Regulators of aerobic and anaerobic methane oxidation in two pristine temperate peatland types
Peatlands are important ecosystems and have long been recognized to play a critical role in the global carbon (C) cycle (Armentano and Menges , Matthews and Fung , Moore and Knowles ). These waterlogged environments facilitate the accumulation of C in the form of partially decomposed plant material called peat (Clymo ). Peatlands store approximately one-third of the world's soil C, despite covering only 3%–4% of the Earth's land surface and their generally low primary productivity (UNEP ). Their high C sequestration capacity is attributed to low temperature and waterlogged, nutrient-poor conditions, which result in slow decomposition rates. At the same time, peatlands are a major source of greenhouse gases, including methane (CH 4 ), contributing ~30% to global CH 4 emissions (Saunois et al. , Canadell et al. ). Given that the global warming potential of CH 4 is 81.2 times higher than that of carbon dioxide (CO 2 ) on a 20-year timescale (Smith et al. ), even small changes in the atmospheric CH 4 concentration can have large effects on Earth's climate. The atmospheric levels of CH 4 nearly tripled since the onset of the industrial revolution ~270 years ago, with an alarming increase in the last 100 years (Lan et al. ). The emission of CH 4 from peatlands and other natural systems is mitigated by the activity of methanotrophs, microorganisms that utilize CH 4 as a source of C and energy (Conrad ). Methane is oxidized via both aerobic and anaerobic processes, depending on redox conditions and the availability of oxygen or alternative electron acceptors (Cui et al. , Dedysh and Knief , Guerrero-Cruz et al. ). Up to 50 to >90% of CH 4 produced by methanogens in anoxic soil zones can be consumed by methanotrophic bacteria and archaea before it reaches the atmosphere (Oremland and Culbertson , Reeburgh , Segarra et al. ). In anoxic zones, CH 4 can be oxidized by anaerobic CH 4 -oxidizing archaea (ANME) using electron acceptors such as nitrite (NO 2 − ), nitrate (NO 3 − ), sulfate (SO 4 2− ), ferric iron (Fe 3+ ), or humic substances, either independently or in association with a syntrophic partner (Beal et al. , Knittel and Boetius , Cui et al. , Scheller et al. , Valenzuela et al. ). Moreover, anaerobic bacterial methanotrophs in the order Methylomirabiliales perform anaerobic oxidation of CH 4 (AOM) using internally produced oxygen from the disproportionation of nitric oxide stemming from the reduction of NO 2 − /NO 3 − (Raghoebarsing et al. , Ettwig et al. ). Eventually, CH 4 that reaches the oxic zones can be oxidized by aerobic methanotrophic bacteria belonging to Alphaproteobacteria, Gammaproteobacteria, Methylacidiphilae , and Actinomycetes (so far only Candidatus Mycobacterium methanotrophicum proven as a bona fide methanotroph), which use oxygen as a terminal electron acceptor (Guerrero-Cruz et al. , van Spanning et al. ). Even though the importance of AOM has been well-researched and recognized in marine systems (Barnes and Goldberg , Reeburgh , Liang et al. , Aromokeye et al. , Xiao et al. ), the occurrence and importance of anaerobic CH 4 oxidizers in non-marine systems, especially peatlands, has received little attention. Consequently, our understanding of AOM in these ecosystems remains limited (Smemo and Yavitt , Zhu et al. , Zhong et al. , Shi et al. ). In particular, the terminal electron acceptors associated with AOM in peatlands are unresolved. As observed in other systems, SO 4 2− may be a favourable electron acceptor in peatlands (Liang et al. , Xiao et al. ). Some studies have identified SO 4 2− as an important electron acceptor for AOM in peatlands (Pester et al. , Segarra et al. , Miller et al. ), but there are also reports of no discernible impact or even inhibition of AOM following SO 4 2− addition to various peatland types (Blazewicz et al. , Gupta et al. ). Ferric iron is another candidate for a favourable electron acceptor in peat. It is prevalent in minerotrophic peatlands connected to groundwater flow in contact with iron-bearing rocks such as hematite, ferrihydrite, pyrite, and magnetite (Küsel et al. , Bhattacharyya et al. ). Furthermore, Fe³⁺ reduction has been reported to stimulate peat AOM in in vitro anoxic incubations (Miller et al. ); however, there have also been reports of AOM inhibition or no effect upon the addition of Fe³⁺ (Blazewicz et al. , Gupta et al. ). Additionally, organic compounds, such as humic acids, which are commonly found in peatlands, have been suggested as potential electron acceptors (Keller et al. , Rush et al. ). Even though the deposition of reactive nitrogen (N) in Europe has decreased since 1990, Central Europe still has deposition levels that are highly likely to cause harmful effects (Waldner et al. , EEA ). About 50% of the forested areas in the Czech Republic, including the sites studied here, receive >1 g N m −2 annually (Hůnová et al. ), exceeding the critical threshold for protecting Central European forests (Bobbink and Roelofs ). This significantly threatens natural, undisturbed peatlands (Bragazza ). Studies on methanotrophs in peatlands have reported inhibition of aerobic methane oxidation upon ammonium (NH 4 + ) addition (Crill et al. , Kravchenko , , Saari et al. ), but records of stimulation or no effect on potential CH 4 oxidation rates were also reported (Kravchenko , Nykänen et al. ). These inconsistent findings demonstrate that it is likely that the threshold at which reactive N (NH 4 + in particular) begins to affect aerobic CH 4 oxidation depends on a suite of additional environmental factors and community interactions (Ho et al. ). Here, we investigated the responses of aerobic and anaerobic methanotrophic communities to elevated N availability and to the addition of alternative electron acceptors, specifically Fe 3+ and SO 4 2− , in a fen and a spruce swamp peat. Both peatlands are pristine sites located in the Bohemian Forest National Park of the Czech Republic. They were chosen due to their difference in chemistry. Namely, the peat of fens from this region has been shown to contain more SO 4 2− and ferrous iron than peat of closely located swamps (Urbanová et al. ). Our study aimed to determine if (i) NH 4 + addition stimulates aerobic methanotrophic activities, (ii) the addition of SO 4 2− and Fe 3+ as alternative electron acceptors stimulates anaerobic methanotrophic activities, and (iii) to determine the community structures of aerobic and anaerobic CH 4 oxidizers. We hypothesized that (i) NH 4 + addition will stimulate aerobic CH 4 consumption rates in both peat types due to the low natural N availability in these ecosystems. Furthermore, we hypothesized that (ii) AOM will be more stimulated by the addition of SO 4 2− and Fe 3+ in the swamp than in the fen peat due to the lower natural availability of these electron acceptors in the swamp peat.   Peat sampling and chemical analyses Peat samples were collected from two pristine, distinct peatlands: a fen (49.0173572 N, 13.5673000E) and a spruce swamp forest (hereafter swamp; 49.0166411 N, 13.5685650E). These sites are located in the Bohemian Forest National Park in the southwestern part of the Czech Republic. The fen site is dominated by Sphagnum mats with Carex rostrata and Trientalis europaea . In contrast, the swamp site is mainly covered by Eriophorum spp., Sphagnum spp., Vaccinium myrtillus, Polytrichum spp., and scattered Picea abies . Previous comprehensive descriptions of these peat sites have been provided by Urbanová and Bárta . The sampling was conducted in June (summer) and October (autumn) 2021. The temperature of each peat core was recorded in situ using a digital thermometer. Three replicate cores per site were taken using a box corer (6.5 × 5.5 cm inner dimension) along a 20 m transect. In summer, we sampled the upper 10 cm peat layer, which is typically above the water table, after removing the top live vegetation that had not decomposed to study aerobic CH 4 oxidation. We collected the lower 10–30 cm layer, which is typically water-logged throughout the year, in autumn to investigate AOM. The soil from replicate peat cores was sealed in plastic bags to prevent water loss during cooled transportation to the laboratory, where they were homogenized. Peat surface water was collected during autumn sampling at the fen site using sterile glass bottles. After taking the material needed for the peat chemical analyses (see below), ~5 g of homogenized peat from each sample was stored at –20°C for DNA extraction. The remaining peat samples were stored in sealed plastic bags at 4°C for 2 weeks until used for incubation. The moisture content was determined by the weight loss method using 5 g of fresh peat (Burt ). The pH was measured by shaking 5 g of each peat sample in 15 ml of deionized water for two hours before measuring with a pH meter (Inolab pH 720, wtw 82 362). The organic matter content was determined by loss on ignition at 550°C (Ball ). The NH 4 + , SO 4 2− , NO 2 − , and NO 3 + concentrations were measured in peat extracts using a flow injection analyzer (FIA Lachat QC8500, Lachat Instruments). Peat extracts were obtained by shaking fresh peat in a 2 M KCl solution (3:8; wt:vol) at 120 r/m for one hour, followed by centrifugation (5000 × g for 20 min) and filtration through a 20 µm nylon mesh. Oxic peat incubations From each replicate soil core ( n = 6; 3 per site) we established two microcosms with 20 g of field-moist, homogenized peat from the upper layer (0–10 cm) in sterile 500 ml glass bottles capped with bromobutyl rubber stoppers (DWK Life Sciences) and plastic screw caps. This resulted in six microcosms per site. They were pre-incubated in the dark at 24 C (a typical summer temperature of this layer) for 48 h to allow for temperature acclimatization. Afterwards, all bottles were opened and three bottles per site were then amended with 460 µl of 20 mM ammonium chloride solution to raise the concentration in the peat pore water by 0.5 mM. Thereby we set up triplicate ammonium-amended and non-amended microcosms per site. Subsequently, the bottles were sealed again, and CH 4 was injected into all twelve bottles to obtain a headspace concentration of 0.5%. The bottles were then incubated for 5 days at 24°C in the dark without agitation. The CH 4 concentration in the headspace was quantified immediately after CH 4 addition and on every following incubation day using a gas chromatograph (Agilent Technologies 6850 GC System) equipped with a flame ionization detector and a capillary column (Model no. J & W 19095P-Q03, fused silica, 40 µm). The potential CH 4 oxidation rates [rate constants ( k )] were determined by applying the differential rate law for a first-order reaction. Anoxic peat incubations From each replicate soil core ( n = 6; 3 per site) we established three microcosms with 5 g of field-moist peat from the lower layer (10–30 cm) of the fen and swamp sites in sterile 130 ml glass bottles ( n = 18; 9 per site). Additionally, we established microcosms for two different sets of controls: viable peat ( n = 4; 2 per site) and non-viable peat microcosms ( n = 2; 1 per site). For the viable peat controls 5 g of peat was used and for the non-viable controls 5 g of a composite sample per site, obtained by combining equal amounts from the three replicate peat cores, was autoclaved for 60 min. To reproduce the high water table conditions observed at the fen site, which caused water-logged conditions below 10 cm, 10 ml of filtered fen peat water (0.2 µm polycarbonate membrane filters, ThermoFisher Scientific) were added to the bottles containing fen peat. This addition compensated for water loss, mimicking the submerged conditions observed in the fen's lower layer. Subsequently, all bottles were tightly closed with bromobutyl rubber stoppers and capped with plastic screw caps. Then, the bottles were evacuated and purged with N 2 (100% vol/vol) repeatedly for 20 min to remove ambient air, kept at 0.5 bar overpressure after the last N 2 flush, and placed at 4°C for 5 days before electron acceptor amendment. Triplicate bottles with viable fen and swamp peat were amended either with 1 ml of sterile 100 mM (fen) or 10 mM (swamp) solution of either ferric ammonium citrate or sterile sodium sulfate to keep the approximate natural level of difference in SO 4 2− and Fe 3+ concentrations between fen and swamp peat as determined here and by others (Urbanová et al. ), or left unamended. Bottles were pre-incubated for 4 days at 19°C for peat acclimatization and then flushed again with N 2 as described above. Then, 13 CH 4 (99 atom % 13 C, Campro Scientific) was injected into the sets of triplicate microcosms amended with SO 4 2− , Fe 3+ , or nothing to achieve a final headspace concentration of 0.5% vol/vol. No 13 CH 4 was injected into the headspaces of the controls. Thereby, the following five treatments were established per site: (i) triplicate microcosms with viable peat amended with Fe 3+ and 13 CH 4 , (ii) triplicate microcosms with viable peat amended with SO 4 2− and 13 CH 4 , (iii) triplicate microcosms with viable peat and 13 CH 4 , (iv) duplicate microcosms with viable peat, and (v) a microcosm with autoclaved non-viable peat. All bottles were incubated statically in the dark at 19°C (a typical summer temperature of this layer) for 50 days. Gas samples from the headspace were collected on days 4, 16, and 50 of incubation to measure the quantity of CO 2 and determine its stable C isotope composition (δ 13 C-CO 2 ). To collect the headspace samples, 3 ml of headspace gas samples from each time point were injected into 12 ml helium-flushed exetainers. Also, from the same samples, the concentrations of CH 4 were measured with a gas chromatograph (Agilent Technologies 6850 GC System) equipped with a flame ionization detector and a ShinCarbon ST Micropacked column (2 m, 0.53 mm ID, Restek). δ 13 CO 2 values were determined using a GasBench II system connected to a MAT253 Plus isotope ratio mass spectrometer via a Conflo IV interface (ThermoFisher Scientific). From each exetainer, triplicate injections (250 µl) serving as technical replicates were injected onto a Carboxen 1010 PLOT column (0.53 mm ID; Supelco, Bellefonte, USA) at 70°C to isolate CO 2 from other atmospheric gasses, with He as the carrier gas. The 13/12 C ratios in CO 2 were expressed in conventional delta-notation (δ 13 CO 2 ) after normalization of values to the Vienna Pee Dee Belemnite scale using international reference materials (IAEA-603 and IAEA-303), laboratory carbonate standards, and working gasses (0.3% CO 2 in He). The analytical error was <0.2‰. The δ 13 CO 2 values from control treatment without introduced 13 CH 4 were averaged per peat type and per timepoint ( n = 2) and resulting background δ 13 CO 2 values were subtracted from respective δ 13 CO 2 values in treatments with added 13 CH 4 in the headspace. At the end of the anoxic incubation, 5 g of peat from each bottle was stored at –20°C for DNA extraction. DNA extraction, PCR amplification, and sequencing Total DNA was extracted from 0.5 g of the upper (before oxic incubation) and lower (before and after anoxic incubation) layer peat samples, according to Angel et al. . Two independent extractions per sample were performed and pooled together. The DNA extracts were purified with the OneStep TM PCR inhibitor removal kit (Zymo Research) and quantified with the Quant-iT™ PicoGreen® dsDNA Assay Kit (ThermoFisher) according to the manufacturer's instructions. The 16S rRNA gene in each sample was amplified by polymerase chain reaction (PCR) in triplicates using the 515F-mod/806R-mod primers with CS1/CS2 tags (Walters et al. ). The PCR products were examined for a single band of the correct size on a 1.5% agarose gel and afterwards pooled. Library preparation and amplicon sequencing on an Illumina MiniSeq platform (2 × 150 bp cycle configuration; V2 reagent kit, Illumina) was performed at the DNA Services Facility of the University of Illinois, Chicago, USA. Sequence data processing and statistical analysis Primer sequences were removed from the raw amplicon sequence data using cutadapt v3.5 (Martin ). Further bioinformatic and statistical analyses were processed in R version 4.2 (R Core Team ). Quality trimming, dereplication, inferring sample composition, merging paired reads, chimera removal, and clustering into amplicon sequence variants (ASVs) were carried out using the DADA2 pipeline (v1.26.0) (Callahan et al. ). Taxonomic classification of the 16S rRNA ASVs was performed using the GTDB database (GTDB_bac120_arc122_ssu_r95) (Parks et al. ). To account for differences in sequencing depth between samples, the ASV count data was normalized using the sum total method (Gloor et al. ). Methanotroph ASVs were identified based on affiliation with known methanotrophic species and genera. To compare the methanotroph community structures, a non-metric multidimensional scaling analysis based on Bray–Curtis distances of relative ASV abundances was performed (Bray and Curtis , Oksanen et al. ). Furthermore, tests for significant differences between methanotroph community structures based on the Bray–Curtis distances were conducted using the adonis2 and pairwiseAdonis2 function of the package ‘vegan’ v2.6–4 (Oksanen et al. , Arbizu ), which calculates a permutational multivariate analyses of variance (PERMANOVA) with 999 permutations. To compare the mean potential aerobic CH 4 oxidation rates between sites and treatments, a one-way analysis of variance (ANOVA) followed by Tukey's HSD post-hoc test was performed using the ‘aov’ function from the base R stats package. Additionally, to compare the means of the edaphic properties between the two peat types, a paired t -test was employed using the ‘t.test’ function from the same package. Phylogenetic analysis To acquire a suitable set of representative full-length 16S rRNA gene sequences, all genomes of cultured and uncultured methanotrophs from the detected genera were downloaded from the databases of the National Center for Biotechnology Information (NCBI). Anvi'o (v7.1) and the below-detailed scripts were used to extract the 16S rRNA gene sequences from the downloaded genomes (Eren et al. ). The names of genome fasta files were simplified using anvi-script-reformat-fasta . Subsequently, contig databases were generated from the reformatted fasta files using anvi-gen-contigs-database . Gene annotation and extraction of full-length 16S rRNA gene sequences within the contigs database were carried out using anvi-run-hmms and anvi-get-sequences-for-hmm-hits , leveraging HMMER (Eddy ) and Prodigal (Hyatt et al. ). Of the downloaded 353 genomes belonging to 8 methanotroph genera , only 144 genomes contained 16S rRNA gene sequences. The extracted sequences were then used to form a local database using makeblastdb ( https://www.ncbi.nlm.nih.gov/books/NBK569841/ ). Representative ASV sequences (Table S2) were used to query the database using blastn ( https://www.ncbi.nlm.nih.gov/books/NBK279690/ ). Representative sequences for phylogenetic analyses were selected based on the top six BLASTN hits (≥340 bits alignment score, ≥90% sequence identity) of both cultured and uncultured methanotroph species, when present, for each representative ASV (for the full results of query see ). All sequences were aligned using MAFFT, adopting the L-INS-i method with the auto option (Katoh and Frith ), followed by trimming using trimAl (Capella-Gutiérrez et al. ). The final alignment contained 68 sequences (including the 16S rRNA gene sequence of Candidatus Methylacidiphilum infernorum IT5 (GCA_017310525.1), used as an outgroup) and included 252 and 1545 positions for the selected ASVs and representative full-length 16S rRNA gene sequences, respectively, after manual trimming with Jalview (Waterhouse et al. ). Subsequently, an approximate maximum likelihood phylogenetic tree was constructed using IQ-TREE (v2) (Minh et al. ) with 1000 ultrafast bootstrap iterations (Hoang et al. ) and using the TN+F+R3 model selected by ModelFinder (Kalyaanamoorthy et al. ). The final tree was visualized and refined using the iTOL software (Letunic and Bork ).   Peat samples were collected from two pristine, distinct peatlands: a fen (49.0173572 N, 13.5673000E) and a spruce swamp forest (hereafter swamp; 49.0166411 N, 13.5685650E). These sites are located in the Bohemian Forest National Park in the southwestern part of the Czech Republic. The fen site is dominated by Sphagnum mats with Carex rostrata and Trientalis europaea . In contrast, the swamp site is mainly covered by Eriophorum spp., Sphagnum spp., Vaccinium myrtillus, Polytrichum spp., and scattered Picea abies . Previous comprehensive descriptions of these peat sites have been provided by Urbanová and Bárta . The sampling was conducted in June (summer) and October (autumn) 2021. The temperature of each peat core was recorded in situ using a digital thermometer. Three replicate cores per site were taken using a box corer (6.5 × 5.5 cm inner dimension) along a 20 m transect. In summer, we sampled the upper 10 cm peat layer, which is typically above the water table, after removing the top live vegetation that had not decomposed to study aerobic CH 4 oxidation. We collected the lower 10–30 cm layer, which is typically water-logged throughout the year, in autumn to investigate AOM. The soil from replicate peat cores was sealed in plastic bags to prevent water loss during cooled transportation to the laboratory, where they were homogenized. Peat surface water was collected during autumn sampling at the fen site using sterile glass bottles. After taking the material needed for the peat chemical analyses (see below), ~5 g of homogenized peat from each sample was stored at –20°C for DNA extraction. The remaining peat samples were stored in sealed plastic bags at 4°C for 2 weeks until used for incubation. The moisture content was determined by the weight loss method using 5 g of fresh peat (Burt ). The pH was measured by shaking 5 g of each peat sample in 15 ml of deionized water for two hours before measuring with a pH meter (Inolab pH 720, wtw 82 362). The organic matter content was determined by loss on ignition at 550°C (Ball ). The NH 4 + , SO 4 2− , NO 2 − , and NO 3 + concentrations were measured in peat extracts using a flow injection analyzer (FIA Lachat QC8500, Lachat Instruments). Peat extracts were obtained by shaking fresh peat in a 2 M KCl solution (3:8; wt:vol) at 120 r/m for one hour, followed by centrifugation (5000 × g for 20 min) and filtration through a 20 µm nylon mesh.   From each replicate soil core ( n = 6; 3 per site) we established two microcosms with 20 g of field-moist, homogenized peat from the upper layer (0–10 cm) in sterile 500 ml glass bottles capped with bromobutyl rubber stoppers (DWK Life Sciences) and plastic screw caps. This resulted in six microcosms per site. They were pre-incubated in the dark at 24 C (a typical summer temperature of this layer) for 48 h to allow for temperature acclimatization. Afterwards, all bottles were opened and three bottles per site were then amended with 460 µl of 20 mM ammonium chloride solution to raise the concentration in the peat pore water by 0.5 mM. Thereby we set up triplicate ammonium-amended and non-amended microcosms per site. Subsequently, the bottles were sealed again, and CH 4 was injected into all twelve bottles to obtain a headspace concentration of 0.5%. The bottles were then incubated for 5 days at 24°C in the dark without agitation. The CH 4 concentration in the headspace was quantified immediately after CH 4 addition and on every following incubation day using a gas chromatograph (Agilent Technologies 6850 GC System) equipped with a flame ionization detector and a capillary column (Model no. J & W 19095P-Q03, fused silica, 40 µm). The potential CH 4 oxidation rates [rate constants ( k )] were determined by applying the differential rate law for a first-order reaction.   From each replicate soil core ( n = 6; 3 per site) we established three microcosms with 5 g of field-moist peat from the lower layer (10–30 cm) of the fen and swamp sites in sterile 130 ml glass bottles ( n = 18; 9 per site). Additionally, we established microcosms for two different sets of controls: viable peat ( n = 4; 2 per site) and non-viable peat microcosms ( n = 2; 1 per site). For the viable peat controls 5 g of peat was used and for the non-viable controls 5 g of a composite sample per site, obtained by combining equal amounts from the three replicate peat cores, was autoclaved for 60 min. To reproduce the high water table conditions observed at the fen site, which caused water-logged conditions below 10 cm, 10 ml of filtered fen peat water (0.2 µm polycarbonate membrane filters, ThermoFisher Scientific) were added to the bottles containing fen peat. This addition compensated for water loss, mimicking the submerged conditions observed in the fen's lower layer. Subsequently, all bottles were tightly closed with bromobutyl rubber stoppers and capped with plastic screw caps. Then, the bottles were evacuated and purged with N 2 (100% vol/vol) repeatedly for 20 min to remove ambient air, kept at 0.5 bar overpressure after the last N 2 flush, and placed at 4°C for 5 days before electron acceptor amendment. Triplicate bottles with viable fen and swamp peat were amended either with 1 ml of sterile 100 mM (fen) or 10 mM (swamp) solution of either ferric ammonium citrate or sterile sodium sulfate to keep the approximate natural level of difference in SO 4 2− and Fe 3+ concentrations between fen and swamp peat as determined here and by others (Urbanová et al. ), or left unamended. Bottles were pre-incubated for 4 days at 19°C for peat acclimatization and then flushed again with N 2 as described above. Then, 13 CH 4 (99 atom % 13 C, Campro Scientific) was injected into the sets of triplicate microcosms amended with SO 4 2− , Fe 3+ , or nothing to achieve a final headspace concentration of 0.5% vol/vol. No 13 CH 4 was injected into the headspaces of the controls. Thereby, the following five treatments were established per site: (i) triplicate microcosms with viable peat amended with Fe 3+ and 13 CH 4 , (ii) triplicate microcosms with viable peat amended with SO 4 2− and 13 CH 4 , (iii) triplicate microcosms with viable peat and 13 CH 4 , (iv) duplicate microcosms with viable peat, and (v) a microcosm with autoclaved non-viable peat. All bottles were incubated statically in the dark at 19°C (a typical summer temperature of this layer) for 50 days. Gas samples from the headspace were collected on days 4, 16, and 50 of incubation to measure the quantity of CO 2 and determine its stable C isotope composition (δ 13 C-CO 2 ). To collect the headspace samples, 3 ml of headspace gas samples from each time point were injected into 12 ml helium-flushed exetainers. Also, from the same samples, the concentrations of CH 4 were measured with a gas chromatograph (Agilent Technologies 6850 GC System) equipped with a flame ionization detector and a ShinCarbon ST Micropacked column (2 m, 0.53 mm ID, Restek). δ 13 CO 2 values were determined using a GasBench II system connected to a MAT253 Plus isotope ratio mass spectrometer via a Conflo IV interface (ThermoFisher Scientific). From each exetainer, triplicate injections (250 µl) serving as technical replicates were injected onto a Carboxen 1010 PLOT column (0.53 mm ID; Supelco, Bellefonte, USA) at 70°C to isolate CO 2 from other atmospheric gasses, with He as the carrier gas. The 13/12 C ratios in CO 2 were expressed in conventional delta-notation (δ 13 CO 2 ) after normalization of values to the Vienna Pee Dee Belemnite scale using international reference materials (IAEA-603 and IAEA-303), laboratory carbonate standards, and working gasses (0.3% CO 2 in He). The analytical error was <0.2‰. The δ 13 CO 2 values from control treatment without introduced 13 CH 4 were averaged per peat type and per timepoint ( n = 2) and resulting background δ 13 CO 2 values were subtracted from respective δ 13 CO 2 values in treatments with added 13 CH 4 in the headspace. At the end of the anoxic incubation, 5 g of peat from each bottle was stored at –20°C for DNA extraction.   Total DNA was extracted from 0.5 g of the upper (before oxic incubation) and lower (before and after anoxic incubation) layer peat samples, according to Angel et al. . Two independent extractions per sample were performed and pooled together. The DNA extracts were purified with the OneStep TM PCR inhibitor removal kit (Zymo Research) and quantified with the Quant-iT™ PicoGreen® dsDNA Assay Kit (ThermoFisher) according to the manufacturer's instructions. The 16S rRNA gene in each sample was amplified by polymerase chain reaction (PCR) in triplicates using the 515F-mod/806R-mod primers with CS1/CS2 tags (Walters et al. ). The PCR products were examined for a single band of the correct size on a 1.5% agarose gel and afterwards pooled. Library preparation and amplicon sequencing on an Illumina MiniSeq platform (2 × 150 bp cycle configuration; V2 reagent kit, Illumina) was performed at the DNA Services Facility of the University of Illinois, Chicago, USA.   Primer sequences were removed from the raw amplicon sequence data using cutadapt v3.5 (Martin ). Further bioinformatic and statistical analyses were processed in R version 4.2 (R Core Team ). Quality trimming, dereplication, inferring sample composition, merging paired reads, chimera removal, and clustering into amplicon sequence variants (ASVs) were carried out using the DADA2 pipeline (v1.26.0) (Callahan et al. ). Taxonomic classification of the 16S rRNA ASVs was performed using the GTDB database (GTDB_bac120_arc122_ssu_r95) (Parks et al. ). To account for differences in sequencing depth between samples, the ASV count data was normalized using the sum total method (Gloor et al. ). Methanotroph ASVs were identified based on affiliation with known methanotrophic species and genera. To compare the methanotroph community structures, a non-metric multidimensional scaling analysis based on Bray–Curtis distances of relative ASV abundances was performed (Bray and Curtis , Oksanen et al. ). Furthermore, tests for significant differences between methanotroph community structures based on the Bray–Curtis distances were conducted using the adonis2 and pairwiseAdonis2 function of the package ‘vegan’ v2.6–4 (Oksanen et al. , Arbizu ), which calculates a permutational multivariate analyses of variance (PERMANOVA) with 999 permutations. To compare the mean potential aerobic CH 4 oxidation rates between sites and treatments, a one-way analysis of variance (ANOVA) followed by Tukey's HSD post-hoc test was performed using the ‘aov’ function from the base R stats package. Additionally, to compare the means of the edaphic properties between the two peat types, a paired t -test was employed using the ‘t.test’ function from the same package.   To acquire a suitable set of representative full-length 16S rRNA gene sequences, all genomes of cultured and uncultured methanotrophs from the detected genera were downloaded from the databases of the National Center for Biotechnology Information (NCBI). Anvi'o (v7.1) and the below-detailed scripts were used to extract the 16S rRNA gene sequences from the downloaded genomes (Eren et al. ). The names of genome fasta files were simplified using anvi-script-reformat-fasta . Subsequently, contig databases were generated from the reformatted fasta files using anvi-gen-contigs-database . Gene annotation and extraction of full-length 16S rRNA gene sequences within the contigs database were carried out using anvi-run-hmms and anvi-get-sequences-for-hmm-hits , leveraging HMMER (Eddy ) and Prodigal (Hyatt et al. ). Of the downloaded 353 genomes belonging to 8 methanotroph genera , only 144 genomes contained 16S rRNA gene sequences. The extracted sequences were then used to form a local database using makeblastdb ( https://www.ncbi.nlm.nih.gov/books/NBK569841/ ). Representative ASV sequences (Table S2) were used to query the database using blastn ( https://www.ncbi.nlm.nih.gov/books/NBK279690/ ). Representative sequences for phylogenetic analyses were selected based on the top six BLASTN hits (≥340 bits alignment score, ≥90% sequence identity) of both cultured and uncultured methanotroph species, when present, for each representative ASV (for the full results of query see ). All sequences were aligned using MAFFT, adopting the L-INS-i method with the auto option (Katoh and Frith ), followed by trimming using trimAl (Capella-Gutiérrez et al. ). The final alignment contained 68 sequences (including the 16S rRNA gene sequence of Candidatus Methylacidiphilum infernorum IT5 (GCA_017310525.1), used as an outgroup) and included 252 and 1545 positions for the selected ASVs and representative full-length 16S rRNA gene sequences, respectively, after manual trimming with Jalview (Waterhouse et al. ). Subsequently, an approximate maximum likelihood phylogenetic tree was constructed using IQ-TREE (v2) (Minh et al. ) with 1000 ultrafast bootstrap iterations (Hoang et al. ) and using the TN+F+R3 model selected by ModelFinder (Kalyaanamoorthy et al. ). The final tree was visualized and refined using the iTOL software (Letunic and Bork ).   Edaphic properties The pH of both the fen and the swamp peat was in the acidic range (3.94–4.51) and comparable between both layers at each site (Table ). Temperature and organic matter content were not significantly different between the upper layers of the fen and swamp sites. The lower peat layer of the swamp site contained significantly less water and SO 4 2 than the same layer of the fen site. Effect of elevated N availability on aerobic CH 4 oxidation potential To test whether aerobic methanotrophs were N-limited in the upper peat layers, we amended controlled microcosm incubations with CH 4 only or with both CH 4 and NH 4 + . In the fen and swamp peat incubations without NH 4 + amendment, 91% and 48% of the supplied CH 4 were consumed, respectively, over the 5-day incubation period (Fig. ). Methane consumption was consistently faster in the fen peat (rate constants of −0.54 ± 0.10 and −0.38 ± 0.11 day −1 , without and with NH 4 + addition, respectively) than in the swamp peat (−0.10 ± 0.04 and −0.17 ± 0.07 day −1 , without and with NH 4 + addition, respectively). However, there was no significant effect of NH 4 + amendment to either peat. Methanotrophic community structure in the upper peat layers The relative abundances of methanotrophs constituted 5.3% and 2.6% of the total prokaryotic community in the upper layer of the fen and swamp peat, respectively (Fig. ). We identified methanotrophic bacteria belonging to the three orders Methylococcales, Methylomirabilales , and Rhizobiales in both peat samples. Within these orders, six methanotrophic bacterial genera were identified: Methylocella, Methylocystis, Methyloferula, Methylobacter, Methylomonas , and Methylovulum next to ASVs without an assigned genus belonging to the orders Methylomirabilales , and Methylococcales. A detailed overview of the phylogeny of the detected methanotrophic ASVs is given in the supplement . Members of Methylocystis and Methylocella were the methanotrophs with the highest relative abundance in both peat samples, with the fen peat showing nearly twice as high relative abundances of Methylocystis - and Methylocella -like ASVs (3.04% and 1.29%, respectively) than the swamp peat (1.70% and 0.79%, respectively). The methanotrophic community structures did not significantly differ between the fen and swamp peat samples according to multivariate analysis of variance (PERMANOVA, R 2 = 0.03, P = .06). However, they exhibited a clear separation in the ordination analysis (non-metric multidimensional scaling, Fig. ), with swamp methanotroph communities being more similar to each other than the fen communities. A general overview of the total detected prokaryotic community is provided in the supplement . Effect of electron acceptor addition on anaerobic CH 4 oxidation potential Anoxic incubations of the peat samples from the lower layer with and without the addition of 13 CH 4 and two alternative electron acceptors were conducted to investigate the peat's potential for AOM. In the absence of an added electron acceptor, production of 13 CO 2 was detectable only in the fen peat incubations from day 16 onwards, demonstrating the anaerobic oxidation of 13 CH 4 [Fig. (iii)]. In the presence of added SO 4 2− , first signs of AOM were detectable on day 16 of the incubation in the fen peat and already on day four in the the swamp peat. However, the extent of AOM activity in the fen peat was not significantly different from the non-amended control incubation. Adding Fe 3+ stimulated AOM in swamp peat while inhibiting AOM in fen peat [Fig. ]. No 13 CO 2 formation was observed in viable controls, and no CH 4 production was detected in non-viable controls (data not shown). However, we did observe CH 4 production in the viable controls of the fen peat to a concentration of ~61 mmol CH 4 g dw −1 at the end of the incubation on day 50. Methanotrophic community structure in the lower peat layers before and after anoxic incubation Field-fresh, lower-layer peat from the fen and swamp sites contained methanotrophic bacteria of the orders Methylococcales, Methylomirabilales , and Rhizobiales from the phyla Methylomirabilota and Pseudomonadota (Fig. ; ). Within these orders, the same methanotrophic genera were identified as in the upper peat layers (see also ). Likewise, members of Methylocystis and Methylocella showed the highest average relative abundance in both the fen (4.59% and 2.00%, respectively) and in the swamp peat (6.79% and 2.16%, respectively). Similar to the upper layer, the methanotroph communities in the lower layers differed significantly between the two sites as determined by analyses of variance (PERMANOVA, R 2 = 0.15, P = .001) and by ordination analysis (Fig. ). We additionally observed significant differences in methanotroph community structures before compared to after the anoxic incubation for nearly all treatments in both peat types (supplementary ). Merely the community in swamp peat incubated with SO 4 2− and CH 4 did not change in community structure. Moreover, there were apparent increases in the average relative abundance of some specific aerobic methanotrophs. For instance, ASVs assigned to the genus Methylomonas had significantly increased at the end of the fen peat incubation of only 13 CH 4 -amended and non-amended viable controls (eight-fold; P = .001; Fig. ). Likewise, ASVs of the Methylocystis genus increased in relative abundance in the SO 4 2− -amended treatment of fen peat (1.4-fold; P = .04). A general overview of the total detected prokaryotic community is given in the supplement .   The pH of both the fen and the swamp peat was in the acidic range (3.94–4.51) and comparable between both layers at each site (Table ). Temperature and organic matter content were not significantly different between the upper layers of the fen and swamp sites. The lower peat layer of the swamp site contained significantly less water and SO 4 2 than the same layer of the fen site.   4 oxidation potential To test whether aerobic methanotrophs were N-limited in the upper peat layers, we amended controlled microcosm incubations with CH 4 only or with both CH 4 and NH 4 + . In the fen and swamp peat incubations without NH 4 + amendment, 91% and 48% of the supplied CH 4 were consumed, respectively, over the 5-day incubation period (Fig. ). Methane consumption was consistently faster in the fen peat (rate constants of −0.54 ± 0.10 and −0.38 ± 0.11 day −1 , without and with NH 4 + addition, respectively) than in the swamp peat (−0.10 ± 0.04 and −0.17 ± 0.07 day −1 , without and with NH 4 + addition, respectively). However, there was no significant effect of NH 4 + amendment to either peat.   The relative abundances of methanotrophs constituted 5.3% and 2.6% of the total prokaryotic community in the upper layer of the fen and swamp peat, respectively (Fig. ). We identified methanotrophic bacteria belonging to the three orders Methylococcales, Methylomirabilales , and Rhizobiales in both peat samples. Within these orders, six methanotrophic bacterial genera were identified: Methylocella, Methylocystis, Methyloferula, Methylobacter, Methylomonas , and Methylovulum next to ASVs without an assigned genus belonging to the orders Methylomirabilales , and Methylococcales. A detailed overview of the phylogeny of the detected methanotrophic ASVs is given in the supplement . Members of Methylocystis and Methylocella were the methanotrophs with the highest relative abundance in both peat samples, with the fen peat showing nearly twice as high relative abundances of Methylocystis - and Methylocella -like ASVs (3.04% and 1.29%, respectively) than the swamp peat (1.70% and 0.79%, respectively). The methanotrophic community structures did not significantly differ between the fen and swamp peat samples according to multivariate analysis of variance (PERMANOVA, R 2 = 0.03, P = .06). However, they exhibited a clear separation in the ordination analysis (non-metric multidimensional scaling, Fig. ), with swamp methanotroph communities being more similar to each other than the fen communities. A general overview of the total detected prokaryotic community is provided in the supplement .   4 oxidation potential Anoxic incubations of the peat samples from the lower layer with and without the addition of 13 CH 4 and two alternative electron acceptors were conducted to investigate the peat's potential for AOM. In the absence of an added electron acceptor, production of 13 CO 2 was detectable only in the fen peat incubations from day 16 onwards, demonstrating the anaerobic oxidation of 13 CH 4 [Fig. (iii)]. In the presence of added SO 4 2− , first signs of AOM were detectable on day 16 of the incubation in the fen peat and already on day four in the the swamp peat. However, the extent of AOM activity in the fen peat was not significantly different from the non-amended control incubation. Adding Fe 3+ stimulated AOM in swamp peat while inhibiting AOM in fen peat [Fig. ]. No 13 CO 2 formation was observed in viable controls, and no CH 4 production was detected in non-viable controls (data not shown). However, we did observe CH 4 production in the viable controls of the fen peat to a concentration of ~61 mmol CH 4 g dw −1 at the end of the incubation on day 50.   Field-fresh, lower-layer peat from the fen and swamp sites contained methanotrophic bacteria of the orders Methylococcales, Methylomirabilales , and Rhizobiales from the phyla Methylomirabilota and Pseudomonadota (Fig. ; ). Within these orders, the same methanotrophic genera were identified as in the upper peat layers (see also ). Likewise, members of Methylocystis and Methylocella showed the highest average relative abundance in both the fen (4.59% and 2.00%, respectively) and in the swamp peat (6.79% and 2.16%, respectively). Similar to the upper layer, the methanotroph communities in the lower layers differed significantly between the two sites as determined by analyses of variance (PERMANOVA, R 2 = 0.15, P = .001) and by ordination analysis (Fig. ). We additionally observed significant differences in methanotroph community structures before compared to after the anoxic incubation for nearly all treatments in both peat types (supplementary ). Merely the community in swamp peat incubated with SO 4 2− and CH 4 did not change in community structure. Moreover, there were apparent increases in the average relative abundance of some specific aerobic methanotrophs. For instance, ASVs assigned to the genus Methylomonas had significantly increased at the end of the fen peat incubation of only 13 CH 4 -amended and non-amended viable controls (eight-fold; P = .001; Fig. ). Likewise, ASVs of the Methylocystis genus increased in relative abundance in the SO 4 2− -amended treatment of fen peat (1.4-fold; P = .04). A general overview of the total detected prokaryotic community is given in the supplement .   Methanotrophic communities in upper and lower peat layers Down to 30 cm into the peat, alphaproteobacterial methanotrophs of the genera Methylocystis and Methylocella were the dominant methanotrophs, which is consistent with prior research in the same region (Urbanová and Bárta , Chroňáková et al. ). The observed similarity of methanotrophic community structure over this depth may reflect similarity in peat conditions. This could be explained, in part, by substantial water level fluctuations known to occur at the sites, which can lead to the equalization of redox conditions and microbiological processes in the affected peat layers (Urbanová et al. ). The high relative abundance of species from the Methylocystis and Methylocella genera is likely enabled by their high level of adaptation to low pH, as methanotrophs of the same genera have been found to be abundant and active in other acidic peatlands and under acidic culturing conditions (Dedysh et al. , , , , Chen et al. , Jaatinen et al. , Putkinen et al. ). An additional factor may be the capacity of Methylocystis and Methylocella to grow mixotrophically by co-oxidizing hydrogen and CH 4 , allowing them to utilize available resources effectively in nutrient-limited environments (Carere et al. , Hakobyan et al. ). Furthermore, they possess the ability to use single-carbon compounds other than CH 4 , as well as multi-carbon compounds and hydrogen sulfide for growth (Dedysh et al. , Im et al. , Farhan Ul Haque et al. , Jung et al. , Picone et al. , Awala et al. , Schmitz et al. , ). We detected two ASVs of anaerobic bacteria from the order Methylomirabilales in both studied peat lower layers. Bacteria of the only described genus in this order, Methylomirabilis , are known as intra-aerobic methanotrophs that couple CH 4 -oxidation to the reduction of NO 3 − , NO 2 − , and potentially iodate in the absence of oxygen and possess all genes needed for aerobic CH 4 oxidation (Ettwig et al. , , Zhu et al. , , Yao et al. ). They regulate emissions of greenhouse gases from anoxic ecosystems by oxidizing CH 4 into CO 2 and converting NO 2 − directly into dinitrogen, which minimizes nitrous oxide production (Ettwig et al. ). The two detected ASVs (Seq_6811 and Seq_929) share 96% and 98% identity with the 16S rRNA gene sequence in the metagenome-assembled genome of the uncultured Candidatus Methylomirabilis spp. SMAG_U2938 retrieved from agricultural soil  as their closest identified relative over ~250 aligned nucleotide positions. Within this closely related metagenome-assembled genome (CheckM completeness 79%), there is no CH 4 monooxygenase identified, indicating that these microorganisms may not partake in CH 4 oxidation. Furthermore, both ASV sequences exhibit <91% identity and cluster separately from the cultured representatives in the genus Methylomirabilis , making it likely that they stem from a different family or genus . However, it is impossible to speculate about the role of the detected Methylomirabilales organisms in this study for AOM before more genomic information is available. In other temperate peatlands, ANME archaea were found (de Jong et al. , Weil et al. ), but we did not detect any ANME archaeal members in our study, implying their absence or very low general abundance in the studied peat. The primers used in this study target a large fraction of the known ANME archaeal diversity (Walters et al. , Su et al. ); however, we cannot wholly exclude the detection limits of the method. The absence of detectable ANME archaea and the low presence of Methylomirabilales , which moreover likely belong to a different family or genus than the known anaerobic CH 4 oxidizers, suggests that the observed AOM may at least in part be driven by yet unidentified methanotrophs. Aerobic CH 4 oxidation potential and the impact of N-addition The CH 4 oxidation potential in peat from the fen site was ~1.5-fold higher than in peat from the swamp site. Likely, this higher CH 4 oxidation potential is supported by the relatively more abundant and diverse methanotrophs in the fen peat. Moreover, this aligns well with higher CH 4 and CO 2 production potentials reported for fens in previous studies on the same region (Urbanová et al. , Urbanová et al. ). Additionally, spruce swamps were shown to suppress methanogens and, in turn, limit methanotroph activity and growth due to their lower water table and greater aeration, as also is the case for our study site (Chroňáková et al. ). In contrast to our initial hypothesis, NH 4 + addition had no significant effect on CH 4 oxidation potentials in peat from either peatland type. First, this result shows that CH 4 oxidation in the studied peatlands was not limited by N availability. Secondly, it implicates that neither were the CH 4 monooxygenase enzymes of aerobic methanotrophs competitively inhibited by the added NH 4 + nor were the methanotrophs subjected to toxic levels of nitrification products upon the addition of NH 4 + , as seen previously (Kravchenko ). Anaerobic oxidation of CH 4 and electron acceptor preferences AOM is known to occur in peatlands, where oxygen often becomes a limited resource due to prolonged waterlogging, which makes it a vital process for reducing CH 4 emissions from these systems. Different electron acceptors, such as NO 3 − , NO 2 − , Fe 3+ , SO 4 2− , and humic substances, have been found to support AOM in terrestrial systems (Raghoebarsing et al. , Keller et al. , Pozdnyakov et al. , Zhu et al. , Segarra et al. , Miller et al. 2019 , Rush et al. , Fan et al. ). In this study, anoxic incubations with added SO 4 2− and 13 C-CH 4 showed AOM in peat from both peatland types. However, the AOM activity in the fen peat amended with SO 4 2− was comparable to that of unamended control incubations. Therefore, it remains unclear whether AOM was driven by the added SO 4 2− , the SO 4 2− that is naturally present in the fen peat, by other internally available electron acceptors such as NO 3 − , or forms of organic matter, or perhaps by a mixture of these. In the same peat, AOM was inhibited by the addition of Fe 3+ . This dynamic has been observed before (Blazewicz et al. ) and could be due to a change in redox conditions leading to thermodynamic shifts favouring metabolic processes other than AOM. In contrast, AOM in the swamp peat was stimulated by the additions of both SO 4 2− and Fe 3+ , which partially confirms our hypothesis of AOM in this peat being more stimulated by the addition of alternative electron acceptors. Natural insoluble sources of Fe 3+ (e.g. ferrihydrite) typically release this electron acceptor less readily and have been shown to support AOM in the long term (He et al. ). In contrast, the soluble ferric ammonium citrate used in this experiment likely stimulated AOM activity instantly. Naturally occurring AOM coupled to Fe 3+ reduction may therefore take place at a slower rate than observed in our laboratory experiment. We can rule out the support of internal electron acceptors for AOM in the swamp peat, as the incubation control amended with only 13 CH 4 showed no AOM activity. We, therefore, conclude that both peatland types hold the potential for AOM but that the absence of suitable electron acceptors in the swamp peat severely limits AOM. The only detected potential anaerobic methanotrophs in both peatlands were members of the order Methylomirabilales , albeit in low relative abundance and with lower than genus-level 16S rRNA gene sequence identity to known methanotrophs of the order. They did not exhibit a discernable pattern related to treatment in the anoxic incubations, and their involvement in the observed AOM remains unresolved. The relative abundance of Methylocystis -like and Methylomonas -like methanotrophs increased in the fen peat amended with SO 4 2− , with 13 CH 4 only, and in the viable control incubations (no addition of electron acceptors nor 13 CH 4 ). Considering that AOM occurred in fen incubations even without electron acceptor amendment and that CH 4 was produced in the viable controls, it is tempting to speculate about the involvement of these classically considered aerobic methanotrophs in the observed AOM activity. Recent findings support this notion as both genera have been described to be capable of anaerobically oxidizing CH 4 using NO 3 − or NO 2 − as terminal electron acceptors (Kits et al. , Liu et al. ). Additionally, there are reports of aerobic CH 4 oxidizers, including Methylomonas spp., growing during and participating in AOM by switching to other electron acceptors such as Fe 3+ and SO 4 2− , either alone or in association with metal-reducing bacteria (Oswald et al. , , Cabrol et al. , Zheng et al. ). We did note an 8- and 17-fold increase in the mean relative abundance of the SO 4 2− reducer Desulforhopalu s and the dissimilatory Fe 3+ reducer Telmatospirillum in fen incubations amended with SO 4 2− and Fe 3+ , respectively . However, linking their potential activity with CH 4 oxidation by (aerobic) methanotrophs via extracellular electron transfer goes beyond the scope of this work. Finally, recent works have shown that methanobactin, a small copper-binding peptide, which is produced by Methylocystis among other methanotrophs, can function to generate O 2 via metal ion reduction and thereby enable AOM (Dershwitz et al. ).   Down to 30 cm into the peat, alphaproteobacterial methanotrophs of the genera Methylocystis and Methylocella were the dominant methanotrophs, which is consistent with prior research in the same region (Urbanová and Bárta , Chroňáková et al. ). The observed similarity of methanotrophic community structure over this depth may reflect similarity in peat conditions. This could be explained, in part, by substantial water level fluctuations known to occur at the sites, which can lead to the equalization of redox conditions and microbiological processes in the affected peat layers (Urbanová et al. ). The high relative abundance of species from the Methylocystis and Methylocella genera is likely enabled by their high level of adaptation to low pH, as methanotrophs of the same genera have been found to be abundant and active in other acidic peatlands and under acidic culturing conditions (Dedysh et al. , , , , Chen et al. , Jaatinen et al. , Putkinen et al. ). An additional factor may be the capacity of Methylocystis and Methylocella to grow mixotrophically by co-oxidizing hydrogen and CH 4 , allowing them to utilize available resources effectively in nutrient-limited environments (Carere et al. , Hakobyan et al. ). Furthermore, they possess the ability to use single-carbon compounds other than CH 4 , as well as multi-carbon compounds and hydrogen sulfide for growth (Dedysh et al. , Im et al. , Farhan Ul Haque et al. , Jung et al. , Picone et al. , Awala et al. , Schmitz et al. , ). We detected two ASVs of anaerobic bacteria from the order Methylomirabilales in both studied peat lower layers. Bacteria of the only described genus in this order, Methylomirabilis , are known as intra-aerobic methanotrophs that couple CH 4 -oxidation to the reduction of NO 3 − , NO 2 − , and potentially iodate in the absence of oxygen and possess all genes needed for aerobic CH 4 oxidation (Ettwig et al. , , Zhu et al. , , Yao et al. ). They regulate emissions of greenhouse gases from anoxic ecosystems by oxidizing CH 4 into CO 2 and converting NO 2 − directly into dinitrogen, which minimizes nitrous oxide production (Ettwig et al. ). The two detected ASVs (Seq_6811 and Seq_929) share 96% and 98% identity with the 16S rRNA gene sequence in the metagenome-assembled genome of the uncultured Candidatus Methylomirabilis spp. SMAG_U2938 retrieved from agricultural soil  as their closest identified relative over ~250 aligned nucleotide positions. Within this closely related metagenome-assembled genome (CheckM completeness 79%), there is no CH 4 monooxygenase identified, indicating that these microorganisms may not partake in CH 4 oxidation. Furthermore, both ASV sequences exhibit <91% identity and cluster separately from the cultured representatives in the genus Methylomirabilis , making it likely that they stem from a different family or genus . However, it is impossible to speculate about the role of the detected Methylomirabilales organisms in this study for AOM before more genomic information is available. In other temperate peatlands, ANME archaea were found (de Jong et al. , Weil et al. ), but we did not detect any ANME archaeal members in our study, implying their absence or very low general abundance in the studied peat. The primers used in this study target a large fraction of the known ANME archaeal diversity (Walters et al. , Su et al. ); however, we cannot wholly exclude the detection limits of the method. The absence of detectable ANME archaea and the low presence of Methylomirabilales , which moreover likely belong to a different family or genus than the known anaerobic CH 4 oxidizers, suggests that the observed AOM may at least in part be driven by yet unidentified methanotrophs.   4 oxidation potential and the impact of N-addition The CH 4 oxidation potential in peat from the fen site was ~1.5-fold higher than in peat from the swamp site. Likely, this higher CH 4 oxidation potential is supported by the relatively more abundant and diverse methanotrophs in the fen peat. Moreover, this aligns well with higher CH 4 and CO 2 production potentials reported for fens in previous studies on the same region (Urbanová et al. , Urbanová et al. ). Additionally, spruce swamps were shown to suppress methanogens and, in turn, limit methanotroph activity and growth due to their lower water table and greater aeration, as also is the case for our study site (Chroňáková et al. ). In contrast to our initial hypothesis, NH 4 + addition had no significant effect on CH 4 oxidation potentials in peat from either peatland type. First, this result shows that CH 4 oxidation in the studied peatlands was not limited by N availability. Secondly, it implicates that neither were the CH 4 monooxygenase enzymes of aerobic methanotrophs competitively inhibited by the added NH 4 + nor were the methanotrophs subjected to toxic levels of nitrification products upon the addition of NH 4 + , as seen previously (Kravchenko ).   4 and electron acceptor preferences AOM is known to occur in peatlands, where oxygen often becomes a limited resource due to prolonged waterlogging, which makes it a vital process for reducing CH 4 emissions from these systems. Different electron acceptors, such as NO 3 − , NO 2 − , Fe 3+ , SO 4 2− , and humic substances, have been found to support AOM in terrestrial systems (Raghoebarsing et al. , Keller et al. , Pozdnyakov et al. , Zhu et al. , Segarra et al. , Miller et al. 2019 , Rush et al. , Fan et al. ). In this study, anoxic incubations with added SO 4 2− and 13 C-CH 4 showed AOM in peat from both peatland types. However, the AOM activity in the fen peat amended with SO 4 2− was comparable to that of unamended control incubations. Therefore, it remains unclear whether AOM was driven by the added SO 4 2− , the SO 4 2− that is naturally present in the fen peat, by other internally available electron acceptors such as NO 3 − , or forms of organic matter, or perhaps by a mixture of these. In the same peat, AOM was inhibited by the addition of Fe 3+ . This dynamic has been observed before (Blazewicz et al. ) and could be due to a change in redox conditions leading to thermodynamic shifts favouring metabolic processes other than AOM. In contrast, AOM in the swamp peat was stimulated by the additions of both SO 4 2− and Fe 3+ , which partially confirms our hypothesis of AOM in this peat being more stimulated by the addition of alternative electron acceptors. Natural insoluble sources of Fe 3+ (e.g. ferrihydrite) typically release this electron acceptor less readily and have been shown to support AOM in the long term (He et al. ). In contrast, the soluble ferric ammonium citrate used in this experiment likely stimulated AOM activity instantly. Naturally occurring AOM coupled to Fe 3+ reduction may therefore take place at a slower rate than observed in our laboratory experiment. We can rule out the support of internal electron acceptors for AOM in the swamp peat, as the incubation control amended with only 13 CH 4 showed no AOM activity. We, therefore, conclude that both peatland types hold the potential for AOM but that the absence of suitable electron acceptors in the swamp peat severely limits AOM. The only detected potential anaerobic methanotrophs in both peatlands were members of the order Methylomirabilales , albeit in low relative abundance and with lower than genus-level 16S rRNA gene sequence identity to known methanotrophs of the order. They did not exhibit a discernable pattern related to treatment in the anoxic incubations, and their involvement in the observed AOM remains unresolved. The relative abundance of Methylocystis -like and Methylomonas -like methanotrophs increased in the fen peat amended with SO 4 2− , with 13 CH 4 only, and in the viable control incubations (no addition of electron acceptors nor 13 CH 4 ). Considering that AOM occurred in fen incubations even without electron acceptor amendment and that CH 4 was produced in the viable controls, it is tempting to speculate about the involvement of these classically considered aerobic methanotrophs in the observed AOM activity. Recent findings support this notion as both genera have been described to be capable of anaerobically oxidizing CH 4 using NO 3 − or NO 2 − as terminal electron acceptors (Kits et al. , Liu et al. ). Additionally, there are reports of aerobic CH 4 oxidizers, including Methylomonas spp., growing during and participating in AOM by switching to other electron acceptors such as Fe 3+ and SO 4 2− , either alone or in association with metal-reducing bacteria (Oswald et al. , , Cabrol et al. , Zheng et al. ). We did note an 8- and 17-fold increase in the mean relative abundance of the SO 4 2− reducer Desulforhopalu s and the dissimilatory Fe 3+ reducer Telmatospirillum in fen incubations amended with SO 4 2− and Fe 3+ , respectively . However, linking their potential activity with CH 4 oxidation by (aerobic) methanotrophs via extracellular electron transfer goes beyond the scope of this work. Finally, recent works have shown that methanobactin, a small copper-binding peptide, which is produced by Methylocystis among other methanotrophs, can function to generate O 2 via metal ion reduction and thereby enable AOM (Dershwitz et al. ).   Our data show that both pristine peatland types hold the potential for aerobic and anaerobic CH 4 oxidation supported by diverse methanotrophs, mainly of the orders Rhizobiales, Methylococcales , and possibly Methylomirabilales . This highlights the ecosystem's capacity for CH 4 emission mitigation. Aerobic CH 4 oxidation was not N-limited and commenced readily upon the addition of CH 4 , regardless of the addition of NH 4 + . Although the potential for AOM was present in peat from both peatland types, in swamp peat it was only realized upon adding either Fe 3+ or SO 4 2− . Contrastingly, in the fen peat, no stimulation of AOM by addition of electron acceptors was observed, and even an inhibition of AOM by Fe 3+ addition was found. This indicates an ongoing contribution of AOM to CH 4 oxidation in fen peat, while in the swamp peat CH 4 emissions may only be mitigated via aerobic oxidation.   fiae153_Supplemental_Files   