Weathering of geological parent material contributes very marginally to mineral nitrogen (N) content of soils (Berg and Matzner 1997; Parton et al. 2007). Thus, decomposition of litter and mineralisation inhere particular importance for N availability in forest ecosystems (Parton et al. 2007). Decomposition has been described as a three-stage process: (i) the early stage, (ii) the late stage, and (iii) the humus-near stage, latter is characterised by a decomposition rate close to zero (Berg and McClaugherty 2003; Berg and Matzner 1997).

During the early stage, initial litter N concentrations primarily drive net N release from litter (Parton et al. 2007). A rapid mass loss of native litter N but a simultaneous net increase in N and lignin contents of the decaying litter during the early stage was generally reported (Berg and McClaugherty 2003; Laganière et al. 2010; Melillo et al. 1982; Zeller et al. 2000). Fahey et al. (2011) observed a continuous decrease in native litter 15N of 20–30 % in the first year after an addition of 15N-labelled litter to a sugar maple stand. The loss, however, was balanced by a large N incorporation from external sources into the litter that resulted in an increase in N contents of 26–32 %. External N incorporation balanced N loss over a 2-years period in three European beech forest soils and net release of N was measured in the third year (Zeller et al. 2000; Zeller et al. 2001). Translocation of N from the soil to the decaying litter by fungi accounted for 25–35 % of the exogenous N, whereas atmospheric N deposition accounted for 30–50 % in a French beech forest (Zeller et al. 2000).

In the late stage of decomposition, condensation of degradation products combined with import of nutrients produces recalcitrant humic compounds, which accumulate (Berg and McClaugherty 2003). Berg and Matzner (1997) reported a rate-retarding effect of high initial litter N and soil N concentrations during the late stage of decomposition, which was ascribed to (i) the suppressing effect of N on the formation of ligninase in white-rot fungi, and (ii) condensation reactions in which lignin incorporates N, resulting in stable chemical structures. In forest ecosystems exposed to chronic high N depositions, leaching of organic N from the decaying litter but simultaneously reduced fungal N transport, due to sufficient supply of external N by depositions, may further promote N sequestration in soil organic matter (SOM; Fahey et al. 2011). In highly N enriched acid forest soils, absence of earthworms and bioturbation, which are key factors for decomposition, additionally reduce N loss (e.g. Schaefer et al. 2009). An increase in organic layer and soil N has indeed been reported from forests exposed to chronic high N and acid depositions (e.g. Meiwes et al. 2009).

Beech leaf litter is generally rich in N, water soluble compounds, tannin, and lignin (e.g. Berg and McClaugherty 2003; Lorenz et al. 2004; Melillo et al. 1982) and European beech forests have often been exposed to chronic high N and acid depositions. Thus, rapid decomposition in the early stage but slow decomposition in the late stage and temporary N sequestration may be expected (Berg and McClaugherty 2003).

Mineral N, which is released from the decaying litter, may be utilised by micro-organisms for nitrification and denitrification processes during which nitrous oxide (N2O) may be produced either as a by-product or intermediate (Davidson et al. 2000). Nitrous oxide is a potent greenhouse gas and contributes to the catalytic decomposition of stratospheric ozone (Crutzen 1979; IPCC 2007). High emissions of N2O have been reported from European beech forest soils, which have been exposed to chronic high N depositions (Brumme and Beese 1992; Papen and Butterbach-Bahl 1999; Zechmeister-Boltenstern et al. 2002). However, one of our previous studies revealed that despite the high N2O emissions of 2.6 kg N ha−1 yr−1 from an acid beech forest soil in Germany, only 6 % of the total emissions were derived from N depositions of the recent year whereas the remaining part was derived from the soil N pool (Eickenscheidt et al. 2011). Therefore the question arises as to what extent the turnover of leaf litter N contributes to N2O emissions.

Turnover of litter N is mostly studied through mass loss observations (e.g. Berg and McClaugherty 2003) and only few studies have used stable isotope tracers (e.g. Fahey et al. 2011; Zeller and Dambrine 2011), where the emphasis has been on the early-stage of decomposition. Long-term 15N tracer studies are required to improve our present knowledge about the processes of litter N dynamics and potential long-term N sequestration in the soil.

The objective of our study was to assess the contribution of beech leaf litter N to N mineralisation, N2O emissions, and N sequestration in the soil 2–3 years and 10–11 years after litterfall application. We conducted a 15N isotope litter exchange experiment in a beech stand exposed to chronic high N depositions at the Solling, Germany.

Materials and methods

Study area

The study was conducted in a 160-yr-old beech stand (Fagus sylvatica) at the Solling in Central Germany (51° 46′ N, 9° 34′ E, 500 m a.s.l., mean temperature: 6.97 °C, annual precipitation: 1193 mm, stand deposition: 20 kg N ha−1 yr−1). The soil type is a Dystric Cambisol (pHH2O ≤ 3.6; silty loam), which derived from Triassic sandstone covered by loess. The surface organic layer is a typical moder consisting of an L (litter) layer, F (fermentation) layer, and H (humus) layer with a thickness of 1–2 cm, 4–5 cm, and 2 cm, respectively. The organic layer in total stores 2790 g C m−2 and 129 g N m−2 (Table 1). Leaf litter production amounted to 277 ± 25 g m−2 yr−1 (standard deviation sd) from 1991 to 2003 with an average N content of 1.24 ± 0.16 % (sd), which corresponded to 3.4 ± 0.4 g N m−2 yr−1 (sd; Khanna et al. 2009). The N content of live foliage was 2.46 ± 0.11 % (sd) from 1995 to 2002 and thus, withdrawal of N of live foliage during leaf senescence amounted to 51 % (Khanna et al. 2009). A detailed site description can be found in Brumme and Khanna (2009) and in Eickenscheidt et al. (2011).

Table 1 Characteristics of the surface organic layer and upper 2 cm of the Ah horizon of the beech forest at Solling, Germany

Experimental design

In December 1997, six chamber bases (area: 0.0559 m2) were installed. The bases were inserted into the organic layer. We removed the freshly fallen beech leaf litter of the recent year and replaced it by 17 g 15N-labelled beech leaf litter (0.71 ± 0.01 % N; 3.57 ± 0.07 atom% 15N), which corresponded to the average annual leaf litter production. The N content of the replaced litter was lower than of the nativ leaf litter, which was 1.2 % N in 1997, since the 15N-labelled litter was derived from beech trees that were grown in a greenhouse and irrigated by a 15N solution. Three of the chamber bases were only used for soil sampling in 1999 and the other three bases were used for N2O and 15N-N2O gas flux measurements in 1999 and 2008 as well as for soil sampling in 2008. Three additional chamber bases, which contained the native unlabelled litter fall (no exchange), were installed as control chambers for 15N-N2O flux measurements. Control soil samples for natural 15N abundances were taken randomly at the study site where no litter exchange was conducted.

N2O and 15N-N2O flux measurements

We measured N2O and 15N-N2O fluxes from January 1999 to December 2000 and from May 2007 to July 2008 using the closed chamber method (chamber vol.: 5.4 L). Three N2O gas samples were collected in evacuated glass flasks (100 mL) at three regular time intervals following chamber closure (36 min at most) using a portable gas sampler equipped with a pressure sensor (Loftfield et al. 1997). Analysis was done with a gas chromatograph equipped with an EC detector (GC 14A, Shimadzu, Duisburg, Germany). Samples for 15N-N2O measurements were removed at 0.5 and 36 min (1999–2000) and at 0.5 and 20 min (2007–08) and were stored in evacuated 100 mL glass bottles. An isotope ratio mass spectrometer coupled with a preconcentration unit was used for 15N analysis (PreCon-GC-IRMS, Thermo Finnigan MAT, Bremen, Germany). Isotope abundances were expressed as absolute proportion [atom%]. We used the Keeling plot approach to calculate the 15N-N2O abundance of the N2O emitted from the soil (Tilsner et al. 2003), which was described in detail in Eickenscheidt et al. (2011).

Soil sampling and in situ incubation

After gas flux measurements, soil sampling and an in-situ incubation of the soil was conducted within the chamber bases to determine net rates of ammonification, nitrification, and 15Nexcess ammonification in summer 2008. For each chamber, two soil samples of the surface 10 cm of the soil were taken with a stainless steel core (diameter: 5.5 cm). The sample included the surface organic layer and the surface mineral soil (Table 1). Soil from one core was removed and divided into 2 cm sections, of which soil extracts for mineral N (Nmin = NH +4 -N + NO 3 -N) and 15Nmin analyses were prepared within few hours (t0 sample). The other soil core (t1 sample) was put into a PE bag and was returned into the hole for undisturbed in situ incubation, which lasted 16 days. During the incubation, the mean temperature in 10 cm soil depth, which was logged every 2 h, was 15.1 ± 1.2 °C (standard deviation). After the incubation, soil cores were subject to the same treatment as the t0 samples. Mineral N extracts were obtained after shaking 30 g fresh mineral soil with 100 mL 0.5 M K2SO4 solution or 15 g of the organic layer with 50 mL K2SO4 solution. We analysed Nmin by using continuous flow injection colorimetry (Cenco/Skalar Instruments, Breda, The Netherlands). A modified diffusion method (Corre et al. 2003; Stark and Hart 1996) was applied for the determination of 15N in the extracted NH +4 and nitrate (NO 3 ). However, net 15Nexcess nitrification rates were not available due to failure of the isotope ratio mass spectrometer. In addition, total carbon (C) and nitrogen (Nt) contents (CNS Elemental Analyzer, Heraeus Elementar Vario EL, Hanau, Germany) and 15N abundance in Nt were determined in 1999 and 2008. Analyses of 15N abundance (in Nt and NH +4 ) were conducted with a Delta C plus IRMS (Finnigan MAT, Bremen, Germany).

Calculation of the 15N recovery and of the mean residence time

The recovery of the 15N label of the exchanged leaf litter in N2O, Nt, and net ammonification were calculated using 15Nexcess values. For calculations of 15Nexcess values, natural 15N abundances of unlabelled samples (N2O: Fig. 4; Nt: Table 1; NH +4 : 0.3641 atom% (organic layer), 0.3666 atom% (0–5 cm mineral soil)) were subtracted from total 15N abundances obtained from chambers with 15N-labelled leaf litter exchange. On average 69.8 mg 15Nexcess m−2 were applied by the leaf litter exchange in 1997 (the natural abundance of the native leaf litter was 0.365 atom% 15N in 1997). For recovery assessments the ratio of 15Nexcess masses of N2O, Nt, and net ammonification, respectively, to masses of applied 15Nexcess by litter exchange were calculated.

The mean residence time (MRT) of N (and C) in the surface organic layer (0–8 cm) was assessed using the measured mass of N (C) in the surface organic layer divided by the mass of N (C) of the annual aboveground total litterfall (6.8 g N m−2 yr−1 and 258 g C m−2 yr−1; Khanna et al. 2009).

Statistical analyses

Statistical analyses were conducted using R 2.12.0 (R Development Core Team 2010). We applied linear mixed effects models on time series data (Eickenscheidt et al. 2011). Results were given as mean ± 1 standard error and effects were considered significant where P ≤ 0.05.


Recovery and movement of 15N in the soil

Two years after 15N-labelled beech litter addition, 92 % of the 15N was recovered in the surface 10 cm of the soil, most of which (80 % of the total recovery) was found in the upper part of the F layer of the surface organic layer at 2–4 cm soil depth (Fig. 1a). Eleven years after the addition, 27 % of the 15N label was recovered with the highest recovery (15 % of the total recovery) in 4–6 cm soil depth, the lower part of the F layer (Fig. 1b). The changes of 15N-Nt-excess contents in the soil profile from 1999 to 2008, with a decrease in 0–2 cm (P = 0.038) and 2–4 cm soil depths (P = 0.006) and an increase in 4–6 cm soil depth (P = 0.001), were significant. The ratio of 15N-Nt-excess to native Nt (14 N + natural 15N) within the surface organic layer (0–8 cm) decreased from 0.03 % after 2 years to 0.01 % after 11 years. The mean residence time (MRT) of N derived from the annual aboveground litterfall in the surface organic layer averaged to approx. 19 years and was higher than the MRT of C (10.5 years) in the surface organic layer.

Fig. 1
figure 1

Distribution of soil 15N in the surface 10 cm of the soil 2 years (a) and 11 years (b) after 15N-labelled litter exchange in the beech stand. Among soil depths, means (± SE; n = 3) followed by the same letter were not significantly different (ANOVA, Tukey HSD-test at P ≤ 0.05). On the right hand side, the recovery (Rec.) of the 15N-labelled litter in soil Nt is given for each layer

Mineralisation of N and 15N

Net ammonification and mineralisation, measured at about 15 °C in the field, was low in 0–2 cm soil depth (L layer), highest in 2–4 cm soil depth (upper F layer; Fig. 2a), and decreased gradually below this depth. Net NH +4 immobilisation occurred in 8–10 cm soil depth (mineral soil). Net nitrification increased with soil depth within the organic layer but differences were not statistically significant (Fig. 2a). In total, net mineralisation amounted to 58 mg N m−2 d−1 within the surface 10 cm of the soil, half of which was nitrified.

Fig. 2
figure 2

Net ammonification and nitrification rates (a) and net 15N ammonification rates (b) of the surface 10 cm of the soil in the beech stand. On the right hand side of Fig. 2b, the daily recovery (Rec.) of the 15N-labelled litter in net ammonification is given for each soil layer. The undisturbed in situ incubation lasted 16 days and was conducted within the chamber bases of the litter exchange experiment in summer 2008. Among soil depths, means (±SE; n = 3) followed by the same letter were not significantly different (ANOVA, Tukey HSD-test at P ≤ 0.05). Letters in figure a) belong to net ammonification rates and no significant differences occurred among net nitrification rates in different soil depths. Net 15N nitrification rates were not available

Net 15Nexcess ammonification in the surface 10 cm of the soil in general followed the trend of total net ammonification and in total amounted to 6 ± 1 μg 15N m−2 d−1 (Fig. 2b). Half of the total 15N ammonification could be attributed to the 2–4 cm soil depth. During the incubation period, 0.0086 % of the 15N label was net ammonified in the surface 10 cm of the soil per day (Fig. 2b). Assuming a similar contribution of the 15N label to net nitrification than to net ammonification and assuming annual net N mineralisation rates between 0.7 and 1.0 g N m−2 yr−1 (Brumme et al. 2009), 2–3 % of the litter N was net mineralised 11 years after the litter exchange.

About 0.6 % of total net N ammonification could be ascribed to the exchanged leaf litter, assuming that the native N contained in the 15N-labelled litter participated equally in net ammonification as the 15N label. Data for net 15Nexcess nitrification were not available due to technical problems.

On average, the daily net 15Nexcess ammonification of the surface 10 cm of the soil accounted for 0.030 ± 0.012 % of the 15N-Nt-excess. The proportion was two times higher than the proportion of net native N ammonification per day of the native Nt regarding the same depth interval, which was 0.015 ± 0.005 %. However, the differences were not statistically significant. Also, no significant differences between these proportions occurred when the different depth intervals of the soil were considered separately (Fig. 3).

Fig. 3
figure 3

Ratio of daily net ammonification to Nt given for 15Nexcess derived from labelled beech litter and for native N (14 N + natural 15N) for the five depth intervals. Among soil depths, means (±SE; n = 3) followed by the same letter were not significantly different regarding 15N or native N, respectively (ANOVA, Tukey HSD-test at P ≤ 0.05). No significant differences occurred between 15N and native N regarding the different soil depths (Student’s t-test at P ≤ 0.05)

N2O and 15N-N2O fluxes

Fluxes of N2O displayed pronounced seasonality with high emissions during the summer months and low emissions during the winter months (Fig. 4). The cumulative N2O emission amounted to 1.33 ± 0.26 kg N ha−1 yr−1 in 1999, 1.54 ± 0.88 kg N ha−1 yr−1 in 2000, and 1.35 ± 0.57 kg N ha−1 yr−1 in 2007–08.

Fig. 4
figure 4

Mean (± SE; n = 3) N2O fluxes in the litter exchange experiment of the Solling beech stand from January 1999 to December 2000 (a) and from May 2007 to July 2008 (b) and mean (±SE; n = 3) 15N abundances of N2O emitted from the soil from the chambers where litter was exchanged and from the control chambers for the measurements in 1999–2000 (c) and in 2007–08 (d)

In 1999–2000 15N-N2Oemitted abundances were significantly higher in the litter exchange chambers than in the control chambers (Table 2, Fig. 4; P = 0.036). Cumulative 15N-N2Oexcess emissions were 0.29 g 15N ha−1 yr−1 in 2000 which was 0.04 % of the added 15N in leaf litter. Eleven years after the litter application, the recovery of 15N in emitted N2O decreased to 0.13 g 15N ha−1 yr−1 in 2007–08, which was 0.02 % of the added 15N in leaf litter. However, the abundances of 15N in N2O emitted from the soil in 2007–08 did not significantly differ between the litter exchange chambers and the control chambers (Table 2, Fig. 4).

Table 2 Abundances of 15N in N2O emitted from the soil of the beech stand during the measurements in 1999–2000 and 2007–08

Assuming that native N contained in the exchanged leaf litter participated equally in N2O emissions as the 15N label, then 0.3 % (2007–08) and 0.6 % (1999–00) of total annual N2O emissions were derived from total N of a single litterfall year.

In 2007–08, the daily 15N-N2Oexcess emissions amounted to 0.0002 % of 15N-Nt-excess of the surface 10 cm of the soil, which was the same as the ratio of native N2O emissions to native Nt of the surface 10 cm of the soil assuming N2O only derived from the top 10 cm of the soil.


Relocation and sequestration of litter N in the soil

Recovery of 92 % of the 15N label 2 years after the 15N-labelled beech litter application was in line with other studies and pointed to low N losses during the first years of decomposition. Zeller et al. (2000) found between 80 % and 85 % of 15N-labelled beech leaf litter 3 years after the litter application in the surface 30 cm of an acid soil that had a poorly developed moder type humus. Fahey et al. (2011) reported a recovery of over 90 % for 15N derived from sugar maple leaf litter application in the surface organic layer and surface 10 cm of the soil 1 year after litter application. However 11 years after litter application, 73 % of the initially added 15N had been mineralised, taken up by trees, leached, or emitted as gaseous N compounds in our study. Uptake of litter N by trees probably represented the main sink for litter N, since total annual N uptake by trees amount to approximately 11.0 g N m−2 compared to N leaching of 0.17 g and gaseous N emissions of 0.19 g N2O-N (Brumme and Khanna 2009) and of <0.05 g NOx-N (unpublished data) at the Solling beech stand. The low turnover rate of the N pool of the surface organic layer expressed by a high MRT of 19 years, in contrast to a MRT of 10.5 years for C, indicated high recovery of 27 % of the 15N of exchanged litter 11 years after litter addition. The annual amount of litter-N sequestration thus equals to about 0.9 g N m−2 yr−1 and significantly contributes to the mean total N sequestration of 2.1 g N m−2 yr−1 in the surface organic layer observed over a period of 35 years at the studied site (Meiwes et al. 2009). The site is characterized by high load of acid and nitrogen deposition (Meesenburg et al. 2009), which has led to an N-enriched but very acid soil with a low decomposition of the litter material and the absence of deep borrowing earthworms (Meiwes et al. 2009; Schaefer et al. 2009). Although the C/N ratio of the surface organic layer was constant over the last 35 years (Meiwes et al. 2009) the soil still is accumulating N as indicated by the low leaching losses of 0.17 g N m−2 yr−1 and input–output balances of 2.1 g N m−2 yr−1 (Brumme and Khanna 2009). The most reasonable explanation is the stabilisation process which was induced by N depositions either directly by microbial N immobilisation or indirectly by plant uptake and the production of N-enriched litter including a big increase in fructification products (Khanna et al. 2009; Brumme and Khanna 2009).

Only limited amount of the applied 15N had been relocated to the lower part of the F layer (Fig. 1), despite the huge amount of 2.8 kg C m−2 and 0.07 kg N m−2 which had been deposited through total aboveground litterfall since the beginning of the labelling experiment (Khanna et al. 2009). The surprisingly constant accumulation of 15N in the H layer over the 11 years of observation revealed the lack of bioturbation, though leaching of soluble organic and inorganic compounds between the F and H layer would have occurred. Labelled 15N accumulated in the H layer was probably derived from initially high leaching of soluble organic N during the early-stage decomposition (first 2 years) as was also observed by Zeller et al. (2000). However, the downward movement of 15N into the lower F layer might primarily be caused by litter deposition and decomposition. Limited relocation of litter 15N from the organic horizons to the mineral soil within 4–5 years after litter application was also reported from three European beech forest soils which accumulated 0.7–1.3 kg C on the mineral soil (Zeller and Dambrine 2011; Zeller et al. 2001). The slow incorporation rate of litter N into the mineral-bound soil fraction was attributed to the absence of earthworms. Bird and Torn (2006) recovered very little 13C and 15N from pine needle litter below a 8 cm thick O-horizon in which the litter was placed 2 years after the beginning of the experiment. In contrast, Fahey et al. (2011) measured a comparable fast relocation of litter 15N in the soil of a sugar maple forest with low earthworm activities. In their study, already 1 year after litter application the greatest proportion of recovered 15N was found in 0–5 cm mineral soil depth. The mineral soil was covered by only 0.36 kg C indicating a rapid incorporation of litter into the mineral soil.

Beside downward movement there is evidence that fungal hyphae may be responsible for an upward movement of N and C (Zeller et al. 2000). We observed that a significant proportion of applied 15N was still present in the L layer 2 years after litter exchange (Fig. 1a). This may be due to fungal hyphae transporting N from decaying 15N-labelled litter to fresh litter. Fungal transport of N amounted to 25–35 % of exogenous N that was incorporated into decaying litter in an acid beech forest soil with low throughfall N depositions of 6–8 kg ha−1 yr−1 whereas 30–50 % of the incorporated N might have come from deposited N (Zeller et al. 2000). In the sugar maple stand investigated by Fahey et al. (2011), 18 % of external incorporated N in fresh litter was derived from the litter of the previous year. Bird and Torn (2006) suggested that as a result of fungal N translocation, organic horizons may provide a sink for leachable litter N. However, 11 years after litter exchange this process was of less importance in the present study as was indicated by a non-significant 15N enrichment in the L layer (Fig. 1b).

Mineralisation of leaf litter N

The summed net N mineralisation of the surface 10 cm of the soil amounted to 58 mg N m−2 d−1 (Fig. 2a), similar as was reported for this site in previous studies (Brumme et al. 2009). The surface organic layer of the investigated stand was the primary horizon for N mineralisation and contributed to 77 % to total net N mineralisation of the surface 20 cm of the soil (Brumme et al. 2009). The present study showed an increased in the net ammonification and mineralisation rates within the soil profile from the L layer (0–2 cm) to the upper F layer (2–4 cm; Fig. 2), which was the most important layer for N turnover. The change in the N turnover could be attributed to the decrease in the C/N ratio from 28 to 22 (Table 1). Litter-bag studies with 15N labelled beech litter by Zeller et al. (2000) showed that the shift from prevailing immobilisation to net mineralisation of N occurred after the first 2 years of decomposition when the C/N ratio had lowered to 25. In the present study ammonification and mineralisation continuously decreased with increasing soil depth below 2–4 cm depth (Fig. 2). Due to small changes in the C/N ratio other mechanisms might be responsible for this decrease. Respiration measurements conducted with surface organic matter of an adjacent beech stand showed a strong decrease in the release of CO2 from 59 μg C ha−1 g−1 in the L layer to 12 and 3.9 μg C ha−1 g−1 in the F and H layer, respectively (Chodak et al. 2003). This clearly indicated a considerably decreased of mineralisation in the lower part of the surface organic layer probably caused by chemical stabilisation processes since physical stabilisation processes should be of minor importance in organic horizons. At the same beech stand, 15N pool dilution studies conducted by Corre et al. (2003) revealed a gross mineralisation rate that was approx. four times higher than our net mineralisation rate indicating a high microbial recycling of N. Similar results were reported from other forest soils but their relevance for the N sequestration is still not clear (Venterea et al. 2004).

About 0.6 % of total net N ammonification could be ascribed to the exchanged leaf litter and approximately 2–3 % of the litter N was net mineralised 11 years after litterfall, indicating that the (labelled) leaf litter N still participated in measureable amounts in N turnover. Compared to the native N twice as much labelled N was mineralised from the labelled N pool in the lower F layer and the H layer (Fig. 3). This demonstrated that the N from the comparably new 15N-labelled beech litter was not preferentially stabilised compared to “older” native N, which was decomposed more slowly. Results from three European beech forests, where mineralisation of litter 15N was 2–4-fold higher than mineralisation of native soil N 4–5 years after litter application, were in line with our findings (Zeller and Dambrine 2011). The 15N released from the decaying litter was mainly recovered in the coarse particulate organic matter fraction (>200 μm; Zeller and Dambrine 2011). Bird et al. (2008) also reported that 58 % of the 15N of needle litter was recovered in the light fraction, which represents the SOM fraction with the highest turnover, in a Scots pine forest during the initial 1.5 years of their study.

N2O emissions and contribution of leaf litter N

Emissions measured during our experiment were similar to the mean value of a 10-year measuring period at the Solling beech stand, which amounted to 1.9 kg N ha−1 yr−1 (Brumme and Borken 2009). The beech stand belonged to the seasonal emission type where due to low oxygen availability in the soil, caused by high microbial and root respiration and low gas diffusivity in the surface organic layer, denitrification was promoted during the summer months (Brumme et al. 1999; Wolf and Brumme 2002). Low gas diffusivity was the result of laminar undecomposed beech leaves in the L layer of the surface organic layer combined with high precipitation rate (Ball et al. 1997; Brumme and Borken 2009). Leaching of NO 3 to the transitional zone between the H layer and the surface mineral soil presumably provided the substrate for denitrification. Brumme (1995) reported that high N2O emissions occurred at temperatures of >10 °C and water tensions <200 hPa. Hence, years with high summer precipitation, as was the case for 2007 but not for 2008, exhibit high N2O emissions (compare annual fluxes of Fig. 4).

The 15N experiment indicated that the 15N-labelled beech litter N measurably contributed to total N2O fluxes. The contribution decreased from the 2–3 years measurement to the 10–11 years measurement after litter exchange due to the decrease in 15N-labelled beech litter mass. Eleven years after the litter exchange, about 1 % of the net mineralised litter N was emitted as N2O. Probably less than 1 % of the litter N was lost as N2O during 2 and 11 years after litter exchange.

The proportions of N2O-N of total N were the same for labelled 15N and native N in the surface 10 cm of the soil, despite the two times higher relative ammonification of labelled 15N than that of native N 11 years after the exchange. This result may be due to a significant contribution of native N2O produced below 10 cm depth. However, we have no information about the proportion of 15Nexcess nitrification of 15N-Nt-excess but which probably could further clarify the observation.

Beech leaf litter N of a single year contributed between 0.3 % (eleventh year) and 0.6 % (second year) of the total annual N2O emissions after litterfall. Thus, accumulated leaf litter N of the second to eleventh litterfall year only contributed approx. 5 % to total annual N2O emissions using an average emission of 0.46 % (harmonic mean) for a single litterfall year.

The observation that neither leaf litter N during the 2–11 years period nor inorganic N depositions of the recent year (contribution of 6 % to total N2O emissions; Eickenscheidt et al. 2011) were primarily responsible for high N2O emissions measured at the Solling beech stand, still raised the question of the origin of these emissions. The other sources of N2O emissions may include the first year litterfall N with possibly a rapid initial release of 15N as also was reported by Zeller et al. (2000). However, we suppose that the older N pool of accumulated older litter material and accumulated deposited N may represents the major source for N2O emissions from this soil.


Our study demonstrated that one quarter of the 15N label of initially added beech leaf litter was still present in the surface 10 cm of the soil 11 years after litter addition. The high recovery of added 15N indicated a tight microbial N cycle. Movement of beech leaf litter N through the surface organic layer was very slow. Loss of litter N from the chamber bases was mainly attributed to tree uptake whereas losses by gaseous N2O and NOx emissions and by N leaching probably accounted for less than 1 % of the litter N during the 10 years of observation. The 15N-labelled beech leaf litter still participated measurably in N turnover and N2O production and obviously was not completely stabilised 11 years after litter exchange. The contribution of beech leaf litter of a single year to total annual N2O emissions ranged between 0.3 % (eleventh year) and 0.6 % (second year). We suggest that the high N2O emissions observed in the investigated beech stand resulted from accumulated N in older humus or soil N pools, since neither leaf litterfall of the last 2 to 11 years nor N depositions of the recent year represented a significant source for N2O.