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Connecting litter quality, microbial community and nitrogen transfer mechanisms in decomposing litter mixtures

Dennis Lummer , Stefan Scheu and Olaf Butenschoen

D. Lummer, S. Scheu and O. Butenschoen (obutens@gwdg.de), J. F. Blumenbach Inst. of Zoology and Anthropology, Georg-August-Univ. G ö ttingen, Berliner Str. 28, DE-37073 G ö ttingen, Germany.

Synergistic eff ects on decomposition in litter mixtures have been suggested to be due to the transfer of nitrogen from N-rich to N-poor species. However, the dominant pathway and the underlying mechanisms remain to be elucidated. We conducted an experiment to investigate and quantify the control mechanisms for nitrogen transfer between two litter spe-cies of contrasting nitrogen status ( 15 N labeled and unlabeled Fagus sylvatica and Fraxinus excelsior ) in presence and absence of micro-arthropods. We found that 15 N was predominantly transferred actively aboveground by saprotrophic fungi, rather than belowground or passively by leaching. However, litter decomposition remained unaff ected by N-dynamics and was poorly aff ected by micro-arthropods, suggesting that synergistic eff ects in litter mixtures depend on complex environmen-tal interrelationships. Remarkably, more 15 N was transferred from N-poor beech than N-rich ash litter. Moreover, the low transfer of 15 N from ash litter was insensitive to destination species whereas the transfer of 15 N from labeled beech litter to unlabeled beech was signifi cantly greater than the amount of 15 N transferred to unlabeled ash suggesting that processes of nitrogen transfer fundamentally diff er between litter species of diff erent nitrogen status. Microbial analyses suggest that nitrogen of N-rich litter is entirely controlled by bacteria that hamper nitrogen capture of microbes in the environment supporting the source-theory. In contrast, nitrogen of N-poor fungal dominated litter is less protected and transferable depending on the nitrogen status and the transfer capacity of the microbial community of the co-occurring litter spe-cies supporting the gradient-theory. Th us, our results challenge the traditional view regarding the role of N-rich litter in decomposing litter mixtures. We rather suggest that N-rich litter is only a poor nitrogen source, whereas N-poor litter, can act as an important nitrogen source in litter mixtures. Consequently both absolute and relative diff erences in initial litter C/N ratios of co-occurring litter species need to be considered for understanding nitrogen dynamics in decomposing litter mixtures.

Decomposition of leaf litter is a key process in terrestrial ecosystems, controlling nutrient and carbon cycling, pri-mary productivity and fundamentally contributes to the maintenance of soil fertility and ecosystem functioning (Aber and Melillo 2001, Moore et al. 2004). Litter decom-position has been studied for decades and the main factors controlling decomposition processes have been identifi ed to be climate, litter quality and the decomposer commu-nity (Aerts 1997, Berg et al. 2000, Cornwell et al. 2008). However, most previous studies focused on decomposition of single litter species which does not represent natural ter-restrial ecosystems where typically several diff erent species coexist and consequently diff erent litter species decompose in mixtures rather than in monocultures. Only recently an increasing number of studies have investigated litter decom-position dynamics in litter mixtures (Wardle et al. 1997, Hoorens et al. 2003, 2010, Gartner and Cardon 2004) and it has been shown that the mass loss of litter mixtures often deviates from the additive mass loss estimated from single litter species (Madritch and Cardinale 2007, Jonsson and Wardle 2008, Perez Harguindeguy et al. 2008). For

example, reviewing 30 studies investigating decomposition of litter mixtures Gartner and Cardon (2004) found that only in about 30% of the studies litter decomposed in accordance to an additive model. Non-additive litter decom-position in mixtures may be either synergistic or antago-nistic resulting in faster or slower decomposition of litter in mixtures as compared to the additive model. It has been suggested that primarily chemical and physical interactions between litter species contribute to non-additive eff ects (Hoorens et al. 2003, Epps et al. 2007, Liu et al. 2007).For example, according to H ä ttenschwiler et al. (2005) the transfer of inhibitory litter constitutes and nutrients caneither decelerate or accelerate the decomposition of co-occurring litter species within litter mixtures. Moreover, leaf litter structural components can alter micro-environmental conditions in litter mixtures resulting in increased or decreased decomposition rates, and synergistic or antago-nistic eff ects can occur by interactions across trophic levels. Th e under lying mechanisms are still debated. Complicatingthe picture the contribution of each mechanism to non-additive litter decomposition processes in litter mixtures

Oikos 000: 001–007, 2012

doi: 10.1111/j.1600-0706.2011.20073.x

© 2011 Th e Authors. Oikos © 2012 Nordic Society Oikos

Subject Editor: Heikki Setälä. Accepted 13 October 2011

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likely depends on climatic conditions and on traits of the litter species studied. Synergistic eff ects on decomposition in litter mixtures have been suggested to be mainly due to the transfer of nitrogen from litter species high in nitrogen to litter species low in nitrogen, but this has not been proven convincingly. For example, in a recent study Schimel and H ä ttenschwiler (2007) investigated the transfer of nitrogen between six tropical litter species of diff erent N status and, indeed, found that large amounts of nitrogen were trans-ferred between litter species. However, they were unable to identify the underlying mechanism, since nitrogen can either be transferred actively by fungi via mycelia (Frey et al. 2000, 2003) or passively by leaching and diff usion (Currie and Aber 1997, Berg and McClaugherty 2003). Recently, Tiunov (2009) proposed that the transfer of nitrogen in lit-ter mixtures is predominantly active; however, this assump-tion needs further support.

Beside litter quality and climate, the decomposer com-munity, i.e. soil microorganisms and fauna, is an important component aff ecting litter decomposition (Swift et al. 1979, H ä ttenschwiler et al. 2005). Detritivores aff ect litter decom-position in mixtures by selective feeding on certain litter species whereas secondary decomposers indirectly aff ect lit-ter decomposition via changing microbial activity (Kaneko and Salamanca 1999, H ä ttenschwiler and Gasser 2005, De Oliveira et al. 2010). In particular soil micro-arthropods such as Collembola are known to preferentially feed on mycorrhizal and saprophytic fungi (Johnson et al. 2005, Jonas et al. 2007), thereby most likely reducing the fl ow of carbon and nutrients through fungal networks and impacting on litter decomposition dynamics. However, pre-vious studies investigating nitrogen transfer pathways in lit-ter mixtures often neglected the role of soil invertebrates in litter N transfer.

Th e aims of the present study were to identify the main N transfer pathway in litter mixtures, to quantify the amount of N transferred between litter species of diff erent N status, to investigate the contribution of N transfer to litter decom-position in litter mixtures and to quantify the role of soil micro-arthropods for litter N transfer and its consequences for litter decomposition. We hypothesized that 1) N will be predominantly transferred actively by fungal hyphae, 2) N will be transferred to a greater extent from high-N litter to low-N litter, 3) mixing high-N litter with low-N litter will result in increased litter mass loss of low-N litter, and 4) microbivorous soil micro-arthropods decrease litter decomposition by cutting off hyphal connections and reduc-ing N transfer between litter compartments.

Material and methods

Soil and leaf litter

Th e soil was taken from a 130-year old beech forest on limestone near G ö ttingen (Germany, southern Lower Saxony). Long-term annual mean temperature in G ö ttingen is 7.9 ° C and annual rainfall is 720 mm. Th e soil is shallow and of Rendzina type with 29.6% clay, 61.4% silt, 6.3% fi ne sand (63 – 250 μ m) and 2.7% coarse sand ( � 250 μ m). It was of neutral pH (7.0; 0.01 M CaCl 2 , 1/2.5 w/v) with a

content of organic C of 13.0% and of organic N of 1.0%. Samples were taken from the upper 10 cm of the soil and sieved (4 mm) to remove stones, large plant residues and macrofauna invertebrates. Half of the soil was defaunated by three freezing/thawing cycles by varying the temperature between � 28 ° C and � 20 ° C (Huhta et al. 1998).

Natural leaf litter of two temperate tree species of contrasting quality based on litter C-to-N ratio (beech, Fagussylvatica , C/N ratio 41.47, 0.365 atom% 15 N; and ash, Fraxinus excelsior , C/N ratio 23.05, 0.365 atom% 15 N) were collected in the Hainich National Park (Th uringia,Germany). Briefl y, nets were placed under beech and ash trees and freshly fallen senesced leaf litter was collected over a period of one week. 15 N labelled leaf litter of beech (C/N ratio 31.48, 14.08 atom% 15 N) and ash (C/N ratio 20.89, 10.50 atom% 15 N) was obtained by growing young trees in PVC containers in a climate controlled greenhouse and watering them with 15 NH4

� (99 atom% 15 N) over a period of two years. In late autumn naturally senesced leaves were collected. Petioles of ash leaves were removed and leafl ets were retained. Unlabelled and labelled leaf litter was dried at 65 ° C for three days before being used in the experiment.

Experimental set up

Th e experiment was set up in 128 microcosms consisting of polyethylene tubes (height 350 mm, diameter 150 mm) fi lled either with 1.8 kg fresh weight (fw) defaunated or non-defaunated soil up to a height of 120 mm. To separate N transfer pathways four treatments were set up. One fourth of the microcosms (32) was equipped with vertical plas-tic sheets (120 � 150 mm) sealed to the inner PVC tube and the bottom of the microcosms with fungicide free silicone, thereby separating the soil column into two halves and allowing only aboveground transfer of nutrients (treat-ment AT). Th e second 32 microcosms were equipped with plastic sheets (70 � 150 mm) which were sealed to the inner cylinder of the PVC tubes above the soil surface to sepa-rate the litter layer vertically into two halves allowing only belowground transfer of nutrients (BT). Th e third 32 micro-cosms were equipped with plastic rings covered with gauze (5 mm mesh size) which were fi xed to the inner side of the PVC tubes 50 mm above the soil surface; the physical separation of the leaf litter from the soil surface prevented fungi to transfer N from the soil into the litter, thereby allowing to quantify the transfer of N by leaching (LT). Th e fourth 32 microcosms represented the natural condi-tions without sheets and rings, i.e. litter was in direct con-tact to the soil freely accessible to fungi from the soil and the litter (ABT). A total of 2.4 g dry weight (dw) leaf lit-ter was added to the microcosms. Unlabelled and labelled beech and ash leaf litter was added in all possible com-binations: 1) 15 N labelled beech leaves with non-labelled beech leaves ( 15 beech/beech); 2) 15 N labelled beech leaves with non-labelled ash leaves ( 15 beech/ash); 3) 15 N labelled ash leaves with non-labelled ash leaves ( 15 ash/ash); 4) 15 N labelled ash leaves with non-labelled beech leaves ( 15 ash/beech). Th e litter species were either separated horizontally by sheets (BT) or vertically by rings (LT). In the LT treat-ments only labelled litter was placed on the gauze covering

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the ring, whereas the co-occurring unlabelled litter species was placed on the soil surface below. To avoid diff usion of nutrients into other litter species, litter species added on the surface of the two belowground separated halves of the soil column of the AT and ABT treatments were spatially separated (4 cm distance between litter species). Eight addi-tional microcosms, fi lled either with 1.8 kg fresh weight (fw) defaunated or non-defaunated soil, served as control to assess eff ects of aboveground resources on soil community biomass and composition. Microcosms were closed by lids at the top and incubated at 15 ° C in a climate chamber. To simulate precipitation and facilitate leaching of nutrients, microcosms were watered with 100 ml distilled water in weekly intervals. Microcosms were equipped with ceramic lysimeters connected to a vacuum pump to drain the soil under semi-natural conditions ( � 200 to � 400 hPa).

Harvest and analytical procedure

After 104 days microcosms were destructively sampled. Th e leaf litter remaining was removed and adhering soil was brushed off carefully; then, the litter was weighed and stored at � 20 ° C until microbial analyses. Soil samples were taken from each microcosm. In the microcosms of the LT, BT and ABT treatments one soil sample was taken with a soil corer (diameter 50 mm), whereas in the AT treatment one soil samples was taken (diameter 25 mm) from each half of the separated soil column and mixed. Soil samples were sepa-rated into two subsamples of which one was used for animal extraction the other one for microbial analyses. Invertebrates were extracted by heat and stored in saturated salt solution (NaCl) at 10 ° C (Macfayden 1961, Kempson et al. 1963). Animals were counted and identifi ed using light micros-copy. Microbial biomass in the second soil subsample and litter was measured using the substrate-induced respiration(SIR) method (Anderson and Domsch 1978). Th e micro-bial respiratory response was measured in an electrolytic O 2 -microcompensation apparatus at hourly intervals for 24 h at 22 ° C (Scheu 1992). Microbial biomass was mea-sured after the addition of glucose as substrate to saturate the catabolic activity of microorganisms. Th e maximun initial respiratory response (MIRR; μ l O 2 156 g �1 dw h �1 ) was calculated as the mean of the lowest three readings within the fi rst 10 h and microbial biomass was calculated as Cmic � 38 � MIRR ( μ g Cmic g �1 dw; Beck et al. 1997). Microbial community composition in selected leaf litter subamples was measured by phospholipids fatty acid (PLFA) analysis. Extraction of lipids was performed on 2 g fresh weight litter material according to Frosteg å rd et al. (1993). Separated phospholipid fatty acid methyl-esters were identi-fi ed by chromatographic retention time and mass spectral comparison with a mixture of standard qualitative bacterialacid methyl-ester and fatty acid methyl-ester that ranged from C11 to C20. For each sample the abundance of indi-vidual phospholipid fatty acid methyl-esters was expressed per unit dry weight. Th e fatty acids i15:0, a15:0, 15:0, i16:0, 16:1 ω 7, 17:0, i17:0, cy17:0, 18:1 ω 7 and cy19:0 were assumed to represent bacteria (Zelles 1999), the con-centration of the fungal specifi c fatty acid 18:2 ω 6,9 was used as an indicator of fungal biomass (Frosteg å rd and B å å th 1996). Th e ratio of fungal-to-bacterial PLFAs was

calculated to estimate the relative abundance of these two microbial groups. Leaf litter mass loss ( D M) was calculated as D M (%) � ((m 0 � m 1 )/m 0 ) � 100, where m 0 is the dry weight of initial leaf litter and m 1 the dry weight of leaf litter at harvest. For the analysis of nitrogen concentrations and nitrogen stable isotope ratios, dried leaf litter was ground to powder and approximately 3 mg were weighed into tin cap-sules. Nitrogen concentrations and isotope ratios ( 15 N/ 14 N) of leaf litter were measured by a system of an elemental anal-yser coupled with a trapping box (type CN) and an isotope ratio mass spectrometer (MAT 251) (Reineking et al. 1993). Atmospheric N 2 served as a standard for 15 N. Acetanilide (C 8 H 9 NO) was used for internal calibration. Th e amount of 15 N excess in leaf litter were calculated by subtracting natural signatures of retained leaf litter from the measured 15 N atom% at the end of the experiment.

Statistical analysis

Data on mass loss and 15 N excess in target litter species was analyzed using four-way analysis of variance (ANOVA), with the factors litter species (ash and beech), soil status (defaunated, and non-defaunated), nitrogen translocation treatment (AT, BT, ABT and LT) and co-occurring source litter species (labelled ash and labelled beech). Data on soilmicrobial biomass and abundance of soil invertebrates were analyzed with three-way ANOVA using soil status, nitrogen translocation treatment and litter mixture ( 15 beech/beech, 15 beech/ash, 15 ash/ash and 15 ash/beech) as factors; this was done because soil samples were taken and analyzed per microcosm and soil microbial biomass and abundance of soil fauna were aff ected by both litter species. Microbial biomass data of target and source litter species were ana-lyzed with four-way ANOVA using litter species, soil sta-tus, nitrogen translocation treatment and co-occurring litter species as factors. One-way ANOVA with the factor litter mixture was used to analyze the amount of bacterial and fungal PLFAs and the fungal-to-bacterial PLFA ratio in lit-ter species. Data were log(x � 1) transformed when required to satisfy the assumption of ANOVA. Percentage data were square root arcsin-transformed. Means presented in text and fi gures represent back-transformed means of the log(x � 1) and square root arcsin-transformed data. Statistical data analyses were performed using SAS 9.2 (SAS Inst.).

Results

Litter mass loss

After 104 days mass loss of the unlabelled target species signifi cantly diff ered between beech and ash litter with mass losses of 25 and 51%, respectively (F 1,77 � 119.67, p � 0.0001). Neither the nitrogen translocation treatment, nor the labelled source litter species aff ected mass loss of the target litter species. In contrast, soil fauna aff ected litter mass loss, but this eff ect depended on the identity of the target litter species (F 1,77 � 8.26, p � 0.005); soil fauna did not aff ect mass loss of beech litter, but mass loss of ash litter was on average 24% higher in presence of soil fauna com-pared to treatments without soil fauna.

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and source litter species did not aff ect microbial biomass of the target litter species. Microbial biomass of the source litter species signifi cantly diff ered between beech and ash litter with 10 900 and 15 200 μ g Cmic g �1 litter dw, res-pectively (F 1,80 � 20.18, p � 0.0001). In addition, micro-bial biomass of the source litter species was aff ected by the nitrogen translocation treatment (F 3,80 � 9.76, p � 0.0001) and signifi cantly diff ered between BT (9600 μ g C mic g �1 litter dw) and the three others (average of ∼ 13 600 μ g C mic g �1 litter dw.

Microbial community composition

We analyzed the PLFA pattern of labelled and unlabelled leaf litter in the nitrogen translocation treatments with substantial translocation of 15 N (AT and ABT) to char-acterize the microbial community composition. In total, 30 PLFAs with a chain length up to C24 were detected. Th e amount of PLFAs ranged from 579 nmol g �1 dw in labelled beech litter in presence of unlabeled ash litter to 1790 nmol g �1 dw in unlabelled beech litter in presence of labelled ash litter, with an overall amount of 1260 nmol g �1 dw. Bacterial PLFAs (as percentages of the amount of total PLFAs detected) did not diff er between treatments, whereas the fungal specifi c fatty acid 18:2 ω 6,9 varied signi-fi cantly between treatments (F 7,8 � 46.83, p � 0.0001) and was highest in the target beech litter. As a conse-quence, the fungal-to-bacterial ratios signifi cantly diff ered between ash and source beech litter on the one side and unla-belled beech litter on the other (F 7,8 � 16.75, p � 0.0007; Fig. 2).

Soil fauna

At the end of the experiment soil fauna abundance signi-fi cantly diff ered between defaunated and non-defaunated microcosms with 730 and 14 200 ind. m �2 , respectively (F 1,124 � 158.66, p � 0.0001). Nitrogen transformation treat-ments signifi cantly aff ected soil fauna abundance; the density of soil animals in the ABT treatment (22 300 ind. m �2 ) exceeded that in the other three treatments by a fac-

Nitrogen dynamics

Nitrogen dynamics depended on the litter species and the fauna treatment (F 1,83 � 7.22, p � 0.009 for the interac-tion between factors). Independent of soil fauna, nitrogen concentration of beech litter increased by 36%, whereas nitrogen concentration of ash litter decreased by 34 and 12% in presence respectively absence of soil fauna.

15 N transfer depended on target litter species, the nitro-gen translocation treatment and the source litter species (F 3,83 � 21.90, p � 0.0001 for the interaction between these three factors). Independent of the source litter species, more 15 N was transferred into ash litter in the AT and ABT treat-ments compared to the LT and BT treatments (Fig. 1). In contrast, the source litter species did not aff ect the amount of 15 N transferred into beech litter in the LT and BT treat-ments, but signifi cantly more 15 N was transferred into beech litter in the AT and ABT treatments with labelled beech than in those with labelled ash litter (Fig. 1). Soil fauna did not aff ect 15 N transfer into the target litter species.

Microbial biomass

Soil fauna signifi cantly aff ected soil microbial biomass, but this depended on the nitrogen translocation treatment (F 3,88 � 4.18, p � 0.008). In BT (1888 μ g C mic g �1 soil dw) and AT treatments (1254 μ g C mic g �1 soil dw) soil fauna did not aff ect soil microbial biomass but signifi cantly increased it in ABT and LT treatments from 1142 to 1417 and from 1309 to 1532 μ g C mic g �1 soil dw, respectively. Surprisingly, soil microbial biomass neither diff ered between litter treatments nor between treatments with and without litter addition.

Litter microbial biomass of the target species varied signifi cantly with the nitrogen translocation treatment (F 3,74 � 5.72, p � 0.001); it increased in the order BT � LT � AT � ABT from 11 800 to 13 900 to 14 000 and 16 000 μ g C mic g �1 litter dw, respectively. However, only the BT and ABT treatments diff ered signifi cantly. Soil fauna

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Figure 1. 15 N excess (mg g �1 dw litter) in unlabelled target beech (open bars) and ash litter (closed bares) in diff erent nitrogen translocation treatments (LT, leaching; BT, belowground transfer; AT, aboveground transfer; ABT, above- and belowground transfer) originating from 15 N labelled (source) beech ( 15 beech) or ash litter ( 15 ash). Means of eight replicates (combination of both fauna treatments); means sharing the same letter do not diff er signifi -cantly (Tukey ’ s HSD test, α � 0.05).

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Figure 2. Fungal-to-bacterial ratios in labelled source litter (open bars) and unlabelled target litter (closed bars) in diff erent mixtures of ash and beech litter (the labelled species in each litter mixture is marked by ‘ 15 ’ ). Means of two replicates; means sharing the same letter do not diff er signifi cantly (Tukey ’ s HSD test, α � 0.05).

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Nitrogen transfer between litter species of different quality

In contrast to our second hypothesis, signifi cantly more 15 N was transferred from low-N beech than high-N ash litter. Remarkably, the low transfer of 15 N from ash litter was unspecifi c with almost the same amount transferred into (target) beech and ash litter. In contrast, signifi cantly more 15 N was transferred into beech than ash litter from (source) beech litter. Th is suggests that processes of N transfer fundamentally diff er between litter species of diff erent N status, presumably due to diff erences in microbial commu-nity composition.

According to the concept of Schimel and Bennett (2004) microorganisms in our experiment fi rst accessed the high-N ash litter and controlled N availability to other micro-organisms. Indeed, the high mass loss of ash litter and the low translocation of 15 N into other litter in the present study suggest high-N assimilation of bacteria dominating the microbial community of ash litter. Since the N status and the microbial community composition of unlabelled and labelled ash litter were similar, only small amounts of 15 N were translocated between them. In contrast, the amount of 15 N transferred into beech litter, though small, likely resulted from the high N assimilation of bacteria in ash litter allowing only small amounts of 15 N to be transferred by fungi dominating the microbial community of beech lit-ter. Th is suggests that the capture of N from ash litter by the fungal community of beech litter was hampered by the bacteria colonizing the ash litter, presumably by the pro-duction of inhibitory compounds (Mille-Lindblom and Tranvik 2003, Mille-Lindblom et al. 2006, Romani et al. 2006). Overall, the results suggest that the transfer of N from high-N litter is predominantly under the control of the indigenous microbial community supporting the source-control theory of Schimel and H ä ttenschwiler (2007).

In contrast to the transfer of 15 N from high-N ash litter, the transfer from low-N beech litter depended on N concentrations of the (target) litter species, supporting the gradient theory (Hillel 1998, Schimel and H ä ttenschwiler2007). As expected, only small amounts of 15 N were transferred into ash litter, suggesting that both high N avail-ability and low nutrient translocation capacity of the bacte-rial community reduced 15 N capture from beech litter. In contrast, high amounts of 15 N transferred from (source) beech into (target) beech litter, suggest higher nutrient defi ciency and/or higher translocation capacity in the lat-ter. Indeed, N concentrations of the labelled (source) beech litter exceeded that of the unlabelled (target) beech litter, whereas the opposite was true for the fungal biomass. Th is suggests that the more pronounced transfer of 15 N from the labelled to the unlabelled beech litter was due to diff erences in N concentrations determining microbial community composition and activity.

In the present study, high-N litter was generally a poor N source, whereas N translocation was more pro-nounced between litter species low in N, presumably due to the dominance of bacteria in the former and fungi in the latter. Th is suggests that both relative diff erences in litter C/N ratios but also absolute initial litter C/N ratios that control the composition and nutrient translocation

tor of about two (average of ∼ 12 000 ind. m �2 ; F 4,58 � 8.37, p � 0.0001). Surprisingly, soil fauna abundance did not diff er between litter treatments nor between treatments with and without litter addition. Th e soil fauna community was dominated by micro-arthropods of the order Collem-bola ( Isotomidae 30.6%, Onychiuridae 10.6%, Sminthuridae 10.0%) and Acari [ Phthiracaridae 9.6%, Oppiidae 6.0%, Suctobelbidae 5.3% (Oribatida); Gamasida 11.2%].

Discussion

Nitrogen translocation pathway between litter species

Synergistic eff ects of litter mixtures previously observed in litter decomposition studies (Gartner and Cardon 2004) have been commonly assigned to the transfer of N from lit-ter species rich in N to litter species low in N (Seastedt 1984, Chapman et al. 1988, McTiernan et al. 1997). However, the mechanisms responsible for the transfer of N between litter species and subsequent synergistic eff ects on litter decomposition are little understood. Knowledge of the con-tribution of the biological and physicochemical process of N transfer between litter species is needed to explain confl icting results of litter mixture experiments on litter decomposition. Further, knowledge of the dominant pathway of N transfer is likely to improve predictions on litter decomposition in litter mixtures in ecosystems of diff erent climate (e.g. precipitation) and decomposer com-munity composition (e.g. fungal-dominated versus bacterial-dominated systems). Results of this study suggest that nitro-gen is predominantly transferred actively by microorganisms rather than by physical forces via leaching, validating our fi rst hypothesis and confi rming previous suggestions (McTiernan et al. 1997, Tiunov 2009). A substantial amount of 15 N was transferred between spatially separated litter spe cies predominately horizontally rather than vertically by leaching, indicating that the physicochemical loss by leaching is only of minor importance as N transfer mechanism between litter species in mixtures.

In addition to separating biological and physicochemi-cal processes of N translocation, the present study allowed the quantifi cation of the relative importance of above- and belowground transfer processes of N in litter mixtures. As expected, N was predominately transferred aboveground whereas belowground transfer was low. Th is and the high fungal biomass in treatments with substantial transfer of 15 N suggest the transfer of N between litter species by sap-rotrophic fungi via their mycelia. Saprotrophic fungi gen-erally dominate microbial processes in litter in particular in forest ecosystems (Rayner and Boddy 1988) by produc-ing lignocellulolytic enzymes essential for the degradation of recalcitrant plant materials such as cellulose and lignin (Osono and Takeda 2002). In contrast to bacteria, they are capable of transferring nutrients between patches via their mycelia (Hart and Firestone 1991, Frey et al. 2000, 2003). Th us, our results suggest that the transfer of N, and pre-sumably that of other nutrients, between decomposing lit-ter species is predominantly due to saprotrophic fungi in the litter layer.

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forest soil used. Unexpectedly, micro-arthropods aff ected litter mass loss by directly consuming litter. However, this was restricted to the high-N ash litter suggesting that direct eff ects of soil micro-arthropods are driven by litter quality.

Acknowledgements – We thank Verena Ei ß feller and the Research Training Group 1086 at the Univ. of G ö ttingen for providing labelled ash litter and Reinhard Langel of the Centre of Stable Isotope Research and Analysis at the Univ. of G ö ttingen for stable isotope analysis.

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capacity of the microbial community in the respective lit-ter species need to be considered for understanding litter decomposition in litter mixtures.

Effect of litter mixing and nitrogen transfer on litter mass loss

We hypothesized that mixing of low quality beech litter with high quality ash litter results in increased litter mass loss of beech litter. In contrast to this expectation, litter mass loss of beech and also that of ash litter remained unaf-fected by the source litter species, although in some mix-tures large amounts of N were transferred between litter materials. Previous studies indicated that the early stage of decomposition is limited predominantly by the availability of N (Taylor et al. 1989, Berg 2000). However, neutral or even negative eff ects of increased availability of N have also been reported (Knorr et al. 2005, Kazakou et al. 2009, Liu et al. 2010), suggesting that diff erences in initial litter N concentrations or increased exogenous N availability are insuffi cient to explain synergistic mixture eff ects on litter decomposition.

As suggested previously (H ä ttenschwiler et al. 2005), mechanisms other than nutrient availability substantially aff ect decomposition processes in litter mixtures. For exam-ple, litter compounds such as phenolics form resistant complexes with proteins thereby inhibiting microbial activity and decreasing the decomposition of co-occurring litter species (Schimel et al. 1998, H ä ttenschwiler and Vitousek 2000). Further, the structure of litter species can alter micro-environmental conditions within litter mixtures and can result in increased or decreased decomposition of lit-ter species in mixtures (Hansen and Coleman 1998, Hansen 2000, Wardle et al. 2003). Th e physical separation of litter species in our experiment preclude conclusions on the role of spatial separation of litter species for litter decomposition; future studies need to address this issue.

Effect of micro-arthropods on litter mass loss

Th e soil animal community forms an important component of the soil system, aff ecting key ecosystem processes, such as primary production, litter decomposition and carbon and nutrient cycling via direct and indirect eff ects (Seastedt 1984, Lavelle and Spain 2001, Berg and Laskowski 2006). In particular micro-arthropods signifi cantly alter rates of decomposition, and carbon and nutrient cycling in terrestrial ecosystems via interactions with the microbial community (Hanlon and Anderson 1979, Petersen and Luxton 1982, Teuben and Roelofsma 1990). Th erefore, we hypothesized that soil invertebrates indirectly modify litter decomposition by grazing on microorganisms, in particular saprotrophic fungi, thereby disrupting N transfer between decompos-ing litter species. Unexpectedly, micro-arthropods neither aff ected microorganisms in the litter layer nor N transfer between litter species. However, micro-arthropods slightly increased soil microbial biomass in two of the N transloca-tion treatments but this was independent of litter addition. Th is suggests that micro-arthropods in the present study predominately relied on resources from the soil layer, such as partly decomposed organic matter present in the high carbon

7

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