Anaerobic oxidation of methane and its potential role in heavy metal(loid) speciation in wetland soils: occurrence, mechanisms and environmental implications

Abstract

Wetlands deliver essential ecological services but are increasingly threatened by heavy metal(loid)s (HMs) due to anthropogenic activities. The speciation of HMs, which dictates their toxicity and mobility, can be transformed by microbial redox processes like anaerobic oxidation of methane (AOM). However, the contribution of AOM to HM speciation transformation in wetland soils remains inadequately assessed. This review summarizes the current understanding of how AOM couples with the reduction of various elements to directly or indirectly affect HM speciation. These elements include arsenate, chromate, selenate/selenite, antimonate, vanadate, Fe(III), Mn(IV), and sulfate. We examine the responsible microorganisms and their electron transfer pathways, and evaluate the potential for applying these AOM processes in the remediation of HM contamination. These AOM processes can potentially influence the reduction, mobilization, or immobilization of HMs, thereby regulating their biogeochemical cycles in wetland soils. Future research priorities include determining the role of aerobic methanotrophs in these processes, clarifying the impacts of environmental conditions and HM forms, and developing targeted AOM regulation strategies for remediating contaminated wetlands. This work advances the mechanistic understanding of the interactions between HMs and AOM, and provides theoretical insights for developing remediation strategies for HM-contaminated wetland soils.

---

Wei Ye

Wei Ye received his BS in Biotechnology from Wuhan Institute of Technology, Wuhan, China, in 2023. He is currently pursuing his Master's degree in Biotechnology and Engineering at the same institution. His research focuses on anaerobic oxidation of methane and its environmental effects.

Lihu Liu

Dr Lihu Liu is an associate professor of Environmental Science at Wuhan Institute of Technology, with more than ten years of research experience in heavy metal(loid) biogeochemistry. He has published over 60 peer-reviewed articles in leading journals including Environmental Science & Technology, Water Research, and Geochimica et Cosmochimica Acta, focusing on heavy metal pollution and remediation.

Huaiying Yao

Dr Huaiying Yao is a professor at Wuhan Institute of Technology. His research specializes in soil element cycling and microbial mechanisms, particularly the regulatory factors and processes governing greenhouse gas emissions in typical environments, as well as the microbial-mediated distribution, migration, and transformation of typical pollutants. He has authored more than 200 scientific publications and is a recipient of the National Science Fund for Distinguished Young Scholars of China. He also serves on the editorial boards of Agriculture, Ecosystems & Environment, Soil Ecology Letters, and Journal of Soils and Sediments.

---

Environmental significance

Wetlands, critical ecosystems often described as the “kidneys of the Earth”, face increasing contamination by heavy metal(loid)s (HMs) due to rapid urbanization and industrialization. Anaerobic oxidation of methane (AOM) potentially influences HM speciation. Arsenate, chromate, selenate/selenite, antimonate, and vanadate can act as electron acceptors and be directly reduced during AOM. Moreover, AOM coupled with the reduction of Fe(III), Mn(IV), or sulfate can indirectly alter HM speciation, owing to the strong affinity of iron/manganese oxides and sulfides for HMs. A thorough understanding of these HM-associated AOM processes is essential for developing effective remediation strategies for contaminated wetland soils. This review systematically examines the occurrence, mechanisms, and environmental implications of HM-related AOM in wetland ecosystems, emphasizing the need to identify key microorganisms, quantify process rates, and clarify environmental controls under natural conditions.

1. Introduction

Anthropogenic activities, driven by rapid urbanization and industrialization, have caused substantial inputs of heavy metal(loid)s (HMs) into wetland ecosystems.¹ HM contamination poses multiple threats, including endangering plant and human health, and damaging ecosystem functions.^2–5 Over half of the world's internationally recognized wetlands are polluted by HMs.¹ Evidence from diverse regions highlights its severity. For instance, arsenic (As) levels exceed safety thresholds in coastal wetlands of China's Yellow River Delta.⁶ Extreme chromium (Cr) contamination (up to 3142 mg per kg total Cr) occurs in soils contaminated by tannery wastewater in Kanpur, India.⁷ Selenium (Se) bioaccumulation in aquatic insects is markedly higher from agricultural drainage areas (4.4 µg g⁻¹) than that from non-agricultural sites (1.5 µg g⁻¹) in South Dakota, USA.⁸

The bioavailability and mobility of HMs are primarily determined by their chemical speciation. Anaerobic oxidation of methane (AOM) is a widespread microbial redox process that occurs in wetland soils under hypoxic or anoxic conditions.⁹ In this process, methane (CH₄) is oxidized to carbon dioxide (CO₂) by transferring electrons to terminal acceptors such as nitrite (NO₂⁻), nitrate (NO₃⁻), sulfate (SO₄²⁻) or HMs.¹⁰ AOM plays a critical role in mitigating atmospheric methane levels, consuming up to 200 Tg annually and reducing potential emissions by over 50% in freshwater wetlands.^11,12 Moreover, AOM potentially affects HM speciation in wetland environments, according to the redox potential (Eh) and Gibbs free energy (Fig. 1). AOM can directly reduce HM oxyanions like chromate (Cr(VI)), arsenate (As(V)), antimonate (Sb(V)), and selenate/selenite (Se(VI)/Se(IV)), by utilizing them as electron acceptors.^13–17 In addition to these direct interactions, AOM indirectly influences HM mobility through coupled secondary processes. Given that Fe and Mn oxides can bind over 80% of total HMs in soils,¹⁸ AOM-driven Fe(III) or Mn(IV) reduction may promote the release of associated HMs, enhancing their bioavailability. In contrast, when AOM couples with sulfate reduction, the produced sulfide (S²⁻) can precipitate with HMs such as Zn(II), Pb(II), Cd(II), and As(III) to form insoluble sulfides. This process reduces their toxicity and mobility.^19,20 These dual roles highlight the critical yet complex function of AOM in regulating HM speciation and behavior in wetland soils.

Fig. 1 Standard electrode potentials for selected redox couples and the corresponding Gibbs free energy of the related reactions. Potential and Gibbs free energy values are adopted from ref. 13, 15, 21, 26, 47, 54 and 127–131. The standard electrode potential is the equilibrium potential relative to the standard hydrogen electrode (with a potential set at 0 V) under standard conditions (gas pressure of 101.325 kPa, metal ion activity of 1.0, and temperature of 25 °C).¹³²

Previous reviews have focused on the reaction mechanisms of AOM itself and the influence of aerobic oxidation of methane (AeOM) on HM speciation.^21,22 Karthikeyan et al.²² summarized the reaction mechanisms and key enzymes involved in AeOM, alongside advances in the speciation transformation of toxic HMs such as Cr, Se, and mercury (Hg), highlighting its notable influence on HM biogeochemistry.²² However, in many environments, AOM contributes more substantially to methane consumption than aerobic oxidation. For instance, AOM accounts for approximately 90% and over 50% of methane consumption in marine and wetland ecosystems, respectively.^11,23 Evidence has suggested that AOM can directly or indirectly influence HM fate in water.^14,24,25 For instance, Dang et al.²⁶ examined the direct reduction of HMs coupled with methane oxidation in wastewater and groundwater systems under aerobic, microaerobic, and anaerobic conditions. The review outlined key microorganisms, electron transfer mechanisms, and the potential of methane-driven processes in HM removal and recovery, further corroborating the role of AOM in shaping HM speciation.²⁶ However, the specific contribution of AOM to HM biogeochemical cycling in wetlands has not been thoroughly evaluated.

Here, we summarize current knowledge on the transformation of HM speciation associated with AOM processes in wetland soils. We focus on coastal wetlands, freshwater wetlands, paddy soils, and constructed wetlands, which are recognized as critical environments where both AOM activity and HM accumulation often coincide.^9,27,28 The content of this review encompasses: (i) AOM coupled directly with HM reduction and AOM with Fe(III), Mn(IV), and sulfate as electron acceptors; (ii) the key microorganisms and underlying mechanisms, including electron transfer pathways and functional genes; (iii) the effect of these AOM processes on HM speciation, and (iv) the environmental factors that regulate these AOM processes in wetland systems.

2. Mechanisms and electron transfer pathways of HM-associated AOM

2.1 Mechanisms

Two main mechanistic models have been proposed for HM-associated AOM:¹⁴ (i) the independent model, in which anaerobic methanotrophic archaea (ANME) conduct both methane oxidation and HM reduction alone; and (ii) the syntrophic model, in which ANME oxidize methane through reverse methanogenesis and shuttle electrons to bacterial partners for HM reduction. The reverse methanogenesis pathway serves as the core metabolic route in AOM.²¹ Essentially the reverse of the methanogenesis process, this pathway oxidizes CH₄ to CO₂ by reversing the catalytic direction of methanogenic enzyme systems, while transferring electrons to terminal acceptors to generate energy (Fig. 2).²¹ The process begins with the conversion of methane to CH₃-CoM by methyl-coenzyme M reductase (Mcr).²⁹ The methyl group is then transferred to tetrahydromethanopterin (H₄MPT), forming CH₃–H₄MPT, which is subsequently converted to CH₂=H₄MPT under the catalysis of N⁵,N¹⁰-methylene-H₄MPT reductase (Mer).²⁹ Through further enzymatic steps involving dehydrogenation, hydrolysis, and group transfer, CH₂=H₄MPT is ultimately oxidized to CO₂ (Fig. 2).²⁹ Three possible electron transfer pathways are involved in the reverse methanogenesis process: (i) heterodisulfide reductase (HdrDE): HdrDE oxidizes the coenzyme M (HSCoM)/coenzyme B (HS-CoB) and transfers electrons to redox shuttles such as 9,10-anthraquinone-2,6-disulfonate (AQDS), thereby driving HM reduction; (ii) F₄₂₀H₂ dehydrogenase (Fpo): Fpo transfers electrons derived from F₄₂₀H₂ to the electron transport chain, which ultimately reduces HMs; (iii) multi-heme cytochromes (MHCs): MHCs mediate extracellular electron transfer, directly reducing HMs outside the cell (Fig. 2).^30,31

Fig. 2 Reverse methanogenesis pathway of HM-associated AOM. HMRM: heavy metal(loid)-reducing microorganisms; Mcr: methyl coenzyme M reductase; Mtr: tetrahydromethanopterin S-methyl-transferase; Mer: N⁵,N¹⁰-methylene-H₄MPT reductase; Mtd: F₄₂₀H₂-dependent methylene-H₄MPT dehydrogenase; Mch: N⁵,N¹⁰-methenyl-H₄MPT cyclohydrolase; Ftr: formylmethanofuran (CHO-MFR)-H₄MPT formyltransferase; Fmd: CHO-MFR dehydrogenase; MHCs: multi-heme cytochromes; Hdr: heterodisulfide reductase; Fpo: F₄₂₀H₂ dehydrogenase; ATPase: ATP synthase; FdX_O: oxidized ferredoxin; FdX_R: reduced ferredoxin; CoB: Coenzyme B; CoMS-SCoB: Coenzyme B-coenzyme M heterodisulfide.

2.2 Electron transfer

Extracellular electron transfer (EET) is essential for HM-associated AOM, as it allows methane-oxidizing microorganisms (MOMs) or their syntrophic partners to deliver electrons from intracellular methane oxidation to extracellular electron acceptors.^26,32 As illustrated in Fig. 3, EET mechanisms can be broadly divided into direct and indirect pathways.²¹ In direct EET, MHCs exposed on the membrane of ANME facilitate electron transfer through an “electron hopping” mechanism, in which iron ions within heme centers relay electrons from the cell to external acceptors.⁹ Some microorganisms also produce conductive nanowires, wire-like appendages that form physical bridges for electron transport to extracellular substrates or between cells.³² Indirect EET relies on redox-active compounds such as humic substances and AQDS, which function as electron shuttles by reversibly accepting and donating electrons.²⁶ Additionally, MOMs can generate soluble metabolic intermediates (e.g., hydrogen, formate, acetate) that serve as electron donors for HM-reducing microorganisms (HMRMs).²¹

Fig. 3 Proposed extracellular electron transfer mechanisms of HM-associated AOM.

3. Direct transformation of HM speciation associated with AOM

As early as 2008, Caldwell et al.³³ proposed that certain HMs, such as Fe(III), Mn(IV), As(V), and Se(IV/VI), could potentially act as electron acceptors for AOM. Subsequent research has validated a broader range of HMs in this role, significantly increasing interest in these coupled processes. This section therefore summarizes recent advances in AOM directly linked to the reduction of As(V), Cr(VI), Se(IV/VI), Sb(V), and V(V) in wetland soils. We discuss the consequent effects on the speciation, toxicity, and mobility of the related HMs. Table 1 presents the key functional microorganisms, process rates, and study settings involved in these processes.

Table 1 Key functional microorganisms, rates, and settings involved in AOM-AsR, AOM-CrR, AOM-SeR, AOM-SbR and AOM-VR

Type	Culture medium	Mainly functional microorganisms	HM reduction rate (µmol L^−1 d^−1)	Ref.
AOM-AsR	Wetland soil	ANME-2a–c, Sulfurospirillum and Geobacter	9.97 ± 0.71	14
	Paddy soil	ANME, Methylobacter, Methylobacter and/or As(V) reducing microbe	0.128	122
	Paddy soil	ANME-2d and/or As(V) reducing microbe	0.176	35
	Paddy soil	ANME-2a–c, ANME-2d, ANME-3 and Geobacteraceae	0.365	36
AOM-CrR	Sludge	ANME-2d	5.78	13
	Sludge	Ca. Methanoperedens and/or Cr(VI) reducing microbe	≥6.0	39
	Mineral medium	Methanobacterium, Methanosarcina, Ca Methanoperedens and Meiothermus	5.38	40
	Trace element solution	Methanothrix, Methanomicrobiales, Methanosarcina, Methanobacterium and Methanoculleus	296.7 ± 18.7	41
	Mineral medium	Ca Methanoperedens and Ca Methanoperedens	90 ± 4	45
AOM-SeR	Mineral medium	Methylomonas, Methylophilus, Methylocystis, Comamonadaceae and Azospira	12 ± 0.5	50
	Mineral medium	Ca. Methanoperedens, Ca. Methylomirabilis and/or Se(VI) reducing microbe	12.4	16
	Mineral medium	Methylococcus, Ca. Methanoperedens, Ca. Methylomirabilis and Denitratisoma	8.23	47
	Freshwater wetland sediment	Methanosarcina, Methylocystis, Denitratisoma and SR-FBR-L83_norank	28.9 ± 0.5	48
	Freshwater wetland sediment	Methylocystis, Pseudoxanthomonas, Piscinibacter and Rhodocyclaceae	1.15 ± 0.04	49
AOM-SbR	Mineral medium	Methanosarcina, Methylomonas and Methylophilus	51.25	15
	Mineral medium	Methylomonas, Solibacteres, Chitinophagaceae, Chitinophagaceae and Sporomusacea	10.2	53
AOM-VR	Mineral medium	Methanosarcina and Methanobacterium	12.96	54
	Sludge	Methanobacterium, Methylomonas, Stenotrophomonas and Steroidobacter	321.4 ± 17.9	25

---

3.1 As(V) reduction

The reaction for AOM-coupled arsenate reduction (AOM-AsR) is presented in eqn (1) (Fig. 1).²¹ Key microorganisms identified in this process include various archaeal groups (e.g., ANME-1, ANME-2a–c, ANME-2d, and ANME-3) and bacteria such as Sulfurospirillum and Geobacter.^14,34–36 Evidence from both laboratory and field studies confirms the occurrence of AOM-AsR in wetland soils. For instance, Shi et al.¹⁴ conducted ¹³CH₄ incubation experiments with As-contaminated paddy soils from seven provinces across China. Their results confirmed the coupling of AOM with As(V) reduction, with observed rates approaching the theoretical stoichiometry of 4 : 1 (As : methane).¹⁴ Field evidence from paddy soils shows increasing total As and As(III) concentrations in porewater over time, with ¹³CH₄ treatments exhibiting higher levels than controls.³⁵

Metagenomic and polymerase chain reaction (PCR) analyses have confirmed the transcriptional activity of key functional genes during AOM-AsR, including mcrA (for reverse methanogenesis) and arrA (encoding respiratory arsenate reductase (ArrAB)).¹⁴ This genetic evidence suggests that ANME oxidize methane through the mcrA-dependent reverse methanogenesis pathway, possibly shuttling electrons extracellularly through MHCs.³⁷ Concurrent As(V) reduction is likely mediated through the ArrAB pathway.¹⁴ Based on current evidence, it is speculated that AOM-AsR could be conducted solely by ANME or through a syntrophic partnership (Fig. 4). However, further research is needed to validate the precise mechanism with highly enriched cultures or pure strains.

Fig. 4 Metabolic mechanisms of AOM-AsR, AOM-CrR, AOM-SeR, AOM-SbR, and AOM-VR. MOM: methane-oxidizing microorganism; HMRM: heavy metal(loid)-reducing microorganism.

AOM-AsR reduces As(V) to the more mobile and toxic As(III), thereby mobilizing As in wetland soils.³⁸ Isotopic tracing and incubation experiments have quantified the significance of this process. For instance, AOM-AsR was found to account for up to 49.2% of As release in cultured paddy soils.¹⁴ Tracer studies further demonstrated its contribution to be 80.8% under laboratory conditions and 33.9% in field settings.³⁵ Given its major role in regulating As bioavailability, strategies that decouple AOM from As(V) reduction are promising for remediating contaminated wetlands.

Multiple interventions can effectively decouple AOM from As(V) reduction, thereby alleviating As mobilization. The application of a passivator containing Fe oxide and CaSO₄ reduced dissolved As by 78.0% while increasing the AOM rate by 55.0% in contaminated paddy soil, primarily by impeding electron transfer from methane to As(V).³⁴ Low-molecular-weight organic acids (LMWOAs; e.g., citric, oxalic, and acetic acid) exuded by rice roots can also inhibit AOM-AsR. Their addition reduced porewater As(III) by 35.1–65.7% within 14 days in anaerobic incubations, likely through suppressing the activity of Candidatus “Methanoperedens” (ANME-2d) and arrA-associated microorganisms.³⁶

3.2 Cr(VI) reduction

AOM-coupled Cr(VI) reduction (AOM-CrR) is shown in eqn (2) (Fig. 1).¹³ Reported microorganisms involved in this process include ANME-2d, Methanobacterium, Methanosarcina, Methanothrix, Methanomicrobiales, Methanoculleus, and Meiothermus.^39–41 Under certain conditions, methanogens may perform “trace methane oxidation” to generate organic metabolic intermediates that serve as substrates for other microorganisms.²⁹ Additionally, some aerobic MOMs (e.g., Methylosinus and Methylocystis) might participate in AOM-CrR by potential cooperation with Cr(VI)-reducing bacteria.^41–43 Although evidence comes largely from laboratory systems, it confirms that CH₄ can serve as the sole electron donor for Cr(VI) reduction, supporting the occurrence of AOM-CrR in wetlands.^13,44

Similar to AOM-AsR, two primary mechanisms have been proposed for AOM-CrR (Fig. 4). Dong et al.⁴⁰ cultivated a Cr(VI)-reducing biofilm under anaerobic conditions in a methane-based membrane biofilm batch reactor (MBBR). Microbial community analysis revealed a predominance of methanogens (Methanobacterium and Methanosarcina) and the Cr(VI)-reducing genus Meiothermus, along with the presence of ANME-2d archaea.⁴⁰ Based on these findings, they proposed that these archaea anaerobically oxidize CH₄ and subsequently transfer the electrons to Meiothermus species for Cr(VI) reduction.⁴⁰ In contrast, other research indicates that ANME-2d can reduce Cr(VI) independently. Proteomic analysis revealed a significant upregulation of MHCs, pointing to a dominant extracellular Cr(VI) reduction pathway; meanwhile, the upregulation of a formate-dependent nitrite reductase suggested a possible intracellular route.⁴⁵ Extracellular respiration was estimated to account for approximately 94.0% of the Cr(VI) reduction.⁴⁵ These studies have successfully demonstrated key mechanisms and exhibited a high potential for Cr pollution remediation.¹³ However, the rates and relative contributions of AOM-CrR in natural wetlands, where multiple electron acceptors compete, remain largely unknown.

AOM-CrR is influenced by multiple factors, with electron acceptor competition playing a dual regulatory role.^43,44 Low concentrations of SO₄²⁻ promote Cr(VI) reduction, partly due to the involvement of certain sulfate-reducing bacteria (SRB).⁴³ Thermodynamically, the Cr(VI)/Cr(III) couple has a higher Eh than the SO₄²⁻/S²⁻ couple (Fig. 1), favoring electron flow to Cr(VI).⁴³ However, high concentrations of SO₄²⁻ inhibit Cr(VI) reduction by competing for electron donors.⁴³ Furthermore, H₂S, the product of SO₄²⁻ reduction, can inhibit the activity of methane-oxidizing and other microorganisms, further lowering Cr(VI) bioreduction efficiency.⁴⁴ The addition of Fe(III) can enhance AOM-CrR. In a methane-driven microbial fuel cell-granular sludge system, Fe(III) increased the Cr(VI) removal rate from 16.5% to 46.3%.⁴⁶ This improvement is mainly attributed to Fe(III) acting as an electron acceptor to stimulate CH₄ oxidation, the resulting Fe(II) then abiotically reduces Cr(VI).⁴⁶

By converting toxic Cr(VI) to less toxic, bioavailable, and mobile Cr(III), AOM-CrR can alleviate Cr impacts in wetlands. However, investigations into its application in natural wetland settings remain limited. Examples include a membrane biofilm reactor (MBfR) achieving more than 95% Cr(VI) removal using methane,⁴² switching from nitrate to Cr(VI) as an electron acceptor in an MBfR after long-term enrichment,³⁹ and a bioelectrochemical system coupling Cr(VI) reduction with current generation (99.2% Cr(VI) removal).⁴¹ Natural wetlands present greater complexity due to fluctuating redox conditions and competing electron acceptors. Despite this, enhancing AOM-CrR remains a promising strategy for mitigating Cr pollution in wetland soils.

3.3 Se(IV) and Se(VI) reduction

The reactions for AOM coupled with selenate (SeO₄²⁻) and selenite (SeO₃²⁻) reduction (AOM-SeR) are shown in eqn (3) and (4), respectively (Fig. 1).⁴⁷ Microorganisms implicated in AOM-SeR include ANME-2d, Candidatus “Methylomirabilis”, Methanosarcina, Denitratisoma, SR-FBR-L83_norank, Pseudoxanthomonas, Piscinibacter, and Rhodocyclaceae.^16,48,49 Furthermore, some aerobic MOMs (e.g., Methylomonas, Methylococcus) may also participate in this process.^47,48,50

Both direct and indirect mechanisms are involved in AOM-SeR (Fig. 4). Direct reduction is suggested by studies where nitrate-reducing AOM cultures switched to selenate reduction in a MBfR.¹⁶ Chemical analyses confirmed the conversion of soluble selenate to nanoparticulate Se⁰.¹⁶ 16S rRNA gene sequencing revealed the persistence of only AOM-associated archaea like Candidatus “Methylomirabilis” and ANME-2d, with no known specialist selenate-reducers detected.¹⁶ In contrast, the indirect mechanism involves syntrophic partnerships. For example, in a selenite-reducing system traced with ¹³CH₄, the typical AOM archaea (ANME-2d and Candidatus “Methylomirabilis”) declined while the aerobic genus Methylococcus became dominant, likely adapting via syntrophy.⁴⁷ Another study described a defined consortium where the archaeon Methanosarcina oxidized CH₄ through reverse methanogenesis, providing electrons to selenate-reducing bacteria. Meanwhile, Methylocystis produced organic acids (acetate, formate) that fueled heterotrophic reducers like Pseudoxanthomonas and Piscinibacter.^48,49

These microbial interactions are underpinned by specific functional genes. Genomes of ANME-2d and Candidatus “Methylomirabilis” encode MHCs for extracellular electron transfer.¹⁶ ANME-2d exhibits metabolic versatility, coupling AOM to the reduction of various acceptors including nitrite, nitrate, sulfate, and Fe/Mn.^9,51,52 The presence of the narG gene (encoding the α subunit of Nar) in Methylomonas supports a potential direct link between methane oxidation and selenate reduction.⁵⁰ Moreover, genera like Pseudoxanthomonas, Piscinibacter, and Rhodocyclaceae carry genes for periplasmic nitrate reductases (Nap), which are essential for selenate respiration.⁴⁹

Several factors influence AOM-SeR efficiency. Thermodynamically, selenate is a more favorable electron acceptor than selenite, though some species only perform partial reduction.⁴⁸ A major practical challenge is self-inhibition due to the accumulation of insoluble Se⁰ precipitates, which can coat cells and biofilms, hindering electron transfer.^16,47,48 This is a key common problem that may need to be overcome in its practical applications.⁴⁸ Additionally, the inherent toxicity of Se species to microorganisms and competition from other electron acceptors (e.g., nitrate) may further suppress activity.⁴⁷ Therefore, strategies such as regulating microbial community structure and separating Se⁰ precipitates could enhance AOM-SeR efficiency, offering a promising approach for mitigating Se contamination in wetland soils in the future.

3.4 Sb(V) reduction and V(V) reduction

AOM coupled with antimonate (Sb(OH)₆⁻) reduction (AOM-SbR) is described by eqn (5) (Fig. 1).¹⁵ Research has relied mainly on laboratory enrichments, identifying key microorganisms such as Methanosarcina, Methanolobus, Solibacteres, Chitinophagaceae, and Sporomusacea.^15,53 Aerobic MOMs (e.g., Methylomonas, Methylophilus) may also be involved.¹⁵

The proposed mechanism of AOM-SbR involves methane-oxidizing archaea carrying out reverse methanogenesis. The resulting organic metabolic intermediates can serve as electron donors for Sb(V) reduction by associated heterotrophic bacteria (Fig. 4). This syntrophic coupling was demonstrated in an MBBR biofilm enriched for 12 months with methane and Sb(V), where Sb₂O₃ was identified as the main reduction product.¹⁵ In the MBBR biofilm, although the relative abundances of aerobic MOMs Methylomonas and Methylophilus declined during AOM-Sb(V), they remained a substantial component of the biofilm community.¹⁵

AOM coupled with vanadate reduction (AOM-VR) follows eqn (6) (Fig. 1).⁵⁴ Observed in river sediments and anaerobic sludge, key microorganisms include Methanosarcina, Methanobacterium, Stenotrophomonas, Steroidobacter, Anaeromyxobacter, Geobacter, Methylococcus, and Methylomonas.^25,30,54

In one bioreactor study, AOM-VR produced a V(IV)-containing precipitate and was linked to the enrichment of Methanobacterium (85.2%) and the production of volatile fatty acids (e.g., acetate).²⁵ This supports a hypothesis where Methanobacterium oxidizes methane to formate through reverse methanogenesis, with heterotrophs then using these intermediates to reduce V(V).²⁵ Independent V(V) reduction by methane oxidizers is also plausible. Yan et al.³⁰ demonstrated V(V) bioreduction using acetate as the sole electron donor in V(V)-contaminated river sediments. DNA stable isotope probing (DNA-SIP) and metagenomic analysis identified Methanosarcina, Anaeromyxobacter and Geobacter (carrying narG and napA genes) as potential V(V)-reducers.³⁰Methanosarcina possesses genes for both reverse methanogenesis and metal reduction, suggesting its potential to oxidize methane through reverse methanogenesis and directly reduce V(V).³⁰

¹³CH₄ tracing microcosm incubation experiments of Sb-contaminated paddy soil showed that Sb(V) addition strongly inhibited AOM (about 97.3%).⁵⁵ Sb may reduce microbial diversity and impair key syntrophic partnerships, disrupting electron transfer.⁵⁶ This highlights a major obstacle for AOM-SbR in natural settings. Research on both AOM-SbR and AOM-VR remains preliminary, with a weak evidence base and a notable lack of in situ validation. The strong inhibition of AOM by Sb(V) suggests that AOM-SbR may not occur spontaneously in highly contaminated sites. Therefore, evaluating the environmental significance of these processes requires priority investigation into the concentration thresholds at which pollutants inhibit them.

4. Indirect transformation of HM speciation associated with AOM

Given the strong affinity of Fe/Mn oxides and sulfides for HMs, this section comprehensively reviews recent advances in AOM coupled with the reduction of Fe(III), Mn(IV), and sulfate in wetland soils. It also examines the potential implications of these processes for HM speciation. Details on the key microorganisms, reaction rates and environmental settings are summarized in Table 2.

Table 2 Key functional microorganisms, rates, and settings involved in Fe-AOM, Mn-AOM, and S-AOM

Type	Culture medium	Mainly functional microorganisms	Activity (µmol L^−1 d^−1)	Ref.
Fe-AOM	Freshwater river sediment	Methanosarcinales	7.4–98.6	76
	Freshwater lake sediment	Ca. Methanoperedens ferrireducens	2.74–8.35	82
	Paddy soil	Methanoperedenaceae, Geothermobacter and Proteiniclasticum	8.11 ± 0.44	118
Mn-AOM	Constructed wetland	ANME-2d, Deferrisoma, Anaeromyxobacter and Geobacter	167.5	100
	Freshwater sediment	Ca. Methanoperedens manganicus and Ca. Methanoperedens manganireducens	44.5	98
	Mineral medium	Ca. Methanoperedens manganicus	25–49	108
S-AOM	Paddy soil	Methanoperedenaceae and Desulfobulbaceae	6.9 ± 0.3	118
	Freshwater sediment	ANME-2d	0.99 ± 0.13	77

---

4.1 Fe-mediated AOM

Fe is abundant and widely distributed in the Earth's crust, averaging about 3.5% in global soils.⁵⁷ In natural environments, Fe occurs primarily as hydroxides, hydrated oxides, and oxides, collectively termed Fe oxides.⁵⁸ Wetland soils typically exhibit high Fe contents. For example, total Fe contents range from 10.1 to 56.1 g kg⁻¹ in paddy soils of southern China,⁵⁹ and from 229 to 414 g kg⁻¹ in flocs from Icelandic wetlands.⁶⁰ Fe oxides possess high specific surface areas and abundant hydroxyl functional groups, which enable effective adsorption of HMs through electrostatic interactions and ligand exchange.⁶¹ For instance, As can be adsorbed by amorphous Fe oxides or associated with crystalline Fe oxides.^62,63 In As-contaminated soils, over 80 wt% of As(V) may be bound to amorphous Fe oxides,⁶⁴ and roughly 95% of total As in mine tailings is associated with Fe oxides.⁶⁵ Cr(VI) primarily coordinates with Fe oxides through the formation of inner- and outer-sphere complexes.⁶⁶ In a typical Cr-contaminated soil, approximately 11% of total Cr is associated with crystalline Fe oxides and 7% with amorphous Fe oxides.⁶⁷ Se can also be immobilized on Fe oxides through electrostatic adsorption or ion-exchange,⁶⁸ forming inner-sphere complexes on hematite and mixed outer- and inner-sphere complexes on goethite and hydrous ferric oxides.⁶⁹ A survey of 58 Japanese agricultural soils found that Fe-bound Se constituted up to 55.9% of extractable Se.⁷⁰

Fe-dependent AOM (Fe-AOM) is represented by eqn (7) (Fig. 1).²⁶ It was first identified by Beal et al.⁷¹ in marine sediments, where ANME were shown to couple methane oxidation with Fe(III) reduction. Subsequent studies have confirmed Fe-AOM in diverse environments, including marine sediments,^72,73 coastal sediments,^74,75 freshwater lakes,^76,77 lake sediments,^78,79 paddy soils,⁸⁰ and riverbeds,⁸¹ indicating a broad distribution across wetland ecosystems.

Reported microorganisms involved in Fe-AOM include Candidatus “Methanoperedens nitroreducens”, Methanosarcina acetivorans, Candidatus “Methanoperedens ferrireducens”, Methanomassiliicoccaceae, pGrfC26, Geobacter, and Geothrix.^31,76,82,83 The process can occur via syntrophy, where archaea (such as Methanomassiliicoccaceae and pGrfC26) oxidize methane through a reverse methanogenesis pathway and transfer electrons to iron-reducing bacteria (IRB) (e.g., Geobacter, Geothrix).⁸³ Alternatively, Some ANME can also perform Fe-AOM independently. For example, Candidatus “Methanoperedens ferrireducens” (ANME-2d) can directly perform Fe-AOM through a combination of reverse methanogenesis and MHC-mediated extracellular electron transfer.⁸²

In wetlands, where Fe oxides often immobilize HMs, Fe(III) reduction coupled with AOM may lead to the release of associated HMs (Fig. 5A).⁸⁴ For instance, Fe-AOM can promote Fe(III) reduction and liberate phosphate adsorbed on Fe oxides into groundwater.⁸⁵ Similarly, Fe-AOM drives the reductive dissolution of As-bearing Fe(III) minerals, upregulates methane oxidation genes, and mobilizes As into aqueous phases.²⁴

Fig. 5 Indirect transformation of HM speciation mediated by AOM. (A) Indirect transformation mediated by Fe-/Mn-AOM. (B) Indirect transformation mediated by S-AOM. ANME: anaerobic methanotrophic archaea; MRB: metal-reducing bacteria; SRB: sulfate-reducing bacteria.

The efficiency of Fe-AOM is influenced by Fe(III) bioavailability. In a long-term paddy soil incubation (420 days), both soluble EDTA-Fe(III) and insoluble ferrihydrite served as electron acceptors, but EDTA-Fe(III) ultimately suppressed activity over time.⁸³ A comparison of four Fe oxides (ferrihydrite, goethite, hematite, and magnetite) by He et al. showed that ferrihydrite supported the highest short-term Fe-AOM rate, while goethite showed the highest rate after 380 days.⁸⁶ In porewater and sediments of the Bothnian Sea, high concentrations of sulfate inhibited methanogenesis by consuming organic matter, and Fe(III) was reduced abiotically by organics and biogenic sulfide, diminishing Fe-AOM potential.⁷⁴

Although Fe-AOM-induced As release has been confirmed, its influence on the speciation and mobility of other HMs in wetland soils needs further study. Notably, Fe(II) released from Fe(III) reduction during Fe-AOM may be re-oxidized to Fe(III) oxides through abiotic or microbial pathways.⁸⁷ These regenerated Fe(III) oxides can re-adsorb or co-precipitate HMs, creating a dynamic equilibrium between HM mobilization and immobilization.⁸⁸ This cycle is crucial for accurately predicting the fate and bioavailability of HMs in wetland ecosystems.

4.2 Mn-mediated AOM

Mn is a redox-sensitive element and an essential trace nutrient for most organisms.⁸⁹ It constitutes approximately 0.1% of the Earth's crust and is nearly ubiquitous in soils and sediments worldwide.⁹⁰ Mn concentrations in wetlands vary considerably across locations, depending on geological background, land use, and other environmental factors. For instance, Mn levels can reach up to 679 mg kg⁻¹ in paddy soils of Changsha, Hunan Province, China.⁹¹ More than 30 types of Mn oxides/hydroxides (collectively termed Mn oxides) have been identified.⁹⁰ These minerals typically possess a low point of zero charge and provide abundant cation adsorption sites, making them key regulators of HM partitioning between solid and aqueous phases in natural systems.⁹² Mn oxides can oxidize As(III) to As(V) and also adsorb As(V) through surface complexation.^93,94 In paddy soils from Zhangping, Fujian Province, China, Mn oxide-bound As represented 27.7% to 31.4% of total As.⁵⁹ Similarly, Mn oxides effectively immobilize cationic HMs such as cadmium (Cd) and lead (Pb) through inner- and outer-sphere complexation.^95,96 As reported, Mn oxides retained 11.6% of the total adsorbed Cd in a fluvaquent soil from Galicia, Spain.⁹⁷ The strong adsorption of HMs by Mn oxides greatly lowers their bioavailability and mobility in soils.

Mn-dependent AOM (Mn-AOM) is one of the earliest recognized forms of metal-dependent AOM (eqn (8) in Fig. 1).¹⁷ Besides laboratory bioreactors,⁹⁸ Mn-AOM has been documented in various natural and engineered anaerobic settings, including marine sediments,^71,99 constructed wetlands,^100–102 and freshwater sediments.^76,103,104

Microorganisms known to mediate Mn-AOM include Candidatus “Methanoperedens nitroreducens” (e.g., Candidatus “Methanoperedens sp. BLZ1”), Candidatus “Methanoperedens manganicus”, Candidatus “Methanoperedens manganireducens”, and L. pratensis.^76,98,105 Mn-AOM can be carried out solely by ANME. For instance, Candidatus “Methanoperedens manganicus” and Candidatus “Methanoperedens manganireducens” perform direct extracellular electron transfer through MHCs, enabling independent reduction of Mn oxides.⁹⁸ Alternatively, a syntrophic pathway has been proposed in which Candidatus “Methanoperedens sp. BLZ1” oxidizes methane and L. pratensis assists in extracellular electron transfer and Mn(IV) reduction.¹⁰⁵

Like Fe-AOM, the reductive dissolution of Mn oxides during Mn-AOM can release associated HMs (Fig. 5A).¹⁰⁶ Although the direct impact of Mn-AOM on co-occurring HM speciation remains poorly studied, Mn oxide reduction has been shown to mobilize associated contaminants in various wetland soils. For instance, under mildly reducing conditions in German floodplain topsoils, reductive dissolution of Mn oxides resulted in the release of cobalt (Co) and molybdenum (Mo).¹⁰⁷

Only a few studies have examined factors influencing Mn-AOM. Xie et al.¹⁰⁸ reported that humic acid doubled the Mn-AOM rate, likely by serving as an electron shuttle to facilitate long-range extracellular electron transfer between ANME and Mn oxides. So far, no systematic investigation has addressed how Mn form (e.g., soluble, solid-phase, and crystallinity) affects Mn-AOM activity. Given its potential role in HM mobilization, Mn-AOM may represent an important yet overlooked biogeochemical process controlling the fate of HMs in wetlands.

Mn(II) produced during Mn-AOM can also be re-oxidized to Mn(IV) oxides under oxic or suboxic conditions.¹⁰⁹ The regenerated Mn oxides can re-immobilize HMs through surface complexation or oxidation-precipitation.¹¹⁰ This dynamic equilibrium between Mn oxide reduction and re-formation must be considered when evaluating the long-term impact of Mn-AOM on HM speciation in wetlands.

4.3 SO₄²⁻-mediated AOM

Sulfur (S) comprises 0.03–0.1% of the Earth's crust and is widely distributed across terrestrial, atmospheric, and marine environments.¹¹¹ Sulfate-dependent AOM (S-AOM) was first discovered in marine sediments. In this process, ANME oxidizes methane through several proposed pathways, including reverse methanogenesis,²⁹ acetogenesis,²⁹ and methanogenesis.¹¹² The electrons generated are transferred to sulfate through dissimilatory sulfate reduction,¹¹³ direct electron transfer,¹¹⁴ or intercellular nanowire conduction,¹¹⁵ ultimately reducing sulfate to hydrogen sulfide (eqn (9), Fig. 1).²⁶

S-AOM serves as a major methane sink and is prevalent in sulfate-rich wetlands. For instance, in sediments of the Beidagang coastal wetland (Tianjin, China), S-AOM accounted for 71.01–95.91% of methane consumption and 37.03–96.41% of sulfate consumption in the sulfate-methane transition zone.¹¹⁶ Although sulfate concentrations are typically low in freshwater wetlands, rapid sulfur recycling can sustain high sulfate reduction rates.¹¹ In three freshwater wetlands in the United States, AOM rates averaged 20 µmol L⁻¹ d⁻¹, comparable to those in marine systems, with sulfate reduction likely serving as the dominant electron-accepting process.¹¹ S-AOM is also relevant in paddy soils, where sulfate is often introduced through fertilizers such as ammonium sulfate.¹¹⁷ Following amendment with iron and sulfate, roughly 47% of AOM in paddy soil was estimated to couple with sulfate reduction.¹¹⁸

Under anaerobic conditions, sulfate in S-AOM is reduced by SRB to sulfide, which can form stable sulfide-bound species with HMs (Fig. 5B). Although direct evidence linking S-AOM to HM speciation in wetland soils remains limited, sulfate reduction is widely recognized as an effective strategy for immobilizing HMs. For instance, adding exogenous sulfate to paddy soils from Shangyu and Tongling reduced porewater As concentrations by 51.1% and 29.2%, respectively.¹¹⁷ This treatment enriched SRB and promoted the precipitation of As sulfides.¹¹⁷ Similarly, amending paddy soil with CaSO₄ enhanced sulfate reduction and decreased Cd solubility, with in situ X-ray absorption spectroscopy confirming precipitation of CdS.¹¹⁹ In aqueous systems, Cr(VI) can also react with H₂S to form Cr(OH)₃ and elemental sulfur.¹²⁰

Despite the theoretical potential of S-AOM-driven sulfate reduction to immobilize HMs through sulfide formation, direct evidence tracing how sulfide derived specifically from S-AOM participates in HM sulfidation is still lacking. Moreover, in sulfate-limited freshwater wetlands, competition between S-AOM and other sulfate-consuming processes for sulfate may be a key factor controlling its efficiency in HM fixation.

5. Effect of environmental factors

Environmental factors, including pH, Eh, and organic matter content, are key regulators of AOM activity and subsequent HM speciation in natural wetland soils.¹²¹

pH primarily influences AOM by modulating the activity of functional microorganisms. For example, the decrease of pH from 6.5 to 5.5 in a paddy soil suspension inhibited AOM-AsR and decreased As(III) production by 85%.¹²² ANME typically exhibit optimal activity under neutral conditions, while the abundance of the mcrA gene declines significantly under acidic pH.¹²² pH also affects the protonation and speciation of HMs. For instance, As(V) predominates under neutral to alkaline conditions, whereas As(III) is more stable in acidic environments.¹²³

Eh determines the preferred electron acceptors for AOM, thereby shaping HM transformation pathways. In environments with higher Eh, such as drained rice fields, AOM tends to couple with Fe(III) or Mn(IV) reduction, promoting the reductive dissolution of Fe/Mn oxides and the release of associated HMs.^24,124 In contrast, under low-Eh conditions (e.g., waterlogged wetlands), S-AOM is favored, where sulfate reduction leads to sulfide precipitation and HM immobilization.^19,124 Fluctuations in redox conditions can shift AOM pathways, resulting in dynamic changes in HM toxicity and mobility.

Soil organic matter influences AOM and the fate of HMs through two primary mechanisms: serving as an electron shuttle and acting as a chelating agent for HMs. For example, humic acid can accelerate AOM rates by shuttling electrons between anaerobic microorganisms and electron acceptors.¹⁰⁸ Conversely, recalcitrant soil organic matter may form complexes with HMs or coat Fe/Mn oxides, thereby limiting HM accessibility to AOM-active microorganisms and lowering transformation efficiency.⁹⁰

Environmental factors such as pH, Eh, and soil organic matter interact to regulate AOM activity and HM behavior, driving complex spatiotemporal variations in HM speciation across natural wetlands. These factors also operate in feedback loops; for instance, AOM can alter local pH and Eh conditions, which in turn influences the dissolution–precipitation dynamics of Fe/Mn oxides and the complexation capacity of sediment organic matter.^125,126 Such indirect effects further modulate HM adsorption, desorption, and speciation transformation.

6. Summary and future perspectives

This review synthesizes current understanding of the role of AOM in transforming HM speciation and its environmental implications in wetland soils. We have outlined the key microorganisms involved across various AOM pathways, predominantly ANME and methanogens, with a noted potential contribution from aerobic methanotrophs under hypoxic conditions. Relevant functional genes and enzymes involved in these processes are also discussed. Two principal mechanisms underpin HM-associated AOM: a syntrophic process, where archaea oxidize methane through reverse methanogenesis and transfer electrons to HM-reducing bacteria, and a direct process mediated solely by MOMs.

A key insight is the dual role of AOM in regulating HM fate. The AOM-driven direct reduction of Cr(VI), Se(IV/VI), and V(V) converts these contaminants into less toxic, mobile, and bioavailable forms, indicating a potential detoxification route. Conversely, the direct microbial reduction of As(V) to As(III) by AOM increases its toxicity and mobility. The reduction of Sb(V) yields the more toxic but generally less mobile Sb(III), illustrating a complex trade-off between toxicity and transport behavior. Theoretically, Fe/Mn-dependent AOM may mobilize HMs through the reductive dissolution of metal oxides, while S-dependent AOM could immobilize HMs through sulfide precipitation. However, direct experimental evidence from natural wetland settings remains limited, revealing a significant gap between theoretical models and field observations.

Major knowledge gaps hinder a predictive understanding of AOM mediated HM biogeochemistry in wetland environments. The key microorganisms involved in these processes are not fully characterized. Mechanistic understanding of electron transfer pathways, key genes, and enzyme functions is limited by the lack of pure cultures and defined synthetic consortia. The effects of environmental factors like temperature, pH, Eh, and HM form on these processes in natural soils are poorly understood. Important unresolved questions include how aerobic methanotrophs sustain activity under anaerobic conditions and their quantitative contribution to HM transformation, as well as the electron acceptor preferences of AOM in environments with multiple coexisting HMs. Moreover, most reported reaction rates are derived from laboratory cultures or bioreactors. These rates likely differ from those in natural wetlands, where conditions are dynamic due to redox fluctuations, competition among multiple electron acceptors, and biotic interactions. Therefore, quantifying the in situ rates of HM associated AOM processes is also one of the primary future challenges.

Advances in techniques such as metagenomic sequencing, stable isotope probing, and metatranscriptomics provide powerful tools to address these gaps. Future research should integrate multi-omics data with geochemical analyses to identify active microorganisms, elucidate in situ reaction pathways, and quantify process rates under environmentally relevant conditions. These insights are fundamental for developing targeted strategies to exploit AOM for remediating HM-contaminated wetland soils.

For HM-contaminated wetlands, translating this knowledge into practical remediation technology is essential. Two promising strategies are: (i) bioaugmentation, involving the inoculation of sites with selected, high performance AOM microbial consortia that combine efficient methane oxidation with HM tolerance; and (ii) environmental regulation, through managing parameters like water level (to control Eh) and amending with slow-release electron donors or acceptors (e.g., sulfate) to stimulate microbial pathways that promote HM stabilization.

Author contributions

Wei Ye: writing – original draft, visualization. Lihu Liu: conceptualization, funding acquisition, project administration, supervision, writing – review & editing. Zhaozhi Zheng: writing – review & editing. Shengwen Xu: writing – review & editing. Yongxiang Yu: writing – review & editing. Ningguo Zheng: writing – review & editing. Yongbao Zhang: writing – review & editing. Huaiying Yao: conceptualization, funding acquisition, project administration, supervision, writing – review & editing.

Conflicts of interest

The authors declare no competing interest.

Data availability

No primary research results, software or code have been included and no new data were generated or analyzed as part of this review.

Acknowledgements

We are thankful for the financial support from the National Natural Science Foundation of China (U24A20626), the Outstanding Young and Middle-aged Science and Technology Innovation Team Project in Colleges and Universities of Hubei Province (T2023008), the Science Research Foundation of Wuhan Institute of Technology (K2023075), and the Tracking Support Project for Postdoctoral Pioneer Talents of Hubei Province (2024HBBHXF063).

References

1. T. Xu, B. Weng, D. Yan, K. Wang, X. Li and W. Bi, et al., Wetlands of international importance: status, threats, and future protection, Int. J. Environ. Res. Publ. Health, 2019, 16(10), 1818.
2. Z. Fu and S. Xi, The effects of heavy metals on human metabolism, Toxicol. Mech. Methods, 2020, 30(3), 167–176.
3. R. Riyazuddin, N. Nisha, B. Ejaz, M. I. R. Khan, M. Kumar and P. W. Ramteke, et al., A comprehensive review on the heavy metal toxicity and sequestration in plants, Biomolecules, 2021, 12(1), 43.
4. S. Cheng, Effects of heavy metals on plants and resistance mechanisms, Environ. Sci. Pollut. Res., 2003, 10(4), 256–264.
5. C. Zhang, S. Nie, J. Liang, G. Zeng, H. Wu and S. Hua, et al., Effects of heavy metals and soil physicochemical properties on wetland soil microbial biomass and bacterial community structure, Sci. Total Environ., 2016, 557–558, 785–790.
6. J. Bai, Q. Zhao, W. Wang, X. Wang, J. Jia and B. Cui, et al., Arsenic and heavy metals pollution along a salinity gradient in drained coastal wetland soils: depth distributions, sources and toxic risks, Ecol. Indic., 2019, 96, 91–98.
7. D. Paul, B. Choudhary, T. Gupta and M. T. Jose, Spatial distribution and the extent of heavy metal and hexavalent chromium pollution in agricultural soils from Jajmau, India, Environ. Earth Sci., 2015, 73, 3565–3577.
8. B. L. Henry, J. S. Wesner and J. L. Kerby, Cross-ecosystem effects of agricultural tile drainage, surface runoff, and selenium in the prairie pothole region, Wetlands, 2020, 40(3), 527–538.
9. Q. Zhao and Y. Lu, Anaerobic oxidation of methane in terrestrial wetlands: the rate, identity and metabolism, Sci. Total Environ., 2023, 902, 166049.
10. M. Cui, A. Ma, H. Qi, X. Zhuang and G. Zhuang, Anaerobic oxidation of methane: an “active” microbial process, MicrobiologyOpen, 2015, 4(1), 1–11.
11. K. E. A. Segarra, F. Schubotz, V. Samarkin, M. Y. Yoshinaga, K.-U. Hinrichs and S. B. Joye, High rates of anaerobic methane oxidation in freshwater wetlands reduce potential atmospheric methane emissions, Nat. Commun., 2015, 6(1), 7477.
12. L. Liu, N. Zheng, Y. Yu, Z. Zheng and H. Yao, Soil carbon and nitrogen cycles driven by iron redox: a review, Sci. Total Environ., 2024, 918, 170660.
13. Y.-Z. Lu, L. Fu, J. Ding, Z.-W. Ding, N. Li and R. J. Zeng, Cr(VI) reduction coupled with anaerobic oxidation of methane in a laboratory reactor, Water Res., 2016, 102, 445–452.
14. L.-D. Shi, T. Guo, P.-L. Lv, Z.-F. Niu, Y.-J. Zhou and X.-J. Tang, et al., Coupled anaerobic methane oxidation and reductive arsenic mobilization in wetland soils, Nat. Geosci., 2020, 13(12), 799–805.
15. C.-Y. Lai, Q.-Y. Dong, B. E. Rittmann and H.-P. Zhao, Bio-reduction of antimonate by anaerobic methane oxidation in a membrane biofilm batch reactor, Environ. Sci. Technol., 2018, 52(15), 8693–8700.
16. J. Luo, H. Chen, S. Hu, C. Cai, Z. Yuan and J. Guo, Microbial selenate reduction driven by a denitrifying anaerobic methane oxidation biofilm, Environ. Sci. Technol., 2018, 52(7), 4006–4012.
17. Z. He, Q. Zhang, Y. Feng, H. Luo, X. Pan and G. M. Gadd, Microbiological and environmental significance ofmetal-dependent anaerobic oxidation ofmethane, Sci. Total Environ., 2018, 610–611, 759–768.
18. T. Lu, W. Wang, L. Liu, L. Wang, J. Hu and X. Li, et al., Remediation of cadmium-polluted weakly alkaline dryland soils using iron and manganese oxides for immobilized wheat uptake, J. Clean. Prod., 2022, 365, 132794.
19. Y.-N. Xu and Y. Chen, Advances in heavy metal removal by sulfate-reducing bacteria, Water Sci. Technol., 2020, 81(9), 1797–1827.
20. S. K. Hwang and E. H. Jho, Heavy metal and sulfate removal from sulfate-rich synthetic mine drainages using sulfate reducing bacteria, Sci. Total Environ., 2018, 635, 1308–1316.
21. Y. Zhao, Y. Liu, S. Cao, Q. Hao, C. Liu and Y. Li, Anaerobic oxidation of methane driven by different electron acceptors: a review, Sci. Total Environ., 2024, 946, 174287.
22. O. P. Karthikeyan, T. J. Smith, S. U. Dandare, K. S. Parwin, H. Singh and H. X. Loh, et al., Metal(loid) speciation and transformation by aerobic methanotrophs, Microbiome, 2021, 9, 1–18.
23. J. Ding, Y.-Z. Lu, L. Fu, Z.-W. Ding, Y. Mu and S. H. Cheng, et al., Decoupling of DAMO archaea from DAMO bacteria in a methane-driven microbial fuel cell, Water Res., 2017, 110, 112–119.
24. M. Glodowska, E. Stopelli, M. Schneider, B. Rathi, D. Straub and A. Lightfoot, et al., Arsenic mobilization by anaerobic iron-dependent methane oxidation, Commun. Earth Environ., 2020, 1(1), 42.
25. B. Zhang, Y. Jiang, K. Zuo, C. He, Y. Dai and Z. J. Ren, Microbial vanadate and nitrate reductions coupled with anaerobic methane oxidation in groundwater, J. Hazard. Mater., 2020, 382, 121228.
26. C.-C. Dang, G.-J. Xie, B.-F. Liu, D.-F. Xing, J. Ding and N.-Q. Ren, Heavy metal reduction coupled to methane oxidation: mechanisms, recent advances and future perspectives, J. Hazard. Mater., 2021, 405, 124076.
27. C. Li, H. Wang, X. Liao, R. Xiao, K. Liu and J. Bai, et al., Heavy metal pollution in coastal wetlands: a systematic review of studies globally over the past three decades, J. Hazard. Mater., 2022, 424, 127312.
28. Q. Lu, J. Bai, G. Zhang and J. Wu, Effects of coastal reclamation history on heavy metals in different types of wetland soils in the Pearl River Delta: levels, sources and ecological risks, J. Clean. Prod., 2020, 272, 122668.
29. P. H. A. Timmers, C. U. Welte, J. J. Koehorst, C. M. Plugge, M. S. M. Jetten and A. J. M. Stams, Reverse methanogenesis and respiration in methanotrophic archaea, Archaea, 2017, 2017(1), 1654237.
30. G. Yan, X. Sun, Y. Dong, W. Gao, P. Gao and B. Li, et al., Vanadate reducing bacteria and archaea may use different mechanisms to reduce vanadate in vanadium contaminated riverine ecosystems as revealed by the combination of DNA-SIP and metagenomic-binning, Water Res., 2022, 226, 119247.
31. Z. Yan, P. Joshi, C. A. Gorski and J. G. Ferry, A biochemical framework for anaerobic oxidation of methane driven by Fe(III)-dependent respiration, Nat. Commun., 2018, 9(1), 1642.
32. R. Li, B. Xi, X. Wang, Y. Li, Y. Yuan and W. Tan, Anaerobic oxidation of methane in landfill and adjacent groundwater environments: occurrence, mechanisms, and potential applications, Water Res., 2024, 255, 121498.
33. S. L. Caldwell, J. R. Laidler, E. A. Brewer, J. O. Eberly, S. C. Sandborgh and F. S. Colwell, Anaerobic oxidation of methane: mechanisms, bioenergetics, and the ecology of associated microorganisms, Environ. Sci. Technol., 2008, 42(18), 6791–6799.
34. J. Yang, L. Zou, L. Zheng, Z. Yuan, K. Huang and W. Gustave, et al., Iron-based passivator mitigates the coupling process of anaerobic methane oxidation and arsenate reduction in paddy soils, Environ. Pollut., 2022, 313, 120182.
35. Y. Zhou, T. Guo, W. Gustave, Z. Yuan, J. Yang and D. Chen, et al., Anaerobic methane oxidation coupled to arsenate reduction in paddy soils: insights from laboratory and field studies, Chemosphere, 2023, 311(2), 137055.
36. Y. Zhang, D. Tong, L. Zou, H. Ji, X. Zhou and W. Gustave, et al., Low-molecular-weight organic acids inhibit the methane-dependent arsenate reduction process in paddy soils, Ecotoxicol. Environ. Saf., 2024, 282, 116716.
37. X. Zhang, Z. Yuan and S. Hu, Anaerobic oxidation of methane mediated by microbial extracellular respiration, Environ. Microbiol. Rep., 2021, 13(6), 790–804.
38. K. Zeng, Z. Zheng, L. Liu, Y. Yu, S. Xu and H. Yao, Mn(II) inhibits rather than promotes As(III) oxidation during co-oxidation of As(III) and Fe(II) by oxygen, J. Environ. Manag., 2025, 393, 127082.
39. J.-H. Luo, M. Wu, J. Liu, G. Qian, Z. Yuan and J. Guo, Microbial chromate reduction coupled with anaerobic oxidation of methane in a membrane biofilm reactor, Environ. Int., 2019, 130, 104926.
40. Q.-Y. Dong, Z. Wang, L.-D. Shi, C.-Y. Lai and H.-P. Zhao, Anaerobic methane oxidation coupled to chromate reduction in a methane-based membrane biofilm batch reactor, Environ. Sci. Pollut. Res., 2019, 26, 26286–26292.
41. X. Zheng, L. Chen, S. Zhang and Z. Xie, Cathodic Cr(VI) removal in a methane-powered bioelectrochemical system, J. Water Process Eng., 2024, 67, 106217.
42. C.-Y. Lai, L. Zhong, Y. Zhang, J.-X. Chen, L.-L. Wen and L.-D. Shi, et al., Bioreduction of chromate in a methane-based membrane biofilm reactor, Environ. Sci. Technol., 2016, 50(11), 5832–5839.
43. P.-L. Lv, L. Zhong, Q.-Y. Dong, S.-L. Yang, W.-W. Shen and Q.-S. Zhu, et al., The effect of electron competition on chromate reduction using methane as electron donor, Environ. Sci. Pollut. Res., 2017, 25, 6609–6618.
44. R. Qin, X. Dai, Y. Xian, Y. Zhou, C. Su and Z. Chen, et al., Assessing the effect of sulfate on the anaerobic oxidation of methane coupled with Cr(VI) bioreduction by sludge characteristic and metagenomics analysis, J. Environ. Manag., 2024, 349, 119398.
45. S. Wang, X. Zhang, D. Tian, J. Zhao, H. Rabiee and F. Cai, et al., Anaerobic oxidation of methane coupled to reductive immobilization of hexavalent chromium by “Candidatus Methanoperedens”, J. Hazard. Mater., 2024, 480, 136020.
46. C. Su, Y. Xian, R. Qin, Y. Zhou, M. Lu and X. Wan, et al., Fe(III) enhances Cr(VI) bioreduction in a MFC-granular sludge coupling system: experimental evidence and metagenomics analysis, Water Res., 2023, 235, 119863.
47. Y.-N. Bai, X.-N. Wang, Y.-Z. Lu, L. Fu, F. Zhang and T.-C. Lau, et al., Microbial selenite reduction coupled to anaerobic oxidation of methane, Sci. Total Environ., 2019, 669, 168–174.
48. L.-D. Shi, P.-L. Lv, M. Wang, C.-Y. Lai and H.-P. Zhao, A mixed consortium of methanotrophic archaea and bacteria boosts methane-dependent selenate reduction, Sci. Total Environ., 2020, 732, 139310.
49. L.-D. Shi, P.-L. Lv, S. J. McIlroy, Z. Wang, X.-L. Dong and A. Kouris, et al., Methane-dependent selenate reduction by a bacterial consortium, ISME J., 2021, 15(12), 3683–3692.
50. C.-Y. Lai, L.-L. Wen, L.-D. Shi, K.-K. Zhao, Y.-Q. Wang and X. Yang, et al., Selenate and nitrate bio-reductions using methane as the electron donor in a membrane biofilm reactor, Environ. Sci. Technol., 2016, 50(18), 10179–10186.
51. A. Vaksmaa, S. Guerrero-Cruz, T. A. van Alen, G. Cremers, K. F. Ettwig and C. Lüke, et al., Enrichment of anaerobic nitrate-dependent methanotrophic ‘Candidatus Methanoperedens nitroreducens’ archaea from an Italian paddy field soil, Appl. Microbiol. Biotechnol., 2017, 101(18), 7075–7084.
52. L.-d. Shen, C.-y. Geng, B.-j. Ren, J.-h. Jin, H.-c. Huang and X. Liu, et al., Detection and quantification of candidatus methanoperedens-like archaea in freshwater wetland soils, Microb. Ecol., 2023, 85(2), 441–453.
53. L.-D. Shi, M. Wang, Y.-L. Han, C.-Y. Lai, J. P. Shapleigh and H.-P. Zhao, Multi-omics reveal various potential antimonate reductases from phylogenetically diverse microorganisms, Appl. Microbiol. Biotechnol., 2019, 103, 9119–9129.
54. Z. Wang, L.-D. Shi, C.-Y. Lai and H.-P. Zhao, Methane oxidation coupled to vanadate reduction in a membrane biofilm batch reactor under hypoxic condition, Biodegradation, 2019, 30, 457–466.
55. M. Zhang, G. Lu, Z. Li, F. Xu, N. Yang and X. Sun, et al., Effects of antimony on anaerobic methane oxidization and microbial community in an antimony-contaminated paddy soil: a microcosm study, Sci. Total Environ., 2021, 784, 147239.
56. S. Koechler, J. Farasin, J. Cleiss-Arnold and F. Arsene-Ploetze, Toxic metal resistance in biofilms: diversity of microbial responses and their evolution, Res. Microbiol., 2015, 166(10), 764–773.
57. L. Liu, A. Li, M. Cao, J. Ma, W. Tan and S. L. Suib, et al., Photoinduced self-organized precipitation in leachate for remediation of heavy metal contaminated soils, ACS ES&T Eng., 2022, 2(8), 1376–1385.
58. H. Guo and A. S. Barnard, Naturally occurring iron oxide nanoparticles: morphology, surface chemistry and environmental stability, J. Mater. Chem. A, 2013, 1(1), 27–42.
59. X. Xu, C. Chen, P. Wang, R. Kretzschmar and F.-J. Zhao, Control of arsenic mobilization in paddy soils by manganese and iron oxides, Environ. Pollut., 2017, 231, 37–47.
60. L. K. ThomasArrigo, L. Notini, J. Shuster, T. Nydegger, S. Vontobel and S. Fischer, et al., Mineral characterization and composition of Fe-rich flocs from wetlands of Iceland: implications for Fe, C and trace element export, Sci. Total Environ., 2022, 816, 151567.
61. L. Liu, Z. Jia, W. Tan, S. L. Suib, L. Ge and G. Qiu, et al., Abiotic photomineralization and transformation of iron oxide nanominerals in aqueous systems, Environ. Sci. Nano, 2018, 5(5), 1169–1178.
62. K. Zeng, L. Liu, N. Zheng, Y. Yu, S. Xu and H. Yao, Iron at the helm: steering arsenic speciation through redox processes in soils, Environ. Res., 2025, 274, 121327.
63. C. M. van Genuchten, A. Finger, J. R. van der Meer and J. Peña, Bacterial bioreporter detection of arsenic associated with iron oxides, Environ. Sci.: Processes Impacts, 2018, 20(6), 913–922.
64. B. Cances, F. Juillot, G. Morin, V. Laperche, L. Alvarez and O. Proux, et al., XAS evidence of As(V) association with iron oxyhydroxides in a contaminated soil at a former arsenical pesticide processing plant, Environ. Sci. Technol., 2005, 39(24), 9398–9405.
65. E. J. Kim, J.-C. Lee and K. Baek, Abiotic reductive extraction of arsenic from contaminated soils enhanced by complexation: arsenic extraction by reducing agents and combination of reducing and chelating agents, J. Hazard. Mater., 2015, 283, 454–461.
66. X. Li, C. Guo, X. Jin, C. He, Q. Yao and G. Lu, et al., Mechanisms of Cr(VI) adsorption on schwertmannite under environmental disturbance: changes in surface complex structures, J. Hazard. Mater., 2021, 416, 125781.
67. F. X. Han, Y. Su, B. M. Sridhar and D. L. Monts, Distribution, transformation and bioavailability of trivalent and hexavalent chromium in contaminated soil, Plant Soil, 2004, 265, 243–252.
68. J. Fan, Y. Zeng and J. Sun, The transformation and migration of selenium in soil under different Eh conditions, J. Soils Sediments, 2018, 18, 2935–2947.
69. D. Peak and D. Sparks, Mechanisms of selenate adsorption on iron oxides and hydroxides, Environ. Sci. Technol., 2002, 36(7), 1460–1466.
70. Y. Nakamaru, K. Tagami and S. Uchida, Distribution coefficient of selenium in Japanese agricultural soils, Chemosphere, 2005, 58(10), 1347–1354.
71. E. J. Beal, C. H. House and V. J. Orphan, Manganese- and iron-dependent marine methane oxidation, Science, 2009, 325(5937), 184–187.
72. N. Riedinger, M. J. Formolo, T. W. Lyons, S. Henkel, A. Beck and S. Kasten, An inorganic geochemical argument for coupled anaerobic oxidation of methane and iron reduction in marine sediments, Geobiology, 2014, 12(2), 172–181.
73. H. Yang, S. Yu and H. Lu, Iron-coupled anaerobic oxidation of methane in marine sediments: a review, J. Mar. Sci. Eng., 2021, 9(8), 875.
74. J. Rooze, M. Egger, I. Tsandev and C. P. Slomp, Iron-dependent anaerobic oxidation of methane in coastal surface sediments: potential controls and impact, Limnol. Oceanogr., 2016, 61(S1), S267–S282.
75. M. Egger, O. Rasigraf, C. J. Sapart, T. Jilbert, M. S. M. Jetten and R. Rockmann, et al., Iron-mediated anaerobic oxidation of methane in brackish coastal sediments, Environ. Sci. Technol., 2015, 49(1), 277–283.
76. K. F. Ettwig, B. Zhu, D. Speth, J. T. Keltjens, M. S. Jetten and B. Kartal, Archaea catalyze iron-dependent anaerobic oxidation of methane, Proc. Natl. Acad. Sci. U. S. A., 2016, 113(45), 12792–12796.
77. A. Mostovaya, M. Wind-Hansen, P. Rousteau, L. A. Bristow and B. Thamdrup, Sulfate-and iron-dependent anaerobic methane oxidation occurring side-by-side in freshwater lake sediment, Limnol. Oceanogr., 2022, 67(1), 231–246.
78. O. Sivan, M. Adler, A. Pearson, F. Gelman, I. Bar-Or and S. G. John, et al., Geochemical evidence for iron-mediated anaerobic oxidation of methane, Limnol. Oceanogr., 2011, 56(4), 1536–1544.
79. I. Bar-Or, M. Elvert, W. Eckert, A. Kushmaro, H. Vigderovich and Q. Zhu, et al., Iron-coupled anaerobic oxidation of methane performed by a mixed bacterial-archaeal community based on poorly-reactive minerals, Environ. Sci. Technol., 2017, 51(21), 12293–12301.
80. D. Luo, X. Meng, N. Zheng, Y. Li, H. Yao and S. J. Chapman, The anaerobic oxidation of methane in paddy soil by ferric iron and nitrate, and the microbial communities involved, Sci. Total Environ., 2021, 788, 147773.
81. L.-d. Shen, L. Ouyang, Y. Zhu and M. Trimmer, Active pathways of anaerobic methane oxidation across contrasting riverbeds, ISME J., 2019, 13(3), 752–766.
82. C. Cai, A. O. Leu, G.-J. Xie, J. Guo, Y. Feng and J.-X. Zhao, et al., A methanotrophic archaeon couples anaerobic oxidation of methane to Fe(III) reduction, ISME J., 2018, 12(8), 1929–1939.
83. Z. He, Y. Xu, Y. Zhu, J. Feng, D. Zhang and X. Pan, Long-term effects of soluble and insoluble ferric irons on anaerobic oxidation of methane in paddy soil, Chemosphere, 2023, 317, 137901.
84. Y. N. Vodyanitskii and S. A. Shoba, Biogeochemistry of carbon, iron, and heavy metals in wetlands (analytical review), Moscow Univ. Soil Sci. Bull., 2015, 70, 89–97.
85. Y. Du, Y. Xiong, Y. Deng, Y. Tao, H. Tian and Y. Zhang, et al., Geogenic phosphorus enrichment in groundwater due to anaerobic methane oxidation-coupled Fe(III) oxide reduction, Environ. Sci. Technol., 2024, 58(18), 8032–8042.
86. Z. He, Y. Zhu, J. Feng, Q. Ji, X. Chen and X. Pan, Long-term effects of four environment-related iron minerals on microbial anaerobic oxidation of methane in paddy soil: a previously overlooked role of widespread goethite, Soil Biol. Biochem., 2021, 161, 108387.
87. L. Wei, X. Rong, H. Yuan, Y. Li, X. Fan and N. Ling, et al., Carbon stabilization by iron plaque on rice roots: The role of oxygen loss, Soil Biol. Biochem., 2025, 109947.
88. M. Shi, X. Min, Y. Ke, Z. Lin, Z. Yang and S. Wang, et al., Recent progress in understanding the mechanism of heavy metals retention by iron (oxyhydr) oxides, Sci. Total Environ., 2021, 752, 141930.
89. C. G. Fraga, Relevance, essentiality and toxicity of trace elements in human health, Mol. Aspects Med., 2005, 26(4–5), 235–244.
90. H. Li, F. Santos, K. Butler and E. Herndon, A critical review on the multiple roles of manganese in stabilizing and destabilizing soil organic matter, Environ. Sci. Technol., 2021, 55(18), 12136–12152.
91. S. C. Maguffin, L. Abu-Ali, R. V. Tappero, J. Pena, J. S. Rohila and A. M. McClung, et al., Influence of manganese abundances on iron and arsenic solubility in rice paddy soils, Geochim. Cosmochim. Acta, 2020, 276, 50–69.
92. Z. Michálková, M. Komárek, M. Vítková, M. Řečínská and V. Ettler, Stability, transformations and stabilizing potential of an amorphous manganese oxide and its surface-modified form in contaminated soils, Appl. Geochem., 2016, 75, 125–136.
93. S. Ouvrard, P. De Donato, M.-O. Simonnot, S. Begin, J. Ghanbaja and M. e. Alnot, et al., Natural manganese oxide: combined analytical approach for solid characterization and arsenic retention, Geochim. Cosmochim. Acta, 2005, 69(11), 2715–2724.
94. L. Liu, M. Zhang, S. L. Suib and G. Qiu, Rapid photooxidation and removal of As(III) from drinking water using Fe-Mn composite oxide, Water Res., 2022, 226, 119297.
95. Q. Su, B. Pan, S. Wan, W. Zhang and L. Lv, Use of hydrous manganese dioxide as a potential sorbent for selective removal of lead, cadmium, and zinc ions from water, J. Colloid Interface Sci., 2010, 349(2), 607–612.
96. C. M. van Genuchten and J. Peña, Sorption selectivity of birnessite particle edges: a d-PDF analysis of Cd(II) and Pb(II) sorption by δ-MnO₂ and ferrihydrite, Environ. Sci.: Processes Impacts, 2016, 18(8), 1030–1041.
97. D. Arenas-Lago, F. A. Vega, L. F. O. Silva and M. L. Andrade, Soil interaction and fractionation of added cadmium in some Galician soils, Microchem. J., 2013, 110, 681–690.
98. A. O. Leu, C. Cai, S. J. McIlroy, G. Southam, V. J. Orphan and Z. Yuan, et al., Anaerobic methane oxidation coupled to manganese reduction by members of the Methanoperedenaceae, ISME J., 2020, 14(4), 1030–1041.
99. X. Xiao, M. Luo, C. Zhang, T. Zhang, X. Yin and X. Wu, et al., Metal-driven anaerobic oxidation of methane as an important methane sink in methanic cold seep sediments, Microbiol. Spectr., 2023, 11(2), e05337–05322.
100. W. Liu, H. Xiao, H. Ma, Y. Li, T. M. Adyel and J. Zhai, Reduction of methane emissions from manganese-rich constructed wetlands: role of manganese-dependent anaerobic methane oxidation, Chem. Eng. J., 2020, 387, 123402.
101. K. Zhang, X. Wu, J. Chen, W. Wang, H. Luo and W. Chen, et al., The role and related microbial processes of Mn-dependent anaerobic methane oxidation in reducing methane emissions from constructed wetland-microbial fuel cell, J. Environ. Manag., 2021, 294, 112935.
102. K. Zhang, X. Wu, W. Wang, J. Chen, H. Luo and W. Chen, et al., Anaerobic oxidation of methane (AOM) driven by multiple electron acceptors in constructed wetland and the related mechanisms of carbon, nitrogen, sulfur cycles, Chem. Eng. J., 2022, 433, 133663.
103. N. T. Torres, L. M. Och, P. C. Hauser, G. Furrer, H. Brandl and E. Vologina, et al., Early diagenetic processes generate iron and manganese oxide layers in the sediments of Lake Baikal, Siberia, Environ. Sci.: Processes Impacts, 2014, 16(4), 615–944.
104. G. Su, J. Zopfi, H. Yao, L. Steinle, H. Niemann and M. F. Lehmann, Manganese/iron-supported sulfate-dependent anaerobic oxidation of methane by archaea in lake sediments, Limnol. Oceanogr., 2020, 65(4), 863–875.
105. W. Liu, S. Xu, H. Ma, Y. Li, J. Mąkinia and J. Zhai, Anaerobic consortia mediate Mn(IV)-dependent anaerobic oxidation of methane, Chem. Eng. J., 2023, 468, 143478.
106. D. Gasparatos, Sequestration of heavy metals from soil with Fe–Mn concretions and nodules, Environ. Chem. Lett., 2013, 11, 1–9.
107. I. Hindersmann and T. Mansfeldt, Trace element solubility in a multimetal-contaminated soil as affected by redox conditions, Water, Air, Soil Pollut., 2014, 225, 1–20.
108. M. Xie, X. Zhang, S. Li, N. Maulani, F. Cai and Y. Zheng, et al., Humic substances as electron acceptor for anaerobic oxidation of methane (AOM) and electron shuttle in Mn(IV)-dependent AOM, Sci. Total Environ., 2024, 912, 169576.
109. Y. Huang, X. Zheng, Z. Zhao, J. Tao, T. Hu and Z. Han, et al., Integration of manganese ores with activated carbon into constructed wetland for greenhouse gas emissions reduction, J. Environ. Manag., 2025, 375, 124205.
110. H. Zhou and C. Fu, Manganese-oxidizing microbes and biogenic manganese oxides: characterization, Mn(II) oxidation mechanism and environmental relevance, Rev. Environ. Sci. Bio/Technol., 2020, 19(3), 489–507.
111. Z. R. Todd, Sources of nitrogen-, sulfur-, and phosphorus-containing feedstocks for prebiotic chemistry in the planetary environment, Life, 2022, 12(8), 1268.
112. J. J. Moran, E. J. Beal, J. M. Vrentas, V. J. Orphan, K. H. Freeman and C. H. House, Methyl sulfides as intermediates in the anaerobic oxidation of methane, Environ. Microbiol., 2008, 10(1), 162–173.
113. J. Milucka, T. G. Ferdelman, L. Polerecky, D. Franzke, G. Wegener and M. Schmid, et al., Zero-valent sulphur is a key intermediate in marine methane oxidation, Nature, 2012, 491(7425), 541–546.
114. S. E. McGlynn, G. L. Chadwick, C. P. Kempes and V. J. Orphan, Single cell activity reveals direct electron transfer in methanotrophic consortia, Nature, 2015, 526(7574), 531–535.
115. G. Wegener, V. Krukenberg, D. Riedel, H. E. Tegetmeyer and A. Boetius, Intercellular wiring enables electron transfer between methanotrophic archaea and bacteria, Nature, 2015, 526(7574), 587–590.
116. W. La, X. Han, C.-Q. Liu, H. Ding, M. Liu and F. Sun, et al., Sulfate concentrations affect sulfate reduction pathways and methane consumption in coastal wetlands, Water Res., 2022, 217, 118441.
117. S. Yan, J. Yang, Y. Si, X. Tang, Y. Ma and W. Ye, Arsenic and cadmium bioavailability to rice (Oryza sativa L.) plant in paddy soil: influence of sulfate application, Chemosphere, 2022, 307, 135641.
118. Z. He, J. Shen, Y. Zhu, J. Feng and X. Pan, Enhanced anaerobic oxidation of methane with the coexistence of iron oxides and sulfate fertilizer in paddy soil, Chemosphere, 2023, 329, 138623.
119. B. Fulda, A. Voegelin and R. Kretzschmar, Redox-controlled changes in cadmium solubility and solid-phase speciation in a paddy soil As affected by reducible sulfate and copper, Environ. Sci. Technol., 2013, 47(22), 12775–12783.
120. C. Kim, Q. Zhou, B. Deng, E. C. Thornton and H. Xu, Chromium(VI) reduction by hydrogen sulfide in aqueous media: stoichiometry and kinetics, Environ. Sci. Technol., 2001, 35(11), 2219–2225.
121. P. Marschner, Processes in submerged soils–linking redox potential, soil organic matter turnover and plants to nutrient cycling, Plant Soil, 2021, 464(1), 1–12.
122. Z.-F. Yuan, Y.-J. Zhou, L. Zou, Z. Chen, W. Gustave and D. Duan, et al., pH dependence of arsenic speciation in paddy soils: the role of distinct methanotrophs, Environ. Pollut., 2023, 318, 120880.
123. G. Cassone, D. Chillé, C. Foti, O. Giuffré, R. C. Ponterio and J. Sponer, et al., Stability of hydrolytic arsenic species in aqueous solutions: As³⁺vs. As⁵⁺, Phys. Chem. Chem. Phys., 2018, 20(36), 23272–23280.
124. C. Lahiri and G. R. Davidson, Heterogeneous oxygenation of wetland soils with increasing inundation: Redox potential, water depth, and preferential flow paths, Hydrol. Process., 2020, 34(6), 1350–1358.
125. Y. Pan, G. F. Koopmans, L. T. Bonten, J. Song, Y. Luo and E. J. Temminghoff, et al., Influence of pH on the redox chemistry of metal (hydr)oxides and organic matter in paddy soils, J. Soils Sediments, 2014, 14(10), 1713–1726.
126. W. Tan, F. Liu, X. Feng, Q. Huang and X. Li, Adsorption and redox reactions of heavy metals on Fe–Mn nodules from Chinese soils, J. Colloid Interface Sci., 2005, 284(2), 600–605.
127. Z. Xu and D. C. Tsang, Redox-induced transformation of potentially toxic elements with organic carbon in soil, Carbon Res., 2022, 1(1), 9.
128. J. Li, H. Liang, C. Xiao, X. Jia, R. Guo and J. Chen, et al., Enhancing the efficiency and longevity of inverted perovskite solar cells with antimony-doped tin oxides, Nat. Energy, 2024, 9(3), 308–315.
129. S. Seager, M. Schrenk and W. Bains, An astrophysical view of Earth-based metabolic biosignature gases, Astrobiology, 2012, 12(1), 61–82.
130. D. Reynard and H. Girault, Combined hydrogen production and electricity storage using a vanadium-manganese redox dual-flow battery, Cell Rep. Phys. Sci., 2021, 2(9), 100556.
131. S. Hedrich, M. Schlömann and D. B. Johnson, The iron-oxidizing proteobacteria, Microbiology, 2011, 157(6), 1551–1564.
132. G. Jerkiewicz, Standard and reversible hydrogen electrodes: Theory, design, operation, and applications, ACS Catal., 2020, 10(15), 8409–8417.

---