Acidophilic methanotrophs: Occurrence, diversity, and possible bioremediation applications

Abstract

Methanotrophs have been identified and isolated from acidic environments such as wetlands, acidic soils, peat bogs, and groundwater aquifers. Due to their methane (CH₄) utilization as a carbon and energy source, acidophilic methanotrophs are important in controlling the release of atmospheric CH₄, an important greenhouse gas, from acidic wetlands and other environments. Methanotrophs have also played an important role in the biodegradation and bioremediation of a variety of pollutants including chlorinated volatile organic compounds (CVOCs) using CH₄ monooxygenases via a process known as cometabolism. Under neutral pH conditions, anaerobic bioremediation via carbon source addition is a commonly used and highly effective approach to treat CVOCs in groundwater. However, complete dechlorination of CVOCs is typically inhibited at low pH. Acidophilic methanotrophs have recently been observed to degrade a range of CVOCs at pH < 5.5, suggesting that cometabolic treatment may be an option for CVOCs and other contaminants in acidic aquifers. This paper provides an overview of the occurrence, diversity, and physiological activities of methanotrophs in acidic environments and highlights the potential application of these organisms for enhancing contaminant biodegradation and bioremediation.

INTRODUCTION

Methanotrophs, which exist in various environments, are bacteria that grow on methane (CH₄) as a sole source of carbon and energy (Cai et al., 2016; Guerrero‐Cruz et al., 2021; Holmes et al., 1999). Methanotrophs oxidize CH₄ to methanol (CH₃OH), formaldehyde (HCHO), formic acid (HCOOH), and ultimately carbon dioxide (CO₂). CH₄ monooxygenase (MMO), which is critical in C1 metabolism, is responsible for the first step of CH₄ oxidation to CH₃OH (Gesicka et al., 2021; Guerrero‐Cruz et al., 2021; Hanson, 1996). There are two general types of MMO, particulate CH₄ monooxygenase (pMMO), which is an intracytoplasmic membrane‐bound copper‐containing enzyme, and soluble CH₄ monooxygenase (sMMO), which is an iron‐containing cytoplasmic enzyme (Lawton & Rosenzweig, 2016; Lee et al., 2020). Most methanotrophs have pMMO only, some contain both pMMO and sMMO, and very few contain only sMMO.

Methanotrophs have been studied in the field of pollutant biodegradation because they can oxidize environmental contaminants including many different chlorinated volatile organic compounds (CVOCs) using MMO (Chu & Alvarez‐Cohen, 1996; Chu & Alvarez‐Cohen, 1998; Chu & Alvarez‐Cohen, 1999; Oldenhuis et al., 1991; Semrau, 2011). Among the CVOCs oxidized by MMO, trichloroethylene (TCE) is particularly important in that it is widely distributed in groundwater aquifers, and causes negative effects on the immune and central nervous systems as well as being a suspected carcinogen (EPA, 2016). Common anaerobic degradation products of TCE, such as vinyl chloride (VC; a potent carcinogen) and cis‐1,2‐dichloroethene (cis‐DCE), are also degraded by MMO, as are numerous other halogenated organic compounds (Samin & Janssen, 2012; Schäfer et al., 2003). While methanotrophs have been successfully applied for the bioremediation of CVOC‐contaminated aquifers (e.g. Hazen et al., 1994), most in situ remediation of CVOCs is performed via anaerobic bioremediation using carbon source addition, with or without bioaugmentation with dechlorinating consortia containing Dehalococcoides spp., one of two groups of organisms (along with a recently discovered Dehaligenomonas spp.) known to be capable of dehalogenating PCE and TCE all the way to ethene (Chen et al., 2022; Stroo & Ward, 2010; Yang et al., 2017). However, anaerobic bioremediation by Dehalococcoides spp. is largely ineffective at reducing chlorinated ethenes to ethene in naturally acidic aquifers, because it is typically inhibited at pH < ~5.5 (Eaddy, 2008; Lacroix et al., 2014; Rowlands, 2004; Steffan & Vainberg, 2013; Vainberg et al., 2009; Yang, 2012; Yang et al., 2017). Although much more research is required, acidophilic methanotrophs have recently been observed to degrade a range of CVOCs at pH <5.5, suggesting that cometabolic treatment may be an option for CVOCs and other contaminants in acidic aquifers (Choi et al., 2021; Semrau, 2011; Shao et al., 2019; Szwast, 2021).

For this review, an ‘acidophilic methanotroph’ is defined generally as capable of growth at pH 5 or below, with the understanding that these organisms are not necessarily ‘obligate acidophiles’ that require highly acidic pH to survive, consistent with the definition by Madigan (2018). Acidophilic methanotrophs have recently been identified in a wide variety of acidic environments, including peat bogs, wetlands and lakes, thermal soils and springs, and groundwater aquifers among others. Methanotrophs in acidic wetlands are critically important for the control of atmospheric CH₄, which is important in the global carbon cycle (Nguyen et al., 2018; Siljanen et al., 2012), because typically more than 90% of CH₄ produced in wetlands is oxidized by methanotrophs in the surface layers (Oremland & Culbertson, 1992; Siljanen et al., 2012). Moreover, the control of CH₄ fluxes by acidophilic methanotrophs becomes more and more important due to accelerated soil acidification via anthropogenic activities and climate change (Nguyen et al., 2018).

As with neutrophilic methanotrophs, acidophilic methanotrophs contain sMMO and/or pMMO (Belova et al., 2013; Dedysh et al., 2002; Dedysh et al., 2004). For example, Methylocystis bryophila, which has an optimal pH between 6.0 and 6.5 (but grows below pH 5.0), contains both pMMO and sMMO, while Methylocella tundrae only has sMMO, and effectively grows at pH 5.5–6.0 (Belova et al., 2013; Dedysh et al., 2004). A number of recently discovered species including Methylocystis bryophila, Methylocystis heyeri, Methylocapsa aurea, Methylocella palustris, Metylocella silvestris, and Mehtylocella tundrae are facultative methanotrophs (Nazaries et al., 2013). They can use multi‐carbon compounds such as acetate, organic acids, alcohols, ethane, and propane as carbon and energy sources besides CH₄ (Farhan Ul Haque et al., 2020). In addition, methanotrophs belonging to the phylum Verrucomicrobia have recently been isolated and identified from extremely acidic environments (e.g. below pH 2.0 and/or above a temperature of 50°C; Nazaries et al., 2013).

This review summarizes the occurrence, diversity, and physiology of acidophilic methanotrophs. The occurrence and types of acidophilic methanotrophs are described and their roles in acidic environments are also explored in this review. Recent progress and potential applications of methanotrophs for the biodegradation of chlorinated compounds and other pollutants in acidic groundwater and other low pH environments are also discussed. Scientists are just beginning to understand the implications of facultative growth in acidophilic methanotrophs, and very little research has been done to assess the ability of these organisms to cometabolically degrade CVOCs or other pollutants during growth on secondary substrates, as described later in this review.

ACIDOPHILIC METHANOTROPHS

Occurrence

Researchers first began to study acidophilic methanotrophs in natural environments more than 25 years ago. The first evidence of acidophilic methanotrophs was reported by a 16S rRNA‐based study on samples collected from a low pH peat environment (pH 3.6) in 1996 (McDonald et al., 1996). Initially, only a few methanotrophs capable of growth at low pH were isolated in pure culture. This was, in part, due to the use of growth media containing high mineral salt concentrations (1.5–3 g/L) (Dedysh et al., 2002). Peat bogs, where initial enrichment cultures were obtained have not only high acidity but also very low total dissolved solids (TDS) (Dedysh, Panikov, & Tiedje, 1998). Successful isolations of three new acidophilic strains later occurred when low salt medium (containing 50 mg/L mineral salts) was used, and incubation conditions were adjusted to better simulate the oligotrophic and acidic peat bog environment (Dedysh, Panikov, & Tiedje, 1998). Additional strains were subsequently isolated from other acidic environments such as Sphagnum peat bogs (Dedysh et al., 2000; Dedysh et al., 2002) and Sphagnum tundra peatlands (Dedysh et al., 2004).

Following the aforementioned changes in isolation procedure and growth medium, 10 different pure acidophilic methanotrophs were isolated from other environments including forest soil, forest Cambisol, and collapsed palsa soil (Danilova et al., 2013; Dedysh et al., 2000; Dedysh et al., 2002; Dedysh et al., 2004; Dedysh et al., 2007; Dedysh, Didriksen, et al., 2015; Dunfield et al., 2003; Dunfield et al., 2010; Vorobev et al., 2011). Table 1 summarizes the characteristics of acidophilic methanotrophic isolates that have been reported. These pure strains have also been characterized in terms of their morphology, growth conditions, fatty acid profiles, and enzyme activities. The whole genome sequences of seven of the isolates have been reported (Dedysh, Naumoff, et al., 2015; Esson et al., 2016; Han et al., 2018; Kox et al., 2019; Miroshnikov et al., 2017; Oshkin et al., 2019; Ricke et al., 2005), providing basic information for further study.

TABLE 1.

Characteristics of isolated acidophilic methanotrophs.

Genus and species	Source	Optimum pH	Optimum temperature range (°C)	MMO expressed	Phenotype	Multi‐carbon substrate	Phylogenetic affiliation	Formaldehyde assimilation	Reference
Methylocella palustris	Sphagnum peat bogs	5.0–5.5	15–20	sMMO	Obligate		α‐Proteobacteria	Serine	Dedysh et al. (2000)
Methylocella siverstris	Forest cambisol	5.5	15–25	sMMO	Facultative	Organic acids, Alcohols, Ethane, Propane	α‐Proteobacteria	Serine	Dunfield et al. (2003)
Methylocella tundrae	Sphagnum tundra peatlands	5.5–6.0	15	sMMO	Facultative	Organic acids, Alcohols	α‐Proteobacteria	Serine	Dedysh et al. (2004)
Methylocapsa acidiphila	Sphagnum peat bogs	5.0–5.5	20–24	pMMO	Obligate		α‐Proteobacteria	Serine	Dedysh et al. (2002)
Methylocapsa aurea	Forest soil	6.0–6.2	25–30	pMMO	Facultative	Acetate	α‐Proteobacteria	Serine	Dunfield et al. (2010)
Methylocapsa palsarum	Collapsed palsa soil	5.2–6.5	18–25	pMMO	Obligate		α‐Proteobacteria	Serine	Dedysh, Didriksen, et al. (2015), Dedysh, Naumoff, et al. (2015)
Methylocystis heyri	Sphagnum peat‐bog lake	5.8–6.2	25	pMMO and sMMO	Facultative	Acetate	α‐Proteobacteria	Serine	Dedysh et al. (2007)
Methylocystis bryophila	Sphagnum peat‐bog lake	6.0–6.5	25–30	pMMO and sMMO	Facultative	Acetate	α‐Proteobacteria	Serine	Belova et al. (2013)
Methyloferula stellate	Sphagnum peat bogs	4.8–5.2	20–23	sMMO	Obligate		α‐Proteobacteria	Serine and RuMP	Vorobev et al. (2011)
Methylomonas paludis	Sphagnum peat bogs	5.8–6.4	20–25	pMMO	Obligate		γ‐Proteobacteria	RuMP	Danilova et al. (2013)
Methylacidiphilum fumariolicum	Acidic thermal mudpot	0.8–6.0	40–65	pMMO	Facultative	Propane, Ethane	Verrucomicrobia		Schmitz et al. (2021)
Methylacidiphilum infernorum	Acidic thermal soil	1.0–6.0	40–60	pMMO	Obligate		Verrucomicrobia		Schmitz et al. (2021)
Methylacidiphilum kamchatkense	Acidic thermal spring	2.0–5.0	37–60	pMMO	Obligate		Verrucomicrobia		Islam et al. (2016)
Methylacidimicrobium cyclopophantes	Acidic soil	0.6–5.5	44–49	pMMO	Obligate		Verrucomicrobia		van Teeseling et al. (2014)
Methylacidimicrobium fagopyrum	Acidic soil	0.6–5.5	35–39	pMMO	Obligate		Verrucomicrobia		van Teeseling et al. (2014)
Methylacidimicrobium tartarophylax	Acidic soil	0.5–5.5	38–43	pMMO	Obligate		Verrucomicrobia		van Teeseling et al. (2014)

Abbreviations: pMMO, particulate CH₄ monooxygenase; sMMO, soluble CH₄ monooxygenase; RuMP, ribulose monophosphate.

Most acidophilic methanotrophs have been isolated from peat bogs, where CH₄ is produced through anaerobic decay, and peat moss acidifies its surroundings by taking up calcium and magnesium while releasing hydrogen ions (Danilova et al., 2013; Dedysh et al., 2002). All acidophilic isolates are gram‐negative rods or cocci without flagella or pili. The optimal growth temperature for these methanotrophs is below 30°C, typically around 25°C, making them mesophiles. For example, Methylomonas, acidophilic Type 1 organisms, which belong to the γ‐proteobacteria and use the ribulose monophosphate (RuMP) pathway for formaldehyde assimilation, were isolated from an acidic peat bog (Danilova et al., 2013) (Figure 1). Another common methanotroph, Methylobacter, which also belongs to γ‐proteobacteria, also have been isolated from acidic environments, such as forest soils below pH 5.0 (Nguyen et al., 2018). This suggests that some genera may adapt to acidic conditions. However, most acidophilic methanotrophs described to date belong to the α‐proteobacteria as Type II methanotrophs, which use the serine pathway for formaldehyde assimilation (Strong et al., 2015). One Type II example, Methylosinus was the first identified acid‐tolerant methanotroph from an acidic peat lake (Dedysh, Panikov, & Tiedje, 1998). Several other acidophilic methanotrophs belonging to α‐proteobacteria, including Methylocella, Methylocystis, Methylocapsa, and Methyloferula, were discovered in acidic wetlands and peat bogs (Belova et al., 2013; Dedysh et al., 2007). These genera are also known as cold‐tolerant methanotrophs (Dedysh, 2011).

FIGURE 1.

The relationship of acidophilic methanotrophs Types I and II, thermoacidophilic methanotrophs, and facultative methanotrophs. Revised from Figure 1 of Khider et al. (2021). CH₄, methane; CH₃OH, methanol; HCHO, formaldehyde; CO₂, carbon dioxide; RuMP, ribulose monophosphate.

As noted in the introduction, a number of acidophilic methanotrophs have recently been observed to be facultative. Compared to obligate methylotrophs, which only grow on CH₄ and a limited number of C1 compounds, facultative methanotrophs are known to use not only CH₄ but also multi‐carbon compounds (i.e. ethane, propane, acetate, ethanol, succinate, and/or organic acids) as sole carbon and energy sources (Dedysh & Dunfield, 2011; Farhan Ul Haque et al., 2020). Facultative methanotrophs primarily belong to α‐proteobacteria including Methylocystis, Methylocella, Methylocapsa, and Methyloceanibacter (Belova et al., 2013; Dedysh et al., 2005; Dunfield et al., 2010; Vekeman et al., 2016). Crenothrix polyspora, belonging to γ‐proteobacteria, has also been observed to grow on acetate and glucose (Stoecker et al., 2006). Accordingly, facultative methanotrophs might have a competitive advantage over obligate methanotrophs under some conditions due to their metabolic diversity.

Two different genera of methanotrophs (Methylacidiphilum and Methylacidimicrobium) belonging to the phylum Verrucomicrobia have recently been isolated from highly acidic environments (<pH 2.0) (Nazaries et al., 2013; Schmitz et al., 2021; van Teeseling et al., 2014). As shown in Table 1, these methanotrophs can survive under extreme conditions (i.e. at temperatures above 50°C and/or pH below 2.0), and thus are termed ‘thermoacidophilic’. Also, some Verrucombicrobia show an ability to grow on propane, ethane, and H₂ and thus join the expanding group of facultative methanotrophs (Schmitz et al., 2021). One example is Methylacidiphilum fumariolicum, which can grow on multi‐carbon compounds including propane and ethane in addition to CH₄ (Mohammadi et al., 2017; Picone et al., 2020; Pol et al., 2007). These organisms use the Calvin–Benson–Bassham (CBB) cycle to fix carbon dioxide for carbon assimilation unlike Type I and Type II methanotrophs (Op den Camp et al., 2009), and use H₂ as a growth substrate via hydrogen‐oxidizing enzymes (Mohammadi et al., 2019). The Verrucomicrobia phylum has recently been reviewed (Schmitz et al., 2021).

In recent years, the diversity and abundance of acidophilic methanotrophs have been further explored through applications of molecular techniques based on 16S rRNA and/or functional genes (Chen et al., 2008; Ghashghavi et al., 2017). Detection of various novel clones revealed that acidophilic methanotrophs are more widely distributed in the environment than previously thought (Esson et al., 2016; Farhan Ul Haque et al., 2020; Kip et al., 2011; Kip et al., 2012). More discussion of the molecular analysis of acidophilic methanotrophs is provided in ‘Molecular identification’ section and Table 2.

TABLE 2.

Molecular identification of acidophilic methanotrophs based on presence of PCR‐based assays targeting genes.

Acidophilic methanotrophs	PCR‐based assays targeting genes
Genus and species	pmoA	mmoX	mxaF	prmA
Methylocella palustris	/	+	+	/
Methylocella siverstris	/	+	+	+
Methylocella tundrae	/	+	+	+
Methylocapsa acidiphila	+	/	+	/
Methylocapsa aurea	+	/	+	/
Methylocapsa palsarum	+	/	+	+
Methylocystis heyri	+	+	+	/
Methylocystis bryophila	+ a	+	+	/
Methyloferula stellate	/	+	+	/
Methylomonas paludis	+	/	+	/
Methylacidiphilum fumariolicum	+ b	/	+	/
Methylacidiphilum infernorum	+ c	/	+	/
Methylacidiphilum kamchatkense	+ d	/	+	/
Methylacidimicrobium cyclopophantes	+	/	+	+
Methylacidimicrobium fagopyrum	+	/	+	/
Methylacidimicrobium tartarophylax	+	/	+	+

Note: The presence of genes was identified using TBLASTN with representative protein sequences (Gertz et al., 2006) or the Universal Protein Resource (UniProt) database (UniProt Consortium, 2021).

^a pmoA1 and pmoA2,

^b pmoA1 and pmoA3,

^c pmoA1, pmoA2, and pmoA3,

^d pmoA1, pmoA2, pmoA3, and pmoA4.

Molecular identification

Fluorescent oligonucleotide probes based on 16S rRNA (Bourne et al., 2000; Dedysh et al., 2003; Kalyuzhnaya et al., 2006), and PCR‐based assays targeting 16S rRNA, pmoA, mmoX, and methanol dehydrogenase gene, mxaF (Chen et al., 2008; Ghashghavi et al., 2017; Hutchens et al., 2004; Kip et al., 2011; Lau et al., 2007; Redmond et al., 2010) are commonly used to identify and investigate the diversity and abundance of acidophilic methanotrophs in different environments. With the PCR‐based assays, many uncultured clones have been detected, demonstrating the wide existence and diversity of acidophilic methanotrophs (Chen et al., 2008; Kip et al., 2011; Knief, 2015; McDonald et al., 1996). The PCR‐based assays targeting genes for acidophilic methanotrophs are shown in Table 2. Recently, the comparative genomic analysis provides insights into the study of acidophilic methanotrophs (Nguyen et al., 2018). Based on reconstructed genomes of uncultured bacteria, one can analyse genomic features including genome sizes, G + C contents, and number of CDSs, and then compare to the reference genomes to identify the uncultured acidophilic methanotrophs in a given environmental sample (Nguyen et al., 2018).

Characteristics of pure strains

Several pure strains of acidophilic methanotrophs have been isolated as summarized in Table 1. Adapting the molecular identification approach in ‘Molecular identification’, a comparative sequence analysis based on 16S rDNA, pmoA, and mmoX of the isolated acidophilic methanotrophs in Table 1 was performed (Figure 2). Each phylogenetic tree was generated with 16S rDNA, pmoA, and mmoX gene sequences of 10 strains in GenBank using the neighbour‐joining software MEGA. The evolutionary relationship among these 10 strains is provided in Figure 2A. The analysis shows that species belonging to class α‐proteobacteria have a relatively high similarity (73%–87%) in the pmoA gene (Figure 2B) as well as the mmoX gene (80%–86% similarity) (Figure 2C).

FIGURE 2.

Phylogenetic tree based on (A) 16S rDNA, (B) pmoA, and (C) mmoX.

As previously described, three acidophilic methanotrophs were successfully isolated from peat bog environments in 1998 using a medium with low ionic strength (Dedysh, Panikov, Liesack, et al., 1998). These strains all belonged to the genus Methylocella, with Methylocella palustris being the first species identified in this genus (Dedysh et al., 2000). Methylocella palustris was found to only express sMMO because no products were observed during PCR with pMMO‐targeted primers; hybridization with a pmoA probe also was negative. This observation ran counter to the previous belief that all methanotrophs contained the pmoA gene and expressed pMMO (Dedysh et al., 2000). Analysis of their 16S rDNA sequences revealed that these strains may have evolved from the same ancestors, the acidophilic heterotrophic bacteria Beijerinckia indica and Rhodopseudomonas acidophila (Dedysh, Panikov, Liesack, et al., 1998). Also, based on molecular analysis of mmoX genes of the Methylocella strains, they appear more closely related to Methyloferula spp. than to the known Methylosinus‐Methylocystis cluster in the α‐proteobacteria (Figure 2C). The analysis showed that the genus of Methylosinus and Methylocystis are more closely related to each other than to Methylocella.

In 2002, another species belonging to a novel genus, Methylocapsa acidiphila B2^T, was isolated from an acidic Sphagnum peat bog (Dedysh et al., 2002). This bacterium belongs to the α‐proteobacteria and has 97.3% similarity to the 16S rRNA of Methylocella palustris K^T (Figure 2A). However, there was only 7% DNA–DNA hybridization between Methylocapsa acidiphila B2^T and Methylocella palustris K^T. Methylocapsa acidiphila B2^T only expressed pMMO, which was different from Methylocella palustris K^T which only expressed sMMO (Dedysh et al., 2002). Methylocapsa aurea KYG^T, another acidophilic methanotroph in this genus, was observed to be facultative, capable of using acetate as a carbon source in addition to CH₄ (Dunfield et al., 2010; Farhan Ul Haque et al., 2020). This capability distinguishes the strain from closely related Methylocapsa acidiphila B2^T which is unable to grow on non‐C1 substrates. Methylocapsa aurea KYG^T also proved to be more sensitive to pH and salt concentration than other strains in the genus, with optimum pH in a narrow range of 6.0–6.2 and not surviving at pH <5 (Dunfield et al., 2010).

Methylocella silvestris, which was isolated from an acidic forest Cambisol, is morphologically and phenotypically similar to Methylocella palustris K^T (Dunfield et al., 2003). This organism is a facultative methanotroph that possesses only a form of sMMO that is produced by the mmoX gene. As with many methanotrophs that have both sMMO and pMMO, the expression of sMMO is affected by the concentration of copper in the growth medium. However, in this bacterium, the expression of sMMO was not downregulated by copper (Theisen et al., 2005). Methylocella silvestris was the first methanotroph observed to be capable of using propane as a carbon source while constitutively expressing sMMO (Crombie & Murrell, 2014), although other works showed that the expression of mmoX was repressed during growth on acetate (Rahman et al., 2011). During propane oxidation in this strain, sMMO and propane monooxygenase (PrMO) are both expressed (Dunfield & Dedysh, 2014). The substrates that can be utilized by Methylocella silvestris have been expanded to include 2‐propanol, 1,2‐propanediol, acetone, methyl‐acetate, acetol, glycerol, propionate, tetrahydrofuran, and gluconate, as well as the gaseous alkanes ethane and propane (Dunfield & Dedysh, 2014).

Another methanotroph capable of growth at low pH, Methylocella tundrae, uses sMMO to grow on CH₄ (Dedysh et al., 2004), but genome analysis indicates that this bacterium also carries a PrMO gene cluster in its megaplasmids (Kox et al., 2019). While two species of Methylocella possess PrMO allowing them to grow on multi‐carbon substrates, the PrMO gene cluster of Methylocella silvestris is in its genome while that of Methylocella tundrae is in its megaplasmids. A comparative genomic study of Methylocella indicates a close relationship to Beijerinckia indica, which are acidophilic nitrogen‐fixing bacteria but non‐methanotrophs (Tamas et al., 2014). Methylocella and Beijerinckia have been suggested to have evolved from a common obligate methanotroph, but each has expanded beyond just CH₄ as a substrate; in the case of Beijerinckia indica, losing CH₄ metabolism altogether (Dunfield & Dedysh, 2014; Tamas et al., 2014).

Species in the genus Methyloferula are thought to only possess sMMO (Vorobev et al., 2011). Unlike Methylocella silvestris, which contains an additional soluble diiron monooxygenase for propane oxidation, Methyloferula stellata is an obligate methanotroph (Farhan Ul Haque et al., 2020; Vorobev et al., 2011). However, the 16S rRNA gene and sMMO sequence analysis the between two genera suggested that they are closely related (Figure 2A,C).

Methylocystis possess both pMMO and sMMO, and the expression of these MMOs is affected by the concentration of copper as typical of many methanotrophs (Table 2). Methylocystis heyri and Methylocystis bryophila are two moderately acidophilic methanotrophs in this genus (Belova et al., 2013; Dedysh et al., 2007). Interestingly, M. heyri shows a unique profile of phospholipid fatty acids, containing both 18:1ω8c (i.e. a common type of α‐proteobacterial methanotrophs) and 16:1ω8c (i.e. a common type of γ‐proteobacterial methanotrophs) (Dedysh, 2011). These two species are also facultative methanotrophs, and they can express different forms of MMOs, perhaps explaining why they are widespread in the environment (Han et al., 2018; Leng et al., 2015). As an example, M. bryophila contains two different pmoA genes, pmoA1, and pmoA2, responsible for pMMO1 and pMMO2, respectively. The pMMO2 shows a higher CH₄ affinity than pMMO1, potentially allowing growth over a wide range of CH₄ concentrations (Baani & Liesack, 2008).

Methylomonas paludis was the first acidophilic methanotroph discovered that belongs to the γ‐proteobacteria (Danilova et al., 2013). The absence of motility and the ability to grow under acid conditions makes it different from other species in Methylomonas (Danilova et al., 2013). Figure 2A,B shows the long distance between Methylomonas paludis and other acidophilic methanotrophs based on 16S rDNA and pmoA genes, respectively. Based on 16S rRNA sequences, Methylomonas paludis has 80%–90%, homology to the acidophilic methanotrophic species in class α‐proteobacteria (Figure 2A). The pmoA gene sequence of Methylomonas paludis shows 71% homology to that of Methylococcus capsulatus; both species belong to the class of γ‐proteobacteria (Figure 2B). Recently, two novel acid‐tolerant moderately thermophilic methanotrophs, Methylococcaceae strain BFH1 and BFH2, belonging to γ‐proteobacteria have been isolated from tropical soils with CH₄ leakage (Islam et al., 2016).

Mechanisms of survival at low pH

The mechanisms for the survival of acidophilic methanotrophs, particularly those capable of growth under extremely acidic conditions (e.g. pH < 3) are not well understood. However, the general literature on acidophilic organisms indicates a number of different potential strategies that allow existence under highly acidic conditions, including reversed membrane potentials, extremely impermeable membranes, and the occurrence of numerous secondary transporters (Baker‐Austin & Dopson, 2007). The majority of these hypotheses are derived from genome and biochemical analyses. The composition of fatty acids and lipids in the cell membrane of acidophiles has been observed to differ from more neutrophilic organisms supporting the critical nature of membrane structure in low pH survival (Sharma et al., 2016). Other proposed, but as yet unproven, mechanisms to maintain neutral cytoplasmic pH include buffering and sequestration of protons inside the cytoplasm (Sharma et al., 2016). Future studies are needed to specifically investigate mechanisms of pH tolerance in acidophilic methanotrophs and to compare these approaches to those of more widely studied acidophiles.

Heavy metal resistance of acidophilic methanotrophs

Acidophilic microbes often have enhanced heavy metal resistance due to the likelihood of encountering high concentrations of many metals at low pH based on solubility considerations (Nordstrom et al., 2000). Some acidophiles have developed efflux pumping systems for heavy metals and/or expressed heavy metal resistance or reductase genes (Dopson & Holmes, 2014; Mangold et al., 2013). Heavy metal resistance genes include copCD, terB, and merR which are responsible for copper, tellurite, and mercury resistance, respectively. Methylobacter sp. also has genes for specific reductases including arsC (arsenate reductase). Thus, this strain has multiple strategies to protect against the toxicity of heavy metals.

METHANOTROPHIC ACTIVITY IN ACIDIC WETLANDS

Acidic peat bogs, which are dominated by the mosses of the genus Sphagnum, are one of the most extensive types of wetlands (Dedysh, 2009; Kolb & Horn, 2012). These bogs, mostly located from 50°N to 70°N latitude, are a significant source of CH₄, emitting 100–237 Gt per year (Dedysh, 2009; Kolb & Horn, 2012). Type II acidophilic methanotrophs, including Methylocystis, Methylocapsa, and Methylocella, have been most commonly identified in the acidic wetlands and are likely key contributors to the regulation of CH₄ fluxes from these environments, reducing the overall impact on climate change (Dedysh, 2009). Some methanotrophs in acidic bogs are thought to have a symbiotic relationship with Sphagnum, but the interactions between the mosses and methanotrophs are not completely understood (Dedysh, 2011). Methylocystis, one of the active and dominant methanotrophs in acidic wetlands, has been highlighted as a facultative methanotroph, showing the ability to use acetate for growth in addition to CH₄ (Kolb & Horn, 2012). The strain can change its growth substrate to acetate from CH₄ during periods of CH₄ depletion in acidic environments (Dedysh, 2009, 2011). Interestingly, Methylocystis contains a distinct pmoA2 gene, which differs from the pmoA gene, by encoding a pMMO capable of oxidizing CH₄ at very low environmental concentrations (Baani & Liesack, 2008). Due to the high CH₄ affinity of pmoA2, species containing this gene/enzyme are likely to consume CH₄ at atmospheric concentrations (Kolb & Horn, 2012). Thus, this organism appears capable of survival under low and CH₄‐limiting conditions and can switch to an alternate substrate if CH₄ is absent. Type I methanotrophs belonging to γ‐proteobacteria also have been identified in acidic wetlands even though these environments are commonly dominated by the Type II methanotrophs as just described (Dedysh, 2009; Kolb & Horn, 2012; Nguyen et al., 2018). Verrucomicrobial methanotrophs, which have been isolated from extremely acidic conditions, are also found in acidic wetlands, but the overall distribution of these methanotrophs is still unknown (Dedysh, 2009). Further studies are required to better understand the diversity, distribution, and activity of methanotrophs in acidic wetlands.

GROUNDWATER POLLUTION AND METHANOTROPHS

Biodegradation of groundwater contaminants by MMO

Cometabolic biodegradation typically occurs when one or more broad‐specificity enzymes (typically monooxygenases) are induced in bacteria—enzymes that allow such bacteria to grow on a primary substrate (e.g. CH₄, ethane, propane, and butane), yet also to biodegrade a range of other non‐growth compounds, including many contaminants of concern (Alexander, 1994). Numerous different organisms are capable of cometabolic biodegradation including species of Pseudomonas, Burkholderia, and Rhodococcus, but methanotrophs have perhaps received the most study in this regard (Chu & Alvarez‐Cohen, 1998; Halsey et al., 2005; Mahendra & Alvarez‐Cohen, 2006; Singh & Singh, 2017; Wang & Chu, 2017).

In the 1980s, researchers found that CH₄ stimulated TCE degradation in aerobic sediment columns and in a mixed methanotrophic culture (Fogel et al., 1986; Wilson & Wilson, 1985). These studies were important because, at the time, chlorinated solvents such as TCE were perhaps the most important and widespread emerging environmental contaminants, and little was known concerning their biodegradation. In 1988, two pure methanotrophs capable of cometabolically biodegrading TCE were isolated from groundwater samples (Little et al., 1988). The two common monooxygenases possessed by methanotrophs (pMMO and sMMO) each were observed to catalyse the biodegradation of TCE, although rates were found to be appreciably more rapid via sMMO (Lee et al., 2006). During TCE oxidation by either sMMO or pMMO, the initial step is oxidation to TCE‐epoxide, followed by spontaneous and or further enzymatic degradation of the epoxide to multiple products including formate, carbon monoxide, glyoxylic acid, and dichloroacetic acid (Figure 3).

FIGURE 3.

Cometabolic degradation pathway of trichloroethylene (TCE). CH₄, methane; CH₃OH, methanol; pMMO, particulate methane monooxygenase; sMMO, soluble methane monooxygenase.

MMOs (pMMO and sMMO) have been observed to biodegrade a variety of other groundwater contaminants besides TCE, including many different chlorinated aliphatics (e.g. VC, cis‐DCE, 1,2,3‐trichloropropane [TCP], and chloroform [CF]), aromatics such as BTEX, some aromatic hydrocarbons, and numerous other organic pollutants (Bosma & Janssen, 1998; Hand et al., 2015; Im & Semrau, 2011; Malachowsky et al., 1994; Wackett et al., 1989; Wang & Chu, 2017). cis‐1,2‐DCE and VC are degradation intermediates that occur commonly during anaerobic biodegradation of TCE (Schäfer et al., 2003; Semprini et al., 1990; Wüst et al., 1999) (Table 3). TCP was used as a chemical intermediate in organic synthesis, as a solvent, and as an extractive agent. It has been detected in hundreds of surface water and drinking water sources in the US, at levels of 0.1–100 μg/L. Typically, remediation of TCP‐contaminated sites is difficult due to its physiochemical properties (Salter et al., 2010).

TABLE 3.

Representative compounds degraded by monooxygenases.

	pMMO		sMMO		PrMO		Reference
Chlorinated aliphatic hydrocarbons
TCE	+	Methylosinus trichosporium OB3b Methylocystis strain SB2	+	Methylosinus trichosporium OB3b Methylocella spp. a Methylocella palustris a	+	Mycobacterium vaccae JOB5 Rhodococcus jostii RHA1 Rhodococcus rhodochrous	Anderson and McCarty (1997), Hatzinger et al. (2017), Im and Semrau (2011), Malachowsky et al. (1994), Oldenhuis et al. (1991), Shao et al. (2019), Tsien et al. (1989), Wackett et al. (1989)
TCP	N/A		+	Methylosinus trichosporium OB3b	+	Rhodococcus jostii RHA1 Rhodococcus rubber ENV425 Mycobacterium vaccae JOB5 Sphingopyxis sp. AXE‐A	Bosma and Janssen (1998), Wang and Chu (2017)
VC	+	Methylosinus trichosporium OB3b Methylocystis strain SB2 Methylomonas sp. Strain JS1 a	+	Methylosinus trichosporium OB3b Methylocystis sp. Strain MJC1 a	+	Mycobacterium vaccae JOB5 Rhodococcus rhodochrous	Anderson and McCarty (1997), Choi et al. (2021), Hand et al. (2015), Im and Semrau (2011), Malachowsky et al. (1994), Tsien et al. (1989), Wackett et al. (1989)
Chloroform	+	Methylocystis strain SB2	+	Methylosinus trichosporium OB3b	+	Rhodococcus rhodochrous	Im and Semrau (2011), Malachowsky et al. (1994), Oldenhuis et al. (1989)
Aromatic hydrocarbons
Naphthalene	/	Methylosinus trichosporium OB3b	+	Methylosinus trichosporium OB3b	N/A		Brusseau et al. (1990)
BTEX	+	Methylocystis sp. Methylosinus trichosporium OB3b	+	Methylococcus capsulatus Methylosinus trichosporium OB3b	+	Rhodococcus erythropolis	Hesselsoe et al. (2005), Lee et al. (2011), Kulikova and Bezborodov (2000), Wilkins et al. (1994)
Other compounds
1,4‐Dioxane	/	Methylosinus trichosporium OB3b	+ /	Methylosinus trichosporium OB3b	+	Mycobacterium vaccae JOB5 Rhodococcus jostii RHA1	Hand et al. (2015), Mahendra and Alvarez‐Cohen (2006)
MTBE	N/A		+	CH4 grown microbial consortia	+	Rhodococcus sp. Strain ENV425	Hesselsoe et al. (2005), Steffan et al. (1997)
NDMA	/	Methylosinus trichosporium OB3b	+	Methylosinus trichosporium OB3b	+	Mycobacterium vaccae JOB5	Sharp et al. (2005)
‐	‐		‐		‐
Pesticides	+	CH4‐oxidizing cultures	+	CH4‐oxidizing cultures	N/A		Hedegaard et al. (2020)
Pharmaceuticals	+	Methylocystis parvus	+	CH4‐oxidizing cultures	N/A		Benner et al. (2015)

Abbreviations: BTEX, Benzene, toluene, ethylbenzene and xylene; ‐; MTBE, methyl tertiary‐butyl ether; NDMA, N‐nitroso‐dimethylamine; pMMO, particulate CH₄ monooxygenase; PrMO, propane monooxygenase; sMMO, soluble CH₄ monooxygenase; TCE, trichloroethylene; TCP, trichloropropane; VC, vinyl chloride.

^a Acidophilic methanotrophs.

^b Able to degrade; +, Not able to degrade /; N/A, no related literature. Results in the literature differ for 1,4‐dioxane and sMMO, thus the + / designation.

The rates and extents of pollutant degradation by methanotrophs vary based on the MMO expressed. sMMO‐expressing methanotrophs typically degrade more compounds than pMMO‐expressing cells and degrade such compounds at faster initial rates (Alvarez‐Cohen & McCarty, 1991; Anderson & McCarty, 1997; Lee et al., 2006; Lontoh & Semrau, 1998; Shukla et al., 2009; Tsien et al., 1989). Thus, much of the early literature focused on the utility of sMMO‐expressing organisms for pollutant degradation (Alvarez‐Cohen & McCarty, 1991; Lee et al., 2006; McDonald et al., 1997; Shigematsu et al., 1999). For example, sMMO but not pMMO can oxidize naphthalene (Brusseau et al., 1990; Chang et al., 2002). sMMO is also able to oxidize several emerging contaminants, including methyl tertiary‐butyl ether (MTBE) (Hesselsoe et al., 2005; Steffan et al., 1997) and N‐nitrosodimethylamine (NDMA) among others (Sharp et al., 2005). Pharmaceuticals such as sulfamethoxazole (Benner et al., 2015) and ibuprofen (Dawas‐Massalha et al., 2014), as well as a variety of pesticides including bentazone, dichlorprop, and chlorophenoxy herbicide (Hedegaard et al., 2020; Papadopoulou et al., 2019), are also degraded by sMMO. 1,4‐Dioxane, which is commonly used as a stabilizer for chlorinated solvents such as 1,1,1‐trichloroethane (TCA), was originally reported to be oxidized by sMMO (Mahendra & Alvarez‐Cohen, 2006). Later studies with pure cultures and pure sMMO enzyme revealed, however, that 1,4‐dioxane is not a substrate for sMMO and is unlikely to be degraded by methanotrophs in the environment (Hatzinger et al., 2017).

Other studies suggest that pMMO‐expressing methanotrophs survive more readily in environments with complex mixtures and/or high concentrations of pollutants such as chlorinated ethenes (Semrau, 2011). It is thought that sMMO‐expressing methanotrophs accumulate toxic products resulting from pollutant oxidation (e.g. epoxides from chlorinated ethenes) faster than pMMO‐expressing cells, and thus may experience higher rates of cell toxicity due to epoxide‐mediated damage to the sMMO enzyme and other cellular macromolecules (Chu & Alvarez‐Cohen, 1999; Fox et al., 1990; Semrau, 2011). In this instance, the overall pollutant degradation by sMMO‐expressing methanotrophs may be less than observed for those expressing only pMMO (Lee et al., 2006). Due to the potential for cell toxicity for both sMMO and pMMO from chlorinated ethenes, cometabolic treatment of these compounds in groundwater is likely to be more effective when they are present at relatively low concentrations (i.e. tens to hundreds of μg/L). It should also be noted, however, that the toxicity of the relevant epoxides from chlorinated ethenes vary widely by compound.

Contamination and remediation in acidic groundwater

Most of the aforementioned studies of pollutant degradation by methanotrophs have been conducted under neutral pH conditions. There is comparatively little information on the capabilities of acidophilic methanotrophs to biodegrade pollutants, such as TCE or others in groundwater. This is particularly important because low pH groundwater is common throughout the Northern Atlantic Coastal Plain aquifer system in the United States, which occurs from Long Island, New York through most of North Carolina (Denver et al., 2014). Of 419 groundwater samples collected by USGS from this aquifer system, 250 (60%) were reported to have pH values of 5.5 or below (data archive; https://pubs.usgs.gov/circ/1353/). This aquifer system, which includes a number of large military facilities and some large urban areas, is also significantly impacted by CVOCs (Denver et al., 2014). A second aquifer system composed of similarly semi‐consolidated sands with poor buffering capacity is the Gulf of Mexico Coastal Plain aquifer system, running from Georgia, through the panhandle of Florida, and to the southern tip of Texas (DeSimone et al., 2014). This system also has many sites with CVOCs in low‐pH groundwater and/or groundwater with very poor buffering capacity. Many other locations in the United States also have locally acidic groundwater.

As noted in the Introduction, in situ bioremediation of CVOC‐contaminated sites is often performed by adding high concentrations of carbon sources (such as lactate or emulsified vegetable oils) to stimulate natural or introduced Dehalococcoides spp., to anaerobically degrade PCE and TCE to ethene (Chen et al., 2022; Stroo & Ward, 2010; Yang et al., 2017). However, one significant issue with anaerobic bioremediation of CVOCs is that complete reductive dechlorination (i.e. PCE or TCE to ethene) by Dehalococcoides spp. is typically inhibited at pH < ~5.5 (Eaddy, 2008; Lacroix et al., 2014; Rowlands, 2004; Steffan & Vainberg, 2013; Vainberg et al., 2009; Yang, 2012; Yang et al., 2017). Most organisms or consortia capable of reducing TCE have pH optima between ~6.5 and 8 and do not effectively dehalogenate this compound, or other chlorinated ethenes to ethene below pH 5.5. Supporting organisms, such as those that produce cobalamin required by Dehalococcoides spp., also may be inhibited at low pH (Puentes Jácome et al., 2019). In a study by Yang (2012), only one strain was observed to effectively dehalogenate PCE at pH 5.5 (Sulfospirillum multivorans), and this organism degraded PCE to cis‐DCE as a terminal product. VC, a known carcinogen with a U.S. Federal Maximum Contaminant Level (MCL) of 2 μg/L, is also a potential terminal product of a stalled anaerobic biodegradation of chlorinated ethenes in acidic environments. As a result, anaerobic bioremediation is largely ineffective at reducing chlorinated ethenes to ethene in naturally acidic aquifers. In addition, when carbon sources are added to groundwater aquifers in large quantities to promote reductive dechlorination, the formation of organic and inorganic acids can cause pH in poorly buffered aquifers to fall below optimal levels, resulting in incomplete or stalled dechlorination (McCarty et al., 2007).

Aquifer buffering has been attempted in some instances to increase groundwater pH for remediation purposes (Hatzinger et al., 2006; Schaefer et al., 2010), but the amount of buffer required makes this process cost prohibitive for other than small sites. In some instances, a strong base (e.g. NaOH) has been used to increase aquifer pH, but this can easily result in overshooting the desired pH range, and subsequently causing significant precipitation reactions as well as dissolution of natural organics. Because of the difficulty in applying typical in situ anaerobic bioremediation technologies in acidic groundwater, other remediation strategies are required to treat CVOCs in these aquifers. The application of methane with or without exogenous acidophilic methanotrophs may represent an appropriate strategy in many such environments as described below. It is also possible that these organisms are already contributing to the natural attenuation of CVOCs in acidic groundwater, but this process is largely unrecognized.

POTENTIAL BIOREMEDIATION APPLICATIONS IN LOW pH AQUIFERS

The potential for acidophilic methanotrophs to biodegrade pollutants is largely unknown, and evaluation of potential applications of these organisms has just begun in recent years (Choi et al., 2021; Semrau, 2011; Shao et al., 2019; Szwast, 2021). This pursuit is important, particularly for chlorinated solvents because, as previously noted, traditional bioremediation approaches are not as effective at low pH for completely dechlorinating CVOCs, particularly chlorinated ethenes. Secondly, the recent observation that a number of methanotrophs, including six different acidophilic strains, are facultative (Farhan Ul Haque et al., 2020) enhances the potential for natural biodegradation of pollutants (e.g. in environments without CH₄ but with alternate substrates), as well as new ways to enhance pollutant bioremediation in acidic environments. The critical questions are (1) are acidophilic methanotrophs present in the contaminated groundwater environments (2) do they possess forms of MMO capable of biodegrading TCE and other contaminants after growth on CH₄; and (3) are these MMO(s) expressed and active using alternate substrates. One methanotroph, Methylocystis strain SB2, which was isolated from a neutral medium, constitutively expressed pMMO growing with ethanol and successfully degraded VC, trans‐dichloroethylene (t‐DCE), TCE, and 1,1,1‐trichloroethane (1,1,1‐TCA) through cometabolism (Im & Semrau, 2011; Jagadevan & Semrau, 2013).

However, studies on the cometabolism of pollutants by methanotrophs in acidic environments are lacking (Table 3). One initial study by our group demonstrated that Methylocella palustris degraded TCE and several other halogenated organics including 1,2‐dibromoethane (EDB), chloroform, VC, and cis‐DCE, but not perchloroethene (PCE) or 1,4‐dioxane at pH 5.0 (Hatzinger et al., 2017). Further work showed that methanotrophs capable of degrading TCE existed in multiple acidic aquifers (Shao et al., 2019). Using stable‐isotope‐probing techniques, phylogenetically diverse active methanotrophs were detected in low‐pH aquifer microcosms (Shao et al., 2019). The methanotrophs in these microcosms included Methylomonas, Methylocaldum, Methylobacter, Methylosinus, and Methylococcus, which belong to γ‐proteobacteria or α‐proteobacteria, but are not necessarily related to other known acidophilic methanotrophs. It is likely that one or more of these organisms facilitated the observed cometabolic biodegradation of TCE. In addition, a recent study showed the cometabolic biodegradation of VC in acidic environments by isolated acidophilic methanotrophs from acidic peat soils (Choi et al., 2021).

The quantification of natural attenuation of TCE under oxidative conditions has been challenging due to the general absence of easily detected daughter products akin to the production of cis‐DCE, VC, and ethene via reductive dehalogenation. However, a method developed at Clemson University that utilizes ultrapure ¹⁴C‐TCE to quantify the oxidative conversion of this CVOC to ¹⁴CO₂ and soluble ¹⁴C‐labelled daughter products, has overcome this limitation (Mills IV et al., 2018). This technique has recently been applied to estimate TCE degradation rates under aerobic conditions in an acidic aquifer in Maryland with pH ranging from 4.3 to 6.1 (Szwast, 2021). Biostimulation via CH₄ and inorganic nutrient addition also was assessed. First‐order rate constants ranging from 0.012 to 3.0 year⁻¹ were calculated (half‐lives of 0.23–59 years) across the range of microcosms from three locations. The highest rate constants were generally in treatments with CH₄, and nutrients added, but TCE degradation was also observed in treatments representative of in situ conditions (i.e. no additions). These data support the hypothesis that methanotrophs are important but largely unrecognized contributors to aerobic cometabolism of TCE under low pH conditions.

CONCLUSIONS AND FUTURE WORK

This review focuses on the distribution, diversity, and potential bioremediation activities of acidophilic methanotrophs. Since acidophilic methanotrophs were first discovered more than 25 years ago, great progress has been made in describing their characteristics and distribution in low pH environments, including peat bogs, wetlands and lakes, thermal soils, and springs, and more recently, groundwater aquifers. New genera have been discovered with unique physiological characteristics, including the potential for utilizing longer‐chain compounds for substrates. Compared to neutral pH environments, however, there is a relative dearth of information concerning the potential for methanotrophs to biodegrade CVOCs and other pollutants under acidic conditions either naturally or via biostimulation or bioaugmentation. This is particularly important for TCE and many other CVOCs because reductive biodegradation processes are not particularly effective below pH 5.5.

There are a number of important areas that require further research. These include the isolation and identification of pure methanotrophic cultures from low pH groundwater environments, as to date no such organisms are available for study. This is critically important for understanding the fundamental physiological and biodegradative capabilities of methanotrophs in groundwater environments. Another critical area is the assessment of facultative growth of acidophilic methanotrophs in groundwater and whether alternate (i.e. non‐CH₄) substrates can be utilized as carbon sources to promote methanotrophic degradation of TCE and other CVOCs. This question is critical to our understanding of the natural attenuation of these pollutants in aerobic aquifers, a largely unstudied area. Further studies on community diversity and dynamics of methanotrophs and associated organisms (i.e. non‐methanotrophs) that may contribute to pollutant biodegradation are also of interest. Finally, as we come to better understand the potential of these organisms for degrading persistent pollutants, field studies of biostimulation (e.g. CH₄, nutrient, and oxygen addition) for enhanced pollutant remediation in acidic aquifers are required so that this approach can ultimately be optimized and utilized for large‐scale treatment of CVOCs and other pollutants, much the way reductive dehalogenation has been widely applied for CVOC treatment in neutral pH environments.

CONFLICT OF INTEREST STATEMENT

The authors declare no conflicts of interest.

Hwangbo, M. , Shao, Y. , Hatzinger, P.B. & Chu, K.‐H. (2023) Acidophilic methanotrophs: Occurrence, diversity, and possible bioremediation applications. Environmental Microbiology Reports, 15(4), 265–281. Available from: 10.1111/1758-2229.13156