Filamentous fungi are critical to the decomposition of terrestrial organic matter and, consequently, in the global carbon cycle. In particular, their contribution to degradation of recalcitrant lignocellulose complexes has been widely studied. In this review, we focus on the functioning of terrestrial fungal decomposers and examine the factors that affect their activities and community dynamics. In relation to this, impacts of global warming and increased N deposition are discussed. We also address the contribution of fungal decomposer studies to the development of general community ecological concepts such as diversity–functioning relationships, succession, priority effects and home–field advantage. Finally, we indicate several research directions that will lead to a more complete understanding of the ecological roles of terrestrial decomposer fungi such as their importance in turnover of rhizodeposits, the consequences of interactions with other organisms and niche differentiation.
The kingdom Fungi is a monophyletic eukaryotic lineage consisting of chemo-organotrophic organisms with two distinct growth forms: spherical cells (yeasts) and thread-like structures called hyphae (filamentous fungi). The hyphal growth form is of particular importance in terrestrial ecosystems as it enables exploration of soils via bridging of air-filled gaps (pores) and penetration of solid material (Hoffland et al ., 2004 ; Klein & Paschke, 2004 ; Money, 2007 ; Wurzbacher et al ., 2010 ). In addition, hyphae have the ability to translocate nutrients across nutrient-poor patches and to supply growth-limiting elements to zones of metabolic activity (Frey et al ., 2000 ). Fungi have, therefore, been characterized as spatial integrators (Ritz, 2007 ). The mycelial growth form also facilitates biomass recycling, which further increases efficiency in nutrient use in patchy environments (Boddy, 1999 ; Falconer et al ., 2007 ). Due to the success of the hyphal growth form in terrestrial environments, fungi have become important components of terrestrial ecosystem functioning (De Boer et al ., 2005 ), especially with respect to the decomposition of organic matter. Decay of organic matter controls the balance between soil carbon storage and CO 2 release into the atmosphere, and releases mineral nutrients, which are again made available for plant growth. In this review, we will focus on communities of fungi that play a critical role in decomposition processes. Although the link between fungal ecology and carbon cycling is generally acknowledged, the dynamics and interactions of fungal species during decomposition processes are still not fully understood. Topics that have received increasing attention during the last decade are fungal niche differentiation, the relationship between fungal diversity and decomposition, the role of decomposer fungi in the rhizosphere, the impact of climate changes on functioning of fungal communities, incorporation of fungal factors in decomposition models, effects of fungal species on fungal community composition (priority effects) and the selection of a fungal community composition that is specialized in decomposing the litter of the local plant species or vegetation (home-field advantage). Developments in these research topics in relation to decomposition processes will be presented and discussed.
Phylogenetic diversity of saprotrophic fungi
Fungi exhibit a large diversity of lifestyles, with the saprotrophic lifestyle being very likely the original condition. Mutualistic lifestyles (lichens, mycorrhizal fungi) evolved from saprotrophic fungi, and reversal back to saprotrophy from the mycorrhizal life style has not been observed (James et al ., 2006 ; McLaughlin et al ., 2009 ). Four phyla within the kingdom Fungi contain saprotrophic fungi. The polyphyletic phylum Zygomycota contains somewhat over 1000 described species (Kirk et al ., 2008 ). The subphylum Mucoromycotina (Hibbett et al ., 2007 ) contains about 300 described species, many of which are opportunistic saprotrophs (sugar fungi; White et al ., 2006 ; Kirk et al ., 2008 ). The phylum Ascomycota is the largest phylum with about 64 000 described species (Kirk et al ., 2008 ). The saprotrophic capabilities of ascomycetes range from breakdown of simple sugars (sugar fungi) to degradation of the lignocellulose complex ( Xylariales ). There are about 32 000 described species belonging to the phylum Basidiomycota (Kirk et al ., 2008 ). The saprotrophic basidiomycetes occur in the subphylum Agaricomycotina . Until recently, it was believed that the basal Chytridiomycota were mainly aquatic and had virtually no importance for terrestrial ecosystem functioning. However, saprotrophic chytrids can dominate fungal communities in nonvegetated, high-elevation soils (Freeman et al ., 2009 ), and their importance may also have been underestimated for other soils (Gleason et al ., 2012 ).
Fungi and the decomposition processes
Fungi make a major contribution to terrestrial organic matter decomposition, in particular of the more recalcitrant fractions (Dighton, 2003 ; De Boer et al ., 2006 ; Berg & McClaugherty, 2008 ). The ability to decompose these recalcitrant fractions of terrestrial organic matter is based on a combination of morphological characteristics (hyphal growth form) allowing penetration of solid material, and physiological characteristics (extracellular enzymes) allowing degradation of the lignocellulose complex (Money, 2007 ; Baldrian & Valášková, 2008 ; Floudas et al ., 2012 ). In particular, the ability to decompose lignin, a heterologous aromatic polymer, appears to be mainly restricted to Basidiomycota ( Agaricomycotina ) that are known as white-rot fungi (Baldrian, 2008 ; Floudas et al ., 2012 ), although lignin breakdown has been reported for the Xylariales, within the Ascomycota (Worrall et al ., 1997 ; Osono et al ., 2011a , b ). In addition, so-called brown-rot fungi have the ability to modify lignin, thereby gaining access to cellulose which together with hemi-cellulose forms the major energy resource for litter- and wood-degrading fungi (Curling et al ., 2002 ; Yelle et al ., 2008 ; Eastwood et al ., 2011 ; Martinez et al ., 2011 ). Cellulolytic ascomycetes, even though they are not able to degrade or modify lignin like white- and brown-rot fungi, can contribute significantly to the decomposition of lignin-rich organic matter, as thin perforation hyphae of these fungi can reach cellulose-rich layers in woody cell walls (Schmidt, 2006 ). Modification of lignin, as well as decomposition of lignin derivatives, has also been reported for bacteria (Bugg et al ., 2011 ). Such processes may be important for lignin degradation in environments where growth of fungi is restricted, for example, by periodic anoxic conditions (DeAngelis et al ., 2011 ). However, the direct contribution of bacteria to decomposition of natural lignocellulose complexes in terrestrial ecosystems, that is, the attack of these complexes by bacterial enzymes, appears to be minor (Kirby, 2005 ; Floudas et al ., 2012 ; Schneider et al ., 2012 ). However, bacteria may have an important indirect impact on the decomposition of lignocellulose-rich organic material and formation of humus, namely via metabolizing intermediates released by fungal enzymes or via other interactions with fungi (De Boer & Van der Wal, 2008 ).
Although the decomposition of lignocellulose-containing organic material can be considered to be mainly a fungal niche, decomposer fungi are not restricted to these resources (De Boer et al ., 2006 ; Van der Wal et al ., 2006a ; De Graaff, 2010 ). A wide range of organic compounds can be decomposed by more or less specialized fungi (Baldrian et al ., 2011 ). Many of these fungi, which are also called molds, are opportunistic, fast-growing fungi that can produce huge numbers of spores or melanized resting structures of hyphae to survive conditions when no easily degradable carbon sources are available (Van der Wal et al ., 2009a ). As these organisms degrade the same range of labile compounds as bacteria, they have to compete with bacteria for these resources (De Graaff, 2010 ). Environmental conditions, for example, pH and soil moisture, appear to play a major role in determining the relative importance of fungi and bacteria in the decomposition of easily degradable compounds (De Boer et al ., 2006 ). In general, fungi perform better at lower pH and relatively dry conditions (Bapiri et al ., 2010 ; Rousk et al ., 2010a ; Yuste et al ., 2011 ).
Yeasts, unicellular fungi, are well adapted to grow in environments where tolerance against high concentrations of sugars is required, for example, floral nectar and rotting fruits (Gasch, 2007 ; Tekolo et al ., 2010 ). In addition, they can grow anaerobically, using fermentation as the main process of energy generation. Anaerobic decomposition is not very common among fungi but can, for instance, be found for chytrids in cattle rumen (Trinci et al ., 1994 ). Some soil yeasts appear to be specialized in the decomposition of small aromatic compounds, but their role in soil organic matter decomposition remains unclear (Botha, 2011 ).
Like saprotrophs, other important functional groups of fungi such as mycorrhiza-formers, plant- and animal pathogens, endophytes and mycoparasites also obtain their energy by metabolizing organic compounds, that is, they are all chemo-organotrophs. However, as these groups of fungi rely mainly on energy resources from living organisms, they will not be discussed further in this review. It should, however, be realized that there can be considerable overlap between these fungal functional groups. For instance, fungal endophytes can contribute significantly to fungal decomposition processes (Müller et al ., 2001 ; Read & Perez-Moreno, 2003 ; Osono, 2006 ; Purahong & Hyde, 2011 ). Many fungal pathogens of trees use the same mechanisms as genuine saprotrophs to decay wood of living trees, and they often continue their decaying activities, that is, they become real saprotrophs, after the tree has been killed (Schwarze et al ., 2000 ). Fungal pathogens may also indirectly contribute to decomposition by providing material (remainders of killed hosts) for decomposer organisms.
Significant contributions of mycorrhizal fungi, in particular ectomycorrhizal and ericoid mycorrhizal fungi, to decomposition of soil organic matter have been suggested. These fungi provide trees or dwarf shrubs with organically bound N that is released by their enzymatic activities (Read & Perez-Moreno, 2003 ; Lindahl et al ., 2007 ; Van der Wal et al ., 2009b ; Courty et al ., 2010 ). However, their actual contribution to decomposition of soil organic matter is still a matter of debate (Baldrian, 2009 ; Bödeker et al ., 2009 ; Cullings & Courty, 2009 ; Courty et al ., 2010 ). A possible role for ectomycorrhizal fungi in the transformation of stable organic matter and specific mining of nitrogen from this stable humus (Lindahl et al ., 2007 ; Bödeker et al ., 2009 ) deserves more attention.
Spatial distribution of decomposer fungi
Fungal biomass can be estimated via a number of approaches, ranging from microscopical, biochemical, physiological to molecular biological methods (Joergensen & Wichern, 2008 ; Strickland & Rousk, 2010 ; Prévost-Bouré et al ., 2011 ; Baldrian et al ., 2012 ). However, while microscopical and biochemical methods can separate the Glomeromycota from most other fungi (presence or absence of septa, ergosterol, specific phospholipid fatty acids), they cannot distinguish decomposers from ectomycorrhizal and ericoid mycorrhizal fungi. Most fungal biomass measurements therefore give only a first indication of total decomposer fungal density (Strickland & Rousk, 2010 ). These ‘detection’ measurements can be complemented with ‘activity’ measurements, such as the incorporation of 13 C from labeled substrates (SIP) into fungal biomarkers (e.g. ergosterol, specific phospholipid fatty acids, DNA fragments; Malosso et al ., 2004 ; Moore-Kucera & Dick, 2008 ; Drigo et al ., 2010 ; Hannula et al ., 2012a ).
The general picture arising from these measurements is that fungal decomposers are abundant in many terrestrial ecosystems, but their biomass may be poorly represented in freshwater and marine environments, including sediments (Jørgensen & Stepanauskas, 2009 ). The advantages of the hyphal growth form over the unicellular growth form in terrestrial ecosystems is not valid for aquatic ecosystems, as there is less spatial heterogeneity and no air-filled pores to cross (Wurzbacher et al ., 2010 ). In addition, large parts of aquatic ecosystems, in particular sediments, are anaerobic, where the contribution of fungi is restricted to fermentation processes. Especially, the degradation of lignin is very sensitive to oxygen limitation (Ten Have & Teunissen, 2001 ). Furthermore, as plants in an aquatic habitat do not need to generate the same physical support structure, that is, lignocellulose complexes, as plants in a terrestrial habitat, aquatic organic matter is much less recalcitrant for bacterial decomposers than terrestrial organic matter (Hedges & Oades, 1997 ). Interestingly, significant fungal decomposition in aquatic and marine ecosystems can be found where lignocellulose-containing material of terrestrial plants enter these ecosystems for example, leaves and wood in ponds and rivers and in gradient ecosystems, like mangroves (Newell, 1996 ; Hieber & Gessner, 2002 ; Das et al ., 2007 ; Shearer et al ., 2007 ; Gulis et al ., 2008 ; Jobard et al ., 2010 ; Krauss et al ., 2011 ). Even in open oceans, specialized wood-decomposing fungi have been identified on floating tree trunks (Shearer et al ., 2007 ). Aquatic fungi may also be important in the degradation of dissolved aromatic organic compounds of terrestrial origin (Jørgensen & Stepanauskas, 2009 ; Krauss et al ., 2011 ). In addition to the fungal decomposition of terrestrial organic matter in aquatic ecosystems, parasitic fungi may have a major impact on aquatic carbon cycling by supplying bacterial decomposers with remainders of killed phytoplankton (Ibelings et al ., 2004 ; Jobard et al ., 2010 ; Wurzbacher et al ., 2010 ; Rasconi et al ., 2011 ).
In terrestrial ecosystems, fungal biomass is high in habitats that are rich in recalcitrant organic material, for example, forests and heathlands (Frostegård & Bååth, 1996 ; Hättenschwiler et al ., 2005 ; Brant et al ., 2006 ; Osono, 2007 ; Fierer et al ., 2009 ). This is in line with studies showing prevalent stimulation of fungal decomposer activities in soils by adding recalcitrant organic compounds or lignocellulose-rich plant materials (Paterson et al ., 2008 ; Rousk et al ., 2010a ). Several studies have also indicated an inverse relationship between fungal biomass and pH (Bååth & Anderson, 2003 ; Fierer et al ., 2009 ). However, this inverse relationship may be due to covariation of organic matter accumulation and pH. In addition, other fungal functional groups, in particular ericoid mycorrhizal and ectomycorrhizal fungi, contribute strongly to fungal biomass in many acid soils (Strickland & Rousk, 2010 ). Nevertheless, a strong inverse relationship between fungal decomposer activities and pH has been observed under conditions that remove potential confounding effects of organic matter variation and mycorrhizal fungi (Rousk et al ., 2010a , 2011 ). Intensive agricultural management (fertilization, pesticides and tillage) generally has a negative effect on fungal biomass (Strickland & Rousk, 2010 and references therein) and may promote opportunistic decomposer fungi (high growth rate, rapid sporulation) rather than lignocellulolytic fungi, which invest heavily in hyphal networks that are vulnerable to mechanical disruption (Stromberger, 2005 ; Van der Wal et al ., 2006a ). Yet, fungal species richness can still be high in such ecosystems (Hannula et al ., 2010 ; Klaubauf et al ., 2010 ; Xu et al ., 2012 ). The relative importance of fungi in decomposition processes in agricultural soils appears to be influenced by many factors such as type of crop and rotation frequency, crop age, intensity of tillage/fertilization and soil organic matter content (Dick, 1992 ; Stahl et al ., 1999 ; De Vries et al ., 2006 ; Hannula et al ., 2012b ). For example, organic matter content appeared to be much more important than the termination of tillage practices with respect to the recovery of soil-borne fungal biomass (Van der Wal et al ., 2006a , b ). Also, comparisons of no- or reduced-till with conventional tillage practices did not always reveal a negative effect of tillage on fungal biomass (Bailey et al ., 2002 ; Van Groenigen et al ., 2010 ).
Decomposer fungi are not distributed uniformly throughout the soil. In temperate and boreal forest soils with a well-developed organic layer, saprotrophic fungi are most abundant in the upper layers (L and F), where decomposition rates are by far the highest (Cairney, 2005 ; Lindahl et al ., 2007 ; Osono, 2007 ). In addition, high densities of fungi can be present in organic-rich patches in mineral soil (Ritz, 2007 ). Decomposer fungi have developed different strategies to gain access to new patches for example, by forming cords of hyphae that explore the environment (Boddy et al ., 2009 ; Garbeva et al ., 2011 ). Decomposer fungi can also be abundant in the rhizosphere, the zone surrounding roots where microbial activity is influenced by input of root-derived compounds. This has long been considered as a bacteria-dominated habitat because of the rapid growth of bacteria on soluble root exudates (Buée et al ., 2009a ). The high abundance of decomposer fungi in the rhizosphere of natural vegetation may be due to the presence of roots of different ages, as more recalcitrant compounds are released from senescent roots (Hegde & Fletcher, 1996 ). Indeed, a recent study showed that decomposer fungi strongly increase in abundance during flowering and maturation of potato plants (Hannula et al ., 2010 ). Yet, studies using 13 C labeled plants indicated that decomposer fungi may also contribute strongly to the rapid decomposition of simple root exudates (Butler et al ., 2003 ; Treonis et al ., 2004 ; Denef et al ., 2009 ; De Deyn et al ., 2011 ; Hannula et al ., 2012a ). As some saprotrophic fungal species have been reported to penetrate the exterior parts of roots, part of the rapid uptake of labeled carbon by fungi may be directly derived from inside the root (Harman et al ., 2004 ; Vasiliauskas et al ., 2007 ). There may be differentiation among fungal decomposer species in using different carbon resources in the rhizosphere, but this is a research area that needs further exploration (Broeckling et al ., 2008 ; Buée et al ., 2009a ; Paterson et al ., 2009 ; De Graaff, 2010 ; Becklin et al ., 2012 ).
Community level aspects of fungal decomposition
Diversity and decomposition
Natural communities of decomposer fungi usually consist of several or many species. Almost pure monocultures of decomposer fungi typically are only observed for those fungi that enter mutualistic relationships with insects and where the insects grow these fungi for food (e.g. fungus-growing termites, ants, beetles). In such cases of agriculture (or fungi culture) frequency-dependent selection and weeding by the insects cause the gradual loss of species and genotypes (Currie & Stuart, 2001 ; Aanen et al ., 2009 ). Decomposition rates by such monocultures are mainly determined by the metabolic properties of the fungal species, substrate quality and abiotic factors like moisture and temperature. The same factors are important in habitats containing multiple decomposer species, but, in addition, interactions between these species also have an impact on the decomposition process. The type of interactions will largely determine the relationship between fungal diversity and decomposition rates. It has been a standard view of ecology that increased microbial diversity will result in enhanced nutrient cycling because of functional niche complementarity or greater intensity of resource exploitation (Loreau et al ., 2001 ; Hättenschwiler et al ., 2011 ). Higher species richness could result in enhancement of decomposition via additive or synergistic activities for example, different fungal species decomposing different fractions of the substrate without (additive) or with (synergistic) a positive effect on decomposition activities of key species (Hättenschwiler et al ., 2011 ). Indeed, studies on monocultures of decomposer fungi have shown differences between species with respect to their ability to decompose different fractions of wood and litter (Boddy, 2001 ; Cox et al ., 2001 ; Deacon et al ., 2006 ; Osono et al ., 2008 ; Boberg et al ., 2011 ; Fukasawa et al ., 2011 ). In addition, substrate-related niche differentiation (resource partitioning) among soil decomposer fungi has been demonstrated in situ by showing that different substrates induced DNA-synthesizing activity in different fungal taxa (McGuire et al ., 2010 ). Hannula et al . ( 2012a ) also showed that, after pulse-labeling of potato plants, fungal species in the rhizosphere involved in the decomposition of simple exudates differed from those decomposing more recalcitrant compounds.
Positive effects of mixing saprotrophic fungal species on decomposition have been reported (Deacon, 1985 ; Robinson et al ., 1993 ; Setälä & McLean, 2004 ; Treton et al ., 2004 ; Deacon et al ., 2006 ; Costantini & Rossi, 2010 ; LeBauer, 2010 ). Tiunov & Scheu ( 2005 ) reported a positive effect of combining cellulolytic fungi and sugar fungi on decomposition of cellulose, which was attributed to relief of catabolic repression of cellulase production by the consumption of the released sugars by the sugar fungi. In the same study, positive effects of combining species on soil organic matter decomposition were also found, but to a much lesser extent. With larger species numbers, effects became less clear and no consistent pattern emerged (Nielsen et al ., 2011 ). Setälä & McLean ( 2004 ) noted that the diversity–decomposition rate relationship saturated at rather low species levels, however, their best fit in the regression still allowed for the possibility of an increasing decomposition with higher species numbers. However, also other studies have found a rapid saturation of this relationship (Dang et al ., 2005 ).
In all cases where positive effects of fungal diversity on decomposition have been reported, communities were still relatively species-poor (up to 10 species) and no further increase in activity was seen when species richness was further increased (Gessner et al ., 2010 ; Nielsen et al ., 2011 ). Several explanations have been given for this rapid saturation of the diversity–decomposition relationship, including the occurrence of redundancy in metabolic abilities, limited possibilities for facilitation and resource partitioning, intensive competition for space, and interference with antagonistic interactions (Gessner et al ., 2010 ; Hättenschwiler et al ., 2011 ; Kuyper & Giller, 2011 ).
Negative biodiversity–function relationships may occur when competitive interactions between species within a community are stronger than effects of complementarity (Nielsen et al ., 2011 ). Cox et al . ( 2001 ) observed a decrease in litter decomposition when comparing naturally colonized (high diversity treatment) pine needles with those colonized by a single species. Negative diversity–decomposition rate relationships were also reported by Deacon et al . ( 2006 ) and Fukami et al . ( 2010 ). Most of the research on competitive interactions between saprotrophic fungi concerns wood-rot fungi. In her review, Boddy ( 2000 ) concluded that competition is the most common type of interaction between wood-rot fungi and that competition can take place at all stages of wood decomposition. This competition seems to be for space, that is, for occupation of woody surfaces. This capture and defend territorial strategy is clearly visible in larger woody units, for example, logs, where different decay columns can be recognized that are occupied by different individuals of the same or different species (Fig. 1 ). Hence, both intraspecific and interspecific competition can take place. Several mechanisms, including production of nonvolatile and volatile toxins, extracellular enzymes, mycoparasitism and hyphal interference, are used in fungal competition, and the result can be replacement of one fungus by another or deadlock, where the opponents are restricted to their own occupied territory and cannot invade that of the other (Holmer & Stenlid, 1997 ; Boddy, 2000 ; Baldrian, 2004 ; Peiris et al ., 2008 ; Woodward & Boddy, 2008 ). These competitive interactions are thought to incur metabolic costs, and consequently allowing less metabolic energy to be allocated to decomposition (Wells & Boddy, 2002 ). Competitive interactions between distinct groups of fungi can also influence the quality of humic residues after wood or litter decomposition (Fukasawa et al ., 2009 ; Song et al ., 2012 ). For example, while white-rot fungi degrade lignin, cellulose and hemicellulose, brown-rot fungi remove relatively little lignin. Consequently, the type of rot fungus influences the soil residues that remain.
The study of Toljander et al . ( 2006 ) revealed several interesting aspects with respect to community dynamics of wood-rot fungi. They made assemblages of wood-rot fungi of increasing species richness (up to 16) on wood chips and followed species dynamics, fungal biomass and wood decomposition under constant and fluctuating temperature regimes. The persistence of species was strikingly low (1 or 2 species), demonstrating combative exclusion of many species by the strongest competitors.
Given the rapid saturation or even absence of positive diversity–decomposition relationships for fungi, one can ask where and when diversity–functioning relationships will be important in natural environments. When looking at lignin-degrading or lignin-modifying fungi, the diversity can be quite low in ‘patches’ of wood or litter due to combative exclusion (Boddy, 2001 ; Hättenschwiler et al ., 2005 ; Zhang et al ., 2008 ; Kubartova et al ., 2009 ). Decay columns in large woody resource units (e.g. snags, boles, logs) are often occupied by single rot fungal species and the identity of these species strongly affects the rate of wood decomposition in the columns (Fig. 1 ; Boddy, 2001 ; Větrovský et al ., 2011 ). So far, it is not known to what extent these species-dependent decay rates contribute to variations in wood decomposition rates in large woody resource units within forest stands (Müller-Using & Bartsch, 2009 ; Woodall, 2010 ). At a larger scale, the presence of different woody resource units appears to correlate positively with the diversity of wood-decaying fungi (Heilmann-Clausen & Christensen, 2004 ; Hottola et al ., 2009 ; Bassler et al ., 2010 ). Is this increase in diversity also relevant for wood decomposition rates at the scale of forest ecosystems or does it reflects stochastic processes (including the higher chance for rare species to be successful in occupation of a particular woody resource)? Increased species diversity of saprotrophic fungi is not only due to the diversity of woody resource units, but also related to other properties of the forest patch such as size and connectivity (Edman et al ., 2004 ; Jönsson et al ., 2008 ). Such sites, therefore, likely differ in microclimatic factors as well, making it difficult to investigate the relationship between diversity and decomposition rate in such old-growth forests.
For soils, molecular biological techniques have shown that fungal diversity is already high at a small scale (several hundreds of operational taxonomic units per gram of soil), even for agricultural soils with a low fungal biomass (Buée et al ., 2009b ; Hannula et al ., 2010 ; Rousk et al ., 2010b ; Xu et al ., 2012 ). Given the high overlap in metabolic abilities of saprotrophic soil fungal species, it is not to be expected that moderate changes in diversity will have an impact on soil organic matter decomposition rates (Deacon et al ., 2006 ; Kuyper & Giller, 2011 ). It has been observed that shifts in fungal community structure do not necessarily influence decomposition rates. For instance, Höppener-Ogawa et al . ( 2009 ) showed that a shift in abundance of fungal species in grassland soil, which was caused by the introduction of mycophagous bacteria, did not affect the decomposition of added cellulose. Hence, this result is in line with the expected functional redundancy in species-rich communities.
However, strong impacts on soil organic matter decomposition by single species within species-rich decomposer communities do occur. Fairy ring fungi in natural grasslands are examples of such species (Griffith & Roderick, 2008 ). These basidiomycete fungi, belonging to genera such as Agaricus and Marasmius , are visible as ring-like structures (1–300 m diameter) in nutrient-poor grasslands formed by die-back and/or enhanced growth of grasses. Die-back is caused by the production of toxins (HCN in the case of Marasmius oreades , Blenis et al ., 2004 ) by the fungus, whereas enhanced growth is due to mineral nitrogen released by decomposition of soil organic matter by the fungus (Gramms et al ., 2005 ; Griffith & Roderick, 2008 ). Soil organic matter content has been found to be lower inside than outside the rings indicative of a strong decomposing activity, although such effects have not been consistently reported across studies (Edwards, 1988 ; Djajakirana & Joergensen, 1996 ; Gramms et al ., 2005 ). Strong decomposing activity is also supported by high lignocellulolytic enzyme activity within rings (Gramms et al ., 2005 ). Hence, despite the presence of many decomposer fungal species in grassland soils, fairy ring fungi can impact nutrient cycles in grasslands in ways that other species cannot. Presence or absence of active fairy ring fungi has, consequently, a strong impact on the spatial heterogeneity of decomposition processes. These fungi may therefore be considered ‘keystone’ species (Robinson et al ., 2005 ).
In summary, no (uniform) relationship between ecosystem functioning (organic matter decomposition) and fungal decomposer diversity has been demonstrated to date. The predominant type of interactions, the presence of species with extra-ordinary decomposition activities, and the composition of the organic resources, as well as the spatial scale at which decomposition is examined, can determine the nature of this relationship.
Succession, priority effects and home-field advantage
It is well known that fungal community composition changes with time during the decomposition of complex organic matter such as litter and wood (Rayner & Boddy, 1988 ; Frankland, 1998 ; Dighton, 2007 ; Osono, 2007 ; Lindahl & Boberg, 2008 ). This has been coined ‘substratum succession’ to distinguish it from ‘seral succession’ which refers to the occurrence of different fungi in different stages of vegetation succession in terrestrial ecosystems (Frankland, 1998 ; Osono & Trofymow, 2012 ). A major driver of fungal succession in litter is the chemical composition, in particular the content and chemical structure of lignin (Osono, 2007 ). Older studies based on the isolation of fungi suggested a succession from endophytes and primary saprotrophs, mostly ascomycetes, that first decompose sugars and the easily available cellulose fractions of litter to secondary decomposers, mostly basidiomycetes, that attack lignin (Frankland, 1998 ; Dighton, 2007 ). Similar shifts in fungal community composition have been observed during decay of wood (Rayner & Boddy, 1988 ; Olsson et al ., 2011 ). Succession apparently also occurs within the lignin-decomposer community, as ascomycetes with ligninolytic activity appear to dominate during the early stages of litter decomposition followed by ligninolytic basidiomycetes (Osono, 2007 ). However, the general view that lignin is not degraded during the early stages of decomposition (Berg & McClaugherty, 2008 ) has recently been questioned (Koide et al ., 2005 ; Osono et al ., 2009 ; Klotzbücher et al ., 2011 ). So far, molecular biological techniques have not been applied broadly to follow fungal community composition during different stages of decomposition. Poll et al . ( 2010 ) analyzed fungal 18S RNA genes in an agricultural soil in the close proximity of decomposing rye residues. They observed shifts from Mucoromycotina to Ascomycota, with very low frequencies of Basidiomycota . The low recovery of basidiomycete sequences may have been due to the quality of the substrate, the soil origin or the fact that samples were taken outside of litter patches, so-called detritusphere. In a recent study, rRNA was extracted from decaying wood logs ( Picea abies ) to determine the succession of active fungi (Rajala et al ., 2011 ). The results revealed a succession from soft-rot fungi via white- and brown-rot fungi to ectomycorrhizal fungi with progressing decay of logs.
The presence of specific fungi can have a strong impact on the fungal community composition during succession. Heilmann-Clausen & Boddy ( 2005 ) showed that the ability of fungal species associated with advanced decay to colonize partially decayed beech wood was highly dependent on the identity of the initially present fungal species. Similar observations were made by Fukami et al . ( 2010 ) and Dickie et al . ( 2012 ), who showed that pre-inoculation of a wood-rot fungus on wood disks had a strong impact on the composition of fungal communities established from secondary inoculation and responses of species differed for laboratory and field conditions.
These predecessor–successor relationships can be described as priority effects as the predecessor creates conditions that have different (positive or negative) effects on the colonization abilities of potential successor species (Fukami et al ., 2010 ). Experiments with wood-rot fungi mainly suggest negative effects of the predecessor, possibly caused by effective occupation of a territory and the production of toxic secondary metabolites (Woodward & Boddy, 2008 ). Consequently, the successor species that are able to establish tend to be least sensitive to these compounds. Positive effects of predecomposition by primary colonizers on subsequent decomposition by other species have also been shown (Cox et al ., 2001 ; Osono, 2003 ; Osono & Hirose, 2009 ; Oliver et al ., 2010 ). This is probably due to structural disintegration, for example, by partial attack of lignin, making certain fractions of the litter more easily accessible for other fungi.
Besides the apparent consistent temporal shifts in functional groups of fungi during decomposition, a certain degree of specialization of fungal communities toward the decomposition of different litter and wood types is also apparent (Osono, 2007 ). Plant species identity has been indicated as an important factor for the fungal decomposer community composition in both litter and wood (Kulhánková et al ., 2006 ; Kubartova et al ., 2009 ; McGuire et al ., 2010 ; Rajala et al ., 2010 ; Kebli et al ., 2011 ). This is not surprising considering that chemical composition of litter and wood, for example, the amount and structure of lignin, can strongly differ between plant species (Berg & McClaugherty, 2008 ). The selection of specific decomposers by certain litter types has been proposed to result in a so-called home-field advantage: the presence of the best decomposer organisms in soil for a certain litter type as a result of legacies of previously decomposed litter of the same type (Gholz et al ., 2000 ). Indeed, several experiments and observations support this hypothesis (Ayres et al ., 2009 ; Strickland et al ., 2009 ; Keiser et al ., 2011 ). However, other studies have failed to report such advantages or observed effects were limited to only recalcitrant litter types (Wallenstein et al ., 2010 ; Milcu & Manning, 2011 ; Osono et al ., 2011a , b ; St John et al ., 2011 ). These contrasting results indicate that decomposer community composition effects have to be considered in the context of many other factors, for example, abiotic environmental conditions and soil fauna, affecting decomposition rates. For instance, it has been shown in the Netherlands that liming coniferous forests results in a shift in the wood decomposer community, and that many of the species that are characteristic for conifer wood in the limed plots (where liming increased N availability as well), are those that normally occur on wood of deciduous trees (Veerkamp et al ., 1997 ). Recently, Freschet et al . ( 2012 ) proposed that the presence of contrasting qualities within the same litter matrix (the home–field) can lead to a continuum from positive to negative interaction between specific litters and decomposer communities. This so-called substrate quality- matrix quality interaction hypothesis predicts that home-field advantage effects are restricted to situations where the quality (e.g. lignin content) of a given litter type (substrate) is highly similar to the quality of the ecosystem litter layer (matrix).
So far, only limited attention has been given to the relationships between fungal traits or fungal species composition and home-field advantage phenomena. Development of decomposer community specialization was shown by Keiser et al . ( 2011 ), where repeated incubation of soil with the same litter resulted in increased decomposition of that litter. Such conditioning by either hardwood or grass litter resulted in an increase in basidiomycetes (especially Tremellales ) and a general decline of ascomycetes.
Inclusion of fungal community dynamics in decomposition models
The former sections indicate that composition of active decomposer fungi (e.g. presence of fairy rings) and interactions within fungal decomposer communities (e.g. competition between wood-rot fungi) can have impacts on rates of decomposition. Such community aspects of decomposition have not been included in models of organic matter decomposition that are, so far, mainly driven by organic matter quality characteristics and abiotic parameters (Fig. 2 ; Berg & McClaugherty, 2008 ). Moorhead & Sinsabaugh ( 2006 ) argued that not only environmental controls (litter composition, temperature etc.) but also microbial controls of litter decomposition should be included in decomposition models to better predict the impacts of environmental disturbance, for example, increased N input. They developed a guild-based decomposition model, where the guilds represent microbial groups involved in the different stages of decomposition. The activity of these guilds can be differentially influenced by environmental factors, for example, stimulation of activity of opportunists by N. However, the model of Moorhead & Sinsabaugh does not include effects of shifts of community composition within guilds and species interactions within and between guilds. McGuire & Treseder ( 2010 ) suggest that such community-related factors could help to fine-tune decomposition models.
As an example, we show possible influences of composition and interactions of fungi on wood decay rates (Fig. 3 ). Initially, a woody resource is rapidly colonized by opportunistic bacteria and fungi that grow on simple soluble substrates and easily accessible (hemi-) cellulose (De Boer et al ., 2005 ; Van der Wal et al ., 2007 ). The composition of initial colonizers is probably determined by random dispersal as was also indicated for fungal community assembly of senesced tree leaves (Feinstein & Blackwood, 2012 ). The next phase, that is, the colonization of wood by rot-causing basidiomycetes, can take up to 24 months (Nicholas & Crawford, 2003 ). A delay in colonization by wood-rotting basidiomycetes, also known as ‘lag time’ (Harmon et al ., 1986 ), could be due to antagonism expressed by bacteria and micro-fungi that are already present in the wood (De Boer & Van der Wal, 2008 ), resulting in decreased wood decay rates (indicated by ‘a’ in Fig. 3 ). The degree of antagonism against rot fungi may largely depend on the production of antibiotics (composition, amount) and could differ for different communities (identity of the initial colonizers and the rot fung; Greaves, 1971 ; Payne et al ., 2000 ; De Boer et al ., 2003 ). Several rot fungi may be present in adjacent decay columns within large wood samples (Fig. 1 ), which may result in further decreases in decay rates (indicated by ‘b’ in Fig. 3 ).
Little is known about the effects of these interactions on wood decay rates. It may be that their effect only significantly influences wood decay rates on small spatial scales (e.g. within a tree stump or in a forest plot), with only marginal effects on global decomposition rates. Another aspect of fungal community dynamics–functioning relationships that has not been included in models, so far, is the impact of grazing by invertebrates. Fungivory is common among soil invertebrates (Faber, 1991 ), and selective grazing of invertebrate species on specific decomposer fungal species has been demonstrated (Koukol et al ., 2009 ; Crowther et al ., 2011a ). Selective grazing can result in shifts in saprotrophic fungal community composition as well as in changes of decomposition-related enzyme activities and nutrient fluxes (Crowther et al ., 2011a , b ; Tordoff et al ., 2011 ). However, studies to date have mainly focused on single grazer species – fungal species effects. In real communities, multiple interactions of many grazer- and fungal species will take place. Knowledge on such complex dynamics is needed to evaluate the importance of invertebrate grazing on the dynamics and functioning of fungal communities.
Global change and the functioning of decomposer fungi
Decomposition of organic matter is affected by global change. Changes in the decomposition rates of soil organic matter have been observed due to rises in temperature (Davidson & Janssens, 2006 ; Conant et al ., 2008 ; Hartley & Ineson, 2008 ; Osono et al ., 2011a , b ), changes in concentrations of biogenic greenhouse gases (Carney et al ., 2007 ; Singh et al ., 2010 ), and upon nitrogen deposition (Carreiro et al ., 2000 ; Neff et al ., 2002 ; Janssens et al ., 2010 ). However, few studies to date have tried to disentangle the effects that global change scenarios will have on the functioning of fungal decomposer communities from those of total microbial communities. Klamer et al . ( 2002 ) used open-top chambers in a scrub-oak habitat and measured changes in the fungal density and fungal community composition (T-RFLP of ITS sequences) in the soil upon doubling the amount of atmospheric CO 2 for a period of 5 years. They found an increase in fungal biomass as well as a shift in community composition, although diversity was not affected. In a follow-up study, it was found that doubling of atmospheric CO 2 also led to an increased loss of soil carbon, which coincided with increases in the levels of phenol oxidase activity –an enzyme involved in lignin breakdown (Carney et al ., 2007 ). Similarly, a link between soil warming and changes in the functioning of fungal communities has been shown. In a short-term (14 months) forest soil warming experiment lignin-degrading activities increased together with an increase in fungal PLFAs (Feng et al ., 2008 ). A similar result was found in a meta-analysis of functioning of fungal assemblages along a climatic gradient. Fungi with ligninolytic activity in broad-leaved tree species in warmer climates showed the greatest abilities to cause leaf litter mass loss (Osono, 2011 ). Both increased atmospheric CO 2 and soil warming affect the flux and composition of plant-derived labile organic compounds (Lin et al ., 1999 ; Drigo et al ., 2010 ). Such compounds may be important as energy sources for the activity of lignin-degrading fungi (Klotzbücher et al ., 2011 ).
Several lines of evidence have shown that N deposition also affects the functioning of the fungal decomposer community. Studies under controlled laboratory conditions have shown that the expression of certain ligninolytic enzymes is regulated at the level of gene transcription by N concentrations, where limiting levels of N typically de-repress lignin degradation (Tien & Kirk, 1983 ; Li et al ., 1994 ). Also in a field experiment, elevated levels of N led to reduced expression of ligninolytic genes and lower decomposition rates (Edwards et al ., 2011 ). Additional studies have found that, upon N additions, differential extracellular enzymatic responses, as well as changes in the magnitudes of forest soil respiration, could explain both increased and decreased litter decomposition rates (Carreiro et al ., 2000 ; Sinsabaugh et al ., 2002 ; Bowden et al ., 2004 ; Janssens et al ., 2010 ). For example, in forests with low-quality litter, the activity of lignin-degrading phenol oxidase and peroxidase declines substantially in response to N depositions (Carreiro et al ., 2000 ; DeForest et al ., 2004 ). In contrast, reductions in lignin-degrading enzyme activities are not common in ecosystems with high litter quality. However, Saiya-Cork et al . ( 2002 ) reported increased phenol oxidase activity in fresh litter, but decreased phenol oxidase activity in humus in an Acer sacccharum forest soil subjected to nitrogen deposition. Conversely, increases in the activities of cellulases are generally observed in ecosystems independent of litter quality (Carreiro et al ., 2000 ; Gallo et al ., 2004 ). Reductions in fungal biomass and activities have also been reported in N-fertilized plots (DeForest et al ., 2004 ; Rousk et al ., 2011 ), and such a response was accompanied by a significant reduction in the activity of phenol oxidase (Frey et al ., 2004 ). Hence, the physiological status of the fungal decomposer community appears to be altered upon N deposition, and this apparently coincides with differential decomposition rates of soil organic matter.
It has been suggested that the effect of N deposition on saprotrophic species is species-specific, with some species being positively influenced and others negatively affected (Gillet et al ., 2010 ). Allison et al . ( 2007 ) reported a shift in the composition of active fungi in boreal ecosystems upon N fertilization. Another study showed that enhanced N deposition increased the proportion of basidiomycete sequences recovered from litter in an Acer -dominated forest floor, whereas the proportion of ascomycetes in the community was significantly lower under elevated N deposition (Edwards et al ., 2011 ).
In conclusion, it is evident that climate change and N deposition affect fungal decomposer communities by modifying their physiological status, by shifting species composition or by a combination of both. However, our knowledge about the mechanisms and long-term consequences of these global change-induced effects remains limited.
Conclusions and perspectives
Community ecological aspects of decomposer fungi have received increasing attention as they form a well-defined functional group of organisms for which general ecological concepts and hypotheses can be tested. For instance, the relationship between diversity and functioning (lignocellulose decomposition) has been repeatedly studied, revealing the lack of a uniform relationship. The predominant type of interactions, the presence of species with extraordinary decomposition activities, and the composition of the organic resources, as well as the spatial scale at which their decomposition is examined, can determine the nature of this relationship. Other community–functioning relationship concepts such as ‘priority effects’ and ‘home-field advantages’ still require additional research before general conclusions can be drawn. Factors determining initial community assembly and species–area relationships have, so far, only received little attention for fungal decomposers (Feinstein & Blackwood, 2012 ). A better understanding of community ecological aspects of decomposer fungi is not only of basic ecological interest, but may also contribute to improving the reliability of models predicting organic matter decomposition.
The fungal contribution to the decomposition of labile carbon pools, for example, root exudates, has to date received relatively little attention. However, studies using 13 C labeled substrates or plants indicate that this contribution may be much higher than previously thought. A better knowledge of the diversity and ecology of such fungi is essential to the understanding of microbial community dynamics in the rhizosphere that are associated with plant nutrition and health. Saprotrophic fungi may exert control on plant-pathogenic fungi by competing for root exudates (Alabouvette et al ., 2009 ). The possibility that saprotrophic fungi can directly obtain nutrients from living roots by penetrating the outer parts is an interesting aspect of their ecology that deserves more investigation.
A strong increase in our knowledge on diversity of fungi in different terrestrial ecosystems, including agro-ecosystems, is to be expected because of the availability of high-throughput sequencing technologies (Hibbett et al ., 2009 ). Using such techniques, fungal diversities in the range of 100–2000 operational taxonomic units per gram of soil have been reported for soils from different natural and agricultural ecosystems (Buée et al ., 2009a ; Orgiazzi et al ., 2012 ; Xu et al ., 2012 ). As sequencing does not distinguish between functional groups, such appraisals do not however provide direct knowledge of the diversity of fungal decomposers and how this differs across ecosystems. It is likely that comparison of the diversity in different ecosystems within fungal genera will reveal such information (Nagy et al ., 2011 ).
Ideas on niche differentiation among fungal species are currently mainly based on their performance in experiments under strongly controlled conditions and their growth on different organic resources or during different succession stages of organic matter decomposition. However, ongoing developments in molecular biological (comparative genomics, transcriptomics) and biochemical techniques (metabolomics) strongly improve our ability to indicate the identity and metabolic functioning of active fungi in situ (Peiris et al ., 2008 ; Grigoriev et al ., 2011 ; Martin et al ., 2011 ; Ujor et al ., 2012 ). This will give an unprecedented insight into the functioning of terrestrial decomposer fungal communities and will also give a better appreciation of other functional groups involved in decomposition processes. Such baseline understanding can also allow for the examination of decomposition dynamics under different global climate change scenarios.
Although this review has mainly focused on interactions between decomposer fungi, interactions with other organisms (bacteria, archea, arthropods etc.) can also impact on the composition and functioning of fungal communities. Studies of these interactions are making rapid progress, and integrating these results with those of in situ fungal dynamics and activity will be an important and challenging task for fungal ecologists (Boddy et al ., 2010 ).
In summary, future attention to the contribution of fungal species and their intra-and interspecific interactions to decomposition rates under various abiotic and biotic conditions will be required to understand the link between fungal community dynamics and carbon cycling. The past decade has seen a large increase in publications dedicated to the ecology of fungal decomposers. Application of emerging new methods and integrating different disciplines will no doubt continue to fuel this expansion in research intensity.
We thank George Kowalchuk and two anonymous reviewers for their helpful comments. Funding was provided by the Netherlands Organisation for Scientific Research (NWO) in the form of a personal Veni grant to A.v.d.W. This is publication number 5314 of the NIOO-KNAW Netherlands Institute of Ecology.