Soil–litter nitrogen transfer and changes in δ13C and δ15N values in decomposing leaf litter during laboratory incubation

Pedobiologia 56 (2013) 147–152

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Soil–litter nitrogen transfer and changes in ı13 C and ı15 N values in decomposing leaf litter during laboratory incubation Oksana L. Shilenkova, Alexei V. Tiunov ∗ Institute of Ecology and Evolution RAS, Leninsky Prospect 33, 119071 Moscow, Russia

a r t i c l e

i n f o

Article history: Received 6 December 2012 Received in revised form 16 March 2013 Accepted 16 March 2013 Keywords: Stable isotopes Litter decomposition Labile N Mycelial connection Detrital food webs Isotopic baseline

a b s t r a c t Leaf litter is the main source of nutrient and energy input into the soil. Therefore, detailed knowledge on the short-term variations in the isotopic composition of plant litter is needed for correctly estimating the “isotopic baseline” in stable isotope-based studies on detrital foodwebs. In a laboratory experiment, standardized fragments of freshly fallen leaves of Quercus robur, Ulmus glabra, and Populus tremula were incubated on the surface of natural or 15 N-labeled soil during 260 days. At the end of the experiment, the remaining mass represented 62, 53 and 50% of the initial mass for oak, elm and aspen litter, respectively. There was a small decrease in the mean ı13 C values during the initial stages of decomposition, although it was inconsistent among the three litter species tested and did not exceed 1.0‰ during the decomposition period. Calculations based on the total N content, as well as the isotope mixing model suggested that up to 50% of the total litter N was incorporated from the underlying soil. The rates of N transfer from the soil to litter were not affected by the disruption of mycelial connections between the soil and litter. As indicated by ı15 N values of filter paper placed on the soil surface, labile soil N was depleted by 6–8‰ in 15 N relative to bulk soil organic matter. However, in the given experimental settings the input of 15 N-depleted labile N from mineral soil was likely counterbalanced by an increase in the 15 N-enriched microbial biomass and bulk litter ı15 N values changed little. © 2013 Elsevier GmbH. All rights reserved.

Introduction Stable isotope analysis (SIA) became a major tool in soil biology and ecosystem studies, such as those focusing on decomposition. Leaf litter is the main source of nutrient and energy input into detrital food webs (Swift et al. 1979). Knowledge on the short-term variations in the isotopic composition of plant litter is therefore needed for correctly estimating the “isotopic baseline” in the stable isotope based studies on detrital foodwebs (e.g. Scheu and Falca 2000; Chahartaghi et al. 2005). In particular, Ponsard and Arditi (2000) emphasized the importance of a site-specific baseline correction, yet reported pronounced seasonal shifts in the bulk ı15 N values of the L litter layer at some of the study sites. Estimates of site-specific mean ı15 N and ı13 C values of fieldcollected leaf litter and of green leaves of dominant plant species often differ (e.g. Halaj et al. 2005; Okuzaki et al. 2009), although in some other studies they are virtually identical (e.g. Ehleringer et al. 2000; Zeller et al. 2007; Hyodo and Wardle 2009). The difference in ı13 C can be attributed to variation in the isotopic composition of different plant species, often related to the position of leaves within the forest canopy (Brooks et al. 1997; Hyodo et al.

∗ Corresponding author. Tel.: +7 495 958 1449; fax: +7 495 954 5534. E-mail address: a [email protected] (A.V. Tiunov). 0031-4056/$ – see front matter © 2013 Elsevier GmbH. All rights reserved. http://dx.doi.org/10.1016/j.pedobi.2013.03.004

2010b). On the other hand, the isotopic composition of plant litter decomposing on the soil surface can change with time due to a combination of various processes, including (i) changes in the chemical composition of litter, with the proportion of easily degradable compounds (often enriched in 13 C) decreasing whereas that of more 13 C-depleted compounds like lignin increasing (Ågren et al. 1996; Osono et al. 2008); (ii) the accumulation of 13 C with increasing mass loss as a result of isotope fractionation during microbial resˇ ˚ cková et al. 2000a; Feng 2002); (iii) root uptake, piration (Santr uˇ leaching and volatilization of 15 N-depleted labile N forms and the accumulation of 15 N-enriched microbial biomass (Högberg 1997); (iv) nitrogen (and possibly carbon) exchange between plant litter (relatively depleted in 13 C and 15 N) and the underlying mineral soil (relatively enriched in 13 C and 15 N) (Wedin et al. 1995; Frey et al. 2000, 2003). The available data suggest that bulk ı13 C and ı15 N values of leaf litter during middle-term trials (within 12 months) can increase, decrease or remain nearly unchanged (Melillo et al., 1989; Connin et al. 2001; Fernandez et al. 2003; Osono et al. 2008; Bragazza and Iacumin 2009; Ngao and Cotrufo 2011). Given the large differences in the rates and even directions of the above-listed processes that lead to short-term changes in the isotopic composition of decomposing litter, the discrepancies amongst the different experiments and litter species are hardly surprising. In particular, the input of labile N from the mineral soil was only rarely taken into account

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(Zeller et al. 2001). In turn, there are contrasting data on the relative 15 N-depletion (e.g. Makarov et al. 2008) or 15 N-enrichment (e.g. Coyle et al. 2009) of labile soil nitrogen as compared to total soil N. Overall, the amount of information concerning changes in the isotopic composition of litter during middle-term decomposition is very limited. Our preliminary data from several field experiments (Chigineva et al., unpublished data) suggest that, during an ecologically relevant time period (i.e. nine months from the litter fall in early October to a nearly complete litter consumption by soil fauna in late June), changes in ı13 C and ı15 N values in decomposing litter are relatively small. This analysis was, however, hampered by seasonal variations in climatic and edaphic conditions. The aim of the present study was to follow the dynamics of bulk ı13 C and ı15 N in the decomposing litter of different tree species under controlled laboratory conditions during an ecologically relevant time period. Specifically, we hypothesized that changes in bulk ı15 N values in decomposing litter are at least partly induced by the input of soil-derived N. The rates of this input are controlled by passive diffusion but also by active mycelial transport (Berg 1988; Tiunov 2009; Lummer et al. 2012). We therefore suggested that the disruption of mycelial connection between litter and underlying mineral soil should affect ı15 N values of decomposing litter, as well as rates of litter decomposition (cf. Chigineva et al. 2011). The effects of nitrogen translocation from the mineral soil to plant litter were studied using natural and also artificially 15 N-enriched soil.

Materials and methods Elm (Ulmus glabra Huds.) and oak (Quercus robur L.) litter was collected in the Central Forest State Biosphere Nature Reserve (Tver Region, 56◦ 13 N, 32◦ 47 E). Oak leaves were gathered under two trees; these two litter subsets turned out to be different in terms of ı13 C, but not ı15 N values (see below). Freshly fallen aspen (Populus tremula L.) leaves, as well as the uppermost mineral soil from a mixed deciduous forest (0–10 cm) were taken at the Malinki Biological Research Station (Moscow Region, 55◦ 45 N, 37◦ 23 E). The sampled leaves were thoroughly cleared from adhering particles and air-dried. Leaf laminae (excluding larger veins) were cut into fragments of equal dry weight (7.7 ± 0.2, 4.7 ± 0.2 and 10.3 ± 0.2 mg for elm, oak and aspen litter, respectively). Soil was sieved through a screen with a mesh size of 4 mm and defaunated through two consecutive freeze–thaw cycles (from −18 ◦ C to +25 ◦ C). Prior to defaunation, part of the soil was labeled using a 0.1% aqueous solution of 15 N-enriched (10 at%) ammonium sulphate. The amount of the N added corresponded to about 0.04% of the total N content in the initial soil. Defaunated soil was placed into plastic containers (100.5 g dry weight container−1 ) in which the soil surface was covered with a fine plastic net (mesh size of 1.5 mm). Fourteen to sixteen litter fragments of the same species were placed into each microcosm. Each fragment was covered with a 15 mm × 15 mm coverglass to insure tight contact between the litter and soil, and to facilitate the retrieval of decomposed litter. In addition to litter fragments, 10 mm × 10 mm pieces of filter paper (Melior XXI, TU 03-11-03) were placed into several containers. The paper was subsequently used to infer the isotopic composition of labile soil N (cf. Hendricks et al. 2004). Containers with soil were kept in the dark at room temperature (20–22 ◦ C) for 260 days. Soil humidity was kept at 33% of wet weight throughout the experiment. In half of the microcosms, the net with litter fragments was slightly shifted (±2 mm) once a week, whereas in the remaining containers it was left intact. The former treatment intended to destroy the mycelial connection between the decomposing litter and underlying soil. The whole experiment included a total of 12 treatments (3 litter species × 2 labeling treatments × 2

“shifting” treatments) and 16 samplings (some material collected was not subjected to isotopic analyses). About 800 litter fragments were prepared and distributed among 60 microcosms. Samples of decomposing litter were taken every second week starting with week four. Four or five fragments were taken randomly from three microcosms in each treatment. Litter samples were cleared from adhering soil particles, oven-dried for 48 h at 50 ◦ C and weighed (±2 ␮g). During the first 60 days, litter fragments collected at each sampling were used for isotopic analyses. Subsequent analyses were performed at four-week intervals. Due to rapid decomposition, sufficiently clean samples of filter paper could only be collected during 170 days. Litter and soil samples were finely powdered using a ball mill (Retsch MM 200) before isotope composition analyses. Stable isotope analyses were conducted using a Thermo-Finnigan Delta V Plus continuous-flow IRMS coupled with an elemental analyzer (Thermo Flash 1112) in the Joint Usage Center at the Institute of Ecology and Evolution RAS. The isotopic composition of N and C was expressed in the ı-notation relative to the international standard (atmospheric nitrogen or VPDB): ıX(‰) = [(Rsample /Rstandard ) − 1] × 1000, where R is the ratio of the heavier isotope to the lighter isotope. Samples were analyzed with reference gas calibrated against IAEA reference materials USGS 40 and USGS 41 (glutamic acid). The drift was corrected using an internal laboratory standard (acetanilide). The standard deviation of ı13 C and ı15 N values in USGS 40 (n = 8) was <0.15‰. The minimum acceptable signal intensity on mass 28 was set to 500 mV; this decreased the number of accepted ı15 N measurements in the filter paper due to the low total N content. Natural abundance and 15 N-labeled materials were analyzed in separate runs. Within each run, no measurable instrument memory effect was detected. Along with isotopic analyses, nitrogen and carbon contents (as %) were determined in all samples. The initial ı13 C and ı15 N values in the unlabelled soil were −27.7 ± 0.2 and 2.7 ± 0.2‰, respectively (means ± 1 SE). The addition of 15 N-labeled ammonium sulphate resulted in a significant increase in the bulk ı15 N of the soil (14.5 ± 2.4‰). Soil ı13 C and ı15 N values did not change in the course of the experiment (data not shown). The isotopic composition of carbon did not differ significantly among aspen and elm litter, and the first subset of oak litter, though ı13 C values were ca. 2‰ lower in the second subset of oak litter (Table 1). The initial ı15 N values were slightly higher in both subsets of oak litter than in the other two litter species. The rates of litter decomposition were expressed as percentage (%) of the initial mass. The dynamics of litter decomposition were described using an asymptotic model (Berg and Ekbohm 1991): Lt = m × (1 − e−kt/m ), where Lt is the accumulated mass loss (in %), t is time in days, k is the decomposition rate at the beginning of decay while m is the asymptotic level that the accumulated mass loss will ultimately reach. Therefore, the remaining litter mass at time t (Wt ) was approximated as: Wt = 100 − m × (1 − e−kt/m ). Table 1 Initial parameters of soil and litter used in the experiment. Means ± 1 SE, n = 4. C organic (%) Unlabelled soil Labeled soil Aspen litter Elm litter Oak litter subset 1 Oak litter subset 2

3.9 4.0 46.6 41.0 47.9 46.8

± ± ± ± ± ±

0.2 0.3 0.7 0.2 0.2 0.8

N total (%) 0.30 0.31 0.79 1.12 0.92 0.67

± ± ± ± ± ±

0.02 0.02 0.14 0.07 0.10 0.10

ı13 C (‰) −27.7 −27.5 −30.0 −30.2 −29.6 −32.3

± ± ± ± ± ±

ı15 N (‰) 0.2 0.1 0.4 0.4 0.7 0.4

2.7 14.5 −1.1 −1.3 −0.3 0.5

± ± ± ± ± ±

0.2 2.4 0.2 0.1 0.5 0.4

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-28

Excess N =

Nt − Ni , Nt

where Ni and Nt are total N content in an individual leaf fragment (in ␮g) initially and at the time of sampling, respectively. This model assumes the initial N is retained in decomposing litter, thus the proportion of incorporated N is likely to be underestimated. To assess the proportion of soil-derived labile N in the total litter nitrogen, we used a simple mixing model:

δ13C, ‰

The proportion of “excess N” incorporated from the soil into the litter was calculated according to the following equation:

Fig. 1. Mass loss of three litter species as approximated by the asymptotic model. Each point represents a mean value ± 1 SD.

-30 -32

y = -0.06*Ln(x) - 29.9 R 2 = 0.048

100

0

13

δ C, ‰

-28

200

300

B

-30 y = -0.11*Ln(x) - 30.0 R 2 = 0.194

-32 -34 0

-28

100

C

200

y1 = -0.16*Ln(x) - 29.6 R 2 = 0.307

300

Oak 1

-30

13

δ C, ‰

Results At the end of the experiment (day 260) the remaining mass amounted to 62 ± 2, 53 ± 4 and 50 ± 1% of the initial for oak, elm and aspen litter, respectively. The asymptotic model explained from 51% to 84% of the total variation; oak litter decomposed significantly slower than did the two other litter species (Fig. 1). Perturbed (shifted) aspen litter fragments decomposed slightly slower (k = 0.93 ± 0.04) than unperturbed ones (k = 1.29 ± 0.06), but there was no difference in the decomposition rates of the other litter species; nor were significant differences observed in the other parameters studied between the perturbed and unperturbed litter fragments. Data from both treatments were therefore pooled for further analysis. The carbon isotope composition (ı13 C) of individual litter fragments varied considerably at each sampling event (within 3‰ for

A

-34

ı15 Nt − ı15 Ni Fraction of labile N = 15 , ı Npaper − ı15 Ni where ı15 Ni and ı15 Nt are the ı15 N values of litter initially and at the time of sampling, respectively, while ı15 Npaper is the mean of three maximum ı15 N values registered in filter paper placed on the soil (treatments with the labeled soil only). The means were compared using t-test at the P < 0.05 significance level. Changes in the isotopic composition of litter with time were estimated using a non-linear regression procedure and the mean values. Data on individual litter fragments were utilized to calculate the correlation between the ı13 C values and total mass loss. All calculations were performed in Statistica 7.0 (StatSoft, Tulsa, USA).

149

y2 = -0.16*Ln(x) - 32.2 R 2 = 0.376

-32

Oak 2

-34 0

100

200

300

Days Fig. 2. Changes in the carbon isotopic composition of aspen (A), elm (B) and oak (C, two subsets shown) leaf litter during 260 days of laboratory incubation. Each point represents a mean value ± 1 SD.

each litter species). However, there was a small decrease in the mean ı13 C values during the initial stage of decomposition of elm and oak litter, but not in the aspen litter (Fig. 2). In aspen and oak litter, changes in the ı13 C values of individual fragments correlated negatively with the total mass loss, i.e. ı13 C decreased in more decomposed fragments (R = –0.221, P = 0.006 for aspen, R = –0.205, P = 0.163 and R = –0.288, P = 0.021 for two subsets of oak litter). There was no correlation between ı13 C and weight loss in elm litter (R = –0.024, P = 0.798). The total N content in decomposing litter increased substantially during the initial phase of decomposition (50–60 days). At advanced stages of decomposition (>100 days), the N content changed little and averaged 2.2 ± 0.03% in aspen, 2.3 ± 0.1% in elm and 1.9 ± 0.04% in oak litter (cf. Table 1). The proportion of “excess N” imported from the soil reached an average of 28.5 and 34.5% of the total N content in oak and aspen litter, respectively, but only 9.6% in elm litter (Fig. 3). In the filter paper placed on the soil surface, the N content varied between 0.3% and 0.9%. Labeling of soil did not affect the rates of soil-derived N incorporation into litter (data not shown).

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50

Aspen Oak Elm

Excess N, %

40 30 20 10 0

Discussion

100

0

200

300

Days Fig. 3. Changes in the amount of soil-derived excess nitrogen (in % of the total N content) in elm, aspen and oak leaf litter during laboratory incubation. Only treatments with the unlabelled soil are shown for clarity. Each point represents a mean ± 1 SE.

In the treatment with the unlabelled soil, changes in the ı15 N values were small in all litter species (Fig. 4), though there was a positive correlation between ı15 N values of aspen and elm litter and time (R = 0.228, P = 0.047 and R = 0.473, P < 0.01). A strong positive correlation (R = 0.572, P < 0.001) was observed between total mass loss and ı15 N values in elm, but not in aspen litter (P = 0.233). In contrast, the ı15 N values in oak litter tended to decrease with an increased mass loss (R = –0.308, P = 0.021). The soil-derived N accumulated in the filter paper was initially quite depleted in 15 N compared to the bulk ı15 N values of the soil and plant litter, but gradually increased with time. In the treatment with the labeled soil, all three litter species became notably enriched in 15 N already at day 28 (Fig. 4); A

3

soil

2

δ15N, ‰

1 0 -1 -2 -3 -4 -5 -6 -7 50

0

100

150

B

250

200

250

Elm

Aspen

Oak

Paper

300

δ15N, ‰

200 150 100 50 soil

0 0

afterwards the ı15 N values changed little. On average, aspen litter remained the most enriched in 15 N (ı15 N = 105.1 ± 2.4‰), followed by oak (72.5 ± 1.7‰) and elm litter (59.2 ± 1.8‰). The ı15 N values of filter paper at days 28–56 reached about 200‰, presumably reflecting isotopic composition of labile soil N, and subsequently decreased approaching the ı15 N values in plant litter. Two-source mixing model indicated that an average of 52.5%, 36.3% and 29.8% of the total N in aspen, oak and elm litter, respectively, were incorporated from the labile N pool of mineral soil.

50

100

150 Days

200

250

300

Fig. 4. Changes in the nitrogen isotopic composition of plant litter and filter paper incubated on unlabelled (A) and 15 N-labeled (B) soil. The ı15 N values of bulk soil are shown as a dashed line. Points represent the means ± 1 SD for litter and individual measurements for filter paper.

An increase in the ı13 C and ı15 N values from upper to lower litter layers and further down in the upper mineral soil is a nearly universal phenomenon. Although there is a general agreement about the main mechanisms involved in the formation of this pattern, many details are still poorly understood (reviewed in Högberg 1997; Ehleringer et al. 2000; Wynn 2007; Hobbie and Ouimette 2009). In particular, the transformation of the relatively 13 C-depleted “litter” into a 13 C-enriched “soil organic matter” remains somewhat enigmatic, as a majority of long-term studies suggest either no significant change, or a decrease in ı13 C values in decomposing plant material (Melillo et al. 1989; Osono et al. 2008 and references therein). This corroborates the emerging view that lignin and other structural compounds of above-ground litter contribute little to the formation of C pool in mineral soil horizons (Kramer et al. 2003; Rasse et al. 2005; Kramer et al. 2010). Nevertheless, the above-ground plant litter largely fuels detrital foodwebs in the soil, even though decomposer animals rely partly on freshly fixed “root carbon” (Pollierer et al. 2007) and exploit nitrogen derived from mineral soil layers (Caner et al. 2004, see below). The isotopic composition of plant litter C and N is usually used for defining the “isotopic baseline” in the SIA-based studies on the structure of soil animal food webs, but the majority of primary and secondary decomposer soil animals are considerably enriched in 13 C relative to the bulk plant litter (e.g. Ponsard and Arditi 2000; Hyodo et al. 2010a). A possible reason for this is the assimilation of easily available compounds like sugars, starch and cellulose that are enriched in 13 C relative to bulk litter (Bowling et al. 2008; Pollierer et al. 2009). Indeed, the carbon assimilated and subsequently respired during the microbial decomposition of ˇ ˚ cková et al. fresh litter usually has elevated ı13 C values (e.g. Santr uˇ 2000b; Ngao and Cotrufo 2011). A more recalcitrant lignin fraction is 13 C-depleted compared to cellulose. Changes in the relative abundance of different chemical compounds can therefore lead to shifts in the ı13 C values of decomposing litter. An increase in lignin fractions during decomposition has been shown for various litter types (but see Klotzbücher et al. 2011), though it was not necessarily associated with a decrease in the ı13 C values of bulk litter (e.g. Wedin et al. 1995; Fernandez et al. 2003; Gioacchini et al. 2006). In our study, there was a pronounced decrease in ı13 C values during the initial stage of rapid decomposition in oak, but not in aspen litter. The difference could be ascribed to a higher initial content of lignin in oak compared to aspen litter (25.5% and 15.6% respectively, data not shown), however, dissimilar results were obtained in a field study by Osono et al. (2008) where Fagus crenata litter (44% of lignin) showed much less pronounced changes in ı13 C than did Swida controversa litter (16% of lignin). As revealed in numerous experiments, the accumulation of 13 C-depleted lignin in decomposing litter is largely counterbalanced by the processes that favor the accumulation of heavy carbon (Ehleringer et al. 2000). Microbial C is consistently enriched in 13 C compared to the bulk soil or litter (e.g. Dijkstra et al. 2006; Coyle et al. 2009; Hobbie et al., 2012). The proportion of microbial C in decomposing litter can reach several per cent and often increases

O.L. Shilenkova, A.V. Tiunov / Pedobiologia 56 (2013) 147–152

with time, therefore leading to an increase in ı13 C of bulk litter tissue. In addition, some soil-derived C (enriched in 13 C) can be incorporated into microbial biomass (Wedin et al. 1995). In fact, since most of the “saprophagous” soil animals are microbivorous (Swift et al. 1979), the consumption of 13 C-enriched microorganisms is likely to account for the increased ı13 C values in their tissues (Ponsard and Arditi 2000). Overall, our data suggest that decomposition under optimum conditions does not induce strong changes in the bulk ı13 C values of decomposing litter during an ecologically relevant time period. One of the main factors controlling the rates of litter decomposition is the availability of nutrients, particularly nitrogen. The dynamics of the ı15 N values in decomposing litter is driven by several mechanisms including a preferential loss of 15 N-depleted N by root uptake, nitrification and ammonification, as well as by the accumulation of 15 N-enriched microbial biomass (Högberg 1997). The loss of easily degradable N-containing compounds from decaying litter is soon compensated for by the accumulation of soil-derived N. The mechanisms of N translocation into decomposing plant litter likely include an active transport via fungal hyphae (Berg 1988; Frey et al. 2000; Tiunov 2009; Lummer et al. 2012; but see Chigineva et al. 2011). We therefore suggested that littershifting treatment could affect the rates of N translocation and possibly also the ı15 N values of litter. However, these expectations were not supported, as most of the parameters studied were only slightly or not at all affected by the disruption of mycelial connections between the soil and litter. Most likely, this is related to the experimental conditions created in the current experiment, in particular, a high water content and a close contact between litter fragments and underlying soil. Both these factors must have favored direct diffusion of labile mineral and low-molecular organic N compounds, diminishing the significance of hyphal translocation. The changes observed in the total N content in litter, as well as the isotope mixing model, suggest that from 10% to 50% of total N in decomposing litter could be imported from the mineral soil. Such a large proportion of external N can partly be attributed to experimental setup, but a comparably significant exchange of N between leaf litter and the underlying mineral soil was confirmed in other laboratory and field studies (Berg 1988; Zeller et al. 2001). Moreover, detritivorous invertebrates assimilated at most 15% of their total nitrogen content from the decomposing litter in a field experiment by Caner et al. (2004), indicating that litter-dwelling animals largely depend on N imported from the underlying soil layers. Soil organic matter (SOM) in mineral horizons is usually considerably enriched in 15 N relative to above-ground plant litter (Hobbie and Ouimette 2009). The incorporation of soil-derived N should therefore lead to an increase in litter ı15 N values. This enrichment can further be translated into increased ı15 N values of litter-decomposing animals. Nevertheless, no 15 N enrichment of the L litter layer relative to green leaves is usually observed in field studies (Zeller et al. 2007; Hyodo and Wardle 2009; Hyodo et al. 2010b), while the ı15 N values of primary litter decomposers are often lower than expected (Ponsard and Arditi 2000; Scheu and Falca 2000; Klarner et al. 2013). In our experiment, the dynamics of ı15 N in filter paper suggests that mobile soil N initially entering N-devoid organic matter was strongly depleted in 15 N. This conforms to field observations of the dominating mineral N pool (NH4 -N) being depleted in 15 N as compared to bulk SOM (Makarov et al. 2008). On the other hand, the ı15 N values of decomposing filter paper gradually approached those of litter and soil (Fig. 4A), likely reflecting a decreased fraction of labile soil N and an increased proportion of N in microbial biomass. Similarly, the ı15 N values correlated positively with mass loss in the nitrogen-rich elm litter that incorporated the lowest amount of labile soil N. Generally, is seems unlikely that, at the natural levels of ı15 N variation, redistributions

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of nitrogen among different N pools at the soil–litter interface can strongly affect the bulk ı15 N values of decomposing litter until a large proportion of N is removed from the system (e.g. via root uptake, volatilization or leaching; Högberg 1997). To conclude, our study demonstrated a small decrease in bulk ı13 C values during the initial stages of decomposition of leaf litter. This decrease was however inconsistent among the three litter species tested and in any case did not exceed 1.0‰ during the entire 260-day long decomposition period. Calculations based on the total N content, as well as the isotope mixing model suggested a substantial incorporation of labile soil nitrogen into decomposing litter. Given the natural levels of ı15 N variation, the input of 15 N-depleted labile N was likely counterbalanced by an increase in 15 N-enriched microbial biomass while the bulk litter ı15 N values changed little. Acknowledgments This study was supported by the Russian Foundation for Basic Research (Project No. 11-04-00948) and the “Wildlife” Program of the Russian Academy of Sciences. References Ågren, G.I., Bosatta, E., Balesdent, J., 1996. Isotope discrimination during decomposition of organic matter: a theoretical analysis. Soil Science Society of America Journal 60, 1121–1126. Berg, B., 1988. Dynamics of nitrogen (15 N) in decomposing Scots pine (Pinus sylvestris) needle litter. Long-term decomposition in a Scots pine forest VI. Canadian Journal of Botany 66, 1539–1546. Berg, B., Ekbohm, G., 1991. Litter mass loss rates and decomposition patterns in some needle and leaf litter types. Long-term decomposition in a Scots pine forest VII. Canadian Journal of Botany 69, 1449–1456. Bowling, D.R., Pataki, D.E., Randerson, J.T., 2008. Carbon isotopes in terrestrial ecosystem pools and CO2 fluxes. The New Phytologist 178, 24–40. Bragazza, L., Iacumin, P., 2009. Seasonal variation in carbon isotopic composition of bog plant litter during 3 years of field decomposition. Biology and Fertility of Soils 46, 73–77. Brooks, J.R., Flanagan, L.B., Buchmann, N., Ehleringer, J.R., 1997. Carbon isotope composition of boreal plants: functional grouping of life forms. Oecologia 110, 301–311. Caner, L., Zeller, B., Dambrine, E., Ponge, J.F., Chauvat, M., Llanque, C., 2004. Origin of the nitrogen assimilated by soil fauna living in decomposing beech litter. Soil Biology and Biochemistry 36, 1861–1872. Chahartaghi, M., Langel, R., Scheu, S., Ruess, L., 2005. Feeding guilds in Collembola based on nitrogen stable isotope ratios. Soil Biology and Biochemistry 37, 1718–1725. Chigineva, N.I., Aleksandrova, A.V., Marhan, S., Kandeler, E., Tiunov, A.V., 2011. The importance of mycelial connection at the soil–litter interface for nutrient translocation, enzyme activity and litter decomposition. Applied Soil Ecology 51, 35–41. Connin, S.L., Feng, X., Virginia, R.A., 2001. Isotopic discrimination during long-term decomposition in an arid land ecosystem. Soil Biology and Biochemistry 33, 41–51. Coyle, J.S., Dijkstra, P., Doucett, R.R., Schwartz, E., Hart, S.C., Hungate, B.A., 2009. Relationships between C and N availability, substrate age, and natural abundance 13 C and 15 N signatures of soil microbial biomass in a semiarid climate. Soil Biology and Biochemistry 41, 1605–1611. Dijkstra, P., Ishizu, A., Doucett, R., Hart, S.C., Schwartz, E., Menyailo, O.V., Hungate, B.A., 2006. 13 C and 15 N natural abundance of the soil microbial biomass. Soil Biology and Biochemistry 38, 3257–3266. Ehleringer, J.R., Buchmann, N., Flanagan, L.B., 2000. Carbon isotope ratios in belowground carbon cycle processes. Ecological Applications 10, 412–422. Feng, X.H., 2002. A theoretical analysis of carbon isotope evolution of decomposing plant litters and soil organic matter. Global Biogeochemical Cycles 16, Art. No. 1119. Fernandez, I., Mahieu, N., Cadisch, G., 2003. Carbon isotopic fractionation during decomposition of plant materials of different quality. Global Biogeochemical Cycles 17, Art. No. 1075. Frey, S.D., Elliott, E.T., Paustian, K., Peterson, G.A., 2000. Fungal translocation as a mechanism for soil nitrogen inputs to surface residue decomposition in a notillage agroecosystem. Soil Biology and Biochemistry 32, 689–698. Frey, S.D., Six, J., Elliott, E.T., 2003. Reciprocal transfer of carbon and nitrogen by decomposer fungi at the soil–litter interface. Soil Biology and Biochemistry 35, 1001–1004. Gioacchini, P., Masia, A., Canaccini, F., Boldreghini, P., Tonon, G., 2006. Isotopic discrimination during litter decomposition and ı13 C and ı15 N soil profiles in a young artificial stand and in an old floodplain forest. Isotopes in Environmental and Health Studies 42, 135–149.

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