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Linking soil microbial communities to vascular plant abundance along a climate gradient Luca Bragazza1,2,3, Richard D. Bardgett4, Edward A. D. Mitchell5,6 and Alexandre Buttler1,2,7 WSL Swiss Federal Institute for Forest, Snow and Landscape Research, Site Lausanne, Station 2, CH-1015 Lausanne, Switzerland; 2Laboratory of Ecological Systems (ECOS), Ecole
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Polytechnique Federale de Lausanne (EPFL), School of Architecture, Civil and Environmental Engineering (ENAC), Station 2, CH-1015 Lausanne, Switzerland; 3Department of Life Science and Biotechnologies, University of Ferrara, Corso Ercole I d’Este 32, I-44121 Ferrara, Italy; 4Faculty of Life Sciences, University of Manchester, Michael Smith Building, Manchester M13 9PT, UK; 5Laboratory of Soil Biology, University of Neuchatel, Rue Emile-Argand 11, CH-2000 Neuchatel, Switzerland; 6Jardin Botanique de Neuch^atel, Chemin du Perthuis-du-Sault 58, CH-2000 Neuchatel, Switzerland; 7Laboratoire de Chrono-Environnement, UMR 6249 CNRS – INRA, Universite de Franche-Comte, Besancßon, France
Summary Author for correspondence: Luca Bragazza Tel: +41 21 6935750 Email: [email protected] Received: 21 August 2014 Accepted: 20 September 2014
New Phytologist (2014) doi: 10.1111/nph.13116
Key words: bacteria, biomass, bogs, enzymes, ericoids, fungi, phospholipid fatty acid (PLFA), stoichiometry.
The ongoing expansion of shrub cover in response to climate change represents a unique
opportunity to explore the link between soil microbial communities and vegetation changes. This link is particularly important in peatlands where shrub expansion is expected to feed back negatively on the carbon sink capacity of these ecosystems. Microbial community structure and function were measured seasonally in four peatlands located along an altitude gradient representing a natural gradient of climate and associated vascular plant abundance. We show that increased soil temperature and reduced water content are associated with greater vascular plant biomass, in particular that of ericoids, and that this, in turn, is correlated with greater microbial biomass. More specifically, microbial community structure is characterized by an increasing dominance of fungi over bacteria with improved soil oxygenation. We also found that the carbon and nitrogen stoichiometry of microbial biomass differs in relation to soil microbial community structure and that this is ultimately associated with a different investment in extracellular enzymatic activity. Our findings highlight the fact that the determination of the structural identity of microbial communities can help to explain the biogeochemical dynamics of organic matter and provide a better understanding of ecosystem response to environmental changes.
Introduction There is growing evidence that impacts of global environmental change on ecosystem functioning can be driven as much, if not more so, by biotic drivers as abiotic drivers (Hooper et al., 2012; Tilman et al., 2012). Studies relating biotic drivers to ecosystem functioning have mainly focused on primary producers, in particular exploring the effects of changes in plant diversity (Cardinale et al., 2012; Eisenhauer et al., 2013); as a result, our understanding of how changes in below-ground communities influence ecosystem functioning in response to environmental change is still limited (van der Heijden et al., 2008; Chakraborty et al., 2012). One of the main routes by which global environmental change impacts below-ground communities and biogeochemical cycles is indirect, via vegetation change. Such effects, however, can operate over hierarchies of spatial and temporal scales, ranging from individual plant responses to shifts in vegetation dominance at regional and global levels on timescales of years to decades (Bardgett et al., 2013). An opportunity to link above-ground and below-ground responses to global environmental change is Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
currently represented by ‘shrub encroachment’, one of the most dramatic shifts in vegetation occurring worldwide, particularly in alpine and high-latitude regions in response to climate change (Walker et al., 2006; Wookey et al., 2009; Elmendorf et al., 2012). Specifically, peatlands are an ideal system to relate shrub encroachment to soil microbial responses, given that previous studies based on manipulative experiments (Weltzin et al., 2003; Jassey et al., 2013), historical vegetation surveys (Kapfer et al., 2011), and dynamic vegetation modeling (Heijmans et al., 2013) all show that climate warming triggers vegetation change via the expansion of dwarf shrubs (i.e. ericaceous species or ericoids). Peatland ecosystems act, in the long-term, as sinks of atmospheric carbon (C) by accumulating plant litter as peat. Mosses of the genus Sphagnum play a key role in peat accumulation because they compete effectively with vascular plants and produce litter that is recalcitrant to microbial decomposition (Hajek et al., 2011). An increase in vascular plant cover, especially of ericaceous shrubs, in response to climate change is expected to feed back negatively on the C sink capacity of peatlands by reducing litter production of underlying Sphagnum mosses resulting from New Phytologist (2014) 1 www.newphytologist.com
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direct competition for light (Limpens et al., 2011). Given the high degree of host specificity between soil microbes and peatland plants (Artz et al., 2007; Opelt et al., 2007), the importance of feedbacks between plants and soil microbes for nutrient acquisition (Bragazza et al., 2013), and the role of vegetation composition in modulating peatland gas fluxes in response to warming (Ward et al., 2013), we can expect that shrub encroachment will strongly impact below-ground communities in peatlands, but the mechanisms involved are unclear. This represents a serious gap in knowledge, given the importance of microbes in controlling the capacity of peatlands to act as long-term sinks of atmospheric C (Tveit et al., 2012) and, more generally, the known role of differences in microbial physiology in driving C and nutrient cycling in terrestrial ecosystems (Allison et al., 2010). Here, we used a well-characterized altitude gradient (Bragazza et al., 2013) to test the impact of climate warming and associated vegetation change on the structure and function of microbial communities in peatland soil. Specifically, we addressed the following questions. How does below-ground microbial biomass relate to above-ground plant biomass? Is the relative abundance of bacteria and fungi differently affected by warming and related vegetation changes? Does a structural shift in microbial communities affect the nutrient stoichiometry of microbial biomass and, if so, how does this affect microbial metabolism?
Materials and Methods Study sites We selected our study peatlands along an altitude gradient as a ‘space for time’ substitution. The use of climatic gradients to study the climate change impact on natural ecosystems is frequently applied (Dunne et al., 2004; Fukami & Wardle, 2005; K€orner, 2007) and although this approach has advantages and disadvantages, it permits to assess the response of real, intact ecosystems to a set of interacting environmental factors (temperature, precipitation) under equilibrium conditions of the vegetation communities. We then selected four peatlands in Switzerland that fulfilled the following criteria: the Sphagnum layer was characterized by a mixture of Sphagnum magellanicum, Sphagnum capillifolium and Sphagnum fallax; the herbaceous layer was characterized by the presence, in all the sites, of Eriophorum vaginatum, Calluna vulgaris, Andromeda polifolia and Vaccinium species (in particular, Vaccinium oxycoccus and Vaccinium uliginosum); and the sites were located at different altitudes. The selected peatlands were L€ormoos (LM, 585 m above sea level (asl), 46°580 N, 7°240 E; Canton Bern), Praz Rodet (PR, 1040 m asl, 46°330 N, 6°100 E; Canton Vaud), Sortel (SR, 1406 m asl, 46°440 N 7°220 E; Canton Bern), and Hochr€ajen (GR, 1885 m asl, 46°360 N, 7°580 E; Canton Bern). The overall plant species composition and hydrochemical conditions were typical of ombrotrophic (= rain-fed) peatlands in Europe (Bragazza et al., 2013). Based on data collected from the closest meteorological stations, the mean annual temperature (corrected for the altitude) and total precipitation were, respectively: 8.4°C and 1030 mm at L€ormoos; 5.9°C and New Phytologist (2014) www.newphytologist.com
1274 mm at Praz Rodet; 3.9°C and 1346 mm at Sortel; and 1.3°C and 1427 mm at Hochr€ajen. For the four study sites, the amount of precipitation falling during the plant growing season (April–October) was between 57 and 66% of the annual amount. Field sampling During the growing season 2010, three peat and water samples were periodically collected in each of the study peatlands in topographically flat areas characterized by the same Sphagnum species composition. For all statistical analyses, the mean value of the three samples was used. Five sampling campaigns were performed in each peatland from April to October 2010. Three 50-cm-long perforated plastic tubes were installed in representative habitats to periodically measure the water table depth. Peat samples were collected using a knife to a depth between 4 and 14 cm, in order to exclude the upper living portion of Sphagnum plants, but to include the peat depth where the highest fine root production is typically observed. Three pore water samples (= surface water) were collected at each sampling date by means of 10-cmlong moisture samplers (Eijkelkamp Agrisearch Equipment, Giesbeek, the Netherlands). In mid-July 2010, at each peatland site, total above-ground vascular plant biomass and total standing litter were collected in five 25 cm 9 25 cm plots. Peat soil temperature was measured using two data loggers per site inserted at a depth of 5 cm. Peat water content (%) was measured gravimetrically on collected peat samples and calculated as the difference between wet weight and dry weight (DW) (105°C for 24 h) divided by wet weight. Biogeochemical analyses The concentration of dissolved organic carbon (DOC) in pore water was determined by combustion using a Shimadzu analyzer (TOC-V CPH). The ratio of absorbance at 400 and 600 nm (E400/E600) was used as a simple index inversely correlated to the molecular weight of dissolved organic matter (Moore, 1987). Before analyses, pore water samples were filtered through a glass fiber 0.45 lm filter. For the determination of phospholipids (PLFAs), fresh peat samples were extracted with a chloroform : methanol : phosphorus buffer. Lipids were separated into neutral lipids, glycolipids and phospholipids on a silic acid column. The phospholipids were subjected to a mild alkaline methanolysis and the fatty acid methyl esters were detected by GC (Varian CP-3800 GC-MS 1200 L Quadrupole; Santa Clara, CA, USA). The sum of the following PLFA markers was chosen as an abundance – for actinomycetes, 10Me16, 10Me18; for Grampositive bacteria, i15:0, a15:0, i16:0, i17:0, 10Me16, 10Me18; for Gram-negative bacteria, 16:1w7c, cy19:0 – whereas the quantity of 18:2x6 was used as indicator of fungal abundance (Frosteg ard & B a ath, 1996). Total bacteria biomass was calculated as the sum of Gram-positive and Gram-negative PLFA markers. For the determination of microbial biomass carbon (C), nitrogen (N) and phosphorus (P), pairs of c. 20 g of fresh peat were weighed for each replicate and one sample from each pair was immediately extracted in 100 ml solution of 0.5 M K2SO4 for C Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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Data on soil microbial communities were coupled with environmental data using a symmetrical coinertia analysis (Doledec & Chessel, 1994). Briefly, this method computes the covariance matrix crossing the variables of the two data tables and represents the initial principal component analyses (PCAs) applied on each data set in a common ordination space of coinertia axes. To determine significant differences among peatlands for the same factor, a Student’s t-test was applied. Pearson’s correlation (r) and univariate regression (r2) were also used.
Results During the growing season, mean peat temperatures at a depth of 5 cm were 16.8 and 11.4°C at the lowest and highest altitude sites, respectively, with a decreasing trend of the mean peat temperature (°C) in the 3 d preceding the sampling with increasing altitude (m) (y = 18.10.004x, r2 = 0.32, P < 0.01, n = 20). By contrast, mean peat water content (%) increased with increasing Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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altitude (y = 91.7 + 0.002x, r2 = 0.73, P < 0.01, n = 20), ranging from 93.2 to 95.7% at the lowest and highest altitude sites, respectively. The mean water-table depth (cm) decreased with altitude (y = 13.8 + 0.005x; r2 = 0.72, P < 0.05, n = 4), ranging from 10 to 5 cm at the lowest and highest altitude sites, respectively. Total above-ground vascular plant biomass decreased with altitude (r2 = 0.54, P < 0.01, n = 20), being greatest at the lowest altitude site (i.e. 270 37 g m2, n = 5) and lowest at the highest altitude site (i.e. 74 7 g m2, n = 5) (Fig. 1). The main determinant of this change in total above-ground biomass was a decrease in biomass of ericoid dwarf shrubs (y = 128 0.07x, r2 = 0.59, P < 0.01, n = 20), whereas graminoid biomass was not correlated with altitude (r2 = 0.01, P = 0.88, n = 20). We detected a positive relationship between mean aboveground plant biomass and mean total abundance of microbial PLFAs along the altitude gradient (Fig. 1). In addition, aboveground plant biomass was positively related to the concentration of DOC in pore water (Fig. 1). Spectrophotometric properties of pore water showed a positive relationship between DOC concentration and corresponding E400/E600 absorbance ratio (Pearson’s r = 0.79, P < 0.01, n = 20), indicating that higher DOC concentration was characterized by C compounds with a lower molecular weight. The ratio of fungi to bacteria biomass (F : B ratio) was negatively related to altitude (Pearson’s r = 0.74, P < 0.01, n = 20), which mainly resulted from a decreasing trend of fungal biomass with increasing altitude (Pearson’s r = 0.55, P = 0.011, n = 20). The relationship between the F : B ratio and hydrochemical conditions revealed a negative relationship with volumetric peat
Total microbial biomass (µmol g–1)
and N, and Bray–1 (0.03 M NH4F–0.025 M HCl) for P, whereas the other sample was put in a vacuum desiccator and subjected to chloroform vapor. After 1 d of fumigation, the fumigated peat sample was extracted with the same solutions as earlier. Total C and N concentrations in fumigated and unfumigated samples were determined with a Shimadzu total C and N analyzer, whereas P concentration was determined colorimetrically with a continuous-flow autoanalyzer (AutoAnalyzer 3, Seal Analytical, Southampton, UK). Microbial biomass C, N and P were estimated as the differences between the amounts of C, N and P after and before fumigation using an extractability factor of 0.38 for C (Vance et al., 1987), 0.54 for N (Brookes et al., 1985) and 0.38 for P (Brookes et al., 1982). Microbial biomass C, N and P are expressed as mg g–1 oven dry peat (105°C for 24 h). Roots were manually removed from peat samples before any treatment. The potential activity of extracellular hydrolase enzymes was measured by adding 4-methylumbelliferyl-b-D-glucoside for the activity of b-glucosidase (BG), 4-MUF-N-acetyl-b-D-glucosaminide for the activity of b-1,4-N-acetylglucosaminidase (NAG), L-leucine-7-amido-4-methycoumarinhydrochloride for the activity of leucine aminopeptidase (LAP) and 4-MUF-phosphate for the activity of phosphatase (AP) to c. 1 g of fresh peat. After incubation (1 h for BG, NAG and LAP, and 45 min for AP), the fluorescence of the supernatant was measured, after centrifugation, on a microplate reader (BioTek SynergyMX, Lucerne, Switzerland) at 450 nm emission and 330 nm excitation wavelength. To quantify product release and account for quenching effects, a set of standards was prepared using methylumbelliferone (MUF) and 7-amino-4-methylcoumarin (MCU) mixed with peat extract. Enzyme activities were expressed as lmol of substrate (MUF and MCU) converted min–1 g–1 DW of peat (Bragazza et al., 2013). The specific enzyme activity was calculated as the ratio of potential enzyme activity to the correspondent PLFA biomass as determined in each peat sample.
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Above-ground biomass (g m–2) Fig. 1 Relationship between total above-ground vascular plant biomass (mean 1 SE; n = 5) and total microbial biomass (long dashed line), and dissolved organic carbon concentration (DOC; short dashed line) in the four study peatlands located at different altitudes. Microbial biomass is based on phospholipid fatty acids (PLFAs). Each value of microbial biomass and DOC is the mean ( 1 SE) of five seasonal replicates. Linear regression for total microbial biomass, y = 0.068 + 0.0002x, r2 = 0.93, P = 0.037; linear regression for DOC, y = 6.96 + 0.13x, r2 = 0.99, P = 0.004. New Phytologist (2014) www.newphytologist.com
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water content (Fig. 2), but a positive relationship with the E400/ E600 absorbance ratio (Pearson’s r = 0.61, P < 0.05, n = 20) of pore water, indicating a dominance of fungi in the presence of DOC characterized by lower molecular weight. The coinertia analysis showed that along the first axis there was a positive correlation between PLFA markers for fungi and Gram-negative bacteria with environmental variables such as DOC concentration and E400/E600 absorbance ratio (Fig. 3). PLFA markers for actinomycetes were opposite to these variables and related more to the peat water content. Along the second axis there was a negative correlation between peat temperature and PLFA markers associated with Gram-positive bacteria. Carbon, N and P content in soil microbial biomass differed in relation to the structure of microbial communities (Fig. 4). In the two peatlands located at lower altitude (Fig. 4a), where the F : B ratio was greater, the microbial biomass (mean 1 SE) C was significantly greater than in the two higher-altitude sites 1.3
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(9.72 0.42 vs 8.15 0.41 mg g1, Student’s t = 2.8, P = 0.012, n = 10). By contrast, microbial biomass P (Fig. 4c) and, to a greater extent, microbial biomass N (Fig. 4b) were significantly greater in the two higher-altitude sites where the F : B ratio was lower (0.36 0.034 vs 0.48 0.034 mg g1 for P and 0.53 0.025 vs 0.61 0.022 mg g1 for N; Student’s t < 1.87, P < 0.076, n = 10). Differences in nutrient stoichiometry (mean 1 SE) of microbial biomass were also detected with altitude. In particular, the microbial biomass of the two lowest altitude sites had a greater C/N (17.8 0.65 vs 14.2 0.8), C/P (29.6 2.5 vs 17.1 1.2) and N/P (1.86 021 vs 1.34 0.09) ratio than at the two highest altitude sites (Student’s t < 4.3, P < 0.036, n = 10). The specific activity of C and N degrading enzymes increased with increasing microbial biomass C and N (Fig. 5a,b), whereas the specific enzyme activity of phosphatase was not related to the microbial biomass P (Fig. 5c).
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Fig. 2 Relationship between mean fungi to bacteria biomass ratio (F : B; mean 1 SE; n = 5) and peat water content (mean 1 SE; n = 5) in the four study peatlands located at different altitude. The F : B biomass ratio was based on phospholipid fatty acids (PLFAs). Linear regression for F : B ratio, y = 30.90.32x, r2 = 0.97, P < 0.05. Axis 2
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The natural decrease of air temperature and increase of precipitation with altitude were reflected in both a decrease of soil temperature and an increase of soil water content as a result of reduced evapotranspiration. The observed changes in temperature and precipitation along the altitude gradient are consistent with the future climatic scenarios for Switzerland up to 2050 (Frei et al., 2007), thereby making the selected altitude gradient a reliable substitution of time to assess above-ground and below-ground responses in peatland to climate change. Along such a natural climate gradient, above-ground plant biomass, in particular that of low shrubs (i.e. the ericoid dwarf shrubs) was enhanced as a result of higher soil oxygenation, which has previously been shown to be a key factor in controlling biomass allocation of vascular plants in peatlands (Moore et al., 2002; Bubier et al., 2006). The observed increasing cover of vascular plants is in agreement with the results of a vegetation monitoring program of ombrotrophic Swiss peatlands which reported an increase of shrubs during a decadal time frame (Klaus, 2007). We also detected a positive relationship between aboveground plant biomass and below-ground microbial biomass Peat temperature
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Fig. 3 Scatterplot of the microbial community (left) and environmental variables (right) showing their contribution to the coinertia space. Vectors pointing to the same direction within and across both the scatterplots are correlated. Axis 1 has the highest weight (eigenvalue = 2.52), whereas axis 2 is weaker (eigenvalue = 0.8). The correlation value between the two data sets is 0.26 (RV-value, a measure of global similarity between the datasets). DOC, dissolved organic carbon. Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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Fig. 4 Microbial biomass nutrient content for carbon (C) (a), nitrogen (N) (b), and phosphorus (P) (c) in relation to the relative abundance of fungi and bacteria biomass in the four study sites. The fungi to bacteria biomass ratio (F : B ratio) was based on phospholipid fatty acid (PLFA) abundance, whereas the microbial biomass nutrient content was based on chloroform fumigation. Data are means 1 SE (n = 5). Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
Fig. 5 Specific enzyme activity (mean 1 SE, n = 5) and corresponding microbial biomass nutrient for carbon (C) (a), nitrogen (N) (b), and phosphorus (P) (c) in the four study sites. Specific enzyme activity was calculated as the ratio of mean enzyme activity to mean microbial phospholipid fatty acid (PLFA) biomass, whereas the microbial biomass nutrient content was based on chloroform fumigation. Linear regression for C-specific enzyme activity, y = 26.6 + 6.61x, r2 = 0.87, P = 0.068; linear regression for N-specific enzyme activity, y = 12.9 + 63.6x, r2 = 0.88, P = 0.060. New Phytologist (2014) www.newphytologist.com
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(Fig. 1) along the altitude gradient, which is likely explained by the availability of low-molecular-weight C, which can then fuel and sustain a greater microbial biomass (Zak et al., 1994; Paterson et al., 2007; Trinder et al., 2008; Fierer et al., 2009; Gutknecht et al., 2012). Although not measured here, it is well known that above-ground plant biomass in peatlands is positively related to below-ground root biomass (Moore et al., 2002; Murphy et al., 2009; Murphy & Moore, 2010). Likely mechanisms by which greater biomass of vascular plants enhances the availability of low-molecular-weight C are through temperature-induced root exudation (Boone et al., 1998) and the relatively short life span of ericoid fine roots (Valenzuela-Estrada et al., 2008). DOC of lower molecular weight is generally more labile, that is, more easily degradable by soil microbes (Boddy et al., 2008; Saari et al., 2009; Mann et al., 2012). Accordingly, the observation of greater concentration of low-molecular-weight DOC in pore water at lower altitude, where ericoid dwarf shrubs are more abundant and peat temperature is enhanced, supports the hypothesis of a positive link between vascular plant abundance and microbial biomass through the enhanced release of labile C substrate. We detected a significant shift in soil microbial community structure with altitude, being fungi, relative to bacteria, less abundant at higher altitudes. A potential mechanism explaining this response might be the sensitivity of fungi to anoxic conditions (Jaatinen et al., 2007; Peltoniemi et al., 2009; Lin et al., 2012), given that peat water content also increased with altitude. An alternative explanation might be the reduction in ericoid biomass and their associated mycorrhizal fungi with increasing altitude (Perotto et al., 2002; Andersen et al., 2013). However, a more detailed analysis of microbial community structure based on PLFAs revealed a different response of the main microbial groups to changing abiotic and biotic conditions. Indeed, Gram-negative bacteria and, in particular, fungi were more associated with lower-altitude sites (Fig. 3), which is in accordance with their preference for more aerated soils and higher availability of labile C substrate (Jaatinen et al., 2007; Paterson et al., 2007; Garcia-Pausas & Paterson, 2011). The greater abundance of Gram-positive bacteria and actinomycetes at higher-altitude sites seems to confirm previous reports of an adaptability of these microbes to more anaerobic soil conditions as well as to recalcitrant C substrates (Griffiths et al., 1999; Fierer et al., 2003; Paterson et al., 2007; Drigo et al., 2010; Bird et al., 2011). Differences in soil aeration and quality of nutritional resources seem, then, to modulate the ability of different microbial groups to dominate the soil community in relation to their specific physiological requirements (Allison et al., 2010; Waring et al., 2013). A change in the relative abundance of bacteria and fungi in peat soil was associated with shifts in the nutrient content of the microbial biomass. We observed that with increasing F : B ratio the microbial biomass tends to be enriched in C (Fig. 4a) and depleted in N and P (Fig. 4b,c). Such a trend could reflect, in part, differences in microbial nutrient requirements, considering that fungi have a higher C demand compared with bacteria, which appear more constrained by N and P (Fierer et al., 2009; New Phytologist (2014) www.newphytologist.com
Keiblinger et al., 2010). The observed differences in nutrient biomass stoichiometry also explained differences in investment in extracellular enzymes at the community level (Fig. 5a–c). Indeed, a microbial community dominated by fungi seems to have a higher activity of C-degrading enzyme per unit of microbial PLFA biomass (Fig. 5a) compared with a bacteria-dominated community where, instead, microbial investment is towards the acquisition of N compounds (Fig. 5b). Previous studies focusing on C loss in peatlands in response to changes of abiotic conditions reported an increase of peat decomposition with increased soil temperature or reduced soil water content (Dorrepaal et al., 2009; Hardie et al., 2011), a response generally explained by an enhancement of extracellular enzymatic activity (Fenner & Freeman, 2011). Our data, however, demonstrate the important role of the interactive effects of change in vegetation cover and climatic conditions (= abiotic driver) on soil microbial structure and function (Chanton et al., 2008). In particular, our data seem to provide a microbial-based explanation for a previously observed facilitative role of fresh plant-derived C in promoting the decomposition of old organic matter (Fontaine et al., 2008; Hardie et al., 2009; Hartley et al., 2012; van Groenigen et al., 2014; Wild et al., 2014). Indeed, the observed increase of specific activity of C-degrading enzyme under greater vascular plant cover is not simply a response to greater labile C availability (Bragazza et al., 2013), but also reflects an increased C requirement of the new soil microbial biomass. Although some caution is needed for upscaling the observed patterns to different geographic areas or ecosystems, on the whole our findings highlight that the determination of the structural identity of microbial communities can help to explain the biogeochemical dynamics in response to global environmental change and better predict the microbial impacts on organic matter decomposition.
Acknowledgements We thank J. Parisod, M. Lamentovicz, T. Spiegelberger, E. Feldmeyer-Christe, P. Gomis, A. Margot, J. D. Teuscher, E. Rossel, K. Vernez, and J. D. Welch for assistance. The ‘Service des for^ets, de la faune et de la nature’ (Canton de Vaud) and the ‘Service de la promotion de la nature – Office de l’agriculture et de la nature’ (Canton de Berne) are acknowledged for giving permission to access the study sites. This study was supported financially by the Swiss National Science Foundation (project ClimaBog, grant 205321-129981 to L.B.).
References Allison SD, Wallenstein MD, Bradford MA. 2010. Soil-carbon response to warming dependent on microbial physiology. Nature Geoscience 3: 336–340. Andersen R, Chapman SJ, Artz RRE. 2013. Microbial communities in natural and disturbed peatlands: a review. Soil Biology and Biochemistry 57: 979–994. Artz RRE, Anderson IA, Chapman SJ, Hagn A, Schloter M, Potts JM, Campbell CD. 2007. Changes in fungal community composition in response to vegetational succession during the natural regeneration of cutover peatland. Microbial Ecology 54: 508–522. Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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New Phytologist Trinder CJ, Artz RRE, Johnson D. 2008. Contribution of plant photosynthate to soil respiration and dissolved organic carbon in a naturally recolonising cutover peatland. Soil Biology and Biochemistry 40: 1622–1628. Tveit A, Schwacke R, Svenning MM, Urich T. 2012. Organic carbon transformations in high-Arctic peat soils: key functions and microorganisms. The ISME Journal 7: 299–311. Valenzuela-Estrada LR, Vera-Caraballo V, Ruth LE, Eissenstat DM. 2008. Root anatomy, morphology, and longevity among root orders in Vaccinium corymbosum (Ericaceae). American Journal of Botany 95: 1506–1514. Vance ED, Brookes PC, Jenkinson DS. 1987. An extraction method for measuring soil microbial biomass C. Soil Biology and Biochemistry 19: 703–707. Walker MD, Wahren CH, Hollister RD, Henry GHR, Ahlquist LE, Alatalo JM, Syndonia Bret-Harte M, Calef MP, Callaghan TV, Carroll AM et al. 2006. Plant community responses to experimental warming across the tundra biomes. Proceedings of the National Academy of Sciences, USA 103: 1342–1346. Ward SE, Ostle NJ, Oakley S, Quirk H, Henrys PA, Bardgett RD. 2013. Warming effects on greenhouse gas fluxes in peatlands are modulated by vegetation composition. Ecology Letters 16: 1285–1293. Waring BG, Averill C, Hawkes CV. 2013. Differences in fungal and bacterial physiology alter soil carbon and nitrogen cycling: insights from a meta-analysis and theoretical models. Ecology Letters 16: 887–894. Weltzin JF, Bridgham SD, Pastor J, Chen J, Harth C. 2003. Potential effects of warming and drying on peatland plant community composition. Global Change Biology 9: 141–151. Wild B, Schnecker J, Alves RJE, Barsukov P, Ba rta J, Capek P, Gentsch N, Gittel A, Guggenberger G, Lashchinskiy N et al. 2014. Input of easily available organic C and N stimulates microbial decomposition of soil organic matter in arctic permafrost soil. Soil Biology and Biochemistry 75: 143–151. Wookey PA, Aerts R, Bardgett RD, Baptist F, Br athen K, Cornelissen JHC, Gough L, Hartley IP, Hopkins DW, Lavorel S et al. 2009. Ecosystem feedbacks and cascade processes: understanding their role in the responses of Arctic and alpine ecosystems to environmental change. Global Change Biology 15: 1153–1172. Zak DR, Tilman D, Parmenter RR, Rice CW, Fisher FM, Vose J, Milchunas D, Wayne Martin C. 1994. Plant production and soil microorganisms in late-successional ecosystems: a continental scale study. Ecology 75: 2333–2347.
Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
Linking soil microbial communities to vascular plant abundance along a climate gradient Luca Bragazza1,2,3, Richard D. Bardgett4, Edward A. D. Mitchell5,6 and Alexandre Buttler1,2,7 WSL Swiss Federal Institute for Forest, Snow and Landscape Research, Site Lausanne, Station 2, CH-1015 Lausanne, Switzerland; 2Laboratory of Ecological Systems (ECOS), Ecole
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Polytechnique Federale de Lausanne (EPFL), School of Architecture, Civil and Environmental Engineering (ENAC), Station 2, CH-1015 Lausanne, Switzerland; 3Department of Life Science and Biotechnologies, University of Ferrara, Corso Ercole I d’Este 32, I-44121 Ferrara, Italy; 4Faculty of Life Sciences, University of Manchester, Michael Smith Building, Manchester M13 9PT, UK; 5Laboratory of Soil Biology, University of Neuchatel, Rue Emile-Argand 11, CH-2000 Neuchatel, Switzerland; 6Jardin Botanique de Neuch^atel, Chemin du Perthuis-du-Sault 58, CH-2000 Neuchatel, Switzerland; 7Laboratoire de Chrono-Environnement, UMR 6249 CNRS – INRA, Universite de Franche-Comte, Besancßon, France
Summary Author for correspondence: Luca Bragazza Tel: +41 21 6935750 Email: [email protected] Received: 21 August 2014 Accepted: 20 September 2014
New Phytologist (2014) doi: 10.1111/nph.13116
Key words: bacteria, biomass, bogs, enzymes, ericoids, fungi, phospholipid fatty acid (PLFA), stoichiometry.
The ongoing expansion of shrub cover in response to climate change represents a unique
opportunity to explore the link between soil microbial communities and vegetation changes. This link is particularly important in peatlands where shrub expansion is expected to feed back negatively on the carbon sink capacity of these ecosystems. Microbial community structure and function were measured seasonally in four peatlands located along an altitude gradient representing a natural gradient of climate and associated vascular plant abundance. We show that increased soil temperature and reduced water content are associated with greater vascular plant biomass, in particular that of ericoids, and that this, in turn, is correlated with greater microbial biomass. More specifically, microbial community structure is characterized by an increasing dominance of fungi over bacteria with improved soil oxygenation. We also found that the carbon and nitrogen stoichiometry of microbial biomass differs in relation to soil microbial community structure and that this is ultimately associated with a different investment in extracellular enzymatic activity. Our findings highlight the fact that the determination of the structural identity of microbial communities can help to explain the biogeochemical dynamics of organic matter and provide a better understanding of ecosystem response to environmental changes.
Introduction There is growing evidence that impacts of global environmental change on ecosystem functioning can be driven as much, if not more so, by biotic drivers as abiotic drivers (Hooper et al., 2012; Tilman et al., 2012). Studies relating biotic drivers to ecosystem functioning have mainly focused on primary producers, in particular exploring the effects of changes in plant diversity (Cardinale et al., 2012; Eisenhauer et al., 2013); as a result, our understanding of how changes in below-ground communities influence ecosystem functioning in response to environmental change is still limited (van der Heijden et al., 2008; Chakraborty et al., 2012). One of the main routes by which global environmental change impacts below-ground communities and biogeochemical cycles is indirect, via vegetation change. Such effects, however, can operate over hierarchies of spatial and temporal scales, ranging from individual plant responses to shifts in vegetation dominance at regional and global levels on timescales of years to decades (Bardgett et al., 2013). An opportunity to link above-ground and below-ground responses to global environmental change is Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
currently represented by ‘shrub encroachment’, one of the most dramatic shifts in vegetation occurring worldwide, particularly in alpine and high-latitude regions in response to climate change (Walker et al., 2006; Wookey et al., 2009; Elmendorf et al., 2012). Specifically, peatlands are an ideal system to relate shrub encroachment to soil microbial responses, given that previous studies based on manipulative experiments (Weltzin et al., 2003; Jassey et al., 2013), historical vegetation surveys (Kapfer et al., 2011), and dynamic vegetation modeling (Heijmans et al., 2013) all show that climate warming triggers vegetation change via the expansion of dwarf shrubs (i.e. ericaceous species or ericoids). Peatland ecosystems act, in the long-term, as sinks of atmospheric carbon (C) by accumulating plant litter as peat. Mosses of the genus Sphagnum play a key role in peat accumulation because they compete effectively with vascular plants and produce litter that is recalcitrant to microbial decomposition (Hajek et al., 2011). An increase in vascular plant cover, especially of ericaceous shrubs, in response to climate change is expected to feed back negatively on the C sink capacity of peatlands by reducing litter production of underlying Sphagnum mosses resulting from New Phytologist (2014) 1 www.newphytologist.com
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direct competition for light (Limpens et al., 2011). Given the high degree of host specificity between soil microbes and peatland plants (Artz et al., 2007; Opelt et al., 2007), the importance of feedbacks between plants and soil microbes for nutrient acquisition (Bragazza et al., 2013), and the role of vegetation composition in modulating peatland gas fluxes in response to warming (Ward et al., 2013), we can expect that shrub encroachment will strongly impact below-ground communities in peatlands, but the mechanisms involved are unclear. This represents a serious gap in knowledge, given the importance of microbes in controlling the capacity of peatlands to act as long-term sinks of atmospheric C (Tveit et al., 2012) and, more generally, the known role of differences in microbial physiology in driving C and nutrient cycling in terrestrial ecosystems (Allison et al., 2010). Here, we used a well-characterized altitude gradient (Bragazza et al., 2013) to test the impact of climate warming and associated vegetation change on the structure and function of microbial communities in peatland soil. Specifically, we addressed the following questions. How does below-ground microbial biomass relate to above-ground plant biomass? Is the relative abundance of bacteria and fungi differently affected by warming and related vegetation changes? Does a structural shift in microbial communities affect the nutrient stoichiometry of microbial biomass and, if so, how does this affect microbial metabolism?
Materials and Methods Study sites We selected our study peatlands along an altitude gradient as a ‘space for time’ substitution. The use of climatic gradients to study the climate change impact on natural ecosystems is frequently applied (Dunne et al., 2004; Fukami & Wardle, 2005; K€orner, 2007) and although this approach has advantages and disadvantages, it permits to assess the response of real, intact ecosystems to a set of interacting environmental factors (temperature, precipitation) under equilibrium conditions of the vegetation communities. We then selected four peatlands in Switzerland that fulfilled the following criteria: the Sphagnum layer was characterized by a mixture of Sphagnum magellanicum, Sphagnum capillifolium and Sphagnum fallax; the herbaceous layer was characterized by the presence, in all the sites, of Eriophorum vaginatum, Calluna vulgaris, Andromeda polifolia and Vaccinium species (in particular, Vaccinium oxycoccus and Vaccinium uliginosum); and the sites were located at different altitudes. The selected peatlands were L€ormoos (LM, 585 m above sea level (asl), 46°580 N, 7°240 E; Canton Bern), Praz Rodet (PR, 1040 m asl, 46°330 N, 6°100 E; Canton Vaud), Sortel (SR, 1406 m asl, 46°440 N 7°220 E; Canton Bern), and Hochr€ajen (GR, 1885 m asl, 46°360 N, 7°580 E; Canton Bern). The overall plant species composition and hydrochemical conditions were typical of ombrotrophic (= rain-fed) peatlands in Europe (Bragazza et al., 2013). Based on data collected from the closest meteorological stations, the mean annual temperature (corrected for the altitude) and total precipitation were, respectively: 8.4°C and 1030 mm at L€ormoos; 5.9°C and New Phytologist (2014) www.newphytologist.com
1274 mm at Praz Rodet; 3.9°C and 1346 mm at Sortel; and 1.3°C and 1427 mm at Hochr€ajen. For the four study sites, the amount of precipitation falling during the plant growing season (April–October) was between 57 and 66% of the annual amount. Field sampling During the growing season 2010, three peat and water samples were periodically collected in each of the study peatlands in topographically flat areas characterized by the same Sphagnum species composition. For all statistical analyses, the mean value of the three samples was used. Five sampling campaigns were performed in each peatland from April to October 2010. Three 50-cm-long perforated plastic tubes were installed in representative habitats to periodically measure the water table depth. Peat samples were collected using a knife to a depth between 4 and 14 cm, in order to exclude the upper living portion of Sphagnum plants, but to include the peat depth where the highest fine root production is typically observed. Three pore water samples (= surface water) were collected at each sampling date by means of 10-cmlong moisture samplers (Eijkelkamp Agrisearch Equipment, Giesbeek, the Netherlands). In mid-July 2010, at each peatland site, total above-ground vascular plant biomass and total standing litter were collected in five 25 cm 9 25 cm plots. Peat soil temperature was measured using two data loggers per site inserted at a depth of 5 cm. Peat water content (%) was measured gravimetrically on collected peat samples and calculated as the difference between wet weight and dry weight (DW) (105°C for 24 h) divided by wet weight. Biogeochemical analyses The concentration of dissolved organic carbon (DOC) in pore water was determined by combustion using a Shimadzu analyzer (TOC-V CPH). The ratio of absorbance at 400 and 600 nm (E400/E600) was used as a simple index inversely correlated to the molecular weight of dissolved organic matter (Moore, 1987). Before analyses, pore water samples were filtered through a glass fiber 0.45 lm filter. For the determination of phospholipids (PLFAs), fresh peat samples were extracted with a chloroform : methanol : phosphorus buffer. Lipids were separated into neutral lipids, glycolipids and phospholipids on a silic acid column. The phospholipids were subjected to a mild alkaline methanolysis and the fatty acid methyl esters were detected by GC (Varian CP-3800 GC-MS 1200 L Quadrupole; Santa Clara, CA, USA). The sum of the following PLFA markers was chosen as an abundance – for actinomycetes, 10Me16, 10Me18; for Grampositive bacteria, i15:0, a15:0, i16:0, i17:0, 10Me16, 10Me18; for Gram-negative bacteria, 16:1w7c, cy19:0 – whereas the quantity of 18:2x6 was used as indicator of fungal abundance (Frosteg ard & B a ath, 1996). Total bacteria biomass was calculated as the sum of Gram-positive and Gram-negative PLFA markers. For the determination of microbial biomass carbon (C), nitrogen (N) and phosphorus (P), pairs of c. 20 g of fresh peat were weighed for each replicate and one sample from each pair was immediately extracted in 100 ml solution of 0.5 M K2SO4 for C Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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Data on soil microbial communities were coupled with environmental data using a symmetrical coinertia analysis (Doledec & Chessel, 1994). Briefly, this method computes the covariance matrix crossing the variables of the two data tables and represents the initial principal component analyses (PCAs) applied on each data set in a common ordination space of coinertia axes. To determine significant differences among peatlands for the same factor, a Student’s t-test was applied. Pearson’s correlation (r) and univariate regression (r2) were also used.
Results During the growing season, mean peat temperatures at a depth of 5 cm were 16.8 and 11.4°C at the lowest and highest altitude sites, respectively, with a decreasing trend of the mean peat temperature (°C) in the 3 d preceding the sampling with increasing altitude (m) (y = 18.10.004x, r2 = 0.32, P < 0.01, n = 20). By contrast, mean peat water content (%) increased with increasing Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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altitude (y = 91.7 + 0.002x, r2 = 0.73, P < 0.01, n = 20), ranging from 93.2 to 95.7% at the lowest and highest altitude sites, respectively. The mean water-table depth (cm) decreased with altitude (y = 13.8 + 0.005x; r2 = 0.72, P < 0.05, n = 4), ranging from 10 to 5 cm at the lowest and highest altitude sites, respectively. Total above-ground vascular plant biomass decreased with altitude (r2 = 0.54, P < 0.01, n = 20), being greatest at the lowest altitude site (i.e. 270 37 g m2, n = 5) and lowest at the highest altitude site (i.e. 74 7 g m2, n = 5) (Fig. 1). The main determinant of this change in total above-ground biomass was a decrease in biomass of ericoid dwarf shrubs (y = 128 0.07x, r2 = 0.59, P < 0.01, n = 20), whereas graminoid biomass was not correlated with altitude (r2 = 0.01, P = 0.88, n = 20). We detected a positive relationship between mean aboveground plant biomass and mean total abundance of microbial PLFAs along the altitude gradient (Fig. 1). In addition, aboveground plant biomass was positively related to the concentration of DOC in pore water (Fig. 1). Spectrophotometric properties of pore water showed a positive relationship between DOC concentration and corresponding E400/E600 absorbance ratio (Pearson’s r = 0.79, P < 0.01, n = 20), indicating that higher DOC concentration was characterized by C compounds with a lower molecular weight. The ratio of fungi to bacteria biomass (F : B ratio) was negatively related to altitude (Pearson’s r = 0.74, P < 0.01, n = 20), which mainly resulted from a decreasing trend of fungal biomass with increasing altitude (Pearson’s r = 0.55, P = 0.011, n = 20). The relationship between the F : B ratio and hydrochemical conditions revealed a negative relationship with volumetric peat
Total microbial biomass (µmol g–1)
and N, and Bray–1 (0.03 M NH4F–0.025 M HCl) for P, whereas the other sample was put in a vacuum desiccator and subjected to chloroform vapor. After 1 d of fumigation, the fumigated peat sample was extracted with the same solutions as earlier. Total C and N concentrations in fumigated and unfumigated samples were determined with a Shimadzu total C and N analyzer, whereas P concentration was determined colorimetrically with a continuous-flow autoanalyzer (AutoAnalyzer 3, Seal Analytical, Southampton, UK). Microbial biomass C, N and P were estimated as the differences between the amounts of C, N and P after and before fumigation using an extractability factor of 0.38 for C (Vance et al., 1987), 0.54 for N (Brookes et al., 1985) and 0.38 for P (Brookes et al., 1982). Microbial biomass C, N and P are expressed as mg g–1 oven dry peat (105°C for 24 h). Roots were manually removed from peat samples before any treatment. The potential activity of extracellular hydrolase enzymes was measured by adding 4-methylumbelliferyl-b-D-glucoside for the activity of b-glucosidase (BG), 4-MUF-N-acetyl-b-D-glucosaminide for the activity of b-1,4-N-acetylglucosaminidase (NAG), L-leucine-7-amido-4-methycoumarinhydrochloride for the activity of leucine aminopeptidase (LAP) and 4-MUF-phosphate for the activity of phosphatase (AP) to c. 1 g of fresh peat. After incubation (1 h for BG, NAG and LAP, and 45 min for AP), the fluorescence of the supernatant was measured, after centrifugation, on a microplate reader (BioTek SynergyMX, Lucerne, Switzerland) at 450 nm emission and 330 nm excitation wavelength. To quantify product release and account for quenching effects, a set of standards was prepared using methylumbelliferone (MUF) and 7-amino-4-methylcoumarin (MCU) mixed with peat extract. Enzyme activities were expressed as lmol of substrate (MUF and MCU) converted min–1 g–1 DW of peat (Bragazza et al., 2013). The specific enzyme activity was calculated as the ratio of potential enzyme activity to the correspondent PLFA biomass as determined in each peat sample.
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Above-ground biomass (g m–2) Fig. 1 Relationship between total above-ground vascular plant biomass (mean 1 SE; n = 5) and total microbial biomass (long dashed line), and dissolved organic carbon concentration (DOC; short dashed line) in the four study peatlands located at different altitudes. Microbial biomass is based on phospholipid fatty acids (PLFAs). Each value of microbial biomass and DOC is the mean ( 1 SE) of five seasonal replicates. Linear regression for total microbial biomass, y = 0.068 + 0.0002x, r2 = 0.93, P = 0.037; linear regression for DOC, y = 6.96 + 0.13x, r2 = 0.99, P = 0.004. New Phytologist (2014) www.newphytologist.com
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water content (Fig. 2), but a positive relationship with the E400/ E600 absorbance ratio (Pearson’s r = 0.61, P < 0.05, n = 20) of pore water, indicating a dominance of fungi in the presence of DOC characterized by lower molecular weight. The coinertia analysis showed that along the first axis there was a positive correlation between PLFA markers for fungi and Gram-negative bacteria with environmental variables such as DOC concentration and E400/E600 absorbance ratio (Fig. 3). PLFA markers for actinomycetes were opposite to these variables and related more to the peat water content. Along the second axis there was a negative correlation between peat temperature and PLFA markers associated with Gram-positive bacteria. Carbon, N and P content in soil microbial biomass differed in relation to the structure of microbial communities (Fig. 4). In the two peatlands located at lower altitude (Fig. 4a), where the F : B ratio was greater, the microbial biomass (mean 1 SE) C was significantly greater than in the two higher-altitude sites 1.3
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(9.72 0.42 vs 8.15 0.41 mg g1, Student’s t = 2.8, P = 0.012, n = 10). By contrast, microbial biomass P (Fig. 4c) and, to a greater extent, microbial biomass N (Fig. 4b) were significantly greater in the two higher-altitude sites where the F : B ratio was lower (0.36 0.034 vs 0.48 0.034 mg g1 for P and 0.53 0.025 vs 0.61 0.022 mg g1 for N; Student’s t < 1.87, P < 0.076, n = 10). Differences in nutrient stoichiometry (mean 1 SE) of microbial biomass were also detected with altitude. In particular, the microbial biomass of the two lowest altitude sites had a greater C/N (17.8 0.65 vs 14.2 0.8), C/P (29.6 2.5 vs 17.1 1.2) and N/P (1.86 021 vs 1.34 0.09) ratio than at the two highest altitude sites (Student’s t < 4.3, P < 0.036, n = 10). The specific activity of C and N degrading enzymes increased with increasing microbial biomass C and N (Fig. 5a,b), whereas the specific enzyme activity of phosphatase was not related to the microbial biomass P (Fig. 5c).
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Fig. 2 Relationship between mean fungi to bacteria biomass ratio (F : B; mean 1 SE; n = 5) and peat water content (mean 1 SE; n = 5) in the four study peatlands located at different altitude. The F : B biomass ratio was based on phospholipid fatty acids (PLFAs). Linear regression for F : B ratio, y = 30.90.32x, r2 = 0.97, P < 0.05. Axis 2
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The natural decrease of air temperature and increase of precipitation with altitude were reflected in both a decrease of soil temperature and an increase of soil water content as a result of reduced evapotranspiration. The observed changes in temperature and precipitation along the altitude gradient are consistent with the future climatic scenarios for Switzerland up to 2050 (Frei et al., 2007), thereby making the selected altitude gradient a reliable substitution of time to assess above-ground and below-ground responses in peatland to climate change. Along such a natural climate gradient, above-ground plant biomass, in particular that of low shrubs (i.e. the ericoid dwarf shrubs) was enhanced as a result of higher soil oxygenation, which has previously been shown to be a key factor in controlling biomass allocation of vascular plants in peatlands (Moore et al., 2002; Bubier et al., 2006). The observed increasing cover of vascular plants is in agreement with the results of a vegetation monitoring program of ombrotrophic Swiss peatlands which reported an increase of shrubs during a decadal time frame (Klaus, 2007). We also detected a positive relationship between aboveground plant biomass and below-ground microbial biomass Peat temperature
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Fig. 3 Scatterplot of the microbial community (left) and environmental variables (right) showing their contribution to the coinertia space. Vectors pointing to the same direction within and across both the scatterplots are correlated. Axis 1 has the highest weight (eigenvalue = 2.52), whereas axis 2 is weaker (eigenvalue = 0.8). The correlation value between the two data sets is 0.26 (RV-value, a measure of global similarity between the datasets). DOC, dissolved organic carbon. Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
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Fig. 4 Microbial biomass nutrient content for carbon (C) (a), nitrogen (N) (b), and phosphorus (P) (c) in relation to the relative abundance of fungi and bacteria biomass in the four study sites. The fungi to bacteria biomass ratio (F : B ratio) was based on phospholipid fatty acid (PLFA) abundance, whereas the microbial biomass nutrient content was based on chloroform fumigation. Data are means 1 SE (n = 5). Ó 2014 The Authors New Phytologist Ó 2014 New Phytologist Trust
Fig. 5 Specific enzyme activity (mean 1 SE, n = 5) and corresponding microbial biomass nutrient for carbon (C) (a), nitrogen (N) (b), and phosphorus (P) (c) in the four study sites. Specific enzyme activity was calculated as the ratio of mean enzyme activity to mean microbial phospholipid fatty acid (PLFA) biomass, whereas the microbial biomass nutrient content was based on chloroform fumigation. Linear regression for C-specific enzyme activity, y = 26.6 + 6.61x, r2 = 0.87, P = 0.068; linear regression for N-specific enzyme activity, y = 12.9 + 63.6x, r2 = 0.88, P = 0.060. New Phytologist (2014) www.newphytologist.com
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(Fig. 1) along the altitude gradient, which is likely explained by the availability of low-molecular-weight C, which can then fuel and sustain a greater microbial biomass (Zak et al., 1994; Paterson et al., 2007; Trinder et al., 2008; Fierer et al., 2009; Gutknecht et al., 2012). Although not measured here, it is well known that above-ground plant biomass in peatlands is positively related to below-ground root biomass (Moore et al., 2002; Murphy et al., 2009; Murphy & Moore, 2010). Likely mechanisms by which greater biomass of vascular plants enhances the availability of low-molecular-weight C are through temperature-induced root exudation (Boone et al., 1998) and the relatively short life span of ericoid fine roots (Valenzuela-Estrada et al., 2008). DOC of lower molecular weight is generally more labile, that is, more easily degradable by soil microbes (Boddy et al., 2008; Saari et al., 2009; Mann et al., 2012). Accordingly, the observation of greater concentration of low-molecular-weight DOC in pore water at lower altitude, where ericoid dwarf shrubs are more abundant and peat temperature is enhanced, supports the hypothesis of a positive link between vascular plant abundance and microbial biomass through the enhanced release of labile C substrate. We detected a significant shift in soil microbial community structure with altitude, being fungi, relative to bacteria, less abundant at higher altitudes. A potential mechanism explaining this response might be the sensitivity of fungi to anoxic conditions (Jaatinen et al., 2007; Peltoniemi et al., 2009; Lin et al., 2012), given that peat water content also increased with altitude. An alternative explanation might be the reduction in ericoid biomass and their associated mycorrhizal fungi with increasing altitude (Perotto et al., 2002; Andersen et al., 2013). However, a more detailed analysis of microbial community structure based on PLFAs revealed a different response of the main microbial groups to changing abiotic and biotic conditions. Indeed, Gram-negative bacteria and, in particular, fungi were more associated with lower-altitude sites (Fig. 3), which is in accordance with their preference for more aerated soils and higher availability of labile C substrate (Jaatinen et al., 2007; Paterson et al., 2007; Garcia-Pausas & Paterson, 2011). The greater abundance of Gram-positive bacteria and actinomycetes at higher-altitude sites seems to confirm previous reports of an adaptability of these microbes to more anaerobic soil conditions as well as to recalcitrant C substrates (Griffiths et al., 1999; Fierer et al., 2003; Paterson et al., 2007; Drigo et al., 2010; Bird et al., 2011). Differences in soil aeration and quality of nutritional resources seem, then, to modulate the ability of different microbial groups to dominate the soil community in relation to their specific physiological requirements (Allison et al., 2010; Waring et al., 2013). A change in the relative abundance of bacteria and fungi in peat soil was associated with shifts in the nutrient content of the microbial biomass. We observed that with increasing F : B ratio the microbial biomass tends to be enriched in C (Fig. 4a) and depleted in N and P (Fig. 4b,c). Such a trend could reflect, in part, differences in microbial nutrient requirements, considering that fungi have a higher C demand compared with bacteria, which appear more constrained by N and P (Fierer et al., 2009; New Phytologist (2014) www.newphytologist.com
Keiblinger et al., 2010). The observed differences in nutrient biomass stoichiometry also explained differences in investment in extracellular enzymes at the community level (Fig. 5a–c). Indeed, a microbial community dominated by fungi seems to have a higher activity of C-degrading enzyme per unit of microbial PLFA biomass (Fig. 5a) compared with a bacteria-dominated community where, instead, microbial investment is towards the acquisition of N compounds (Fig. 5b). Previous studies focusing on C loss in peatlands in response to changes of abiotic conditions reported an increase of peat decomposition with increased soil temperature or reduced soil water content (Dorrepaal et al., 2009; Hardie et al., 2011), a response generally explained by an enhancement of extracellular enzymatic activity (Fenner & Freeman, 2011). Our data, however, demonstrate the important role of the interactive effects of change in vegetation cover and climatic conditions (= abiotic driver) on soil microbial structure and function (Chanton et al., 2008). In particular, our data seem to provide a microbial-based explanation for a previously observed facilitative role of fresh plant-derived C in promoting the decomposition of old organic matter (Fontaine et al., 2008; Hardie et al., 2009; Hartley et al., 2012; van Groenigen et al., 2014; Wild et al., 2014). Indeed, the observed increase of specific activity of C-degrading enzyme under greater vascular plant cover is not simply a response to greater labile C availability (Bragazza et al., 2013), but also reflects an increased C requirement of the new soil microbial biomass. Although some caution is needed for upscaling the observed patterns to different geographic areas or ecosystems, on the whole our findings highlight that the determination of the structural identity of microbial communities can help to explain the biogeochemical dynamics in response to global environmental change and better predict the microbial impacts on organic matter decomposition.
Acknowledgements We thank J. Parisod, M. Lamentovicz, T. Spiegelberger, E. Feldmeyer-Christe, P. Gomis, A. Margot, J. D. Teuscher, E. Rossel, K. Vernez, and J. D. Welch for assistance. The ‘Service des for^ets, de la faune et de la nature’ (Canton de Vaud) and the ‘Service de la promotion de la nature – Office de l’agriculture et de la nature’ (Canton de Berne) are acknowledged for giving permission to access the study sites. This study was supported financially by the Swiss National Science Foundation (project ClimaBog, grant 205321-129981 to L.B.).
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