Salinity and vegetation shape methylotrophic methanogen diversity and potential activity in coastal sediments

Abstract

Anoxic coastal sediments are significant natural sources of the potent greenhouse gas methane where, methylotrophic methanogenesis may predominate. However, the underlying methanogen community composition as well as the impact of salinity levels and vegetation on their diversity and potential activity remain poorly resolved. To assess methane production and characterize methanogen diversity underlying methanol and trimethylamine (TMA) degradation, we incubated anoxic sediments from three sites along the Medway Estuary (UK)—a high-salinity marine site and two brackish sites, one vegetated (saltmarsh with cordgrass) and one non-vegetated. Saltmarsh sediments exhibited the highest methane yields (46% from TMA, 27% from methanol), whereas we did not observe methanogenesis from methanol in the high-salinity marine samples. Methylotrophic methanogen genera Methanolobus, Methanomethylophilus, and Methanofastidiosum were abundant in all original sediments, where metabolic generalist Methanosarcinales were also present. Methanoculleus were the most abundant taxon (31.5%) in original brackish sediments, while they were only at 1% in saltmarsh sediments. Following the amendment with TMA or methanol, Methanolobus dominated all the sediments with methane production (80%–99.5% relative abundance), however Methanoculleus were still in high relative abundance in brackish incubations with methanol or TMA (4.6% and 10%, respectively). These findings indicate that both salinity and vegetation shape the community composition of methylotrophic methanogens in coastal sediments, yet key methylated substrates are primarily metabolized by Methanolobus, regardless of differences in salinity and vegetation.

Vegetation and salinity affect the methylotrophic methanogen diversity and the methane yield from trimethylamine and methanol in coastal sediments.

Introduction

Coastal ecosystems such as estuaries, saltmarshes, and mangroves are highly productive environments that play a critical role in global carbon cycling (Perillo 2019, Myllykangas et al. 2020). They are complex and dynamic ecosystems, where the mixing between river water and seawater generates moderate to high salinity and continuous fluctuations in physicochemical conditions mainly driven by tidal cycles, sediment transport, nutrient inputs, and changing physical gradients (Allen 2000, Bouma et al. 2009, Cardoso 2021). These environments are recognized as “blue carbon” areas that sequester and store large amounts of carbon, thus contributing to climate change mitigation (Rosentreter et al. 2021). Coastal vegetated ecosystems are estimated to contribute an average 138–226 g C/m² per year to the global carbon sequestration budget (Kuwae et al. 2016). The sequestration capacity of coastal ecosystems increases in the presence of vegetation, which enriches the sediment with organic matter through its biomass and root exudates (Comer-Warner et al. 2022). For instance, although highly variable, global organic carbon accumulation in saltmarsh sediments is ∼12 Tg C per year, while tidal flats are estimated to sequester about 6.8 Tg C per year (Al‐Haj and Fulweiler 2020, Wang et al. 2023).

Although blue carbon ecosystems function as important carbon sinks, recent studies have highlighted that methane (CH₄) emissions from wetland sediments can partially offset their net climate-mitigation benefit (Bridgham et al. 2013, Rosentreter et al. 2021). Methane is an important greenhouse gas and ∼84 times as potent as carbon dioxide (CO₂) at trapping heat in the atmosphere over 20 years, significantly affecting the global temperature and climate system (Myhre et al. 2013). A recent assessment of saltmarshes suggests methane emissions of ∼224 µmol CH₄ per m² per day, corresponding to a median global flux of 0.18 Tg CH₄ per year (Rosentreter et al. 2021). Estuaries have also been identified as significant sources of methane, contributing to a median global flux of 0.17 Tg CH₄ per year (Borges and Abril 2011, Rosentreter et al. 2021). However, these ecosystems are largely understudied and current flux estimates involve substantial uncertainty.

Microbial methane production (methanogenesis) is carried out by methanogenic archaea via three primary pathways: Acetoclastic (acetate dismutation), hydrogenotrophic (CO₂ reduction with hydrogen) and methylotrophic (reduction or dismutation of methylated compounds such as trimethylamine, methanol and dimethylsulfide). Methylotrophic methanogenesis via dismutation produces both methane and CO₂ as end products. Stoichiometrically, three quarters of the methyl groups are reduced to methane, while one quarter is oxidized to CO₂ (Eqs 1 and 2; Yin et al. 2019).

(1)

(2)

Methanosarcina is a metabolically versatile genus, capable of methane production through all three pathways, whereas other major genera are largely pathway-specific (Thauer et al. 2008). For instance, Methanothrix is specialized for acetoclastic methanogenesis, Methanobacterium and Methanobrevibacter for hydrogenotrophic methanogenesis, and Methanomethylovorans and Methanolobus for methylotrophic methanogenesis (Liu and Whitman 2008). The mcrA gene, encoding the α subunit of methyl-coenzyme M reductase, catalyzes the final step of methanogenesis in all known methanogens and is key functional marker for methanogenesis across all pathways.

In coastal ecosystems, most methane production occurs through methylotrophic methanogenesis (Zhuang et al. 2018, Rosentreter et al. 2021). This is because methylotrophic methanogens can remain active alongside other microbial groups, such as sulfate-reducing bacteria, which generally outcompete methanogens for common substrates such as acetate, while lack the ability to metabolize methylated compounds or are weak competitors for them (Zhuang et al. 2016). The most abundant methylated substrates in coastal ecosystems are methanol and trimethylamine (TMA), which originate from the degradation of complex organic matter such as galactin, glycine betaine, pectin, choline, xylan and carbohydrates (Sun et al. 2019, Conrad 2020). TMA can also be excreted from animal tissues or through decomposition of marine organisms and is strongly adsorbed to saline sediments, particularly to those with high organic content (Wang and Lee 1990, Sun et al. 2019). Concentrations of TMA in coastal sediment porewaters can reach up to 29 µM (Lee and Olson 1984). In addition to the fermentation of galactin and pectin, methanol can be introduced to coastal ecosystems from terrestrial sources through air-sea exchange, diffusion, and rainfall (Beale et al. 2013, Dixon et al. 2013). Reported methanol concentrations in coastal sediment porewaters are as high as 78 µM (Zhuang et al. 2014).

To date, a few studies have identified methylotrophic methanogens utilizing methanol and TMA in estuarine and saltmarsh sediments, including members of the genera Methanococcoides (Jameson et al. 2019), Methanolobus (Ticak et al. 2015), and Methanosarcina (McGenity and Sorokin 2019). However, the influence of salinity and vegetation on methylotrophic methanogenesis efficiency and the associated methanogenic diversity remains unclear largely due to inconsistent findings among previous studies (Marton et al. 2012, Ardón et al. 2018, Tsola et al. 2021). To address this knowledge gap, we examined the diversity and methane production potential of methylotrophic methanogens utilizing TMA and methanol in anoxic estuarine and saltmarsh sediments. Samples spanned a high-salinity marine zone and two adjacent brackish zones that differed in vegetation; saltmarsh sediment with high cordgrass density and sediment without (Rooks et al. 2012, Lush et al. 2016). This sampling design allowed us to assess how salinity and vegetation shape the diversity and functional potential of methylotrophic methanogens in anoxic coastal sediments.

Materials and methods

Sediment sampling

Approximately 50 gr of triplicate sediment samples were collected along the Medway Estuary and Saltmarshes (southeast England) in March 2023, during low tide using Perspex corers (3.5 cm diameter). Anoxic sediments were obtained at depths between 4 cm and 10 cm from three distinct sites: a marine site (∼32‰ salinity), a (non-vegetated) brackish site (∼6‰ salinity), and a (vegetated) saltmarsh site adjacent to the brackish site (∼6‰ salinity) with a high abundance of cordgrass Spartina anglica (Fig. 1). In situ sediment temperatures at the time of sampling were ∼6°C. Sediment samples were vacuum sealed in gas-tight bags and transported to the laboratory in a cool box maintained at around 6°C. The sediment samples were used to set up incubations, while subsamples were frozen at −20°C for molecular analyses.

Figure 1.

Sampling locations along the Medway Estuary, UK. Isle of Grain—marine site; Medway Bridge Marina—brackish site and Baty’s marsh—saltmarsh site.

Incubation set-up

Sets of five replicated sediment incubations were established per location and treatment in serum bottles (50 ml). Each serum bottle contained 5 g of wet sediment and 20 ml of artificial seawater (ASW), which included 0.32 M NaCl, 10 mM MgSO₄·7H₂O, 8.8 mM NaNO₃, 3.1 mM CaCl₂·2H₂O, 10 mM MgCl₂·6H₂O, 9 mM Trizma base, and 0.21 mM K₂HPO₄·3H₂O. To simulate natural salinity levels in the incubations, we used neat ASW at 100% for the marine sediment or diluted it 50% with sterilized distilled water for the brackish and saltmarsh sediments. Then, 1 ml of 20 mM NH₄Cl, 0.1 ml of trace element and 0.1 ml of vitamin solution were added and the pH was adjusted to 8 using concentrated HCl (Wyman et al. 1985, Wilson et al. 1996). Final salinities of marine and brackish ASW were ∼32 and ∼16‰, respectively, representing typical marine and brackish conditions although brackish sediments can vary in salinity.

One set of incubations for each site was amended with either TMA or methanol as the carbon and energy source (5 mM final concentration). This concentration was selected to stimulate methanogenic activity under non-limiting substrate conditions. Control incubations, without TMA or methanol additions, were also established to monitor intrinsic methane production. Additionally, abiotic controls with twice-autoclaved sediment samples and without a carbon source, were incubated to assess substrate adsorption by the sediment. To ensure anoxic conditions, the incubation bottles were sealed with gas-tight butyl rubber stoppers and flushed with oxygen-free nitrogen for 10 min. Incubation bottles were kept at ∼21°C in the dark to assess potential methanogenic activity of the sediment without temperature limitation. When the initial TMA or methanol was depleted, additional substrate (5 mM final concentration) was injected through septa. All incubations were terminated once methane production had reached a stationary phase. Sediment from each bottle was then centrifuged at 1000 r/m for 6 min, and the liquid and sediment phases were separated and stored at −20°C until further analysis.

Analytical measurements

Methane and CO₂ production were measured using a gas chromatograph (Agilent Technologies, USA, 6890 N Series) fitted with a Flame Ionization Detector (at 300°C), Porapak q packed stainless steel column (Q 80/100; Supelco, USA) and hot-nickel catalyst to reduce CO₂ to methane (Agilent Technologies, USA; Sanders et al. 2007). Zero-grade nitrogen (British Oxygen Company, BOC, UK) was used as the carrier gas at 30°C with a 14 ml min⁻¹ flow rate. The concentrations of both methane and CO₂ were calculated using certified gas mixture standards (100 ppm methane, 3700 ppm CO₂; BOC, UK) and all samples were corrected for the headspace-water partitioning using Henry’s law constants (Wiesenburg and Guinasso 1979). All the gas concentrations were calculated per gram of wet sediment.

The total amount of CO₂ at the end of the incubation period was calculated by summing the headspace CO₂ and the total dissolved inorganic carbon (ΣDIC) in the water phase in the incubation bottles. For ΣDIC measurements, a 1 ml headspace was created in 3 ml sample vials containing the water phase. This was achieved using a two-way system, where 1 ml of liquid was removed via a syringe while oxygen-free nitrogen was simultaneously introduced into the headspace. Subsequently, 100 µl of 35% HCl was injected through the septa, and the vials were vigorously shaken to ensure complete acidification. Following equilibration, the concentration of CO₂ in the headspace was quantified against an inorganic carbon calibration curve (0.1–8 mM) prepared from sodium carbonate (Trimmer et al. 2009, Shelley et al. 2017).

DNA extraction and PCR

DNA was extracted from 250 mg of sediment sample using the DNeasy PowerSoil kit (Qiagen, NL) according to the manufacturer’s instructions. The mcrA gene, targeting archaeal methyl coenzyme M reductase, was amplified using primers mcrlRD-F and mcrlRD-R (Lever and Teske 2015). The PCR solution consisted of 22 µl of water, primers (1 µl each), 1 µl of DNA sample and 25 µl of 2xMyTaq HS Red Mix (Meridian Bioscience, USA). PCR protocol involved an initial denaturation at 95°C for 5 min, followed by 39 cycles of 95°C for 1 min, 51°C for 1 min, 72°C for 1 min, and a final extension at 72°C for 5 min, with a hold at 10°C. DNA extractions and PCR reactions were assessed using a 1% agarose gel electrophoresis.

Sequencing and data analysis

PCR products were sequenced on an Illumina MiSeq platform (300 bp paired end; Illumina, CA, USA). Data analysis was performed using QIIME 2 (Bolyen et al. 2019) on Queen Mary University of London’s (QMUL) Apocrita High-Performance Computing (HPC) facility (King et al. 2017). Amplicon sequence variants (ASVs) were processed with DADA2 (Callahan et al. 2016). Taxonomic classification of mcrA gene sequences was carried out using a curated database (Tsola et al. 2021). Average read abundances for the ASVs were determined based on five replicates. The sequences are deposited and publicly available at the National Center for Biotechnology Information (NCBI) Read Archive (PRJNA1363573).

Statistical analysis

All statistical analysis was performed using microeco in (R Core Team 2020, Liu et al. 2021). The effects of sediment type (marine, brackish non-vegetated and brackish saltmarsh) and treatment (methanol or TMA amendment) on methane and CO₂ production were assessed using a two-way ANOVA. Prior to analysis, assumptions of normality of residuals and homogeneity of variances were verified using the Shapiro-Wilk test and Levene’s test, respectively. Following a significant ANOVA result, a Tukey’s Honestly Significant Difference (HSD) post-hoc test was conducted to identify specific differences between group means. A significance level of P < 0.05 was applied for all statistical tests.

To assess the statistical significance of differences in diversity, a one-way ANOVA and pairwise PERMANOVA (9999 permutations) were performed. Principal coordinate analysis (PCoA) was applied to the microbial relative abundance data and visualized to simplify the data and identify underlying trends explaining observed variation. Beta-diversity was calculated using the Bray–Curtis index and visualized as principal coordinate analysis (PCoA) ordination plots.

Differential abundance analysis was performed using the LEfSe (Linear discriminant analysis Effect Size) method to identify taxa with significantly different abundances between treatments and quantify the effect size of each discriminating taxon (Segata et al. 2011).

Results

Methane and CO₂ productions in TMA-amended incubations

Methane production started after ten days in the saltmarsh and 11 days in non-vegetated brackish and marine sediments amended with TMA. Saltmarsh sediments exhibited the highest methane production, yielding an average of 316.64 ± 30 μmol methane g⁻¹ wet sediment (Fig. 2A). This corresponded to ∼46% of the theoretical yield based on one mol of TMA produces 2.25 mol of methane (Fig. 2B and C). In brackish incubations, methane production peaked at 202.6 ± 35 μmol g⁻¹ (Fig. 2A) with an approximate yield of 30% (Fig. 2B) after 22 days.

Figure 2.

A) Average cumulative concentrations (μmol g⁻¹ wet sediment) of methane in control, TMA and methanol amended sediment incubations per day throughout the incubation period. B) Average methane yields in the incubations. C) Average total methane and D) Average total carbon dioxide (CO₂) in the incubation bottles at the end of the incubation period. Error bars represent standard error above and below the average, highlighting variability between incubations (n = 5 biological replicates).

The marine sediments with TMA amendment displayed the lowest methane production, reaching 91.85 ±22 μmol methane g⁻¹ wet sediment with ∼20% yield (Fig. 2A and B). No significant methane production or carbon adsorption was detected in the control incubations across any sites.

We measured CO₂ concentrations at the end of the incubation as CO₂ is one of the end products of methylotrophic methanogenesis. The average total CO₂ concentration in saltmarsh sediment incubations was 174.25 ± 9 μmol g⁻¹ wet sediment, whilst marine sediment incubations exhibited the least CO₂ over the incubation period at 87.35 ± 15 μmol g⁻¹ wet sediment (Fig. 2D). However, these were not significantly higher than the unamended controls (132.53 ± 18 μmol g⁻¹ and 36.74 ± 13 μmol g⁻¹ of CO₂ in saltmarsh and marine controls, respectively). In contrast, the brackish sediment incubations produced a total of 143.42 ± 18 μmol CO₂ g⁻¹ wet sediment, which was significantly higher than the controls (7.32 ± 0.5 μmol g⁻¹; Fig. 2D).

Methane and CO₂ productions in methanol-amended sediment incubations

Methane production was observed after an acclimation period of eight days in the saltmarsh incubations and after 15 days in the brackish incubations, whereas no detectable production was observed in the methanol-amended marine incubations (Fig. 2A). Saltmarsh and brackish sediments produced 61.24 ± 10 and 37.96 ± 11 μmol methane g⁻¹ wet sediment, corresponding to ∼27% and 17% of the theoretical yield, respectively, assuming one mole of methanol produces 0.75 mol methane (Fig. 2B and C).

The brackish sediments produced a significantly higher amount of CO₂ (31.41 ± 6 μmol g⁻¹ wet sediment) compared to the controls (7.32 ± 0.5 μmol g⁻¹ wet sediment). In contrast, CO₂ concentration in saltmarsh and marine sediments (144.23 ± 18μmol g⁻¹ and 48.71 ± 16 μmol g⁻¹) did not differ significantly from their controls (132.52 ± 18μmol g⁻¹ and 36.74. ± 13 μmol g⁻¹, respectively, Fig. 2D).

Methanogen diversity in original sediments and incubations

Our mcrA gene sequence data indicate that Methanolobus were the most dominant taxon in the original marine and saltmarsh sediments, comprising 20% and 21.8% of the methanogen community, respectively. In contrast, the brackish sediment was dominated by Methanoculleus (31.5%), although Methanolobus were still present at 17% relative abundance (Fig. 3A). Candidatus Methanomethylophilus were also consistently abundant across all sediments (14%–19%). Notably, both Methanolobus and Methanomethylophilus are strictly methylotrophic methanogens, indicating that methylotrophic lineages comprise the majority of the methanogen community in all three sediment types.

Figure 3.

A) Average relative abundance of the mcrA gene sequences at the genus level in the original, methanol, and TMA incubations from brackish, marine, and saltmarsh sediments with five replicates per site (n = 5). B) Beta-diversity analysis from mrcA gene sequences of methanogens according to treatments and sampling locations. Beta-diversity was calculated using the Bray–Curtis index and visualized as principal coordinate analysis (PCoA) ordination plots.

The amendment of methylotrophic substrates significantly altered the methanogen community composition (P < 0.01), resulting in clear separation from both the original and control groups in all sediments, except for the methanol-amended marine sediment, which did not produce detectable methane. Methanolobus became highly enriched in the saltmarsh and brackish sediment incubations amended with either TMA or methanol, reaching relative abundances of 80%–99.5%. The saltmarsh control incubations without additional carbon also showed elevated Methanolobus abundance (85%). In the marine incubations, Methanolobus increased to 99% following TMA addition but only reached 35.5% with methanol, consistent with the lack of detectable methane production in these incubations (P < 0.01; Fig. 3A). PCoA of the mcrA sequences further supported clustering by both sampling site and substrate amendment (Fig. 3B).

Effects of salinity and vegetation on methanogen diversity in original sediments and incubations

To assess the influence of salinity and vegetation on methanogen community composition in the sediments of Medway Estuary and Saltmarshes, we conducted differential abundance analysis of mcrA gene sequences from the original sediments and the incubations. Vegetation-associated differences were identified by comparing the brackish and saltmarsh sediments, while the influence of salinity was evaluated by comparing the marine and brackish sediments.

The analysis demonstrated clear and significant effects of both salinity and vegetation on methanogen community structure in original sediments. In the brackish sediment, Methanoculleus were dominant at 31.5% relative abundance, but it was only at 1% in the saltmarsh sediments (P < 0.01; Figs 3A and 4A). Candidatus Methanomethylophilus were also significantly more abundant in the brackish sediment (19%) than in the saltmarsh sediment (14.6%; P < 0.01). Methanoculleus are strictly hydrogenotrophic methanogens commonly associated with brackish environments (Dabir et al. 2014), whereas Methanomethylophilus are hydrogen-dependent methylotrophic methanogens. In contrast, the saltmarsh sediment contained Methanosarcina significantly more abundant than the brackish sediment (10.7% and 3.9%, respectively; P < 0.01). Methanosarcina species have broad metabolic flexibility, capable of using multiple methanogenic pathways. Overall, these patterns indicate that vegetation can shape methanogen community composition, resulting in different dominant methanogen taxa and pathways between the sediment types and favouring methylotrophic methanogens in the saltmarsh environment (Fig. 4A).

Figure 4.

Differential abundance analysis of methanogen taxa from mcrA gene sequences in A) brackish and saltmarsh original sediments, B) brackish and saltmarsh methanol incubations, C) brackish and saltmarsh TMA incubations, D) marine and brackish original sediments, E) marine and brackish methanol incubations, and F) marine and brackish TMA incubations to test the effect of vegetation and salinity, respectively. Bar plots show the relative abundance of the taxa significantly different between different samples (P < 0.01 or P < 0.001) identified by Linear Discriminant Analysis Effect Size (LEfSe) analysis. * denotes P < 0.01, ** denotes P < 0.001. Error bars represent the standard error (n = 5 biological replicates).

The effect of salinity was also apparent when comparing the original marine and brackish sediments. Methanoculleus was significantly less abundant in the marine sediment (6%), which instead had significantly more Methanofastidiosum and Methanosarcinales (3.2% and 2.6%, respectively; P < 0.01). The brackish sediment was also more enriched in Methanomethylovorans (3%) and Candidatus Methanomethylophilus (19%; Figs 3 and 4D). Methanofastidiosum and Methanomethylovorans are obligate methylotrophic methanogens. These results indicate that marine and brackish sediments consisted of both metabolically specialist and generalist methanogens, and suggest a shift in methylotrophic methanogen diversity under saline conditions.

Amendment with methanol or TMA resulted in distinct shifts within the sediment methanogen community composition, however, Methanolobus remained the dominant taxon across all incubations (80%–99.5%). In the brackish sediment incubations with methanol or TMA, the relative abundances of Methanoculleus (12.5% and 7%, respectively) and Candidatus Methanomethylophilus (6% and 3%, respectively) declined markedly (Figs 3A and 4B and C). Furthermore, Candidatus Methanomethylophilus and Methanosarcina became undetectable in the saltmarsh incubations amended with methanol or TMA, suggesting that methanogen communities in vegetated (saltmarsh) and non-vegetated (brackish) sediments may respond differently to high concentrations of methylotrophic substrates (Fig. 4B and C). In the marine sediments, Candidatus Methanomethylophilus, Methanosarcina, and Methanoculleus were not detected after TMA amendment, while methanol amendment did not result in methane production or substantial changes in methanogen community composition. These patterns indicate that methanogen community composition significantly changes in TMA- or methanol-amended sediments with different salinities, while Methanolobus is highly adaptable to salinity variations (Fig. 4E and F).

Discussion

Contribution of TMA and methanol to methanogenesis in marine, brackish, and saltmarsh sediments

Significant methane production was observed in the TMA- and methanol-amended incubations, except for the marine sediments amended with methanol, which did not produce detectable methane. In TMA-amended brackish and saltmarsh sediments, methane production began around day 13, whereas methanol-amended saltmarsh sediments showed an earlier onset (day 9) and the brackish sediments a later onset (day 15). This delayed initiation of methanogenesis may be attributed to the relatively high substrate concentrations used in the incubations compared to those naturally found in sediments. Therefore, methanogenic populations with higher affinities for TMA and methanol likely required time to acclimate to the incubation conditions. Our findings align with previous studies on methylotrophic methanogenesis in coastal sediments, where methanogenesis began after extended incubation periods (>12 days; Maltby et al. 2018, Krause and Treude 2021, Tsola et al. 2021). By contrast, more rapid methane production within three days has also been reported in a saltmarsh sediment (Jameson et al. 2019). Together, these results suggest that the potential for TMA- or methanol-dependent methanogenesis is not uniform across anoxic coastal sediments, reflecting combined influence of microbial community structure and sediment geochemistry.

Despite the close spatial proximity of the saltmarsh and non-vegetated brackish sites, the methane yields from TMA differed substantially, with the saltmarsh producing 46% compared to 30% in the non-vegetated sediments. This enhanced TMA-dependent methanogenesis in the saltmarsh sediment is likely linked to the presence of Spartina anglica, as the continuous supply of plant-derived TMA may sustain methylotrophic methanogenesis by allowing the natural methanogenic populations to adapt to TMA under saline conditions (Kim et al. 2020). In contrast, TMA-amended marine sediments exhibited a lower yield of 20%, suggesting that high salinity reduces methylotrophic activity. Stoichiometrically low methane yields from TMA and methanol, along with the absence of methane production in methanol-amended marine sediments, suggest that these compounds were likely metabolized through alternative pathways. In anoxic environments, TMA can undergo acetogenesis or dehydrogenation, producing acetate and dimethylamine, respectively (Sun et al. 2019). If TMA was degraded through acetogenic pathway, the resulting acetate would likely have been metabolized by sulfate-reducing bacteria (SRB). However, these pathways have not been shown to play a major role in marine ecosystems. Methanol can be metabolized through acetogenesis, denitrification, iron reduction or sulfate reduction (Fischer et al. 2021). Under saline conditions, where sulfate concentrations are typically high, SRB may have higher affinities for methanol than methanogens, outcompeting methanogens for this substrate (Sousa et al. 2018, Fischer et al. 2021). Collectively, these findings indicate that substrate type and availability, together with salinity, determine whether TMA and methanol are metabolized by methanogens or diverted to alternative metabolic pathways.

Complete conversion of methanol and TMA during methylotrophic methanogenesis would theoretically yield 0.25 mol CO₂ per mol methanol and 0.75 mol CO₂ per mol TMA. However, final CO₂ concentrations were lower than expected from stoichiometric calculations across all incubations. Further, CO₂ concentrations remained significantly higher than the controls only in the brackish sediment incubations amended with methanol or TMA. This discrepancy indicates that CO₂ produced during methylotrophic methanogenesis was simultaneously consumed. This observation aligns with our previous study, which reported high abundances of Sulfurimonas species in the sediment samples from the same coastal ecosystem (Tsola et al. 2021). It is therefore plausible that Sulfurimonas contributed to CO₂ fixation, reducing the final CO₂ concentrations. Alternatively, in the brackish sediment incubations, part of the CO₂ may have been utilized through hydrogenotrophic methanogenesis, given the presence of hydrogenotrophic methanogens in these sediments. Additionally, methane concentrations in the TMA-amended brackish sediment incubations declined after 22 days, suggesting anaerobic methane oxidation, which may have generated CO₂ as an end product, increasing the final CO₂ concentrations in those incubations. However, our mcrA sequencing data do not provide evidence for microbial communities associated with anaerobic methane oxidation.

Methanogen diversity in marine, non-vegetated brackish, and saltmarsh sediments and their response to TMA and methanol amendment

The original sediments hosted distinct methanogenic communities, all of which comprised the methylotrophic genera Methanolobus and Methanomethylophilus. In addition, the hydrogenotrophic methanogenic genus Methanoculleus predominated in the brackish sediments, suggesting that both hydrogenotrophic and methylotrophic methanogenesis contribute to methane production in brackish sediments. In contrast, Methanoculleus were detected in lower relative abundances in the marine and saltmarsh sediments, where methane production appeared to be driven primarily by Methanolobus through methylotrophic methanogenesis. Methanolobus were previously shown to degrade a wide range of methylated substrates including methanol, TMA and dimethylsulfide, in several coastal ecosystems, showing their significance in carbon cycling and methane production in anoxic coastal sediments (Zhang et al. 2008, Shen et al. 2020, Tsola et al. 2024). Interestingly, our previous study on the same site identified the genera Methanococcoides and Methanomicrobiales as dominant in the marine and brackish sediments (Tsola et al. 2021). While the underlying causes of this shift cannot be directly assessed here, differences in environmental conditions and sediment geochemistry between sampling campaigns may have contributed to the observed variation in methanogen community composition Our findings add to the growing evidence that methylotrophic methanogenesis pathway represents an important contributor to methane production in coastal sediments, emphasizing the need to reassess the complex dynamics governing methane cycling in these ecosystems (Xiao et al. 2018, Seyfferth et al. 2020, Capooci et al. 2024).

Despite the effect of salinity and vegetation on methanogenic populations in the sediments we studied, we did not observe a significant effect on methanogenic community composition when the sediments were incubated with methanol or TMA. Methanolobus emerged as the dominant methylotrophic methanogen across all incubations regardless of the salinity, vegetation or the substrate used. The relatively high TMA and methanol concentrations (5 mM) likely promoted methylotrophic methanogenesis under substrate-replete conditions and may have favoured specific functional groups with relatively high growth rates (up to 0.09 h⁻¹) or lower affinities for TMA and methanol (Liu et al. 1990, Antony et al. 2012). Although detectable methane production occurred in the saltmarsh sediment control incubations, the relative abundance of Methanolobus increased. This may be due to plant tissues directly releasing TMA, TMA precursors such as glycine betaine and choline or dimethylsulfoniopropionate (a common marine osmolyte), which is degraded to dimethylsulfide that can be efficiently metabolized by Methanolobus (Sun et al 2019, Tsola et al. 2021).

In the brackish incubations amended with methanol or TMA, a decline in the relative abundance of Methanoculleus was observed compared to the original samples and controls. As Methanoculleus are hydrogenotrophic methanogens, this reduction is expected following methylated substrate amendment. Previous studies have shown that some Methanoculleus species can utilize alcohols under specific environmental constraints, which may explain their persistence in methanol-amended brackish incubations (Dhillon et al. 2005, Shimizu et al. 2013, Chen et al. 2015). Similarly, the relative abundance of Candidatus Methanomethylophilus decreased significantly in all TMA-amended incubations. This genus, belonging to the order Methanomassiliicoccales, obtains energy by reducing methylated compounds to methane using hydrogen as an electron donor (Borrel et al. 2023). Hydrogen-dependent methanogens were suggested to have a competitive advantage in environments with high hydrogen availability but limited methylated substrates (Feldewert et al. 2020). In our incubations, hydrogen may have been limited due to its consumption by other metabolic pathways such as hydrogenotrophic methanogenesis in brackish sediments, reducing the competitiveness of Methanomethylophilus. Alternatively, Methanomethylophilus species may be outcompeted due their low growth rate (0.026 h⁻¹) compared to Methanolobus (0.09 h⁻¹; Liu et al. 1990, Borrel et al. 2023).

Conclusion

This study helps address the knowledge gap regarding methylotrophic methanogenesis in anoxic coastal environments by demonstrating that both TMA and methanol can drive methane production in coastal sediments. Distinct methanogenic taxa were associated with marine, brackish and saltmarsh sediments, however, Methanolobus consistently dominated the sediment incubations with TMA or methanol, underscoring its adaptability across contrasting environmental conditions. Salinity and vegetation influence both the composition and activity of methanogenic communities, as evidenced by higher methane yields in saltmarsh sediments, likely due to cordgrass-derived substrates, and the absence of methanol-dependent methanogenesis in marine sediments, possibly due to substrate competition from non-methanogenic microbial groups. These findings highlight the complex interplay among salinity, vegetation, and substrate availability in shaping methylotrophic methanogenesis and raise questions surrounding the environmental constraints over methylotrophic methanogen diversity and metabolism. A better understanding of these processes is essential for accurately predicting methane dynamics in coastal ecosystems and their broader implications for global carbon cycling and climate regulation.

Conflicts of interest

None declared.

Funding

This study is financially supported by the Natural Environment Research Council (NE/X011461/1) and a PhD scholarship to A.H-H. by Fundación Politécnico A. C. and Queen Mary University of London.