Fungal diversity in peatlands and its contribution to carbon cycling

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Review

Fungal diversity in peatlands and its contribution to carbon cycling R. Juan-Ovejeroa,*, M.J.I. Brionesa, M. Öpikb a b

Departamento de Ecología y Biología Animal, Universidad de Vigo, 36310 Vigo, Spain Institute of Ecology and Earth Sciences, Department of Botany, University of Tartu, 51005, Tartu, Estonia

ARTICLE INFO

ABSTRACT

Keywords: Abiotic factors Climate change Fungal guilds Pristine peatlands Organic matter decomposition Peatland type

Peatlands are major carbon sinks globally, but it is still unclear what drives their shift from functioning as carbon sink to a source. Fungi rely on soil carbon inputs and play an active role in carbon mobilization and stabilization. Future climate change scenarios predict increases in temperature and lower water tables, which may lead to functional shifts in fungal communities growing in peatlands. Despite their abundance, the impact of fungi on carbon cycling in peatlands is still poorly understood. Therefore, it is crucial to study the dynamics and distribution of fungal communities in pristine peatlands in order to predict the potential changes in peatland ecosystems more accurately. Here, we review the current knowledge about fungal communities in peatlands, including the influence of peatland type, abiotic factors, and temporal and spatial dynamics on fungal diversity. Our overview shows that fungal diversity in peatlands of certain regions such as tropical areas is severely understudied. Furthermore, we examine the ecological roles and the relative contribution to carbon cycling of functional guilds of fungi in peatlands, showing that saprotrophs are the key guild in organic matter decomposition, and ericoid- and ectomycorrhizal fungi have the ability to act as decomposers and vectors of plant carbon input to soil. We suggest that the application of a combination of methodological approaches for taxonomic and functional characterisation of peatland fungal communities is needed to build better understanding about the diversity-functioning relationships in peatland ecosystems. Overall, this review shows that fungi must be considered as important drivers of mechanistic processes underlying biodiversity, plant nutrient uptake and carbon losses in peatlands.

1. Introduction Peatlands have low nutrient availability and high organic matter accumulation that vastly exceeds decomposition and therefore they store approximately one third of the global soil carbon pools (Gorham, 1991; Limpens et al., 2008). Carbon balance depends on gas influx and efflux and the production and release of dissolved organic carbon (DOC) (Fenner et al., 2007), which are also linked to soil biological activities (de Vries et al., 2012; Filser et al., 2016). Microbial community (including fungi, bacteria and archaea) dynamics and structure have been extensively studied in peatland ecosystems, since these organisms are known to play important roles in carbon cycling processes, such as organic matter decomposition and nutrient mineralization or uptake (Mitchell et al., 2003). Strong dominance of bacteria over fungi has been found in one bog dominated by dry hummocks, one poor fen and one rich fen in Canada (Winsborough and Basiliko, 2010). In contrast, higher fungal than bacterial biomass has been observed in bogs with low pH values and high C/N ratios (Golovchenko et al., 2007; Jaatinen et al., 2007), cold mountain bogs

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(Robroek et al., 2013), but also at low altitude bogs (Bragazza et al., 2015). Also, fungi are more adapted to soils with low nutrient availability than bacteria (Wardle et al., 2004). Bacteria and fungi have dissimilar responses to changes in temperature (Bárcenas-Moreno et al., 2009) and moisture (Manzoni et al., 2012), thus a shift in the microbial community structure in a climate change scenario with elevated temperatures and low water tables may affect the amount of carbon losses released to the atmosphere from peatland ecosystems (Kim et al., 2008; Mäkiranta et al., 2009). Proper assessment of the status of peatland biodiversity requires that peatlands are studied in their pristine condition, i.e. not being affected by drainage, afforestation or other anthropogenic activities (Joosten et al., 2012). Peatlands show a low global taxonomic diversity when compared to other terrestrial ecosystems, hosting 5–25% of endemic species (Parish et al., 2008). Fungi are highly specialized in peatlands and fungal diversity strongly depends on soil organic matter chemical composition (Lin et al., 2012, 2014). Importantly, some features of fungi such as ability to degrade complex carbon polymers (Thormann, 2006), including recalcitrant cellulose, hemicellulose and

Corresponding author. E-mail address: [email protected] (R. Juan-Ovejero).

https://doi.org/10.1016/j.apsoil.2019.103393 Received 12 September 2019; Received in revised form 14 October 2019; Accepted 16 October 2019 0929-1393/ © 2019 Elsevier B.V. All rights reserved.

Please cite this article as: R. Juan-Ovejero, M.J.I. Briones and M. Öpik, Applied Soil Ecology, https://doi.org/10.1016/j.apsoil.2019.103393

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polyphenols (Talbot et al., 2008), and to produce methane in aerobic environments (Lenhart et al., 2012), may enhance decomposition rates under global warming. Knowledge about fungal communities in peatlands gained by morphological identification of fruit bodies and cultures has been reviewed before (Thormann and Rice, 2007). Additionally, culturing-based approaches have provided essential understanding about peatland fungal diversity. For instance, fungal assemblages differ between peatlands with different dominant vegetation (Thormann et al., 2004a, 2004b; Stasiöska, 2014; Filippova and Thormann, 2014). Furthermore, vertical stratification of fungal communities along the peat profile was observed, showing differences in community composition between hummocks and hollows of fens and bogs (Nilsson et al., 1992). However, culturing-based approaches do not directly inform about all fungal species growing in peatlands (Andersen et al., 2013), therefore DNAbased identification has become a prevalent approach in the last years. Here, we focus on studies carried out in pristine peatlands, i.e. undrained peatlands that still remain in natural or near-natural conditions. Our review provides an update in knowledge of the peatland fungal diversity gleaned with molecular techniques, and also aims to disentangle the contribution of fungi to carbon cycling in pristine peatland ecosystems.

activities (Table 1). Traditionally, identification of fungal sporocarps has provided information about the fungal biota in these ecosystems (e.g. Stasińska, 2014; Grzesiak and Wolski, 2015; Filippova and Thormann, 2014; Sun et al., 2016) and spore morphology is a classical method used to identify arbuscular mycorrhizal fungi (Morton, 1988; Fuchs and Haselwandter, 2004; Kołaczek et al., 2013). Additionally, establishment of permanent plots to monitor fruiting body surveys is a useful and inexpensive method (Hill et al., 2005), although it only allows to obtain data on the currently fruitbody forming part of the fungal communities. Pure culture based approaches expand the proportion of local fungal diversity to be identified, as non-sporulating but culturable species can also be detected. The first studies used a limited range of culturing media for species isolation (Dooley and Dickinson, 1970; Nilsson et al., 1992), but later application of a wider variety of media and growing conditions enabled to further enlarge the number of fungal species to be detected and identified by culturing (e.g. Thormann et al., 2001, 2004a, 2007). However, the media types and incubation conditions may have benefited fungi species with fast growth rates and species with different growth characteristics and requirements might have remained undetected (Thormann, 2006; Grum-Grzhimaylo et al., 2016). In addition, only a small proportion of soil and peat-inhabiting fungi is known to be amenable to pure culture conditions (Hawksworth, 2001; Allen et al., 2003; O’Brien et al., 2005). Estimation of biomass of broad fungal taxa and other microbial groups can be achieved by fatty-acid based approaches which quantify individual marker lipids in microbial cells (Willers et al., 2015). Namely, phospholipid fatty acid (PLFA) markers are indicators of the

2. Methodological approaches to study peatland fungi A range of methods has been used to identify fungi in peatlands, as well as to quantify fungal biomass and measure their enzymatic

Table 1 Methodological approaches used to identify fungal species, quantify fungal biomass and measure fungal activity in pristine peatlands. Information provided

Method

Specific Method

References

Identification

1) Morphological

a) Sporocarp collection

1a) Stasińska, 2014; Grzesiak and Wolski, 2015; Filippova and Thormann, 2014; Sun et al., 2016 1b) Fuchs and Haselwandter, 2004; Kołaczek et al., 2013 1c) Dooley and Dickinson, 1970; Nilsson et al., 1992; Thormann et al., 2001; Polyakova et al., 2001 Polyakova and Chernov, 2002; Thormann et al., 2004a,b; Rice and Currah, 2006; Thormann et al., 2007; Jassey et al., 2011; Kachalkin and Yurkov, 2012; Terhonen et al., 2014; Grum-Grzhimaylo et al., 2016; Jaiboon et al., 2016; Polburee et al., 2017 2) Polyakova and Chernov, 2002; Wurzburger et al., 2004; Wolfe et al., 2007; Toberman et al., 2008; Kim et al., 2008; Peltoniemi et al., 2009; Kjøller et al., 2010; Straková et al., 2011; Kachalkin and Yurkov, 2012; Lin et al., 2012; Peltoniemi et al., 2012; Preston et al., 2012; Kwon et al., 2013; Hazard et al., 2014; Lin et al., 2014; Terhonen et al., 2014; Grum-Grzhimaylo et al., 2016; Jaiboon et al., 2016; Sun et al., 2016; Asemaninejad et al., 2017a, b; Hiiesalu et al., 2017; Lamit et al., 2017; Polburee et al., 2017; Zhang et al., 2017; Asemaninejad et al., 2018; Jassey et al., 2018; Kennedy et al., 2018; Wang et al., 2019 3a) Sundh et al., 1997; Jaatinen et al., 2007; Mäkiranta et al., 2009; Robroek et al., 2013; Bragazza et al., 2015; Peltoniemi et al., 2015; Robroek et al., 2013; Mpamah et al., 2017; Teurlincx et al., 2018 3b) Olsrud et al., 2007; Peltoniemi et al., 2015 4) Thormann et al., 1999; Mitchell et al., 2003; Fuchs and Haselwandter, 2004; Wurzburger et al., 2004; Golovchenko et al., 2007; Olsrud and Michelsen, 2009; Day and Currah, 2011; Binet et al., 2017; Kołaczek et al., 2013; Chiapusio et al., 2018 5) Winsborough and Basiliko, 2010; Myers et al., 2012; Preston et al., 2012 6) Rice et al.;, 2006; Artz et al., 2007; Toberman et al., 2008; Day and Currah, 2011; Straková et al., 2011; Preston et al., 2012; Kwon et al., 2013; Robroek et al., 2013; Bragazza et al., 2015; Jaiboon et al., 2016; Binet et al., 2017; Jassey et al., 2018; Könönen et al., 2018

b) Spore morphology c) Cultured-based approaches

2) Molecular. DNA-sequenced based approaches

Quantification

3) Chemical markers

4) Staining and microscopy

Activity

a) PLFA

b) Ergosterol

5) Substrate-induced respiration 6) Enzymatic activity

2

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biomass of various groups of microorganisms such as fungi (PLFA markers 18:2ω6,9, 16:1ω5 and 16:1ω9) and bacteria (PLFA markers i15:0, a15:0, 15:0, i16:0, 16;1ω5, 16:1ω7 t, i17:0, a17:0, 17:0, 18:1ω7, cy17:0 and cy19:0) (Frostegård and Bååth, 1996; Olsson, 1999; Frostegård et al., 2011). Application of this method has enabled to quantify the biomass of particular groups of the microbial community of different peatland ecosystems (e.g. Sundh et al., 1997; Jaatinen et al., 2007; Mäkiranta et al., 2009; Bragazza et al., 2015; Teurlincx et al., 2018). Additionally, neutral lipid fatty acid markers (NLFA) are particularly useful to measure arbuscular mycorrhizal fungal (AMF) biomass (e.g. Hammer et al., 2011; Verbruggen et al., 2016). NLFA can provide an indication of AMF energy store, whereas the NLFA/PLFA ratio is used as an indicator of energy consumption or nutrient status of AM fungi (Olsson, 1999; Bååth, 2003). To our knowledge, NLFA biomarkers have not been applied to peatland research. Additionally, quantification of other chemical markers such as chitin and ergosterol can be used to measure total fungal biomass (Baldrian et al., 2013). Chitin is a structural compound, whereas ergosterol is a membrane lipid, and both are found in living and dead fungal cells. In peatlands, ergosterol was successfully used to detect dark septate endophytic fungi but not ericoid mycorrhizal fungi (Olsrud et al., 2007) and labelling ergosterol with a 14C-acetate solution allowed to measure fungal growth in these carbon-rich ecosystems (Peltoniemi et al., 2015). All of these methods give us quantitative information in terms of biomass but they do not allow to identify these organisms at a finer taxonomic scale, and for example, some specific PLFA markers have been found in both bacteria and fungi, which could lead to data misinterpretation (PLFA markers 16:1ω5 and 18:1ω9; Frostegård et al., 2011). Nevertheless, despite their limitations, these methodological approaches provide information about the physiological abilities and activity level of microbial communities (Willers et al., 2015). Finally, staining and microscopy of root samples have been widely used to quantify colonization of fungi associated with plants growing in peatlands (Thormann et al., 1999; Olsrud and Michelsen, 2009). In the case of measuring extracellular enzyme activity of fungi, enzymes first need to be extracted from bulk samples and then enzyme activity is calculated through fluorescence and spectrophotometric methods (Saiya-Cork et al., 2002; Pritsch et al., 2011). Carbon-degrading enzymatic activities have been estimated in several studies performed in peatlands (e.g. Artz et al., 2007; Bragazza et al., 2015; Jassey et al., 2018). Nevertheless, specific methodological developments may be needed to cover all different aspects of enzymatic degradation in peatland ecosystems around the world. For instance, organic molecules and differences in pH-optimum could have caused the inhibition of some enzymatic activities in tropical peatlands, which are characterized by high lignin content and very low pH values (Könönen et al., 2018). More recently, the use of DNA sequencing-based techniques for fungal identification is increasingly replacing morphological and culture-based identification methods. They allow to identify fungi at species and genotype scales, and henceforth allow to measure diversity more precisely. Furthermore, the advantage of the DNA sequencingbased methods is that unculturable organisms can also be detected, and thus a more comprehensive data on diversity can be obtained. Strengths and limitations of DNA-based approaches for fungal identification in natural samples have been summarised elsewhere (Thomsen and Willerslev, 2015; Ruppert et al., 2019). Molecular methods have demonstrated differences in fungal communities between fens and bogs (Peltoniemi et al., 2009; Lin et al., 2012) and in peatlands differing in their time of formation (Zhang et al., 2017). DNA sequencing also showed differences in patterns of vertical distribution of fungal communities in hummocks and hollows (Asemaninejad et al., 2017b). Various marker regions (SSU rRNA gene, ITS region, LSU rRNA gene), as well as different primers, have been employed for fungal identification in these studies, but some fungal groups could be under-

represented depending on the used marker regions and primers (Nilsson et al., 2008; Asemaninejad et al., 2016). Furthermore, taxonomic resolution of fungal identification may differ between studies (Nilsson et al., 2016). Standardizing protocols of DNA analysis by co-amplifying regions (ITS with SSU or LSU rRNA gene regions) (Thomsen and Willerslev, 2015) and the combinations of primers (Lindahl et al., 2013) could be useful to get more complete data on fungal diversity. A combination of both morphological and DNA-based identification has been used recently to study fungal communities in some peatlands (GrumGrzhimaylo et al., 2016; Sun et al., 2016), and it appears to yield the most complete representation of the fungal community. 3. Fungal diversity in pristine peatlands 3.1. Community structure The phyla Ascomycota and Basidiomycota dominate the fungal communities in peatlands, with average relative abundances of 46% and 40%, respectively (Thormann and Rice, 2007). The divisions Zygomycota, Chytridiomycota, and the two recently defined phyla Mucoromycota (including AM-forming subphylum Glomeromycotina) and Zoopagomycota (Spatafora et al., 2016), are much less common than the Basidiomycota and Ascomycota, usually showing relative abundance values lower than 10% (Thormann and Rice, 2007; GrumGrzhimaylo et al., 2016; Zhang et al., 2017). 3.2. Trophic guilds A trophic guild is a species group that exploits the same type of resources in a similar way (Root, 1967; Simberloff and Dayan, 1991). Distinguishing fungal guilds gives us information about fungal community composition based on trophic strategies rather than taxonomical identity, and allows to understand and compare complex communities (Nguyen et al., 2016). The fungal guilds in peatlands are: saprotrophs, ectomycorrhizal (EcM) fungi, ericoid mycorrhizal (ErM) fungi, arbuscular mycorrhizal (AM) fungi, dark septate endophytic (DSE) fungi, yeasts (although sometimes considered as a morphological group rather than as a fungal guild) and pathogens. Saprotrophic fungi are the most abundant guild in most studies describing fungal communities in peatlands, followed by mycorrhizal fungi (Thormann, 2006; Thormann and Rice, 2007). Saprotrophs are free-living filamentous fungi that find in peatlands a suitable environment with abundant organic material to decompose if environmental conditions are favourable (Thormann, 2006). Mycorrhizal fungi are less common in peat soils than in mineral soils due to the stressful environmental conditions, such as anoxia and low pH (Andersen et al., 2013). EcM fungi are found in northern peatlands associated with coniferous host trees such as Picea mariana, Larix laricina, Pinus contorta and Pinus sylvestris (Thormann et al., 1999; Wurzburger et al., 2004; Sun et al., 2016; Hiiesalu et al., 2017). Generally, dominant tree species in peatlands influence the fungal community composition because many EcM fungi are host specific (Kennedy et al., 2018). ErM fungi usually appear in the hair roots of ericaceous plants in pristine peatlands, enhancing plant growth and health (Thormann et al., 1999; Leopold, 2016). The root symbiotic Pezoloma ericaceae aggregate is the most widespread ErM fungus in these ecosystems, although fungi of orders Chaetothryriales, Helotiales and Sebacinales can also be present in association with heathers and dwarf shrubs (Kjøller et al., 2010; Hazard et al., 2014; Peltoniemi et al., 2015). Some fungi species are able to colonize both ericoid- and ectomycorrhizal hosts, and also live as saprotrophs, such as Oidiodendron maius, which can be found associated to ericaceous plants or living as a saprobe in peat formed by Sphagnum mosses (Rice and Currah, 2006). AM fungi are the least abundant fungal guild in peatlands, because 3

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only a limited number of host species are available, including some fen plant species (Thormann and Rice, 2007; Andersen et al., 2013). Though some mosses form AM, Sphagnum mosses that dominate in peatlands do not associate with AM fungi (Read et al., 2000). AM fungi were found colonizing the roots of Lysimachia nummularia, Mentha longifolia and Ranunculus repens in a Polish fen, with percentages higher than 70% for the three plant species (Kołaczek et al., 2013). Interestingly, the plant species Serratula tinctoria and Betonica officinalis growing in a fen showed higher AM fungi colonization than the plant species Drosera intermedia and Lycopodicela inundata that appeared in a bog, but the bog had a higher density of AM fungi spores than the fen (Fuchs and Haselwandter, 2004). DNA-based studies have found AM fungi in some peatlands in very low abundances (Kennedy et al., 2018; Wang et al., 2019), but in other molecular studies AM fungi were not detected (Peltoniemi et al., 2009). Plant mycorrhizal status indicates whether a plant species always (obligate) or only sometimes (facultative) requires to establish a mycorrhizal association (Moora, 2014). Coniferous trees (e.g. Pinus contorta, Pinus ponderosa, Pinus sylvestris, Picea abies) and ericaceous shrubs (e.g. Andromeda polifolia, Erica tetralix, Ledum palustre, Vaccinium myrtillus, Vaccinium oxycoccos) typically growing in peatlands are obligate mycorrhizal species (Harley and Harley, 1987; Wang and Qiu, 2006; Akhmetzhanova et al., 2012). Data on mycorrhizal status about angiosperms show that the majority of species within the family Juncaceae, which commonly appears in peatlands, are obligate or facultative mycorrhizal species, whereas a lower percentage are non-mycorrhizal plants (Wang and Qiu, 2006; Akhmetzhanova et al., 2012). However, within the family Cyperaceae, most of species are either non-mycorrhizal or facultative mycorrhizal ones (Harley and Harley, 1987; Wang and Qiu, 2006; Akhmetzhanova et al., 2012). Moreover, in the family Droseraceae, which comprises carnivorous plants appearing in many bogs, about two thirds of species are facultatively mycorrhizal and one third are non-mycorrhizal species (Harley and Harley, 1987; Wang and Qiu, 2006). Furthermore, the plant mycorrhizal flexibility describes the plant ability to grow with and without mycorrhizal symbiosis (Moora, 2014; Gerz et al., 2018). Flexibly mycorrhizal plants may have a wide niche that allows them to live in different environmental conditions despite of lacking mycorrhizal associations (Gerz et al., 2018). As mentioned above, certain peatland plants such as sedges have been largely considered as non-mycorrhizal (Wang and Qiu, 2006; Lambers and Teste, 2013). However, it has been shown that sedges could associate with mycorrhizal fungi when water table levels are low as a result of drier conditions influencing the fungal interactions with the roots of these plants (Muthukumar et al., 2004; Wolfe et al., 2007). DSE fungi are commonly found in the roots of vascular plants growing in arctic and alpine areas (Day and Currah, 2011), and have been reported to function in a similar way to mycorrhizal fungi, by supporting plant growth and facilitating nutrient uptake (Jumpponen and Trappe, 1998). However, only a few studies recorded their presence in pristine peatlands and it remains unclear whether they are or not a typical and relatively abundant fungal guild in these ecosystems. For example, DSE fungal diversity in Picea abies in undisturbed and drained peatlands and a mineral soil was equally low (Terhonen et al., 2014). Similarly, DSE were present but rare in subarctic peatlands dominated by heather (Olsrud and Michelsen, 2009; Kjøller et al., 2010). DSE were also found in various sedge species of Canadian peatlands (Thormann et al., 1999). The roots of purple-moor grass (Molinia caerulea) showed a high DSE colonization in a fen and a bog. Also, other peatland plants like Carex sp., Serratula tinctoria and Betonica officinalis appeared to be colonized by this fungal guild in the same peatlands, although in a lower percentage than Molinia caerulea (Fuchs and Haselwandter, 2004). Furthermore, DSE fungi were found in a fen in the roots of the species Lysimachia nummularia, Mentha longifolia and Ranunculus repens, though colonization percentage was rather low (Kołaczek et al., 2013). Importantly, the influence of DSE fungi on plant hosts is inconsistent,

with some studies reporting positive effects and others negative and neutral impacts on hosts (i.e. Newsham, 2011; Mayerhofer et al., 2013). Hence, it is necessary to better understand the relationships between different environments, host species and DSE fungi (Terhonen et al., 2014). Yeasts are documented to represent about 10% of the fungal species known in peatlands (Thormann et al., 2007). However, the understanding of global diversity of this group may be incomplete since studies are mainly based on cultured yeasts from peatlands located in Sweden, Canada, Russia and Alaska (Nilsson et al., 1992; Polyakova et al., 2001; Thormann et al., 2007). Moreover, some of the commonly applied culture media are not suitable to grow yeasts, which may lead to under-representations of this group (Thormann et al., 2007). Nevertheless, yeasts have also been found in peatlands in several DNAbased studies. For instance, new yeast species of the genus Candida were reported in Russian bogs (Polyakova and Chernov, 2002; Kachalkin and Yurkov, 2012). Also, two recent studies reported yeasts in peat swamp forests in Thailand through molecular methods (Jaiboon et al., 2016; Polburee et al., 2017). Pathogens are considered a rare fungal guild in peatlands, but the very few studies available have observed interesting patterns. Some pathogenic fungi are restricted to specific Sphagnum mosses (Thormann and Rice, 2007) and plant pathogenic fungal richness is higher in moss carpets of Polytrichum spp. than under coniferous trees growing in peatlands (Hiiesalu et al., 2017). Pathogenic oomycetes, protists with fungal-like lifestyle, were present in Sphagnum-dominated peatlands as well (Singer et al., 2016). Some Sphagnum mosses may host highly specific pathogenic fungal communities driven by refined evolutionary mechanisms between hosts and pathogens, like in the case of the moss Bryophytomyces sphagni, which colonizes spore capsules of Sphagnum and uses them for its own ballistic dispersal (Kostka et al., 2016). In summary, the presence of DSE fungi, yeasts, and pathogens has been documented in various peatlands worldwide, but their abundance and diversity patterns require further studies. 3.3. Influence of peatland type Northern peatlands can be classified as bogs and fens on the basis of pH. Bogs are ombrotrophic and have water pH values below 5.5, whereas fens are minerotrophic and have higher pH values (Wheeler and Proctor, 2000). Bogs are rain-fed, receiving water only in the form of precipitation, whereas fens are mainly fed by groundwater. However, a holistic classification of peatlands takes also into account vegetation growing above-ground (Wheeler and Proctor, 2000; Hájek et al., 2006). Concurrently, microbial community structure differs among pristine fens and bogs (e.g. Jaatinen et al., 2007; Lin et al., 2012; Mpamah et al., 2017). Diversity of culturable fungi isolated from dominant peatland plants was found to differ between fens and bogs (Thormann et al., 2004a; Fig. 1). Thus, addressing a single plant species for fungal surveys may have benefits, such as quick species identification and cost-effectiveness, but this approach may ignore specific fungal diversity associating with other plants. Furthermore, fungal and plant communities may be interrelated as shown in other terrestrial ecosystems (Wardle, 2006; Peay et al., 2013; Hiiesalu et al., 2014; Chen et al., 2018; Neuenkamp et al., 2018), and this is possibly the case for peatlands as well. In relation to this, host plant identity and abundance has been related to richness and abundance of ErM and EcM fungi in bog and fen habitats of a peatland forest (Kennedy et al., 2018). 3.4. Influence of the abiotic environment Relationships between plant and fungal communities are not straightforward and depend on the different ecological niches present in peatlands (Kennedy et al., 2018; Fig. 1). Availability of nutrient resources (Waldrop et al., 2006), soil temperature (Andersen et al., 2013) and soil moisture (Peay et al., 2016) are the most important abiotic 4

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vascular plants able to form mycorrhizal associations in the nutrientpoor fen (Myers et al., 2012). Anoxia is another important abiotic feature of peatlands. Some fungal species are not able to cope with anoxic conditions (Mäkiranta et al., 2009; Lin et al., 2012; Mpamah et al., 2017), therefore low fungal richness can be observed in permanently waterlogged peatlands (Hiiesalu et al., 2017). However, the water table level varies in peatlands, with different amplitude and duration of fluctuations that affects the fungal biodiversity they host (Wheeler and Proctor, 2000; Andersen et al., 2013). Climate change projections indicate that water table levels will decrease in peatlands, bringing more persistent aerobic conditions (Gorham, 1991). Kaisermann et al. (2015) found that fungal communities showed a higher sensitiveness to changes in soil moisture than bacteria. However, effective redistribution of water caused by mycelial network allows fungi to develop under low moisture conditions (Guhr et al., 2015). Drier conditions brought by non-persistent lowering of water table level led to a higher fungal diversity in Finnish peatlands (Peltoniemi et al., 2009, 2012). In relation to this, decreasing the water table level under a tipping point of −24 cm increased fungal richness in a peatland manipulation experiment in Poland (Jassey et al., 2018). In addition, a persistent lowering of water table level increased the activities of carbon-degrading enzymes (possibly from saprotrophic fungi) in bogs than in fens (Straková et al., 2011). In agreement with these results, culturable fungi producing phenol oxidase (an enzyme involved in the polymerization and depolymerization of simple and complex phenolic compounds; Fenner et al., 2005) showed higher enzymatic activity in a bog than in a fen system (Jassey et al., 2011). Controlled warming experiments have shown an increase in DSE fungi colonization in hummocks but not in hollows (Binet et al., 2017), which emphasizes the need for studying changes in abiotic factors at a microhabitat scale in peatlands (Asemaninejad et al., 2017b). However, other studies addressing culturable fungi did not show a common optimal temperature that increased fungal growth rates: some species had faster growing rates at 20 °C, whereas others were indifferent to temperature decreases (Thormann et al., 2004b). Similarly, Peltoniemi et al. (2015) did not find any significant effect of increasing temperatures on fungal growth in fens. However, the two last aforementioned studies did not account for the entire fungal community. The former study used culturing-approaches that may have led to under-estimation of fungal diversity, whereas the latter study applied 14C-acetate in ergosterol techniques that may have not measured growth of mycorrhizal fungi (Bååth, 2001). Moreover, the combined effect of both soil temperature and water table level has been shown to fungal communities (Allison and Treseder, 2008; Strickland and Rousk, 2010; Fig. 1). Peltoniemi et al. (2015) examined the effect of the interaction of both abiotic factors in fungal communities and found that the composition of the fungal communities in boreal fens changed after warming, with some species appearing or disappearing depending on the water table levels.

Fig. 1. Factors influencing fungal diversity in pristine peatlands. This scheme illustrates that fungal diversity is primarily regulated by peatland type, and subsequently, by the abiotic factors characteristic of the particular peatland type and the spatial and temporal changes occurring within the ecosystem.

3.5. Influence of the temporal and spatial scales

factors influencing fungal diversity in terrestrial ecosystems (Fig. 1). Fungal diversity appeared to be higher in fens, which have high nutrient resource supply and pH values, than in low-nutrient bogs. Specifically, fungi that are abundant in relatively nutrient-rich environments, such as fens, have been associated with Carex species and Betula nana, whereas only a few different specialist fungi have been recorded in a bog dominated by Eriophorum vaginatum, Andromeda polifolia and Rubus chamaemorus (Peltoniemi et al., 2009). Similar results of a greater fungal diversity in a fen than in a bog were found in a litterbag experiment at the same study sites (Peltoniemi et al., 2012). Also, the percentage root length colonized by AM fungi was higher in a fen than in a bog, which may indicate that higher pH values benefit AM fungal colonization (Fuchs and Haselwandter, 2004). However, fungal activity was higher in poor fens with low values of pH and mineral carbonates than in rich fens, possibly due to the higher dominance of

The response of soil biodiversity to changes in abiotic factors follows intra- and inter-annual patterns (Bardgett, 2005; Bardgett and van der Putten, 2014; Fig. 1). Studying seasonal dynamics of soil organisms and their relationships with long-term changes in climatic conditions can help us to better understand the responses of microbial communities to future climate changes (Matulich et al., 2015). For example, in boreal and temperate forests, EcM fungi abundance usually peaks in late summer and autumn, whereas saprotrophic fungi abundance is highest in the winter (Davey et al., 2012; Voříšková et al., 2014; Santalahti et al., 2016). However, some individual taxa may not follow the general patterns of the entire fungal community, thereby showing different temporal dynamics (Voříšková et al., 2014; Peršoh et al., 2018). Changes in the fungal diversity of peatlands may occur 5

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Fig. 2. Relative contribution of the different fungal guilds inhabiting pristine peatlands to carbon cycling. Some guilds participate in both decomposition and nutrient uptake, whereas others show a single lifestyle.

throughout the year and between years, and can also be related to changes in abiotic factors (Andersen et al., 2013; Fig. 1). Nevertheless, seasonal dynamics of fungal communities have received little attention. Interestingly, Toberman et al. (2008) found that seasonality rather than summer drought manipulation was the main driver of changes in fungal diversity, resulting in lower values in the summer. Environmental filtering (i.e. when local abiotic conditions select and prevent the establishment and persistence of certain species) by soil nutrients and pH is the main process structuring fungal communities and drives spatial dynamics of fungi at fine scales (Glassman et al., 2017; Fig. 1). Accordingly, despite fungi being generally concentrated in the top 30 cm of the peat profile, a high horizontal variation of fungi in peatlands was attributed to changes in vegetation cover and the chemical content and structure of litter (Lin et al., 2014). Also, fungal communities associated to ericaceous plants followed a spatial structure at very short distances in a subarctic peatland (Kjøller et al., 2010). Vertical spatial partitioning of fungal guilds along the soil profile could overlap, and thereby mycorrhizal and saprotrophic fungi could compete for substrates of the same quality, but having different impacts on organic matter decomposition rates (Bödeker et al., 2016). Preston et al. (2012) found a similar composition of fungal communities at three different peat depths (0–10 cm, 20–30 cm and 30–40 cm) in two bogs and one fen. In other studies, yeasts dominated with increasing depth, whereas the white rot fungi were more abundant in the uppermost layers in both fens and bogs (Lin et al., 2014; Wang et al., 2019). However, saprotrophic fungi can dominate the upper part of the peat profile in a boreal peatland forest (Sun et al., 2016) and in boreal fens (Asemaninejad et al., 2017b, 2018; Wang et al., 2019), but the opposite trend was found in a mesocosm study carried out in a temperate bog (Lamit et al., 2017). As different depths were considered in these studies, it may be necessary to take into account sampling depth when studying the influence of abiotic factors on fungal diversity of peatlands (Peltoniemi et al., 2015).

capacity of tropical peatlands to store peat is thought to be higher than that of northern peatlands (Roulet, 2000; Page et al., 2011). Bogs and fens in the northern hemisphere are usually formed under cold temperatures and partially or permanently waterlogged conditions, vegetation is dominated by mosses or herbs, and, consequently, peat contains high amounts of non-woody dead plant tissues (Rydin and Jeglum, 2015). In contrast, in tropical peatlands (namely peat swamp forests) temperature is usually high, water table varies greatly during the whole year, trees dominate the vegetation, and peat is mainly formed by woody debris and lignin (Posa et al., 2011; Lawson et al., 2014). These differences should be also reflected in differences in the fungal diversity of peatlands in different climatic zones and areas of the world. For example, the fungal community of a near-pristine tropical peatland in Indonesia was highly resistant and able to survive low water levels in a drought incubation experiment (Kwon et al., 2013). Similar resistance to changing water table levels have been found in northern peatlands (i.e. Jaatinen et al., 2007; Jassey et al., 2018). However, there is insufficient data to conclude whether low- and high-latitude peatlands have similar fungal communities. Thus, more focus is needed to survey tropical peatland fungi, considering the specific features of tropical peatlands and aiming to map fungal diversity patterns across peatlands worldwide. 4. Contribution of fungi to carbon cycling in pristine peatlands Fungi have been functionally classified in a gradient of decomposition ability from those that are able to decompose easily degradable compounds to others that are recalcitrant polymer degraders (Thormann, 2006; Treseder and Lennon, 2015; Asemaninejad et al., 2017a; Fig. 2). However, fungi with different carbon degrading abilities are able to cohabit and some groups can degrade both labile and recalcitrant carbon (Treseder, 2016; de Vries and Caruso, 2016). In addition, different fungal groups can show a functional shift in terms of carbon decomposition in peatland ecosystems (Peltoniemi et al., 2012). Fungi are classified as heterotrophic contributors to soil respiration, but the fungal contribution to total soil respiration is unclear, since some mycorrhizal fungi can penetrate the root tissues, thereby producing CO2 inside the roots (Kuzyakov, 2006). Clarifying the functional role of fungi on soil respiration is crucial under climate change conditions, since autotrophic and heterotrophic contributions to CO2 emissions are thought to differ (Wang et al., 2014). In peatlands with low water table levels, root exudates and dissolved organic matter quality are important energy sources stimulating fungal enzymatic activities,

3.6. Tropical vs. temperate peatlands Fungal communities of boreal and temperate peatlands have been studied more thoroughly compared to peatlands in the tropics and the southern hemisphere (Posa et al., 2011). Total tropical peatland area is much smaller (i.e. 11% of the global peatland area) than that of boreal and temperate peatlands (Page et al., 2011). However, the tropical pristine peatland area in South America and Africa is still estimated to be larger than previously thought, with over 1,000,000 km2 occupied by peat-accumulating ecosystems (Gumbricht et al., 2017). Also, the 6

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thereby increasing soil respiration (Jassey et al., 2018). Also, under aerobic conditions, fungi are linked to high CH4 production (Lenhart et al., 2012). Therefore, fungal contribution of soil respiratory losses might increase under oxic environments in peatlands. Linking fungal guilds to carbon-degrading enzymatic activities has proven to be useful in terrestrial ecosystems and -omics techniques could help to better predict the role of fungi in carbon cycling (Talbot et al., 2015). The application of metatranscriptomics and metagenomics has been effective for studying the metabolic potential activities of bacterial communities in Arctic peatlands, indicating a compositional and taxonomical shift in response to elevated temperature that resulted in large increases in CH4 and CO2 emissions (Tveit et al., 2015). However, to our knowledge, -omics techniques have not been applied to study possible shifts in fungal communities in peatlands. Distribution of functional guilds of fungi could provide important information about their respective roles in carbon cycling (Talbot et al., 2015; Schröter et al., 2019). Saprotrophic fungi, commonly considered as the dominant fungal guild in peatlands, are able to decompose both simple and complex organic compounds (Baldrian and Valášková, 2008; Baldrian, 2009; Fig. 2). Importantly, saprotrophs are able to degrade lignin and cellulose and play a crucial role in carbon cycling in Sphagnum peat (Rice et al., 2006). Thus, an increase of the abundance of saprotrophic fungi caused by dry conditions may enhance carbon-degrading enzymatic activities and hence, would drastically increase organic matter decomposition in peatlands (Jassey et al., 2018). In contrast, the role of mycorrhizal fungi in carbon cycling is still under debate (Talbot et al., 2008; Treseder, 2016; de Vries and Caruso, 2016). Categorizing mycorrhizal fungi according to mycorrhizal type is useful in order to obtain more information about the role of the mycorrhizal symbiosis in nutrient cycling (Read, 1991). In this context, ErM and EcM fungi are considered to be organic matter decomposers, whereas AM fungi have been usually assumed not to be very important agents in this process (Read and Perez-Moreno, 2003; Fig. 2). ErM and EcM fungi can act as versatile saprotrophs and biotrophs and therefore, they have a dual lifestyle that allows them to decompose recalcitrant organic matter and actively facilitate the N and P exchange from the soil to the host plant (Read et al., 2004; Martino et al., 2018; Fig. 2). In carbon-rich forest ecosystems with trees forming associations with EcM fungi, the appearance of the fungi fruiting bodies in the autumn and a higher substrate availability promoted CO2 emissions (Heinemeyer et al., 2007, 2012). Additionally, EcM fungal mycelium played a substantial role in dissolved organic carbon release, as it served as substrate for other microorganisms (Högberg and Högberg, 2002). Therefore, peatlands with abundant tree coverage that are dominated by EcM fungal communities (Wurzburger et al., 2004; Hiiesalu et al., 2017; Kennedy et al., 2018) would increase carbon emissions. A common mycorrhizal fungal mycelial network of plant communities with ErM fungi has been observed to increase nutrient recycling in peatland ecosystems with low nutrient availability (Kjøller et al., 2010). Additionally, EcM fungi use extracellular enzymes to promote organic matter decomposition (Talbot et al., 2008), and EcM fungal species with cord forming mycelia (Agerer, 2006) are considered to be an important source of organic matter since they act as a supply of extensive mycelial residues (Högberg and Högberg, 2002; Clemmensen et al., 2013). Mycorrhizal fungal necromass in boreal ecosystems, mainly formed by senesced EcM and ErM fungi, is also an important carbon input that remains in the soil and eventually contributes to soil organic matter formation (Clemmensen et al., 2015). In peatlands, decomposition of fungal necromass may increase under climate change conditions (i.e. elevated temperatures and lowering of water tables) and hence, would decrease the size of their carbon stocks (Fernández et al., 2019). AM fungi contribute to the distribution of assimilated C from host plants to soil microbes (Bonfante and Anca, 2009; Kaiser et al., 2015), so they speed up metabolic activities of the microbial community

although they are not organic matter decomposers (Fig. 2). Even though AM fungi are a minor fungal guild in terms of abundance in peatlands, they are able to improve net primary productivity of plants growing in N- and P-limited ecosystems and to efficiently use photosynthates, thereby enhancing carbon sequestration (Orwin et al., 2011; Treseder, 2016). In addition, priming effects on C decomposition, occurring when high inputs of fresh organic carbon (plant or litter) enter the soil (Talbot et al., 2008; Trinder et al., 2009), should also be taken into account when considering AM fungi as intermediaries, because of their ability to rapidly assimilate labile carbon and to stabilize carbon through reduction of the abundance of other decomposers and decreasing respiratory losses (Verbruggen et al., 2013, 2016). Thus, the priming effect may be highly important in peatlands dominated by plant species producing great amounts of high quality litter, such as the purple-moor grass (Leroy et al., 2017). Only a few studies have investigated the role of DSE fungi, yeasts, and pathogens on organic matter decomposition. DSE fungi have been considered to be extracellular enzyme producers (Caldwell et al., 2000; Fig. 2), and Day and Currah (2011) demonstrated that DSE fungi produce protein-, cellulose-, starch- and soluble polyphenols- degrading enzymes in a temperate mixed forest. Interestingly, colonization by DSE and ErM fungi in the roots of ericaceous shrub Andromeda polifolia has been associated with phenolic production by Sphagnum mosses, which may promote shrub expansion, thereby resulting in potential negative effects on carbon storage in peatlands (Binet et al., 2017; Chiapusio et al., 2018). Regarding yeasts, they are thought to be involved in utilisation of simple carbon compounds (Thormann et al., 2007; Fig. 2) and are generally considered as secondary saprotrophs (Untereiner and Malloch, 2007). However, Jaiboon et al. (2016) reported in a controlled experiment that all 10 yeast species found in a peat swamp forest were able to produce the extracellular enzyme lipase. Data about pathogens is scarce and they are mainly considered to be poor competitors in soil organic matter decomposition (Thormann, 2006; Fig. 2). Henceforth, studies in natural field conditions about the role of DSE fungi, yeasts, and pathogens in peatlands are needed to clarify whether they can access simple or complex carbon compounds. 4.1. Fungal functional traits: an upcoming alternative A functional trait can be defined as a morphological, physiological or phenological characteristic of a living organism that influences its fitness indirectly (by survival, growth and reproduction) and its impact on the environment and on the ecosystem properties (Violle et al., 2007). Changes in the enzymatic activities that regulate soil nitrogen and soil carbon cycling have been successfully predicted by quantifying shifts in fungal traits within a community (Talbot et al., 2015; Kyaschenko et al., 2017), demonstrating that they may help to explain organic matter dynamics better than plant traits (Kyaschenko et al., 2017). In the recent years, a functional trait-based approach for fungi has been used for understanding life histories of mycorrhizal fungi (Chagnon et al., 2013), the role of fungi as a whole at a community level (Crowther et al., 2014), soil aggregation (Rillig et al., 2015) and nitrogen deposition (Treseder et al., 2018). Nevertheless, a deeper and more precise understanding of functional traits of fungi rather than taxonomic classifications, as suggested by de Vries and Caruso (2016), is needed to better understand co-existence patterns of fungi and the role that the fungal community, as a whole, plays in carbon cycling. Suitable fungal functional traits to measure organic matter decomposition may include genetic capacities of fungi to produce enzymes such as lignin peroxidase, β-glucosidase, cellobiohydrolase and lytic polysaccharide monooxygenase (Treseder and Lennon, 2015). Additionally, melanin production has also been seen as an efficient fungal functional trait, since it resists decomposition and it positively correlates to carbon storage in soils (Treseder and Lennon, 2015; Siletti et al., 7

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2017). The next task will be to clarify how these fungal functional traits are helpful in explaining the contribution of fungal communities to carbon cycling in ecosystems with very high organic carbon contents such as pristine peatlands.

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5. Conclusions and challenges This review highlights the huge progress of molecular techniques on understanding the fungal diversity in pristine peatlands over the last few years. Importantly, using a combination of chemical marker-based approaches (fatty-acid, ergosterol and chitin), culturing approaches, fruiting body collection and molecular techniques could give us more precise information about temporal and spatial dynamics of fungal communities, their responses to abiotic factors and the implications on organic matter decomposition in different peatland types. However, global diversity and distribution of fungi across pristine peatlands remains currently unknown and there is a particular shortage of studies from tropical latitudes. Also, DSE fungi, yeasts, and pathogens are little studied fungal guilds, although relatively abundant in peatlands, and deserve more attention in order to clarify their ecological roles in these ecosystems. Categorizing fungal diversity in peatlands worldwide is an important issue that will help us to preserve soil biodiversity and thereby the high carbon storing capacity of these ecosystems. Field experiments and in situ approaches could improve our understanding of the functional roles of fungi under more realistic conditions (Lekberg and Helgason, 2018). A long-term perspective is needed to assess the possible shifts in fungal and plant communities under projected elevated temperatures and lower water table level conditions in peatlands. Additionally, studying the impacts of fungi on labile and recalcitrant carbon pools is challenging but necessary to predict changes in organic matter decomposition rates. In the view of the increasing use of functional traits to study effects and responses to changes in biotic and environmental conditions in terrestrial ecosystems, fungal functional traits are the most suitable tool to link fungal guilds to carbon cycling processes (Talbot et al., 2015; Treseder and Lennon, 2015) and their application in peatlands needs further urgent research. Declaration of Competing Interest None. Acknowledgements We thank D. García de León for comments on an earlier draft of the manuscript and R.R. Granjel for helping with generating ideas for figure design. We are also grateful to two anonymous reviewers who improved this manuscript with their comments. This study received financial support from a MINECO research project (Ref. CGL2014-54861-R) and R.J.-O. was supported by a PhD fellowship (FPI Programme, Ref. BES2015-074461). M.Ö. is funded by the Estonian Research Council (grant IUT20-28), the European Regional Development Fund (Centre of Excellence EcolChange) and ERA-NET Cofund BiodivERsA3 (SoilMan). References Agerer, R., 2006. Fungal relationships and structural identity of their ectomycorrhizae. Mycol. Prog. 5, 67–107. https://doi.org/10.1007/s11557-006-0505-x. Allen, T.R., Millar, T., Berch, S.M., Berbee, M.L., 2003. Culturing and direct DNA extraction find different fungi from the same ericoid mycorrhizal roots. New Phytol. 160, 255–272. https://doi.org/10.1046/j.1469-8137.2003.00885.x. Allison, S.D., Treseder, K.K., 2008. Warming and drying suppress microbial activity and carbon cycling in boreal forest soils. Glob. Change Biol. 14, 2898–2909. https://doi. org/10.1111/j.1365-2486.2008.01716.x. Andersen, R., Chapman, S.J., Artz, R.R.E., 2013. Microbial communities in natural and disturbed peatlands: a review. Soil Biol. Biochem. 57, 979–994. https://doi.org/10. 1016/j.soilbio.2012.10.003. Artz, R.R.E., Anderson, I.C., Chapman, S.J., Hagn, A., Schloter, M., Potts, J.M., Campbell,

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