 1979. [Google Scholar]
  11. Zehnder, A.J.B.; Stumm, W. Biology of Anaerobic Microorganisms; Zehnder, A.J.B., Ed.; Wiley: New York, NY, USA, 1988; pp. 1–38. [Google Scholar]
  12. Amend, J.P.; Shock, E.L. Energetics of overall metabolic reactions of thermophilic and hyperthermophilic Archaea and Bacteria [Review]. Fems Microbiol. Rev. 2001, 25, 175–243. [Google Scholar] [CrossRef]
  13. Conrad, R. Contribution of hydrogen to methane production and control of hydrogen concentrations in methanogenic soils and sediments [review]. Fems Microbiol. Ecol. 1999, 28, 193–202. [Google Scholar] [CrossRef]
  14. Conrad, R. Importance of hydrogenotrophic, aceticlastic, and methylotrophic methanogenesis for methane production in terrestrial, aquatic and other anoxic environments: A mini review. Pedosphere 2020, 30, 25–39. [Google Scholar] [CrossRef]
  15. Zehnder, A.J.B. Water Pollution Microbiology; Mitchell, R., Ed.; Wiley: New York, NY, USA, 1978; pp. 349–376. [Google Scholar]
  16. Schink, B.; Stams, A.J.M. The Prokaryotes: Prokaryotic Communities and Ecophysiology; Rosenberg, E., DeLong, E.F., Lory, S., Stackebrandt, E., Thompson, F., Eds.; Springer: Berlin, Germany, 2013; pp. 471–493. [Google Scholar]
  17. Kimura, M.; Asakawa, S. Handbook of Soil Sciences-Properties and Processes; Huang, P.M., Li, Y., Summer, M.E., Eds.; CRC Press: Boca Raton, FL, USA, 2012; pp. 2632–2640. [Google Scholar]
  18. Wilson, J.M.; Griffin, D.M. Water potential and the respiration of microrganisms in the soil. Soil Biol. Biochem. 1975, 7, 199–204. [Google Scholar] [CrossRef]
  19. Atlas, R.M.; Bartha, R. Microbial Ecology: Fundamentals and Applications; Pearson Education India: New Delhi, India, 1998. [Google Scholar]
  20. Lennon, J.T.; Jones, S.E. Microbial seed banks: The ecological and evolutionary implications of dormancy. Nat. Rev. Microbiol. 2011, 9, 119–130. [Google Scholar] [CrossRef] [PubMed]
  21. Button, D.K. Kinetics of nutrient-limited transport and microbial growth. Microbiol. Rev. 1985, 49, 270–297. [Google Scholar] [CrossRef] [PubMed][Green Version]
  22. Robinson, J.A. Determining microbial kinetic parameters using nonlinear regression analysis. Adv. Microb. Ecol. 1985, 8, 61–114. [Google Scholar]
  23. Lovley, D.R. Minimum threshold for hydrogen metabolism in methanogenic bacteria. Appl. Env. Microbiol. 1985, 49, 1530–1531. [Google Scholar] [CrossRef][Green Version]
  24. Jetten, M.S.M.; Stams, A.J.M.; Zehnder, A.J.B. Acetate threshold and acetate activating enzymes in methanogenic bacteria. Fems Microbiol. Ecol. 1990, 73, 339–344. [Google Scholar] [CrossRef]
  25. Robinson, J.A.; Tiedje, J.M. Competition between sulfate-reducing and methanogenic bacteria for H2 under resting growing conditions. Arch. Microbiol. 1984, 137, 26–32. [Google Scholar] [CrossRef]
  26. Cord-Ruwisch, R.; Seitz, H.J.; Conrad, R. The capacity of hydrogenotrophic anaerobic bacteria to compete for traces of hydrogen depends on the redox potential of the terminal electron acceptor. Arch. Microbiol. 1988, 149, 350–357. [Google Scholar] [CrossRef]
  27. Lovley, D.R.; Goodwin, S. Hydrogen concentrations as an indicator of the predominant terminal electron-accepting reactions in aquatic sediments. Geochim. Cosmochim. Acta 1988, 52, 2993–3003. [Google Scholar] [CrossRef][Green Version]
  28. Conrad, R.; Wetter, B. Influence of temperature on energetics of hydrogen metabolism in homoacetogenic, methanogenic, and other anaerobic bacteria. Arch. Microbiol. 1990, 155, 94–98. [Google Scholar] [CrossRef]
  29. Thauer, R.K.; Kaster, A.K.; Seedorf, H.; Buckel, W.; Hedderich, R. Methanogenic archaea: Ecologically relevant differences in energy conservation. Nat. Rev. Microbiol. 2008, 6, 579–591. [Google Scholar] [CrossRef] [PubMed]
  30. Jetten, M.S.M.; Stams, A.J.M.; Zehnder, A.J.B. Methanogenesis from acetate-A comparison of the acetate metabolism in Methanothrix soehngenii and Methanosarcina spp. Fems Microbiol. Rev. 1992, 88, 181–197. [Google Scholar] [CrossRef]
  31. Fey, A.; Conrad, R. Effect of temperature on carbon and electron flow and on the archaeal community in methanogenic rice field soil. Appl. Env. Microbiol. 2000, 66, 4790–4797. [Google Scholar] [CrossRef][Green Version]
  32. Lueders, T.; Friedrich, M.W. Effects of amendment with ferrihydrite and gypsum on the structure and activity of methanogenic populations in rice field soil. Appl. Env. Microbiol. 2002, 68, 2484–2494. [Google Scholar] [CrossRef] [PubMed][Green Version]
  33. Koebsch, F.; Winkel, M.; Liebner, S.; Liu, B.; Westphal, J.; Schmiedinger, I.; Spitzy, A.; Gehre, M.; Jurasinski, G.; Köhler, S. Sulfate deprivation triggers high methane production in a disturbed and rewetted coastal peatland. Biogeosciences 2019, 16, 1937–1953. [Google Scholar] [CrossRef][Green Version]
  34. Aulakh, M.S.; Wassmann, R.; Rennenberg, H. Methane emissions from rice fields-Quantification, mechanisms, role of management, and mitigation options. Adv. Agron. 2001, 70, 193–260. [Google Scholar]
  35. Timisina, J.; Jat, M.L.; Majumdar, K. Rice-maize systems of South Asia: Current status, future prospects and research priorities for nutrient management [review]. Plant Soil 2010, 335, 65–68. [Google Scholar] [CrossRef]
  36. Sioli, H. The Amazon; Sioli, H., Ed.; Dr. W. Junk Publ.: Dordrecht, The Netherlands, 1984; pp. 127–165. [Google Scholar]
  37. Bartlett, K.B.; Crill, P.M.; Bonassi, J.A.; Richey, J.E.; Harriss, R.C. Methane flux from the Amazon River floodplain: Emissions during rising water. J. Geophys. Res. 1990, 95, 16773–16788. [Google Scholar] [CrossRef]
  38. Bastviken, D.; Tranvik, L.J.; Downing, J.A.; Crill, P.M.; Enrich-Prast, A. Freshwater methane emissions offset the continental carbon sink. Science 2011, 331, 50. [Google Scholar] [CrossRef][Green Version]
  39. Angel, R.; Claus, P.; Conrad, R. Methanogenic archaea are globally ubiquitous in aerated soils and become active under wet anoxic conditions. Isme J. 2012, 6, 847–862. [Google Scholar] [CrossRef][Green Version]
  40. Ponnamperuma, F.N. The chemistry of submerged soils. Adv. Agron. 1972, 24, 29–96. [Google Scholar]
  41. Reddy, K.R.; Patrick, W.H. Effect of alternate aerobic and anaerobic conditions on redox potential, organic matter decomposition and nitrogen loss in a flooded soil. Soil Biol. Biochem. 1975, 7, 87–94. [Google Scholar] [CrossRef]
  42. Yao, H.; Conrad, R.; Wassmann, R.; Neue, H.U. Effect of soil characteristics on sequential reduction and methane production in sixteen rice paddy soils from China, the Philippines, and Italy. Biogeochemistry 1999, 47, 269–295. [Google Scholar] [CrossRef]
  43. Yao, H.; Conrad, R. Thermodynamics of methane production in different rice paddy soils from China, the Philippines and Italy. Soil Biol. Biochem. 1999, 31, 463–473. [Google Scholar] [CrossRef]
  44. Ratering, S.; Conrad, R. Effects of short-term drainage and aeration on the production of methane in submerged rice soil. Glob. Chang. Biol. 1998, 4, 397–407. [Google Scholar] [CrossRef]
  45. Sigren, L.K.; Lewis, S.T.; Fisher, F.M.; Sass, R.L. Effects of field drainage on soil parameters related to methane production and emission from rice paddies. Glob. Biogeochem. Cycles 1997, 11, 151–162. [Google Scholar] [CrossRef]
  46. Achtnich, C.; Bak, F.; Conrad, R. Competition for electron donors among nitrate reducers, ferric iron reducers, sulfate reducers, and methanogens in anoxic paddy soil. Biol. Fertil. Soils 1995, 19, 65–72. [Google Scholar] [CrossRef]
  47. Klüber, H.D.; Conrad, R. Effects of nitrate, nitrite, NO and N2O on methanogenesis and other redox processes in anoxic rice field soil. Fems Microbiol. Ecol. 1998, 25, 301–318. [Google Scholar] [CrossRef]
  48. Glissmann, K.; Conrad, R. Saccharolytic activity and its role as a limiting step in methane formation during the anaerobic degradation of rice straw in rice paddy soil. Biol. Fertil. Soils 2002, 35, 62–67. [Google Scholar] [CrossRef]
  49. Lüdemann, H.; Arth, I.; Liesack, W. Spatial changes in the bacterial community structure along a vertical oxygen gradient in flooded paddy soil cores. Appl. Env. Microbiol. 2000, 66, 754–762. [Google Scholar] [CrossRef][Green Version]
  50. Noll, M.; Matthies, D.; Frenzel, P.; Derakshani, M.; Liesack, W. Succession of bacterial community structure and diversity in a paddy soil oxygen gradient. Env. Microbiol. 2005, 7, 382–395. [Google Scholar] [CrossRef]
  51. Shrestha, P.M.; Kube, M.; Reinhardt, R.; Liesack, W. Transcriptional activity of paddy soil bacterial communities. Env. Microbiol. 2009, 11, 960–970. [Google Scholar] [CrossRef] [PubMed]
  52. Chin, K.J.; Conrad, R. Intermediary metabolism in methanogenic paddy soil and the influence of temperature. Fems Microbiol. Ecol. 1995, 18, 85–102. [Google Scholar] [CrossRef]
  53. Glissmann, K.; Conrad, R. Fermentation pattern of methanogenic degradation of rice straw in anoxic paddy soil. Fems Microbiol. Ecol. 2000, 31, 117–126. [Google Scholar] [CrossRef] [PubMed]
  54. Penning, H.; Conrad, R. Quantification of carbon flow from stable isotope fractionation in rice field soils with different organic matter content. Org. Geochem. 2007, 38, 2058–2069. [Google Scholar] [CrossRef]
  55. Ma, K.; Conrad, R.; Lu, Y. Responses of methanogen mcrA genes and their transcripts to an alternate dry/wet cycle of paddy field soil. Appl. Env. Microbiol. 2012, 78, 445–454. [Google Scholar] [CrossRef][Green Version]
  56. Yuan, Y.; Conrad, R.; Lu, Y. Transcriptional response of methanogen mcrA genes to oxygen exposure of rice field soil. Env. Microbiol. Rep. 2011, 3, 320–328. [Google Scholar] [CrossRef]
  57. Abdallah, R.Z.; Wegner, C.E.; Liesack, W. Community transcriptomics reveals drainage effects on paddy soil microbiome accross all three domains if life. Soil Biol. Biochem. 2019, 132, 131–142. [Google Scholar] [CrossRef]
  58. Watanabe, T.; Kimura, M.; Asakawa, S. Distinct members of a stable methanogenic archaeal community transcribe mcrA genes under flooded and drained conditions in Japanese paddy field soil. Soil Biol. Biochem. 2009, 41, 276–285. [Google Scholar] [CrossRef]
  59. Kiener, A.; Leisinger, T. Oxygen sensitivity of methanogenic bacteria. Syst. Appl. Microbiol. 1983, 4, 305–312. [Google Scholar] [CrossRef]
  60. Fetzer, S.; Bak, F.; Conrad, R. Sensitivity of methanogenic bacteria from paddy soil to oxygen and desiccation. Fems Microbiol. Ecol. 1993, 12, 107–115. [Google Scholar] [CrossRef]
  61. Peters, V.; Conrad, R. Sequential reduction processes and initiation of CH4 production upon flooding of oxic upland soils. Soil Biol. Biochem. 1996, 28, 371–382. [Google Scholar] [CrossRef]
  62. Angel, R.; Matthies, D.; Conrad, R. Activation of methanogenesis in arid biological soil crusts despite the presence of oxygen. PLoS ONE 2011, 6, e20453. [Google Scholar] [CrossRef] [PubMed][Green Version]
  63. Peters, V.; Conrad, R. Methanogenic and other strictly anaerobic bacteria in desert soil and other oxic soils. Appl. Env. Microbiol. 1995, 61, 1673–1676. [Google Scholar] [CrossRef][Green Version]
  64. Leadbetter, J.R.; Breznak, J.A. Physiological ecology of Methanobrevibacter cuticularis sp nov and Methanobrevibacter curvatus sp nov, isolated from the hindgut of the termite Reticulitermes flavipes. Appl. Env. Microbiol. 1996, 62, 3620–3631. [Google Scholar] [CrossRef][Green Version]
  65. Tholen, A.; Pester, M.; Brune, A. Simultaneous methanogenesis and oxygen reduction by Methanobrevibacter cuticularis at low oxygen fluxes. Fems Microbiol. Ecol. 2007, 62, 303–312. [Google Scholar] [CrossRef] [PubMed]
  66. Watanabe, T.; Wang, G.; Lee, C.G.; Murase, J.; Asakawa, S.; Kimura, M. Assimilation of glucose-derived carbon into methanogenic archaea in soil under unflooded condition. Appl. Soil Ecol. 2011, 48, 201–209. [Google Scholar] [CrossRef]
  67. Lee, C.G.; Watanabe, T.; Murase, J.; Asakawa, S.; Kimura, M. Growth of methanogens in an oxic soil microcosm: Elucidation by a DNA-SIP experiment using 13C-labeled dried rice callus. Appl. Soil Ecol. 2012, 58, 37–44. [Google Scholar] [CrossRef]
  68. Lyu, Z.; Lu, Y. Metabolic shift at the class level sheds light on adaptation of methanogens to oxidative environments. Isme J. 2018, 12, 411–423. [Google Scholar] [CrossRef][Green Version]
  69. Conrad, R.; Ji, Y.; Noll, M.; Klose, M.; Claus, P.; Enrich-Prast, A. Response of the methanogenic microbial communities in Amazonian oxbow lake sediments to desiccation stress. Env. Microbiol. 2014, 16, 1682–1694. [Google Scholar] [CrossRef]
  70. Ji, Y.; Angel, R.; Klose, M.; Claus, P.; Marotta, H.; Pinho, L.; Enrich-Prast, A.; Conrad, R. Structure and function of methanogenic microbial communities in sediments of Amazonian lakes with different water types. Env. Microbiol. 2016, 18, 5082–5100. [Google Scholar] [CrossRef] [PubMed]
  71. Reim, A.; Hernandez, M.; Klose, M.; Chidthaisong, A.; Yuttiham, M.; Conrad, R. Response of methanogenic microbial communities to desiccation stress in flooded and rain-fed paddy soil from Thailand. Front. Microbiol. 2017, 8, 785. [Google Scholar] [CrossRef] [PubMed]
  72. Hernandez, M.; Conrad, R.; Klose, M.; Ma, K.; Lu, Y.H. Structure and function of methanogenic microbial communities in soils from flooded rice and upland soybean fields from Sanjiang plain, NE China. Soil Biol. Biochem. 2017, 105, 81–91. [Google Scholar] [CrossRef][Green Version]
  73. Hernandez, M.; Klose, M.; Claus, P.; Bastviken, D.; Marotta, H.; Figueiredo, V.; Enrich-Prast, A.; Conrad, R. Structure, function and resilience to desiccation of methanogenic microbial communities in temporarily inundated soils of the Amazon rainforest (Cunia Reserve, Rondonia). Env. Microbiol. 2019, 21, 1702–1717. [Google Scholar] [CrossRef] [PubMed]
  74. Ji, Y.; Fernandez Scavino, A.; Klose, M.; Claus, P.; Conrad, R. Functional and structural responses of methanogenic microbial comunities in Uruguayan soils to intermittent drainage. Soil Biol. Biochem. 2015, 89, 238–247. [Google Scholar] [CrossRef]
  75. Fernandez Scavino, A.; Ji, Y.; Pump, J.; Klose, M.; Claus, P.; Conrad, R. Structure and function of the methanogenic microbial communities in Uruguayan soils shifted between pasture and irrigated rice fields. Env. Microbiol. 2013, 15, 2588–2602. [Google Scholar] [CrossRef]
  76. Breidenbach, B.; Blaser, M.B.; Klose, M.; Conrad, R. Crop rotation of flooded rice with upland maize impacts the resident and active methanogenic microbial community. Env. Microbiol. 2016, 18, 2868–2885. [Google Scholar] [CrossRef]
  77. Angel, R.; Conrad, R. Elucidating the microbial resuscitation cascade in biological soil crusts following a simulated rain event. Env. Microbiol. 2013, 15, 2799–2815. [Google Scholar] [CrossRef]
  78. Liu, D.; Nishida, M.; Takahashi, T.; Asakawa, S. Transcription of mcrA gene decreases upon prolonged non-flooding period in a methanogenic archaeal community of a paddy-upland rotational field soil. Microb. Ecol. 2018, 75, 751–760. [Google Scholar] [CrossRef]
  79. Atere, C.T.; Ge, T.; Zhu, Z.; Tong, C.; Jones, D.L.; Shibistova, O.; Guggenberger, G.; Wu, J. Rice rhizodeposition and carbon stabilisation in paddy soil are regulated via drying-rewetting cycles and nitrogen fertilisation. Biol. Fertil. Soils 2017, 53, 407–417. [Google Scholar] [CrossRef]
  80. Degens, E.T.; Mopper, K. Early diagenesis of organic matter in marine salts. Soil Sci. 1975, 119, 65–72. [Google Scholar] [CrossRef]
  81. Billen, G. Sediment Microbiology; Nedwell, D.B., Brown, C.M., Eds.; Academic Press: New York, NY, USA, 1982; pp. 15–52. [Google Scholar]
  82. Fey, A.; Conrad, R. Effect of temperature on the rate limiting step in the methanogenic degradation pathway in rice field soil. Soil Biol. Biochem. 2003, 35, 1–8. [Google Scholar] [CrossRef]
  83. Vavilin, V.; Rytov, S.; Conrad, R. Modeling methane formation in sediments of tropical lakes, focusing on syntrophic acetate oxidation: Dynamics and static isotope equations. Ecol. Modeling 2017, 363, 81–95. [Google Scholar] [CrossRef]
  84. Conrad, R.; Klose, M.; Enrich-Prast, A. Acetate turnover and methanogenic pathways in Amazonian lake sediments. Biogeosciences 2020, 17, 1063–1069. [Google Scholar] [CrossRef]
  85. Coates, J.D.; Ellis, D.J.; Blunt-Harris, E.L.; Gaw, C.V.; Roden, E.E.; Lovley, D.R. Recovery of humic-reducing bacteria from a diversity of environments. Appl. Env. Microbiol. 1998, 64, 1504–1509. [Google Scholar] [CrossRef][Green Version]
  86. Keller, J.K.; Weisenhorn, P.B.; Megonigal, J.P. Humic acids as electron acceptors in wetland decomposition. Soil Biol. Biochem. 2009, 41, 1518–1522. [Google Scholar] [CrossRef]
  87. Klüpfel, L.; Piepenbrock, A.; Kappler, A.; Sander, M. Humic substances as fully regenerable electron acceptors in recurrently anoxic environments. Nat. Geosci. 2014, 7, 195–200. [Google Scholar] [CrossRef]
  88. Heitmann, T.; Goldhammer, T.; Beer, J.; Blodau, C. Electron transfer of dissolved organic matter and its potential significance for anaerobic respiration in a northern bog. Glob. Chang. Biol. 2007, 13, 1771–1785. [Google Scholar] [CrossRef]
  89. Gao, C.; Sander, M.; Agethen, S.; Knorr, K.H. Electron accepting capacity of dissolved and particulate organic matter control CO2 and CH4 formation in peat soils. Geochim. Cosmochim. Acta 2019, 245, 266–277. [Google Scholar] [CrossRef]
  90. Lovley, D.R. Happy together: Microbial communities that hook up to swap electrons [review]. Isme J. 2017, 11, 327–336. [Google Scholar] [CrossRef][Green Version]
  91. Kellerman, A.M.; Kothawala, D.N.; Dittmar, T.; Tranvik, L.J. Persistence of dissolved organic matter in lakes related to its molecular characteristics. Nat. Geosci. 2015, 8, 454–457. [Google Scholar] [CrossRef]
  92. Lau, M.P.; Sander, M.; Gelbrecht, J.; Hupfer, M. Solid phases as important electron acceptors in freshwater organic sediments. Biogeochemistry 2015, 123, 49–61. [Google Scholar] [CrossRef]
  93. Medeiros, P.M.; Seidel, M.; Niggemann, J.; Spencer, R.G.M.; Hernes, P.J.; Yager, P.L.; Miller, W.L.; Dittmar, T.; Hansell, D.A. A novel molecular approach for tracing terrigenous dissolved organic matter into the deep ocean. Glob. Biogeochem. Cycles 2016, 30, 689–699. [Google Scholar] [CrossRef]
  94. Mostovaya, A.; Hawkes, J.A.; Koehler, B.; Dittmar, T.; Tranvik, L.J. Emergence of the activity continuum of organic matter from kinetics of a multitude of individual molecular constituents. Env. Sci. Technol. 2017, 51, 11571–11579. [Google Scholar] [CrossRef]
  95. Valle, J.; Gonsior, M.; Harir, M.; Enrich-Prast, A.; Schmitt-Kopplin, P.; Conrad, R.; Hertkorn, N. Extensive processing of sediment pore water dissolved organic matter during anoxic incubation as observed by high-field mass spectrometry (FTICR-MS). Water Res. 2018, 129, 252–263. [Google Scholar] [CrossRef][Green Version]
  96. Hodgkins, S.B.; Tfaily, M.M.; McCalley, C.K.; Logan, T.A.; Crill, P.M.; Saleska, S.R.; Rich, V.I.; Chanton, J.P. Changes in peat chemistry associated with permafrost thaw increase greenhouse gas production. Proc. Natl. Acad. Sci. USA 2014, 111, 5819–5824. [Google Scholar] [CrossRef][Green Version]
  97. Schmidt, F.; Elvert, M.; Koch, B.P.; Witt, M.; Hinrichs, K.U. Molecular characterization of dissolved organic matter in pore water of continental shelf sediments. Geochim. Cosmochim. Acta 2009, 73, 3337–3358. [Google Scholar] [CrossRef]
  98. Seidel, M.; Beck, M.; Riedel, T.; Waska, H.; Suryaputra, I.; Schnetger, B.; Niggemann, J.; Simon, M.; Dittmar, T. Biogeochemistry of dissolved organic matter in an anoxic intertidal creek bank. Geochim. Cosmochim. Acta 2014, 140, 418–434. [Google Scholar] [CrossRef]
  99. Harriss, R.C.; Sebacher, D.I. Methane flux in the Great Dismal Swamp. Nature 1982, 297, 673–674. [Google Scholar] [CrossRef]
  100. Melling, L.; Hatano, R.; Goh, K.J. Methane fluxes from three ecosystems in tropical peatland of Sarawak, Malaysia. Soil Biol. Biochem. 2005, 37, 1445–1453. [Google Scholar] [CrossRef]
  101. Kelley, C.A.; Martens, C.S.; Ussler, W. Methane dynamics across a tidally flooded riverbank margin. Limnol. Oceanogr. 1995, 40, 1112–1129. [Google Scholar] [CrossRef]
  102. Kolb, S.; Horn, M.A. Microbial CH4 and N2O consumption in acidic wetlands [review]. Front. Microbiol. 2012, 3, 7. [Google Scholar] [CrossRef][Green Version]
  103. Bender, M.; Conrad, R. Effect of CH4 concentrations and soil conditions on the induction of CH4 oxidation activity. Soil Biol. Biochem. 1995, 27, 1517–1527. [Google Scholar] [CrossRef]
  104. Henckel, T.; Conrad, R. Characterization of microbial NO production, N2O production and CH4 oxidation initiated by aeration of anoxic rice field soil. Biogeochemistry 1998, 40, 17–36. [Google Scholar] [CrossRef]
  105. DeniervanderGon, H.A.C.; VanBreemen, N.; Neue, H.U.; Lantin, R.S.; Aduna, J.B.; Alberto, M.C.R.; Wassmann, R. Release of entrapped methane from wetland rice fields upon soil drying. Glob. Biogeochem. Cycles 1996, 10, 1–7. [Google Scholar] [CrossRef]
  106. Abao, J.; Bronson, K.F.; Wassmann, R.; Singh, U. Simultaneous records of methane and nitrous oxide emissions in rice-based cropping systems under rainfed conditions. Nutr. Cycl. Agroecosyst. 2000, 58, 131–139. [Google Scholar] [CrossRef]
  107. Weller, S.; Kraus, D.; Ayag, K.R.; Wassmann, R.; Alberto, M.; Butterbach-Bahl, K.; Kiese, R. Methane and nitrous oxide emissions from rice and maize production in diversified rice cropping systems. Nutr. Cycl. Agroecosyst. 2015, 101, 37–53. [Google Scholar] [CrossRef]
  108. Jäckel, U.; Schnell, S.; Conrad, R. Effect of moisture, texture and aggregate size of paddy soil on production and consumption of CH4. Soil Biol. Biochem. 2001, 33, 965–971. [Google Scholar] [CrossRef]
  109. Cai, Y.F.; Zheng, Y.; Bodelier, P.L.E.; Conrad, R.; Jia, Z.J. Conventional methanotrophs are responsible for atmospheric methane oxidation in paddy soils. Nat. Commun. 2016, 7, 11728. [Google Scholar] [CrossRef][Green Version]
  110. Striegl, R.G.; McConnaughey, T.A.; Thorstenson, D.C.; Weeks, E.P.; Woodward, J.C. Consumption of atmospheric methane by desert soils. Nature 1992, 357, 145–147. [Google Scholar] [CrossRef]
  111. Angel, R.; Conrad, R. In situ measurement of methane fluxes and analysis of transcribed particulate methane monooxygenase in desert soils. Env. Microbiol. 2009, 11, 2598–2610. [Google Scholar] [CrossRef] [PubMed]
  112. Lyu, Z.; Shao, N.; Akinyemi, T.; Whitman, W.B. Methanogenesis [review]. Curr. Biol. 2018, 28, R719–R736. [Google Scholar] [CrossRef] [PubMed][Green Version]
Table 1. Effects of desiccation or flooding on microbial populations and methanogenic functions in soils across a gradient from permanent wet to permanent dry conditions. (The symbols denote the following: = constant; ↑ increase; ↓ decrease).
Table 1. Effects of desiccation or flooding on microbial populations and methanogenic functions in soils across a gradient from permanent wet to permanent dry conditions. (The symbols denote the following: = constant; ↑ increase; ↓ decrease).
Permanent WetSeasonal FloodingSeasonal FloodingRotationMostly DryPermanent Dry
lake sedimentsAmazon floodplainpaddy ricerice—upland cropupland soildesert soil crusts
Methanogen numbers after flooding====
Methanogen taxa stimulated by desiccationM’sarcinaceae
or unchanged
Firmicutes after desiccation and reflooding↑ =↑ =
CH4 production after desiccation and reflooding=↑↓
Hydrogenotrophic methanogenesis (%)>50>50<33<33>50>50

Share and Cite

MDPI and ACS Style

Conrad, R. Methane Production in Soil Environments—Anaerobic Biogeochemistry and Microbial Life between Flooding and Desiccation. Microorganisms 2020, 8, 881.

AMA Style

Conrad R. Methane Production in Soil Environments—Anaerobic Biogeochemistry and Microbial Life between Flooding and Desiccation. Microorganisms. 2020; 8(6):881.

Chicago/Turabian Style

Conrad, Ralf. 2020. "Methane Production in Soil Environments—Anaerobic Biogeochemistry and Microbial Life between Flooding and Desiccation" Microorganisms 8, no. 6: 881.

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop