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Perspective

Do Soil Methanotrophs Really Remove About 5% of Atmospheric Methane?

by
Xiaokun Yao
1,
Tao Tao
1,
Wei Li
1,*,
Tingzhen Ming
2 and
Renaud de Richter
3,*
1
Institute for Materials and Processes, School of Engineering, University of Edinburgh, Scotland EH9 3FB, UK
2
School of Civil Engineering and Architecture, Wuhan University of Technology, Wuhan 430070, China
3
Tour-Solaire.Fr, 8 Impasse des Papillons, 34090 Montpellier, France
*
Authors to whom correspondence should be addressed.
Land 2025, 14(9), 1864; https://doi.org/10.3390/land14091864
Submission received: 7 August 2025 / Revised: 4 September 2025 / Accepted: 9 September 2025 / Published: 12 September 2025
(This article belongs to the Section Land, Soil and Water)
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Abstract

It has been experimentally proved that microorganisms in soils are able to remove atmospheric methane (CH4), particularly through experiments with radioelements such as 14CH4. However, a curious question arises: are these microorganisms the only responsible sink for all atmospheric CH4 uptake attributed to soils, or do non-microbial (e.g., chemical) processes also contribute part of it? In this perspective article, we propose that atmospheric methane removal (AMR) in soils may result from a combination of microbial and non-microbial processes. In addition to oxidation by MOB, we analyzed the potential roles of photocatalytic reactions on soil minerals, Fenton-like chemistry in water droplets, chlorine radical pathways in chloride-rich soils and ozone/VOCs-driven •OH generation. These chemical mechanisms may act independently or intertwined with microbial activity under specific environmental conditions. We suggest that future studies use experimental approaches to explore and quantify the relative contributions of these pathways and to help refine our understanding of the soil CH4 sink in the global methane budget.

1. Introduction

Methane (CH4) is a potent greenhouse gas, with atmospheric concentrations currently increasing by about 10 ppb per year. Since preindustrial times, CH4 has contributed up to 0.5 °C of warming, second only to CO2 (0.8 °C) [1,2]. According to the latest ‘Global Methane Budget’, soils act as a net sink for atmospheric methane (AM), removing approximately 32 Tg CH4 per year, which accounts for about 5% of the total AM sink [3].
It has been experimentally proved that microorganisms in soils are able to remove AM, particularly through experiments with radioelements such as 14CH4, which showed that 31–43% of the labelled CH4 was assimilated into microbial biomass, including proteins, nucleic acids, polysaccharides and lipids, while the remainder was oxidized to 14CO2 [4].
It is known that other microbial atmospheric CH4 oxidizers play a role, such as nitrifying bacteria [5] and atmospheric CH4 oxidizing bacteria (atmMOB), which oxidize CH4 as an energy source in unsaturated oxygenated soils [3], and include methanotrophic as well as methylotrophic bacteria [6,7].
However, CH4 oxidation through these microbial processes occurs very slowly when they are exposed only to ambient concentrations of atmospheric CH4 (approximately 2 ppm) [8,9], and their slow growth and sensitivity to environmental factors such as water scarcity and temperature extremes in deserts or high altitudes may restrict their contribution to AMR under certain conditions [10,11].
Moreover, laboratory observations showing that CH4 consumption by methylotrophic bacteria can exceed the supply expected from atmospheric diffusion alone, with CH4 concentrations dropping from 10 ppm to 0.14 ppm within 8 h, suggest an alternative mechanism [12,13]. Similarly, rapid CH4 depletion has been recorded in subsurface environments like caves, often in settings where microbial abundance is minimal [14]. These discrepancies suggest the involvement of additional, potentially nonbiological mechanisms in soil CH4 removal.
Consequently, this raises the question: are these MOBs responsible alone for the approximately 5% of global atmospheric CH4 removal attributed to soils, or might non-microbial processes also contribute to it? Therefore, this perspective article does not dispute that MOB constitute an important sink for AM but questions whether they are responsible for about 5% of AMR, or whether this sink can also be explained by other factors or mechanisms not yet considered in the “Global Methane Budget” [3].
This perspective article aims to suggest that the soil AM sink involves other complex and potentially underexplored mechanisms of AMR that need to be further studied [3]. For instance, rain, fog and dew contain hydrogen peroxide (H2O2), organic peroxides, as well as iron (Fe) [15,16,17]. Considering the fact that water microdroplets can spontaneously generate H2O2 at the air-water interface [18,19,20,21,22,23], this allows the generation of •OH radicals by Fenton-like reactions. We hypothesize that additional AMR mechanisms may coexist in soil.
After explaining our bibliographical methodology, the following sections will examine both microbial and chemical processes of AM, followed by a discussion highlighting the principal shortcomings of the existing publications and proposing experimental approaches to validate or refute the hypotheses outlined here. Through this perspective, we aim to offer alternative frameworks and new research directions to help address the ongoing rise in atmospheric CH4 level and its sinks.

2. Methodology

The starting point of this perspective article is based on the US National Academies’ 2024 report on atmospheric methane removal [1] and the Global Methane Budget 2000–2020 [3]. Both studies highlight several research needs aimed at filling gaps in the fundamental understanding of atmospheric and ecosystem methane sinks.
Therefore, research on the limits of atmospheric CH4 uptake by different microbial communities is necessary [1], and more importantly, the soil CH4 sink may involve other complex and potentially underexplored mechanisms that need to be further studied.
We conducted a targeted search of published, peer-reviewed scientific literature without a specific publication date; therefore, we used the Web of Knowledge and Google Scholar to search for articles published between 1970 and 2025 using the keywords “(CH4 uptake OR CH4 sink) AND (soil).” We then selected an initial list of 250 peer-reviewed publications in English, reporting soils as CH4 sinks, field measurements of CH4 uptake for all soil ecosystem types and computer modelling articles estimating the global soil sink.
For this perspective article, our literature review was neither systematic nor exhaustive but targeted and focused on identifying parameters known to influence the magnitude of the CH4 sink in soils (these are described later in this article). When selecting articles, in addition to the biotic sink due to microbes, we sought to determine whether a possible abiotic sink had been considered by other researchers. Our exclusion criteria included older articles that were redundant with more recent articles in terms of ecosystem soil types and analytical results. No quantitative analysis or data collection was initially included in the study, the focus being on attributing soil CH4 removal to a biotic or abiotic process, separately or jointly, and on identifying abiotic processes.
The objective of this perspective article is to open a discussion and raise awareness about the shortcomings or limitations of some articles taking for granted that the entire soil CH4 sink is due to methanotrophs, and to offer a wide range of alternatives, research needs and challenges to provide a vision and strategy for better implementation of targeted remediation techniques aimed at removing atmospheric methane.
We acknowledge the limitations and uncertainties inherent in this perspective study, in which, to confirm or refute our hypothesis, different forms of knowledge are integrated to seek understanding, overcome preconceived ideas, and support new interdisciplinary research, both to fill gaps in the global balance of CH4 sinks and to develop potential future AMR technologies. Therefore, this perspective, after highlighting the shortcomings of existing studies and emphasizing the novelty and importance of the present study, offers guidance to address research needs.

3. Microbial-Driven Methane Oxidation: Current Status and Limitations

Atmospheric methane oxidizing bacteria (atmMOB), such as Methylocapsa acidiphila, Methylocapsa aurea, Methylocapsa palsarum, Methylocystis rosea and Methylocapsa gorgona are capable of oxidizing CH4 at ambient atmospheric levels (~1.9 ppm) across diverse ecosystems [6,7].
CH4 oxidation in atmMOB begins with its conversion to methanol (CH3OH) by particulate methane monooxygenase (pMMO), followed by sequential oxidation to formaldehyde (CH2O) by methanol dehydrogenase (MDH) and finally converted to CO2 by formate dehydrogenase (FDH) [6,7,24]. Despite their recognized role, MOBs face several limitations. Their slow growth rates [9] reduce their efficiency in extreme environments such as deserts or high altitudes [6,7]. Additionally, CH4 and O2 availability in soil is constrained by gas diffusivity, which significantly influences MOB activity. Modelling studies indicate that a tenfold increase in diffusion rates could nearly triple global soil CH4 uptake [25].
Soil moisture also plays a critical role. Average moisture level of around 20% tends to favour MOB activity in many ecosystems [26,27]. However, both excessively dry, and overly wet conditions can severely restrict microbial oxidation [26,27]. CH4 oxidation by MOBs occurs primarily just below the soil surface [28]. Remarkably, CH4 uptake continues even in harsh environments such as Atacama Desert or Antarctic cryosols, where MOBs barely survive through dormancy or rely on alternative gases. CH4 uptake surprisingly persists and suggests that other oxidation processes besides these bacteria may exist in soils [24]. Soil moisture content of ~20% appears to be optimal for MOBs not only in tropical forests but also in tundra and shortgrass steppe ecosystems [26]. This consistent pattern suggests a shared biophysical condition favourable to MOB activity in these systems. However, this optimum does not apply universally. CH4 uptake has also been observed under non-ideal conditions, including in desert soils [27], snow-covered subalpine soils [29], Arctic soils [30] and permafrost-affected cryosols [31]. These observations further prove the potential limitations of microbial oxidation alone in explaining AMR across all soil types.
Hydrological factors, in conjunction with soil porosity, are key in determining the availability of pore space required for gas exchange processes in soil [28]. CH4 oxidation generally occurs at a depth of 3–15 cm [28]. In this zone, water distribution becomes critical: surface layers often dry out first under typical conditions [32], while deeper layers may remain overly wet [33,34]. Both extremes can reduce CH4 and O2 availability, thus limiting oxidation efficiency. In cold climate soils such as boreal forest, sub-Arctic regions and temperate forests, increased soil moisture is associated with reduced CH4 uptake. Additionally, nitrogen deposition has been shown to significantly suppress CH4 oxidation within just a few hours [5,34]. These findings reinforce the sensitivity of MOBs to both hydrological and nutrient-related environmental factors.
Looking forward for future, grassland soils exposed to elevated atmospheric CO2, warming and drought condition are expected to absorb more atmospheric CH4, mainly due to increased gas diffusivity and reduced soil moisture [35]. These projections suggest that even when microbial oxidation is the dominant pathway, physical constraints such as moisture levels and gas exchange capacity remain central to CH4 uptake.
A study found that the global soil AM sink increased by up to 0.35 Tg in 2020 compared to 2019 and attributed it mainly to warmer soil temperatures in the northern high latitudes, highlighting the role of terrestrial ecosystems in AM remediation and emphasizing the research needs on soil AM sinks in order to better understand AM dynamics [36].
Numerous field and laboratory studies have quantified CH4 oxidation by MOBs under various soil conditions. Table 1 summarizes findings from these studies, highlighting the diversity in removal rates based on soil type, measurement approach and environmental variables. These data demonstrate that although MOB activity dominates in many soil conditions, removal rates vary by land use, temperature, soil moisture and climate.
In the discussion section, we further present the main findings of this perspective study; their explanation and the implications in terms of alternative hypothesis or their limitations underlining research needs.

4. Potential Chemical Mechanisms

In addition to microbial oxidation, several chemical processes occurring in the soils may also contribute to AMR. These include photocatalysis, Fenton-like reactions, chlorine radical (•Cl) pathways and hydroxyl radicals (•OH) generated by ozone (O3) and volatile organic compounds (VOCs) reactions, as well as possible reactions with stable free radicals present in soils and plants. Each of these mechanisms is supported by laboratory or field evidence and may operate under specific environmental conditions.

4.1. Photocatalysis

Photocatalysis involves the use of metal oxides, such as TiO2, ZnO and WO3, commonly found in desert soils to oxidize CH4 under solar radiation [49,50,51,52]. Many semiconductor metal oxides have been explored for CH4 oxidation, either individually or in combination with co-catalysts. Notable systems include TiO2/MoO3, TiO2/H4SiW12O40, TiO2/MoO3/H4SiW12O40, Ag/ZnO, CuO/ZnO, UO22+/MCM-41, V/SiO2, P/SiO2, FeOOH/WO3 and FeOx/TiO2 [49,51]. These catalytic systems have demonstrated potential for CH4 oxidation under various environmental conditions.
Beyond classic photocatalysis, other approaches such as electrocatalysis and thermal catalysis have also been considered [53]. For instance, copper-doped mordenite zeolites, naturally occurring aluminosilicates enriched in sodium, potassium and calcium, have been used for thermal CH4 oxidation [54]. Photocatalysts have been proposed as passive components in large-scale air-moving system for ambient CH4 control [55,56,57]. For instance, silica, aluminosilicates, TiO2 and iron oxides are widespread in desert sands, semi-arid regions and general terrestrial soils areas, all recognized as CH4 sinks [27]. These naturally occurring materials provide abundant reactive surfaces for photocatalytic oxidation, particularly under high-intensity UV exposure. CH4 uptake in desert soils increases with humidity, and CH4 consumption has been observed at depth up to 2 m after rain [27]. However, this mechanism is most effective in arid, well-lit environments where UV radiation is sufficient to drive photocatalysis.

4.2. Fenton-like Reactions

Fenton-like reactions occur when metals, such as iron (Fe) and copper (Cu), react with hydrogen peroxide (H2O2) in atmospheric water, such as rain, fog and dew, to produce •OH [58,59,60]. Organic peroxides present in atmospheric droplets further promote this process [15,16,17]. Water microdroplets can spontaneously generate H2O2, increasing •OH concentration at the air-water interface [18,19,20,21,22,23]. Moreover, tree canopies and trunks can enhance •OH generation under sunlight [38].
These reactions are facilitated in moist soils, particularly under acidic conditions and with sufficient H2O2 availability. In contrast to temperate forest soils where rainfall may suppress microbial activity, desert soils exhibit enhanced CH4 uptake after precipitation, potentially due to the activation of Fenton-type reactions down to a depth of 2 m [27]. Transition metals such as Fe, Cu, Co, Mn and Ti, which are abundant in these soils, may catalyze these reactions especially after wetting events (e.g., rain, fog and dew) [58,59,60].
Additional support comes from early experimental findings: the Sahara Desert has been shown to act as a sink for trace gases such as CCl4 and N2O through photoreactive decomposition on silicates surfaces [61,62]. Organic peroxides and H2O2 can also form photochemically in aqueous suspensions containing sand and oxides [63]. Considering that CH4 has a solubility of 22.7 mg L−1 in water [64] and the enhanced reactivity at the air-water interface [65,66], aqueous-phase CH4 oxidation via Fenton-like reactions is plausible, especially under transient wetting conditions in arid soils.
Furthermore, these reactions may not be limited to soils; evidence suggests they may also occur on plant surfaces [15,16,17]. Recent studies have confirmed the spontaneous generation of H2O2 in water microdroplets [18,19,20,21,22,23]. CH4 uptake by foliage is nearly undetectable at night but approximately doubles under sunlight [38]. AMR has been observed to increase with canopy height, potentially due to increased UV and additional •OH generation via ozone decomposition [67,68]. In Greenland, soil copper content has been identified as a key factor influencing CH4 removal [41]. Although the original article interpreted this as microbial cofactor availability, it may also reflect the catalytic role of copper in Fenton-like reactions [69].
  • Chemical reaction mechanisms involved:
For the Fenton reaction, a multivalent metal such as iron (or copper, manganese, etc.) activates hydrogen peroxide (H2O2) to produce both °OH and HOO° as follows:
Fe2+ + H2O2 → Fe3+ + OH° + OH−
Fe3+ + H2O2 → Fe2+ +HOO°·+ H+
Then, the same mechanism as below is followed, starting by:
CH4 + °OH → °CH3 + H2O

4.3. Chlorine Radical Pathway

Chlorine radicals (•Cl) represent a highly reactive pathway for CH4 oxidation, which reacts ~16 times faster than •OH. In coastal soils and dust aerosols, sunlight promotes photolysis of FeCl3 to generate •Cl [70,71,72]. Similarly, chloride-containing TiO2 surfaces can be photoactivated to release •Cl [73]. •Cl formation has also been observed on icy and snowy surfaces [74,75].
This pathway is most relevant in chloride-rich environments such as marine influenced soils or coastal deserts. In these regions, sea salt deposition combined with acidic pH (pH < 3) and solar radiation facilitates •Cl formation [70,71,73]. Although spatially limited, this mechanism could significantly enhance local CH4 oxidation rates in chloride-enriched ecosystems.
  • Chemical reaction mechanisms involved:
The mechanisms involved in the reaction of methane with hydroxyl radicals or with chlorine radicals proceeds by H abstraction producing in both cases a methyl radical:
CH4 + °OH → °CH3 + H2O
or CH4 + °Cl → °CH3 + HCl
Both these reactions lead to the formation of CO (carbon monoxide) and finally of CO2 through the following oxidation path:
°CH3 + O2 → CH3OO°
CH3OO° + HOO° → CH3OOH + O2
CH3OOH + hv → CH3O° + °OH
CH3O° → CH2O + H°
CH2O + °OH → HCO° + H2O
or CH2O + °Cl → HCO° + HCl
HCO° + O2 → CO + HOO°
CO + 2°OH → CO2 + H2O
Carbon monoxide (CO) has an atmospheric lifetime of about 2 months and is an intermediate compound in the oxidation of all VOCs; therefore, it consumes about 39% of the °OH, while the oxidation of CH4 consumes 12% of the °OH [76].

4.4. Ozone/VOCs Driven Reaction

O3 and VOCs contribute indirectly to CH4 oxidation by generating reactive radicals. Soil and plants serve as O3 sinks, decomposing O3 and releasing NOx and VOCs [77,78,79]. In the presence of sunlight, O3 photolysis forms excited oxygen atoms (O(1D)), which then react with water vapour to produce •OH [75,77,78]. Additionally, VOCs such as isoprene can react with both O3 and •OH, amplifying radical production [79,80,81].
These processes are particularly relevant in the atmospheric boundary layer, well-ventilated soils, porous wood and even indoor environments [67,82,83]. Although dependent on VOC availability and UV intensity, they may contribute to AMR in vegetated, arid or partially shaded settings. While their overall contribution may be limited, they represent an additional oxidation pathway in specific conditions.

4.5. Free Radicals

On plants, stable organic free radicals accumulate in fully senescent leaves [84,85], trees [86], seeds and pollen [87], as well as on lignin and lignin oxidation products [88], and in soils, there are substantial amounts of biogenic persistent free radicals from leaves [89] and from humic acids [90]. Therefore, oxidation reactions of CH4 on those organic long-lived free radicals in soils and plant structures are possible.

4.6. Coexistence and Integration with Microbial Processes

Importantly, these chemical oxidation mechanisms are not mutually exclusive with microbial activity. In fact, CH4 is often rapidly consumed under near-saturated humidity conditions in environments like caves, where microbial presence is minimal [14]. These observations support the idea that chemical processes may act independently or synergistically with MOBs. Figure 1 and Table 2 summarize the various potential non-microbial pathways and their conditions of operation in AMR.

5. Discussion

The lack of studies demonstrating soil AMR by chemical processes is the principal limitation of the research in this field as identified by this perspective article. The main challenges in advancing this research directly, including quantifying the contribution of those chemical processes relative to the microbial sink, are (1) continuous measurements and careful mass balances over appropriate timescales, (2) overcoming the low signal-to-noise ratio, which exceed the accuracy of many chamber systems and sensors and (3) the co-existing microbial and non-microbial processes can often interact with each-other.
Despite these challenges and limitation, it is worth discussing (1) in more depth some additional evidence that strengthens our hypotheses, (2) shortcomings of some of existing publications and (3) the research needs and directions to address the above-mentioned challenges.
  • Radiotracer experiments
Radiotracer experiments using 14CH4 have shown that only 31–43% of CH4 is assimilated into microbial biomass, while the remaining fraction is oxidized to 14CO2 [4]. This raises the possibility that chemical oxidation, such as Fenton-like reactions catalyzed by iron oxides, also contributes to AMR, particularly under conditions where microbial activity is limited.
  • Temperature variations and its effects
Microbial CH4 oxidation in soils is constrained by several environmental factors, including low microbial abundance, slow growth rates of MOBs and vulnerability to water stress, temperature changes and nutrient disturbances [10,11]. Although there exists a thermal acclimatation, the amount of CH4 required to support cell division in methanotrophs is controlled by temperature [91].
During the spring and summer when soil temperatures are higher, soil CH4 consumption is typically higher, but in some studies, it is not much lower during the autumn (−5%) and winter (−14%) [29]. During the cold season when soil temperatures drop below 0 °C, the CH4 soil consumption continues [92]; for instance, in a shortgrass Colorado steppe (with no snow cover) when soil temperature was <−2 °C [29].
Trace gas exchange did not stop throughout the snow-covered period under alpine and sub-alpine snowpacks, as atmospheric CH4 uptake was measured (respectively, 2.9 × 10−6 and 6.7 × 10−6 mol CH4 m−2 d−1), suggesting that the soil serves as a larger CH4 sink while under the snowpack (235 days) than during the snow-free time (130 days) [93]. These observations seem also compatible with a chemical or physical AMR process, since ice-mediated halogen activation is well known both in the Arctic and the Antarctic [75].
  • Soils of arid or desert regions and rain effects
Temperature constrains are especially relevant in arid deserts, permafrost regions and high-altitude soils, where MOB survival often depends on dormancy or alternative substrates such as hydrogen and carbon monoxide [24]. However, these regions still demonstrate measurable CH4 uptake, suggesting that chemical processes, such as photocatalysis, radical reactions initiated by UV radiation or adsorption-driven oxidation on mineral surfaces may also be active.
Some papers show that, contrary to forests [33], rain enhances the CH4 sink in deserts or arid regions [27,94], which could indicate that Fenton reactions may occur, as there is hydrogen peroxide (H2O2) in rain droplets as well as in moisture condensation. In hyper-arid soils (where there is less than 20 mm of rain by year) no CH4 uptake and no methanotrophs or in situ activity could not be detected in the soil crust, which is the biologically most important layer in desert soils, but CH4 uptake was observed in the arid region (where it rains about ∼90 mm year−1) [95]; therefore, biological uptake of AM does not occur in all regions and types of soil. The hypothesis of additional potential energy sources for atmMOB such as sugars, short-chain fatty acids and alcohols present in soils has still to be confirmed [96], but dissolved organic carbon seems to regulate biological CH4 oxidation in semi-arid soils [94], which might also indicate the possibility of radical reactions paths for soil AMR.
In high-altitude desert soils, in all soil types, the CH4 concentrations decreased to less than 1 ppm with soil depth, but the rates of decline with soil depth increased with elevation [97], which seems to indicate that, as UV increases with altitude, UV-induced processes might also intervene.
  • Soil pH modification
Modifying the pH of acidic soils with application of lime (CaO) sometimes increases soil CH4 uptake [98,99] but sometime enhances methanogenesis and CH4 emissions [100]. Meanwhile, dolomite (CaMg(CO3)2) application enhances soil CH4 uptake by up to 15 times [101]. Since the concentration and the growth rate of atmMOB are quite low, such a large increase in AMR appears to be consistent with both an enzymatic and a catalytic process, as dolomite contains many trace metal contaminants.
  • Kinetic isotope effect
In addition to the isotopic composition of CH4 [102,103], one analytical technique used is based on the fact that 12CH4 reacts more quickly than 13CH4, leaving the remaining CH4 enriched in 13C and producing CO depleted in 13C; this is named the carbon kinetic isotope effect (KIE) for CH4 [104]. The 13C/12C KIE of OH oxidation of CH4 is around 5.4‰ at room temperature [105], while the KIE for Cl oxidation is 66‰ [106], and the KIE for soil systems hosting high-affinity methanotrophs is about 1.22‰ [107]. The larger the fraction of CH4 is oxidized by Cl, the larger the depletion of 13C in CO.
The KIE patterns observed in grassland and temperate forest soils differ from those in other known CH4 sinks, suggesting the coexistence of microbial and abiotic oxidation mechanisms [108,109]. These results derived from static flux chamber experiments, indicate that microbial oxidation alone may not account for the observed isotopic signatures and point to potential contributions from radical-based chemical reactions.
An accurate assessment of the relative importance of each process to the total KIE requires confirmation that significant partitioning of 13CH4 and 12CH4 occurs in pore spaces as a result of differences in diffusion rates [107].
  • Diffusion rates in soils
Several researchers [47] have reviewed extensive field measurements of CH4 uptake across various types of ecosystems and have developed process-based biogeochemistry models to simulate soil CH4 uptake [25,26,28,47,110,111,112,113,114]. These models incorporate biogeochemical controls that vary by geographic region, climate zone, ecosystem type and soil characteristics, including texture, density, porosity and organic matter content. Among these impact factors, gas diffusivity is considered the most important factor for CH4 and oxygen (O2) transport under the upper layers of the soil [47], together with soil moisture [33] and microbial oxidation processes [110].
Although it has been calculated that by using “optimal values for all factors,” the maximum intensity of CH4 consumption by natural soils is ≈0.39 mg·m−2·h−1 (9.36 mg CH4·m−2·d−1) [115], and it has been experimentally observed that “from the Amazon floodplain to the Arctic, the fastest rates rarely exceed 6 mg CH4·m−2·d−1” [116], with the observation that “CH4 oxidation by aerobic upland soils is rarely higher than 2.4 mg CH4·m–2·d–1” [117]. In some publications, the capacity of the MOB to consume CH4 exceeds its potential to diffuse from the atmosphere to the consumers [12]. Additional evidence comes from laboratory and modelling studies. Methylotrophic bacteria have been shown to consume CH4 at rate exceeding atmospheric diffusion limits. In one experiment, CH4 concentrations dropped from 10 ppm to 0.14 ppm within 8 h, demonstrating an apparent imbalance between CH4 supply and consumption capacity [12,13]. Similarly, a modelling study demonstrated that increasing gas diffusivity by a factor of ten could nearly triple global soil CH4 uptake, emphasizing the importance of physical gas transport and its potential to mask underlying oxidation mechanisms [25].
  • Diffusion rates in subsurface environments
Cave and karst systems offer a unique natural laboratory for exploring non-microbial AMR. Several studies have documented CH4 oxidation rates in subterranean environments that are equal to or greater than those found in forests, grasslands and tundra ecosystems [48,118,119]. For instance, a comprehensive study of seven Spanish caves, encompassing over 1000 air samples, revealed rapid CH4 depletion under near-saturated humidity and minimal microbial presence [14]. Although initial theories pointed to radon decay-induced radical formation, subsequent research found little supporting evidence [120]. Instead, these observations imply that abiotic processes, possibly driven by surface catalysis or atmospheric ion chemistry, may play a significant role. Similar patterns have been reported in over 20 cave systems, calling attention to enclosed air-rock interfaces in global CH4 budget.
  • Models and AMR attribution to methanotrophic activity
The first evidence of CH4 consumption at atmospheric concentrations by methanotrophs was published only in 2000 [121], while the first imputation of an AM sink to a biological phenomenon was reported in dry marshes in 1982 [122].
Quite often, even in recent publications, only the soil methane uptake was measured, without formal identification of methanotrophs, or metagenomic studies of the soil-microbiome compositions, or any other biological analysis. Some articles do not mention methanotrophy [27,43,97,123], while others do, even global inventories of the soil CH4 sink that review hundreds of published papers and thousands of records of the CH4 uptake by soils worldwide [47,124], even soil methanotrophy models [26], and even the “global methane budget” [3]. These papers seem to assume that CH4 removal from soils is solely due to microorganisms although this is not mentioned in many of the publications they cite.
We noticed that the parameters from publications about methanogenesis (CH4 production) such as in [125,126] are used in a process-based biogeochemistry model of key drivers of soil methanotrophy (CH4 consumption) [127], and that the linear rate constants used by three models of the soil methane sink vary by a factor of about 50 [26,28,110].
AM uptake in soils has been characterized by a low apparent Michaelis kinetic constant which fails to provide information about the rate of substrate uptake at atmospheric CH4 concentrations [9] and is an imprecise measure of substrate affinity as the conditions of the enzyme reaction are not fully known [96].
According to [128], there is a high structural uncertainty in models, which means that more empirical data on soil sinks (and sources) are needed globally, in order to develop complex models to simulate soil water content and shifts from soil sinks to sources.
  • The scale of global air-soil exchange
To contextualize the potential impact of soil AMR and its global implications and magnitude, we can consider the scale of global air-soil exchange. The Earth’s surface spans approximately 510 million km2, of which about 149 million km2 is land. By excluding the principal methane-emitting environments such as natural, impacted and human-made aquatic ecosystems such as lakes, ponds, hydroelectric reservoirs, wetlands and rice paddies [129,130,131], and based on literature estimates of global liquid freshwater covering about 1% (1.5 million  km2) of Earth’s surface and global wetland extents ranging broadly from 5  to 13  million  km2 [132], a total area of 134.5 to 142.5 million km2 of land (including Greenland and Antarctica) can be AM soil sinks. Assuming soils remove ~5% of AM, an estimated 7.5 × 1016 m3 of air may interact annually with the upper 15–20 cm of soil. This means that at least 528 to 558 m3 of air passes through each square metre of soil every year. With ambient CH4 concentrations around 1.9 ppm, this equals approximately 1.25 mg CH4 per m3 of air and about 680 mg CH4 per m2 of land surfaces. These numbers highlight the large but complex capacity of soil to remove AM through both biological and non-biological mechanisms. However, these back of the envelope globally scale estimations have several limitations, since they are based on simplified assumptions, such as uniform soil depth, constant air-soil exchange rates and steady CH4 concentrations, which do not reflect the high variability of soils across different areas.
Many of the reported fluxes come from chamber or incubation studies, which may not fully represent long-term conditions [25,33]. Moreover, the relative roles of microbial oxidation processes remain uncertain [10,11]. Therefore, these estimates should be considered as indicative rather than conclusion, highlighting the need for additional filed measurements and more robust process-based models.
The main findings of this perspective study are that there are still many research needs and that the soil CH4 sink can probably be revisited, for instance, by validating (or not) the H2O2 and Fenton-type reaction hypothesis due to rain, as well as by performing for each soil and climatic zone type, experiments with ‘killed controls’ such as by using a mixture of the broad-spectrum antibiotics, which can demonstrate that all the AM soil uptake is only reduced or on the contrary completely inhibited, possibly helping decide between that none or a significative important part the AM uptake is abiotic or only biological [133].

6. Conclusions and Future

Soil accounts for approximately 5% of the global atmospheric CH4 sink. This contribution appears modest, but it may involve diverse mechanisms, including both biological and chemical domains. While MOBs remain the dominant biological sink, this paper synthesis specifically examined whether MOB alone can account for this global estimate and suggested additional non-biological processes may also drive CH4 oxidation in most soils. In addition to the microbiological activity, in this work, we addressed the possible role of chemical sinks to the about 10% of AM sink attributed to soils and trees, since growing evidence suggests that non-biological oxidation processes may also contribute to soils AMR, including photocatalysis, Fenton-like reactions, chlorine radical and free radical chemistry, and O3/VOCs-driven reaction, also contribute under specific environmental conditions. Quantifying the plausible contribution of chemical sinks in soils relative to the microbial sink by analytical experiments is out of the scope of this perspective article and is the principal limitation, but it highlights one of the principal research needs. But these chemical pathways may be particularly relevant in soils enriched in metal oxides, exposed to strong UV radiation or influenced by salinity and atmospheric deposition. Subterranean systems, mineral surfaces and canopy structures may further expand the spatial domains where such processes occur.
The novelty of this perspective article is to provide an integrated framework that considers both microbial and chemical processes as contributors to the soil CH4 sink while changing the traditional perspective of a simply MOB-driven CH4 oxidation process.
The relative importance of these mechanisms likely varies across soil types, climatic zones and land use systems. Therefore, future research should aim to incorporate both microbial and chemical reaction pathways into unified conceptual and modelling frameworks. Field-experiment validation of chemical CH4 removal will be essential, particularly in chloride-rich, metal-abundant soils and UV intense regions. For instance, in order to verify if one or more chemical reactions are involved, in case several mechanisms are in action simultaneously, or in case of too large incertitude on the measurements, the hydrogen/deuterium KIE of the isotopologues (for instance 13CH4, 12CH3D, 13CH3D, etc.) can be used [134], which requests high-resolution isotope tracing [135], as the isotopic composition of CH4 also depends on its origins: plants, microbes, fossil fuels, abiotic processes [136]. Verifying the hypothesis of Cl radical reaction could be the easiest, as the 13C content of CO can thus be used as a sensitive indicator by making use of the strong carbon KIE in the CH4 + Cl reaction [106]. Incorporating these mechanisms into global modelling could improve the representation of soil CH4 sinks in the global CH4 budget and provide a stronger scientific route for greenhouse gases mitigation and remediation strategies. Advanced techniques such as high-resolution isotope KIE tracing, in situ detection of reactive intermediates and multi-omics approaches can offer deeper insights into how these pathways operate under dynamic environmental conditions.
Moreover, understanding the potential synergy between microbial and chemical processes remains a promising direction. Quantifying their relative contributions will not only refine out representation of CH4 sinks in Earth system models but also identify new leverage points for mitigating atmospheric CH4. However, this perspective study is limited by the lack of direct in situ evidence for abiotic CH4 removal, such as field measurement or laboratory incubation studies as well as the uncertainty of scaling local processes to the global level. As atmospheric CH4 levels are currently increasing by approximately 10 ppb per year, there is an urgent need to improve our understanding of soil-based CH4 removal to address uncertainties through coordinated field studies and interdisciplinary research. As emphasized by the U.S. National Academies [1], such knowledge is essential for refining global CH4 budgets and informing climate mitigation strategies.

Author Contributions

W.L., T.M. and R.d.R. contributed to the design and structuration of the study. Bibliography, data collection and analysis were performed by X.Y., T.T. and R.d.R. R.d.R. wrote the first draft of the manuscript. All authors commented and improved the next versions of the manuscript. W.L. provided overall supervision. All authors have read and agreed to the published version of the manuscript.

Funding

This research was supported by the European Commission H2020 Marie S Curie Research and Innovation Staff Exchange (RISE) award (Grant No. 871998) and the National Key Research and Development Plan (Key Special Project of Inter-governmental National Scientific and Technological Innovation Cooperation, Grant No. 2019YFE0197500).

Data Availability Statement

No new data were created or analyzed in this study. Data sharing is not applicable to this article.

Conflicts of Interest

The authors declare no conflicts of interest.

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Land 14 01864 g001
Figure 1. Summary of CH4 removal mechanisms in soils.
Figure 1. Summary of CH4 removal mechanisms in soils.
Land 14 01864 g001
Table 1. Summary of CH4 removal rates by MOBs across different soil types based on literature reports.
Table 1. Summary of CH4 removal rates by MOBs across different soil types based on literature reports.
Soil TypeMethod TypeRemoval AmountConvertedRef
ForestField measurement (upscaling) 28.7   Tg   C H 4   yr − 1 28.7   Tg   yr − 1 [37]
ForestField measurement 1 − 3   mg   C H 4   m − 2   d − 1 1 − 3   mg   m − 2   d − 1 [34]
ForestField measurement 3.17   mg   C H 4 − C   m − 2   d − 1 4.23   mg   m − 2   d − 1 [5]
ForestField measurement 0.54   ±   0.06   nmol   m − 2   s − 1 0.75   ±   0.08   mg   m − 2   d − 1 [38]
AgricultureField measurement 265 − 444   g   C H 4   ha − 1 Lack of data for conversion[39]
DesertField measurement 0.24 − 4.38   mg   C H 4   m − 2   d − 1 0.24 − 4.38   mg   m − 2   d − 1 [27]
ArableField measurement 0.11   ±   0.06   nmol   m − 2   s − 1 0.152   ±   0.083   mg   m − 2   d − 1 [40]
GrasslandField measurement 0.19   ±   0.09   nmol   m − 2   s − 1 0.263   ±   0.124   mg   m − 2   d − 1 [40]
ArcticField measurement 1.83   ±   0.19   nmol   C H 4   g − 1   dw   d − 1 Lack of data for conversion[41]
ArcticField measurement 0.092   ±   0.011   mg   C H 4   m − 2   h − 1 2.21   ±   0.26   mg   m − 2   d − 1 [30]
Arctic (dry tundra)Field measurement 8.3   ±   3.7   µ mol   m − 2   h − 1 3.19   ±   1.42   mg   m − 2   d − 1 [42]
Arctic (moist tundra)Field measurement 3.1   ±   1.6   µ mol   m − 2   h − 1 1.19   ±   0.61   mg   m − 2   d − 1 [42]
ForestModelling 17.46 − 24.27   Tg   C H 4   yr − 1 17.46 − 24.27   Tg   yr − 1 [43]
GrasslandModelling 17 − 23   Tg   C H 4   yr − 1 17 − 23   Tg   yr − 1 [44]
AgricultureModelling 47 − 60   Tg   C H 4   yr − 1 47 − 60   Tg   yr − 1 [45]
Forest/ArableModelling 43.3   Gg   C H 4   yr − 1 0.0433   Tg   yr − 1 [46]
Global SoilModelling 22   ±   12   Tg   C H 4   yr − 1 22   ±   12   Tg   yr − 1 [47]
ForestLaboratory incubation 50.3 − 163.5   pmol   C   cm − 3   h − 1 Lack of data for conversion[4]
ForestLaboratory incubation 16.32 − 16.38   µ g   C H 4   m − 2   h − 1 0.392 − 0.393   mg   m − 2   d − 1 [32]
ArableLaboratory incubation 11.40 − 14.47   µ g   C H 4   m − 2   h − 1 0.274 − 0.347   mg   m − 2   d − 1 [32]
GrasslandLaboratory incubation 6.74 − 9.30   µ g   C H 4   m − 2   h − 1 0.162 − 0.2232   mg   m − 2   d − 1 [32]
Cave/SubterraneanLaboratory incubation 1.3 − 2.7   mg   C H 4   m − 2   h − 1 1.3 − 2.7   mg   m − 2   d − 1 [48]
Notes: MCH4 = 16 g   mol − 1 ; MC = 12 g   mol − 1 ; 1 nmol   C H 4   m − 2   s − 1 = 1.382 mg   C H 4   m − 2   d − 1 ; 1 µ mol   C H 4   m − 2   h − 1 = 0.384 mg   C H 4   m − 2   d − 1 ; 1 µ g   C H 4   m − 2   h − 1 = 0.024 mg   C H 4   m − 2   d − 1 ; 1 mg   C H 4 − C   m − 2   d − 1   =   1.333   mg   C H 4   m − 2   d − 1 .
Table 2. Potential mechanisms contributing to soil atmospheric methane removal.
Table 2. Potential mechanisms contributing to soil atmospheric methane removal.
MechanismApplicable Soil TypesNecessary ConditionsPotential ContributionEvidence Sources
MOB OxidationAll aerated soilsO2, ~20% moistureMajor[3,4,6,7]
PhotocatalysisDesert sands, Semi-arid soilsUV light, TiO2/ZnO/WO3 Low, local (arid regions)[27,49,50,51,52]
Fenton-like ReactionsForest soils, Desert soilsH2O2 (rain/fog), Fe/Cu, acidic pHModerate, wet soils[18,19,20,21,22,23,38,58,59,60]
•Cl PathwaysCoastal soils, Dust aerosolsChloride (e.g., FeCl3), UVHigh, coastal zones[70,71,72,73]
O3/VOC-Driven ReactionsForest soils, Aerated soilsO3 decomposition, NOx/VOCsLow, widespread[67,77,78,79,80,81,82,83]
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Yao, X.; Tao, T.; Li, W.; Ming, T.; de Richter, R. Do Soil Methanotrophs Really Remove About 5% of Atmospheric Methane? Land 2025, 14, 1864. https://doi.org/10.3390/land14091864

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Yao X, Tao T, Li W, Ming T, de Richter R. Do Soil Methanotrophs Really Remove About 5% of Atmospheric Methane? Land. 2025; 14(9):1864. https://doi.org/10.3390/land14091864

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Yao, Xiaokun, Tao Tao, Wei Li, Tingzhen Ming, and Renaud de Richter. 2025. "Do Soil Methanotrophs Really Remove About 5% of Atmospheric Methane?" Land 14, no. 9: 1864. https://doi.org/10.3390/land14091864

APA Style

Yao, X., Tao, T., Li, W., Ming, T., & de Richter, R. (2025). Do Soil Methanotrophs Really Remove About 5% of Atmospheric Methane? Land, 14(9), 1864. https://doi.org/10.3390/land14091864

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