Draining peatlands for agriculture induces peat decomposition, subsidence, and carbon (C) and nitrogen (N) losses, thereby contributing to soil degradation and climate change. To sustain the agricultural productivity of these organic soils, coverage with mineral soil material has increasingly been used. To evaluate the effect of this practice on the N flows within the plant–soil system, we conducted a 15N tracer experiment on a drained peatland that was managed as an intensive meadow. This peatland was divided into two parts, either without (reference “Ref”) or with ~ 40 cm mineral soil cover (coverage “Cov”). We applied 15NH415NO3 on field plots to follow the fate of 15N in plant–soil system over 11 months. In addition, N mineralization was determined by laboratory incubation. The field experiment showed that Cov lost less 15N (p < 0.05) than Ref, even though plant 15N uptake was similar at both sites. The lower net N loss from the Cov site was accompanied by higher soil 15N retention. The laboratory incubation revealed a ~ 3 times lower N mineralization at Cov than at Ref, whereas the N release per unit soil N was around two times higher at Cov than at Ref, suggesting a faster SOM turnover rate at Cov. Overall, the mineral soil cover increased the retention of fertilizer-N in the soil, thus reducing the system N losses. Our result indicates that agricultural production on drained peatland is less harmful to the environment with mineral soil coverage than using drained peatland directly.
Although peatlands only cover approximately 3% of the terrestrial surface area, they are an essential soil organic matter (SOM) pool and can store 8–14 Gt N globally (Yu et al. 2010; Loisel et al. 2014; Leifeld and Menichetti 2018). However, long-term drainage for agricultural production has already resulted in ~ 51 Mha degraded peatlands worldwide, with the highest share occurring in tropical and temperate regions, where around half of the initial peatland surface has been disturbed due to agricultural production, forestry, or peat extraction (Kasimir et al. 2018; Leifeld and Menichetti 2018). Peatland degradation is typically associated with peat decomposition, which result in C and, to a smaller extent, in N losses, as well as strong soil subsidence. As a consequence, soil C to N ratios decrease (Klemedtsson et al. 2005; Leifeld 2018). The decomposition of peat is a substantial contributor to the N supply for agricultural production in drained peatlands. Therefore, the soil N supply and plant N uptake from drained peatlands might be higher than in mineral soil.
Around 30% of the agriculturally used peatland is managed as grassland globally (Leifeld and Menichetti 2018; Evans et al. 2021). For the temperate zone, plant N uptake in grasslands has been widely explored in both mineral soils and organic soils. It has been reported from grasslands on mineral soil that plant biomass accumulated up to ~ 130 kg N ha−1 year−1 without fertilization in Germany (Bessler et al. 2012). Müller et al. (2011) found that, based on a 38-year field observation in Germany, the aboveground grass N uptake ranges from 50 to 200 kg N ha−1 year−1 with N application of ~ 200 kg N ha−1. In a study on grasslands on organic soil, Sonneveld and Lantinga (2011) reported an aboveground grass N uptake of 342 kg N ha−1 based on a 3-year field experiment in drained peatland without fertilization in the Netherlands. Schothorst (1977) even reported an aboveground grass N uptake of ~ 400 kg N ha−1 from a non-fertilized drained peatland in the Netherlands. These data tentatively suggest that plant N uptake in drained organic soil might be generally higher than in mineral soil, which might be related to the higher soil N supply in drained peatland through organic matter decomposition. Higher soil N mineralization often leads to a supply of N exceeding grass uptake, which consequently results in greater N losses to the environment of grass produced on drained organic soil compared to production on mineral soil (Pijlman et al. 2020). It has been estimated that with the ongoing agricultural use of degraded peatland, 9.7 Mt N year−1 will be released to the environment annually, and c. 2.3 Gt N will be released cumulatively with the full degradation of all currently managed peatland (Leifeld and Menichetti 2018). Therefore, it is vitally important to evaluate how the N losses from drained peatland can be reduced.
In order to compensate for continued soil subsidence of drained organic soils and thereby to maintain agricultural productivity, adding mineral soil as a cover fill with a thicknesses of 0.2–0.5 m on the surface of organic soil has increasingly been adopted by farmers working in Switzerland and other European countries (Schindler and Müller 1999; Ferré et al. 2019). With mineral soil cover, the soil N balance of drained peatlands may change due to various factors. First, the smaller surface soil C and N content in the mineral soil cover material supports smaller microbial biomass and microbial activity (Wardle 1998). This may result in lower SOM mineralization rates with mineral soil coverage compared with the surface soil from non-covered drained organic soil. Second, mineral soil cover might increase fertilizer N retention in drained peatland owing to its overall smaller N content. Third, a cover fill may also change other physical–chemical soil properties (e.g., clay content, soil pore volume, and soil cation exchange capacity) that feedback into soil N dynamics (Barrett and Burke 2002). Finally, for the peat layer underneath the mineral soil coverage, the addition of mineral soil material may compress the peat layer and push it deeper into zones with lower oxygen availability, thereby reducing the mineralization of easily degradable N in those peat layers. A prior study conducted at the same site as studied here proved that mineral soil cover induced a substantial reduction of N2O emissions (Wang et al. 2022), underpinning a strong influence of mineral soil coverage on the N balance in the soil–plant system of the drained peatland. However, a mechanistic understanding of the impact of mineral soil cover on the N cycling in the plant–soil system of drained organic soils is still missing.
In this study, we examined the N dynamics and N loss in plant–soil system in a drained peatland under grassland use both with and without mineral soil coverage. We did so by using isotopically labeled 15N fertilizer in combination with measurements of the corresponding N pools in soil, roots, and harvests. The application of 15N-enriched fertilizer is considered a useful and targeted tool for tracing the fate of applied N in plant–soil systems (Rahman and Parsons 1999; Wesselsperelo et al. 2006; Sebilo et al. 2013; Rowlings et al. 2016; Kalu et al. 2021). The specific objectives of this study were to (1) determine the fertilizer N recovery and fertilizer allocation in the plant–soil system in drained peatland with and without mineral soil coverage; (2) assess the soil mineral N (N and 15 N) release in drained peatland with and without mineral soil coverage; and (3) quantify the impact of mineral soil cover on the total plant–soil system N loss from drained peatland.
Materials and methods
The field experiment was carried out in the Swiss Rhine Valley, at the site Rüthi (47° 17′ N, 9° 32′ E), a drained fen with a peat thickness of ~ 10 m. The site has a cool temperate-moist climate with a mean annual precipitation of 1297 mm and a mean annual temperature of 10.1 °C (1981–2010, https://www.meteoswiss.admin.ch; for precipitation and temperature during the experimental period please see Fig. S1). The site was drained with ditches before 1890 (https://map.geo.admin.ch). In 1973, an intensive drainage system with pumps and pipes was built. The site was used as pasture, and since 2013 as an intensively managed meadow. From 2006 to 2007, one part of the field (~ 2 ha) was covered with mineral soil material (without mixing with the peat underneath) to improve the agricultural usability. We established the field experiment at this mineral soil coverage site (Cov, with mineral soil coverage thickness ~ 40 cm) and used the adjacent drained organic soil (~ 9 ha) without mineral soil coverage as the reference (Ref, see Fig. S2 A). The basic soil properties for both sites are provided in Table 1. Both sites have similar vegetation and identical farming practices with 5–6 cuts per year and ~ 230 kg N ha−1 fertilizer application, both as slurry (applied with a splash plate) and as mineral fertilizer (ammonium nitrate or ammonium sulfate, applied with fertilizer sprayer). The atmospheric N deposition at the study site for 2015 is 20–30 kg N ha−1 year−1 (Rihm and Künzle 2019). Dominant grass species are Lolium perenne, Alopecurus pratensis, Festuca arundinacea, Trifolium spec., and Festuca pratensis.
Experimental design and field management
The study was conducted from July 2020 to July 2021. In July 2020, eight separate plots (four for Cov, four for Ref; size, 3.5 m × 1.5 m) were distributed on the experimental site. Each plot was divided into two subplots (1.5 m × 1.5 m, Fig. S2.B), and the distance between the two subplots was 0.5 m. At each plot, one subplot received 15N double-labeled ammonium nitrate (15NH415NO3) as a treatment plot, and the other one received the same amount of non-labeled ammonium nitrate (NH4NO3) as a control plot. For these treatment plots, 15NH415NO3 was dissolved in water, and the salt solution was applied in three campaigns at the same time the farmer fertilized the overall field. We dissolved 1.35 g, 1.35 g, and 0.8 g 98 atom %15 N 15NH415NO3 in 2.25 L water for each application, which is equivalent to 1 mm precipitation per application. The control plot (i.e., the plot without a label) always received the same amount of NH4NO3 solution. The additional N input rate was chosen based on the typical field fertilizer N application. A total of extra 15N input of 0.57 g N m−2 for each plot were chosen; this was equivalent to ~ 2.5% of the regular field fertilizer N input, and it was assumed that this small extra dose would not cause a major disturbance in the N cycle of the ecosystem. The plot received the same regular fertilizer as the overall field. Extra 15NH415NO3 salt solution was sprayed directly onto the ground on 10 September 2020, 25 March 2021, and 13 May 2021. In order to spread the fertilizer solution homogenously, each subplot was divided into 15 units (0.3 m × 0.5 m, Fig. S2.C), and the same amount of the fertilizer solutions was applied to each unit.
Plant and soil sample collection and analysis
During the experimental period, soil and plant samples were taken 1 day before the regular field harvest events in October of year 2020, May, June, and July of year 2021. In addition, extra soil samples were taken on August 2020 and August 2021. Soil samples from the first sampling were used to determine the background 15N signature over all experimental plots, and those from the last sampling were used to determine the soil bulk density from 0–5 cm to 5–15 cm. For each subplot, composite soil samples from 3 units were collected by using a 6.5-cm-diameter corer for 0–20 cm depth and a 2.6-cm-diameter corer for 20–60 cm depth. Those samples were divided into 4 layers: 0–5 cm, 5–15 cm, 15–30 cm, and 30–60 cm. The samples from Cov were additionally divided at the boundary of mineral soil cover and underlying peat by a stainless steel knife during field sampling. After sampling, soil samples were stored at 4 °C in a cooling room overnight, and visible root and stones were removed from composite soil samples the next day. Soil samples were then dried at 105 °C for 72 h, ground with mortar and pestle, milled in a ball mill (Retsch, MM 400, Germany) at 25 rotation s−1 for 3 min, and finally loaded in a tin capsule to determine the soil N and 15N content via elemental analysis isotope ratio mass spectrometry (EA-IRMS) (vario PYRO cube, Elementar, Germany and isoprime precisION, Elementar, Germany).
For each subplot, aboveground biomass samples from three units were harvested by grass clippers to a height of 3 cm. At the same unit, root samples were collected by taking soil cores with a 6.5-cm-inner-diameter corer down to a depth of 20 cm. Composited grass samples were dried at 60 °C in the oven for 72 h to determine the dry biomass and to calculate the aboveground biomass based on the covered area of the three units (0.15 m2 each). Dried plant samples were cut into small pieces, milled in a ball mill (Retsch, MM 400, Germany) at 25 rotation s−1 for 3 min, and then loaded into a tin capsule to determine the grass N and 15N content with elemental analysis isotope ratio mass spectrometry (EA-IRMS) (vario PYRO cube, Elementar, Germany and isoprime precisION, Elementar, Germany). Roots were extracted from each soil core in the lab. To do so, soil material from the soil core was removed by hand, and the remaining root from the removed soil material was picked out. The left soil core and the roots that were picked out from the soil were submerged in distilled water for 2–3 h, and then put on a fine mesh screen to be washed with a gentle water shower until the residual soil material was removed. The bare roots were dried at 60 °C in the oven until the constant weight (~ 48 h) to determine the dry biomass and calculate the biomass of the root based on the covered area of the three soil cores (0.033 m2 each) for each subplot, and then, the root N and 15 N content was determined by following the same procedure described above.
To determine the net N and 15N mineralization rate of the surface soil at both sites, the 0–5 cm and 5–15 cm soil samples, which were collected in October 2020, were incubated for 28 days. Five duplicated (n = 160) soil samples equivalent to 10 g dry soils were weighted into 50-ml PET containers with soil moisture adjusted to 60% of their water holding capacity. Water holding capacity was determined following Franzluebbers (2020). The PET containers were incubated at 25 °C, and soil moisture was adjusted every 2 days by adding distilled water. After 0, 7, 14, 21, and 28 days of incubation, the soil samples were suspended in 80 ml 0.01 M CaCl2 salt solution to extract soil N and 15N (Steffens et al. 1996), shaken at 160 cycles min−1 for 30 min, and filtered. Total N and 15N from the soil extracts were determined by EA-IRMS (vario TOC cube, Elementar, Germany and iso TOC cube, Elementar, Germany). Daily net N and 15N mineralization rates (Nr_min, mg N kg−1 soil day−1; 15Nr_min, mg 15N kg−1 soil day−1) from two sites and depth were calculated based on the regression slope of the five total dissolved N (mg N kg−1 soil) and 15N (mg 15N kg−1 soil) against their incubation times (day). The specific N and 15N mineralization rate (specific Nr_min, mg N g−1 soil N day−1; specific 15Nr_min, mg 15 N g−1 soil 15 N day−1) was calculated as the regression slope of the five specific dissolved N (mg N kg−1 soil N) and 15 N (mg 15N kg−1 soil 15N) against their incubation times (day). The specific extractable N was determined as the ratio of the total extractable N (mg N kg−1 soil) and the soil N content (%). Correspondingly, the specific extractable 15N was determined as the ratio of the total dissolved 15N (mg 15N kg−1 soil) and the soil 15N content (%).
Isotope calculation and statistics
The N isotope ratios of the samples are presented by using the δ notation (Fry 2006).
where Rsample and Rstandard are the ratios between 15N and 14N of the sample and the standard, respectively. Here, atmospheric N2 is used as a standard with Rstandard = 0.003665 (Mariotti 1983).
The isotope enrichment in the sample from the treatment plot (δ15Nsample) is expressed as 15N enrichment relative to that of the control plot (δ15Ncontrol).
The recovery of the 15N fertilizer in the labeled N pools is calculated as follows:
where %15Nsample is 15N atom percent in the soil sample from the labeled plot; %15Ncontrol is 15N atom percent in the corresponding control plot; Mlabel is the amount of the 15N applied to the treatment plot (g 15N m−2); and %15Nlabel is the 15N atom percent in the labeled fertilizer, and Mpool is the N amount of the labeled pool (g N m−2). In this study, there are three labeled pools, Mpool,grass, Mpool,root and Mpool,soil. Mpool,grass and Mpool,root were determined based on the dry biomass for each plot. Mpool,soil was calculated as follows:
BDi is the soil bulk density (g cm−3) at four different soil depths (0–5 cm, 5–15 cm, 15–30 cm, and 30–60 cm). Soil bulk density from 15–30 cm to 30–60 cm was determined plot wise based on the correlation between soil bulk density and the soil organic carbon from Wang et al. (2021). Li is the thickness of each depth (cm), and Nsample is the total N content of the soil sample from the labeled plot (%).
The mass balances of 15N in the system were used to quantify the 15N recovery in the system; any 15N which was not retained in the plant and soil system was defined as losses. Plot-based 15N losses (Nlosses) were calculated as the difference between 15N input through 15N tracer application, as well as the N output from harvest and the 15N retained in soil and roots. The 15N input through regular fertilizer and atmospheric 15N deposition is accounted for by the 15N abundance from the associated control plot. The cumulative N losses at each harvest event are calculated as follows:
where Nlosses, i is the N losses after the i th harvest event, n = 1, 2, 3, 4; 15Nfer is the cumulative 15N input through fertilization; 15Ngrass is the cumulative 15N uptake through harvest; and Nroot, i and Nsoil, i are the 15N retained in roots and soil at the i th harvest event, respectively.
Statistical analysis and data visualization were performed using the open source software R (version 4.1.3). Significant differences between the two sites for soil and plant N content, δ15N content, 15N enrichment, net N mineralization rate, 15N recovery, and 15N losses were determined using a t-test. Significant differences in the ratio of the specific 15Nr_min: Nr_min, N and 15N release rate, and specific N and 15N mineralization rate in soil layers 0–5 cm and 5–15 cm between the two sites were determined through ANOVA. In case of a significant effect, a Tukey’s HSD test was performed for multiple pairwise comparisons between different sampling dates. The error probability was set as p < 0.05. The results were always reported as mean ± 1 standard error (se).
Effect of mineral soil coverage on plant biomass, N uptake, and plant 15N enrichment
During the experimental period, the cumulated grass yield was not different between sites (Table 2); only in June 2021, the yield at Cov was higher (p < 0.05) than at Ref. The harvested grass took up 274.34 ± 22.78 kg N ha−1 from the mineral soil coverage (Cov) site over four harvest events, which was significantly higher than that taken up from the drained peatland (Ref) site (229.97 ± 10.56 kg N ha−1). The higher grass N uptake of Cov was not observed in all harvest events; it was significant only for the harvest in June 2021, whereas for the rest of the cuts, no significant difference between Cov and Ref was found. The 15N enrichment of the grass varied largely between the different harvest events. It was higher at Ref compared to Cov in October 2020, whereas no significant differences were found for the other harvest events.
For roots, the differences in biomass, N uptake, and 15N enrichment were not constant between the two sites. In June 2021, root biomass and 15N enrichment were significantly higher (p < 0.05) at Cov than at Ref, and a significantly higher root N content was found at Ref in July 2021.
Effect of mineral soil coverage on soil 15N enrichment
Applications of 15N labeled fertilizer induced an increase in the soil 15N signature. At both sites, the highest soil 15N signature was found in July 2021 after the three labeling events were finished, although no 15N tracer was applied directly before that sampling event. At 0–5 cm soil depth, the 15N enrichment was 146.0 ± 13.3‰ at Cov and 49.4 ± 13.7‰ at Ref (Fig. 1D). At 5–15 cm soil depth, the 15N enrichment was 32.7 ± 8.8‰ at Cov and 7.4 ± 1.4‰ at Ref (Fig. 1D). The higher 15N signature was only found at the surface 0–30 cm, below 30 cm depth, and the soil 15N enrichment was similar to the value prior the 15N tracer application, which was near zero (Fig. 1). The surface (0–5 cm, 5–15 cm) soil 15N enrichment was higher (p < 0.05) at Cov than at Ref, whereas below 15 cm, the difference in 15N enrichment between the sites was less pronounced at any sampling date (Fig. 1).
At Cov, the 15N signal moved from the surface soil to the deeper (15–30 cm) layer during the growing season and resulted in a slightly higher 15N enrichment (49.4 ± 13.7‰) at 15–30 cm depth compared with the upper layer (32.7 ± 8.8‰) at the last sampling date (July 2021). However, no such trend was found for Ref (Fig. 1D).
The effect of mineral soil coverage on 15N recovery from drained organic soil
During the experimental period, the recovery of 15N in plants (aboveground biomass and roots) was not different between Cov and Ref. The cumulative tracer exports through aboveground biomass harvest accounted for 32.2 ± 2.2% and 30.0 ± 0.3% of the applied 15N for Cov and Ref, respectively (Fig. 2A). Roots took up 2.5 ± 0.3% and 3.9 ± 0.5% of the applied 15N from Cov and Ref, respectively, after the three labeling events were finished (Fig. 2A). Hence, a significant part of the applied 15N was not used by the plants. A share of 10–20% was incorporated into the soil N pool. At site Cov, 19.8 ± 2.0% of the tracer remained in the soil N pool, more (p < 0.05) than at Ref (9.8 ± 3.2% see Fig. 2B). Overall, site Cov showed smaller N losses (p < 0.05) compared to Ref. At Cov, 45.4 ± 3.0% of the applied labeled mineral fertilizer was lost outside the plant–soil system boundary of the study, whereas at Ref, the loss accounted for 56.2 ± 3.1% (Fig. 2C).
The average amount of soil Nr_min was significantly higher (p < 0.05) at Ref (6.07 ± 0.84 mg N kg−1 day−1) than at Cov (2.10 ± 0.15 mg N kg−1 day−1) at the 0–5 cm depth (Fig. 3A). In addition, in the deeper layer, the average amount of soil Nr_min was significantly higher (p < 0.01) at Ref (5.45 ± 0.26 mg N kg−1 day−1) compared to Cov (1.71 ± 0.13 mg N kg−1 day−1; Fig. 3B). A similar trend to that of the total N release was found for the average release of soil 15N. Ref released 15N at higher rates (0.027 ± 0.001 mg 15N kg−1 day−1 at 0–5 cm; 0.022 ± 0.002 mg N kg−1 day−1 at 5–15 cm) than at Cov (0.009 ± 0.001 mg 15N kg−1 day−1 at 0–5 cm; 0.005 ± 0.0004 mg N kg−1 day−1, at 5–15 cm) (Fig. 3C and D).
In addition, the average amount of N and 15N release showed no difference at the 0–5 cm depth for the two sites (73.67 ± 6.27 mg N m−2 day−1, 0.31 ± 0.02 mg 15N m−2 day−1 at Cov and 77.71 ± 8.08 mg N m−2 day−1, 0.35 ± 0.04 mg 15N m−2 day−1 at Ref, respectively). However, it was significantly (p < 0.05) higher at Ref (204.68 ± 10.53 mg N m−2 day−1, 0.81 ± 0.06 mg 15N m−2 day−1) than at Cov (155.70 ± 15.41 mg N m−2 day−1, 0.48 ± 0.05 mg 15N m−2 day−1) at the 5–15 cm depth (Table 3).
Specific N mineralization
The specific soil N mineralization (specific soil Nsr_min) was significantly higher with mineral soil coverage (Table 3) for both soil layers (0–5 cm and 5–15 cm). Throughout 28 days of incubation, the specific soil Nsr_min at site Cov was 0.60 ± 0.07 mg N g−1 N day−1 and 0.35 ± 0.03 mg N g−1 N day−1 at Ref at a soil depth of 0–5 cm. A similar difference was also found at 5–15 cm soil depth, where Cov released 0.58 ± 0.03 mg N g−1 N day−1, significantly more than Ref (0.36 ± 0.03 mg N g−1 N day−1). In addition, specific soil 15Nsr_min was also significantly higher with mineral soil coverage (Table 3) for both soil layers (0.68 ± 0.10 mg 15N g−1 15N day−1 at 0–5 cm depth and 0.71 ± 0.14 mg 15N g−1 15N day−1 at 5–15 cm depth) than at Ref (0.39 ± 0.02 mg 15N g−1 15N day−1 at 0–5 cm depth and 0.37 ± 0.02 mg 15N g−1 15N day−1 at 5–15 cm depth).
The ratio of the specific soil 15Nsr_min release to the Nr_min release was above one for both layers and sites (Fig. 4). No difference was found between the two sites; however, the ratio of specific 15Nsr_min and Nsr_min from the 5–15 cm soil layer was lower than from the 0–5 cm soil layer. Significant differences among the two soil layers and the two sites are indicated with lowercase letters (ANOVA and Tukey’s honest significant differences)
The effect of mineral soil cover on soil 15N retention and N mineralization
Soil 15N retention
The field 15N tracer experiment showed that 10% of the applied 15N tracer resided in the soil pool in the Ref site, of which more than 90% was found in the top 30 cm of soil. This result was similar to the 15N retention from the drained fen peatland reported by Augustin et al. (1997) who found that 10–20% of the applied labeled 15 N fertilizer were recovered in the soil pool in a 15N tracer experiment from two drained peatlands in Germany, of which more than 90% was located at the 0–20 cm depth. To the best of our knowledge, soil 15N recovery from drained peatland with mineral soil coverage has never been studied. Our results indicate that the soil 15N recovery (~ 20% of the applied 15N tracer) from the Cov site was generally significantly higher than at Ref at the end of the study period, suggesting a better retention of the fertilizer-N though mineral soil coverage. The recovery was at the lower end of the range of data reported from 15N tracer (with 15N enriched fertilizer or slurry) studies in grassland on mineral soil in Europe. This included 20–25% soil 15N retention from a grassland in southern England (Jenkinson et al. 2004), ~ 15% from grassland in the Netherlands (De Vries et al. 2011), and 30–40% from grassland in Germany (Zistl-Schlingmann et al. 2020).
However, we observed a downward movement of the 15N tracer to the deeper layer at Cov, whereas no such trend was found at Ref (Fig. 1D) over the course of the experiment. This indicates that, despite a higher overall recovery in the studied soil layers, fertilizer N might leach faster in the covered mineral soil material at Cov than in the drained peatland at Ref. This may, after longer periods, also change the overall recovery once the leachate leaves the investigated zone of 0–60 cm. The higher 15N leaching from Cov may be attributed to the low absorption rate of the mineral cover material compared to the degraded peat, as the sand content of the mineral soil coverage is much higher than that of the peat at Ref (Table 1). Hence, the low adsorption potential of the sand for anions compared to organic materials with higher anion sorption capacity may have induced a higher N leaching at the mineral soil layer from Cov.
Higher soil 15N recovery at Cov might be due to a higher microbial 15N uptake. Microbial N use is positively correlated with soil substrate availability and soil pH (Elrys et al. 2022). Compared with Ref, the SOM pool from Cov is small (Table 1), but relatively young and labile as previous shown from a 14CO2 measurement on the same site (Wang et al. 2021). The higher availability of labile SOM stimulates soil microbial activity, which ultimately promotes microbial N uptake (Barrett and Burke 2000; Booth et al. 2005; Yang et al. 2022). Moreover, the relative old and stable SOM pool from Ref may exist in forms that the microorganism cannot easily use (Baldock and Skjemstad 2000; Fontaine et al. 2007). This may result in a lower microbial N uptake at Ref than at Cov. In addition, soil pH of the surface layer at Cov was higher than at Ref (Table 1), which may enhance soil microbial N uptake and further induce better soil 15N retention at Cov than at Ref, due to the enhanced microbial activity under high soil pH (Zhang et al. 2017).
Soil N mineralization
The laboratory incubation results showed that soil Nr_min and soil 15Nr_min at 0–5 cm and 5–15 cm depth were significantly higher (p < 0.05) at Ref than at Cov. We attribute this to the overall higher soil N content of the surface peat compared to the mineral soil cover material. In contrast, the specific soil Nsr_min release was higher at Cov than at Ref (Table 3), which indicates that the surface soil organic matter (SOM) at Cov was more labile compared to the Ref site. The labile SOM pool at Cov coincided with the young carbon age at Cov from the former study on the same site (Wang et al. 2021). By definition, the labile SOM pool decomposes very quickly and is easily accessible to plants and microbes (Dungait et al. 2012; Liu et al. 2017).
At both sites, the soil 15Nr_min release was faster than the Nr_min release (Fig. 4), implying that the added 15N turnover rate was higher than the gross N turnover rate. This finding indicates that the newly applied mineral N (15N and 14N), after incorporation into SOM, was preferably stored in the labile soil N pool. This finding is consistent with former studies showing that the exogenous N input is mostly labile (Shevtsova et al. 2003; Mulvaney et al. 2009; Sebilo et al. 2013). This part of the soil organic N pool releases available N for plant uptake in the growing season, but likewise bears the risk of N losses to the environment.
The effect of mineral soil coverage on plant 15N uptake
Over the experimental period, both sites had similar aboveground biomass; however, the aboveground plant N uptake was higher at Cov than at Ref (p < 0.05), suggesting that mineral soil coverage not only sustains the agricultural productivity of the drained peatland, but also increases the fertilizer N use efficiency. However, mineral soil coverage did not influence plant and root 15N content. At both sites, plants took up ~ 30% of the applied 15N fertilizer, similar to the results reported from a meta-analysis of 15N tracer studies, which found that on average, 30% of the applied 15N is taken up by plants in grassland (Templer et al. 2012). It is often assumed that plant N uptake tends to be higher with higher soil N availability (Stevens et al. 2005; Tateno and Takeda 2010). However, we found that the application of mineral soil material, which was relatively poor in N and also released an absolutely smaller amount of N in the incubation experiments, did not reduce the plant N uptake and the plant 15N recovery. The similar aboveground biomass and 15N recovery might be driven by the ample amount of N supplied at both sites and the lack of any N limitations. During the experimental period, ~ 230 kg N ha−1 was applied equally to both sites, and the soil Nr_min results suggest that soil mineralization could supply further N, exceeding the demand for grass production at both sites. Therefore, as N was not limited in the system, the presence of relatively N-poor mineral soil did not impair aboveground yields.
The effect of mineral soil coverage on N loss
Fertilizer N loss reduction
The two sites received ~ 230 kg N ha−1 year−1 fertilizer N input. Together, the fertilizer N input and the soil N supply largely exceed the plant N demand and consequently lead to N losses to the environment, i.e., release into the atmosphere via ammonia volatilization and denitrification as well as via leaching to the groundwater (Robertson and Vitousek 2009; Bowles et al. 2018). At Cov, less of the applied N was lost through the experimental period, considering the storage in the 0–60 cm soil layer. Thus, mineral soil coverage at this site may prevent ~ 25 kg N ha−1 year−1 fertilizer N from being lost to the environment compared with Ref if no substantial leaching below 60 cm will occur.
Effects on peat decomposition
The higher soil N release from Ref at 0–15 cm soil depth (Table 3) also implies a rapid peat decomposition and peatland degradation, as the soil N losses are closely linked to the C losses from the SOM mineralization (Leifeld et al. 2020; Klein et al. 2022). However, we only have evidence for this for the topsoil, whereas the peat underneath the mineral soil coverage was not used for determining soil Nr_min in the incubation experiment due to the reasons below. Firstly, the soil 15N signature underneath 15 cm soil layer was nearly natural abundance (Fig. 1B), which makes it impossible to determine the soil 15N mineralization from the deeper soil layer. Second, the oxygen availability in the deeper soil layer is difficult to simulate in laboratory incubation, and the atmosphere oxygen availability may overestimate the soil N and 15N mineralization rate from the deeper soil layer.
It may be suspected that at Cov, these subsoil organic layers may have a higher Nr_min release than the surface organic soil from Ref site due to two possible mechanisms. First, the covered mineral soil material revealed some N leaching (Fig. 1). This leachate from the mineral soil material may stimulate the decomposition of the peat underneath the mineral soil cover via positive priming (Kuzyakov et al. 2000). Second, the mineral cover enhanced the soil pH of the peat layers underneath (Table 1). As SOM decomposition increases with soil pH (Sinsabaugh et al. 2008), a higher potential for peat decomposition underneath the mineral soil coverage may be possible.
On the other hand, oxygen availability is vital for peat mineralization (Blodau 2002; Tiemeyer et al. 2016). For the peat layer underneath the mineral soil coverage, the oxygen availability for SOM mineralization is reduced (Jørgensen et al. 2012), leading to a presumably lower N release. In addition, the absence of fresh plant residue input into the deeper layers of the organic soil underneath might limit SOM mineralization (Song et al. 2018; Zhang et al. 2021). Fresh plant inputs are the primary source of SOM formation, which could not only determine the chemical composition of SOM, but also impact soil microbial activities. The exclusion of fresh plant inputs may lead to a N limitation for microorganisms and further limit the N mineralization for peat underneath the mineral soil coverage (Mooshammer et al. 2014). Moreover, a former study conducted at the same site found that SOC from mineral soil material contributed greatly to heterotrophic soil respiration at the Cov site (Wang et al. 2021). The contribution of the peat layer underneath the mineral soil coverage was relatively small compared to the contribution of peat C at Ref (Wang et al. 2021). The lower contribution of subsoil peat to C loss from Cov suggests that mineral soil cover might be able to reduce the peat decomposition rate despite incoming N leachate and a higher pH. However, further in situ soil profile-based SOM mineralization experiments are still needed to support this interpretation.
Our findings suggest that mineral soil coverage has the potential to reduce N losses (due to higher soil N retention) from drained peatland and, hence, may make agricultural production on drained peatland less harmful to the environment compared to the continued direct use of these soils. However, we would like to point out that from a nature conservation standpoint as well as climate mitigation strategy, mineral soil coverage does not replace or substitute the mitigating effect that can be achieved by rewetting. Rather, we aim to encourage further research about mineral soil coverage as a peatland management measure in regions, where peatland rewetting is not supported, be it for reasons of national food and feed provision, economic incomes, or political strategies.
Our field 15N trace experiment and laboratory incubation together provide the first insight into how mineral soil coverage influences the N balance of the plant–soil system in agriculturally managed drained peatland. Over the experimental period, mineral soil coverage of drained peatland significantly reduced the system fertilizer N loss. For the deeper peat layer, the effect of mineral soil coverage on peat decomposition and mineralization still needs to be further explored. In summary, the study suggests that mineral soil coverage, a measure used by farmers to counterbalance subsidence, provides an opportunity for reducing the environmental pollution induced by the agricultural use of drained peatland. Furthermore, the reduced N losses with mineral soil coverage from our study also highlighted the need for multiply field observation to evaluate the effectiveness of mineral soil coverage in general terms.
Data from this study are included in the article and supplementary material; further inquiries can be directed to the corresponding author.
Augustin J, Merbach W, Käding H, Schmidt W, Schalitz G, Russow R, Ende HP (1997) N Balance experiments on fen grassland with15N labelled fertilizer. Isot Environ Health Stud 33:31–37. https://doi.org/10.1080/10256019808036356
Baldock JA, Skjemstad JO (2000) Role of the soil matrix and minerals in protecting natural organic materials against biological attack. Org Geochem 31:697–710. https://doi.org/10.1016/S0146-6380(00)00049-8
Barrett JE, Burke IC (2000) Potential nitrogen immobilization in grassland soils across a soil organic matter gradient. Soil Biol Biochem 30:1707–1716. https://doi.org/10.1016/S0038-0717(00)00089-4
Barrett JE, Burke IC (2002) Nitrogen retention in semiarid ecosystems across a soil organic-matter gradient. Ecol Appl 12:878–890. https://doi.org/10.1890/1051-0761(2002)012[0878:NRISEA]2.0.CO;2
Bessler H, Oelmann Y, Roscher C, Buchmann N, Scherer-Lorenzen M, Schulze E-D, Temperton VM, Wilcke W, Engels C (2012) Nitrogen uptake by grassland communities: contribution of N2 fixation, facilitation, complementarity, and species dominance. Plant Soil 358:301–322. https://doi.org/10.1007/s11104-012-1181-z
Blodau C (2002) Carbon cycling in peatlands - a review of processes and controls. Environ Rev 10:111–134. https://doi.org/10.1139/a02-004
Booth MS, Stark JM, Rastetter E (2005) Controls on nitrogen cycling in terrestrial ecosystems: a synthetic analysis of literature data. Ecol Monogr 75:139–157. https://doi.org/10.1890/04-0988
Bowles TM, Atallah SS, Campbell EE, Gaudin ACM, Wieder WR, Grandy AS (2018) Addressing agricultural nitrogen losses in a changing climate. Nat Sustain 1:399–408. https://doi.org/10.1038/s41893-018-0106-0
De Vries FT, Van Groenigen JW, Hoffland E, Bloem J (2011) Nitrogen losses from two grassland soils with different fungal biomass. Soil Biol Biochem 43:997–1005. https://doi.org/10.1016/j.soilbio.2011.01.016
Dungait JJ, Hopkins DW, Gregory AS, Whitmore AP (2012) Soil organic matter turnover is governed by accessibility not recalcitrance. Glob Chang Biol 18:1781–1796. https://doi.org/10.1111/j.1365-2486.2012.02665.x
Elrys AS, Chen Z, Wang J, Uwiragiye Y, Helmy AM, Desoky EM, Cheng Y, Zhang JB, Cai ZC, Muller C (2022) Global patterns of soil gross immobilization of ammonium and nitrate in terrestrial ecosystems. Glob Chang Biol 28:4472–4488. https://doi.org/10.1111/gcb.16202
Evans CD, Peacock M, Baird AJ, Artz RRE, Burden A, Callaghan N, Chapman PJ, Cooper HM, Coyle M, Craig E, Cumming A, Dixon S, Gauci V, Grayson RP, Helfter C, Heppell CM, Holden J, Jones DL, Kaduk J, Levy P, Matthews R, Mcnamara NP, Misselbrook T, Oakley S, Page SE, Rayment M, Ridley LM, Stanley KM, Williamson JL, Worrall F, Morrison R (2021) Overriding water table control on managed peatland greenhouse gas emissions. Nature 593:548–552. https://doi.org/10.1038/s41586-021-03523-1
Ferré M, Muller A, Leifeld J, Bader C, Müller M, Engel S, Wichmann S (2019) Sustainable management of cultivated peatlands in Switzerland: insights, challenges, and opportunities. Land Use Policy 87:104019. https://doi.org/10.1016/j.landusepol.2019.05.038
Fontaine S, Barot S, Barre P, Bdioui N, Mary B, Rumpel C (2007) Stability of organic carbon in deep soil layers controlled by fresh carbon supply. Nature 450:277–280. https://doi.org/10.1038/nature06275
Franzluebbers AJ (2020) Holding water with capacity to target porosity. Agric Environ Lett 5:e20029. https://doi.org/10.1002/ael2.20029
Fry B (2006) Stable isotope ecology. 521 Springer, New York
Jenkinson DS, Poulton PR, Johnston AE, Powlson DS (2004) Turnover of nitrogen-15-labeled fertilizer in old grassland. Soil Sci Soc Am J 68:865–875. https://doi.org/10.2136/sssaj2004.8650
Jørgensen CJ, Struwe S, Elberling B (2012) Temporal trends in N2O flux dynamics in a Danish wetland–effects of plant-mediated gas transport of N2O and O2 following changes in water level and soil mineral-N availability. Glob Change Biol 18:210–222. https://doi.org/10.1111/j.1365-2486.2011.02485.x
Kalu S, Oyekoya GN, Ambus P, Tammeorg P, Simojoki A, Pihlatie M, Karhu K (2021) Effects of two wood-based biochars on the fate of added fertilizer nitrogen—a 15N tracing study. Biol Fertil Soils 57:457–470. https://doi.org/10.1007/s00374-020-01534-0
Kasimir A, He H, Coria J, Norden A (2018) Land use of drained peatlands: greenhouse gas fluxes, plant production, and economics. Glob Chang Biol 24:3302–3316. https://doi.org/10.1111/gcb.13931
Klein K, Schellekens J, Groβ-Schmölders M, Von Sengbusch P, Alewell C, Leifeld J (2022) Characterizing ecosystem-driven chemical composition differences in natural and drained Finnish bogs using pyrolysis-GC/MS. Org Geochem 165:104351. https://doi.org/10.1016/j.orggeochem.2021.104351
Klemedtsson L, Von Arnold K, Weslien P, Gundersen P (2005) Soil CN ratio as a scalar parameter to predict nitrous oxide emissions. Glob Change Biol 11:1142–1147. https://doi.org/10.1111/j.1365-2486.2005.00973.x
Kuzyakov Y, Friedel JK, Stahr K (2000) Review of mechanisms and quantification of priming effects. Soil Biol Biochem 32:1485–1498. https://doi.org/10.1016/S0038-0717(00)00084-5
Leifeld J (2018) Distribution of nitrous oxide emissions from managed organic soils under different land uses estimated by the peat C/N ratio to improve national GHG inventories. Sci Total Environ 631–632:23–26. https://doi.org/10.1016/j.scitotenv.2018.02.328
Leifeld J, Menichetti L (2018) The underappreciated potential of peatlands in global climate change mitigation strategies. Nat Commun 9:1071. https://doi.org/10.1038/s41467-018-03406-6
Leifeld J, Klein K, Wust-Galley C (2020) Soil organic matter stoichiometry as indicator for peatland degradation. Sci Rep 10:7634. https://doi.org/10.1038/s41598-020-64275-y
Liu Y, Wang C, He N, Wen X, Gao Y, Li S, Niu S, Butterbach-Bahl K, Luo Y, Yu G (2017) A global synthesis of the rate and temperature sensitivity of soil nitrogen mineralization: latitudinal patterns and mechanisms. Glob Change Biol 23:455–464. https://doi.org/10.1111/gcb.13372
Loisel J, Yu Z, Beilman DW, Camill P, Alm J, Amesbury MJ, Anderson D, Andersson S, Bochicchio C, Barber K, Belyea LR, Bunbury J, Chambers FM, Charman DJ, De Vleeschouwer F, Fiałkiewicz-Kozieł B, Finkelstein SA, Gałka M, Garneau M, Hammarlund D, Hinchcliffe W, Holmquist J, Hughes P, Jones MC, Klein ES, Kokfelt U, Korhola A, Kuhry P, Lamarre A, Lamentowicz M, Large D, Lavoie M, Macdonald G, Magnan G, Mäkilä M, Mallon G, Mathijssen P, Mauquoy D, Mccarroll J, Moore TR, Nichols J, O’reilly B, Oksanen P, Packalen M, Peteet D, Richard PJH, Robinson S, Ronkainen T, Rundgren M, Sannel ABK, Tarnocai C, Thom T, Tuittila ES, Turetsky M, Väliranta M, Van Der Linden M, Van Geel B, Van Bellen S, Vitt D, Zhao Y, Zhou W (2014) A database and synthesis of northern peatland soil properties and Holocene carbon and nitrogen accumulation. Holocene 24:1028–1042. https://doi.org/10.1177/0959683614538073
Mariotti A (1983) Atmospheric nitrogen is a reliable standard for natural 15N abundance measurements. Nature 5919:685–687. https://doi.org/10.1038/303685a0
Mooshammer M, Wanek W, Zechmeister-Boltenstern S, Richter A (2014) Stoichiometric imbalances between terrestrial decomposer communities and their resources: mechanisms and implications of microbial adaptations to their resources. Front Microbiol 5:22. https://doi.org/10.3389/fmicb.2014.00022
Müller C, Laughlin RJ, Christie P, Watson CJ (2011) Effects of repeated fertilizer and cattle slurry applications over 38 years on N dynamics in a temperate grassland soil. Soil Biol Biochem 43:1362–1371. https://doi.org/10.1016/j.soilbio.2011.03.014
Mulvaney RL, Khan SA, Ellsworth TR (2009) Synthetic nitrogen fertilizers deplete soil nitrogen: a global dilemma for sustainable cereal production. J Environ Qual 38:2295–2314. https://doi.org/10.2134/jeq2008.0527
Pijlman J, Holshof G, Van Den Berg W, Ros GH, Erisman JW, Van Eekeren N (2020) Soil nitrogen supply of peat grasslands estimated by degree days and soil organic matter content. Nutr Cycl Agroecosyst 117:351–365. https://doi.org/10.1007/s10705-020-10071-z
Rahman MK, Parsons JW (1999) Uptake of 15N by wetland rice in response to application of 15N-labelled Sesbania rostrata and urea. Biol Fertil Soils 29:69–73. https://doi.org/10.1007/s003740050526
Rihm B, Künzle T (2019) Mapping nitrogen deposition 2015 for Switzerland. Technical report on the update of critical loads and exceedance, including the years 1990, 2000, 2005 and 2010. Federal Office for the Environment Switzerland Bern. Available at: https://www.bafu.admin.ch. Accessed 06 Jan 2021
Robertson GP, Vitousek PM (2009) Nitrogen in agriculture: balancing the cost of an essential resource. Annu Rev Environ Resour 34:97–125. https://doi.org/10.1146/annurev.environ.032108.105046
Rowlings DW, Scheer C, Liu S, Grace PR (2016) Annual nitrogen dynamics and urea fertilizer recoveries from a dairy pasture using 15N; effect of nitrification inhibitor DMPP and reduced application rates. Agric Ecosyst Environ 216:216–225. https://doi.org/10.1016/j.agee.2015.09.025
Schindler U, Müller L (1999) Rehabilitation of the soil quality of a degraded peat site. In: Stott DE, Mohtar RH, Steinhardt GC (eds) 10th International Soil Conservation Organization Meeting held May 24–29, 1999. Purdue University, West Lafayette, pp 648–654
Schothorst CJ (1977) Subsidence of low moor peat soils in the western Netherlands. Geoderma 17:265–291. https://doi.org/10.1016/0016-7061(77)90089-1
Sebilo M, Mayer B, Nicolardot B, Pinay G, Mariotti A (2013) Long-term fate of nitrate fertilizer in agricultural soils. Proc Natl Acad Sci USA 110:18185–18189. https://doi.org/10.1073/pnas.1305372110
Shevtsova L, Romanenkov V, Sirotenko O, Smith P, Smith JU, Leech P, Kanzyvaa S, Rodionova V (2003) Effect of natural and agricultural factors on long-term soil organic matter dynamics in arable soddy-podzolic soils—modeling and observation. Geoderma 116:165–189. https://doi.org/10.1016/s0016-7061(03)00100-9
Sinsabaugh RL, Lauber CL, Weintraub MN, Ahmed B, Allison SD, Crenshaw C, Contosta AR, Cusack D, Frey S, Gallo ME, Gartner TB, Hobbie SE, Holland K, Keeler BL, Powers JS, Stursova M, Takacs-Vesbach C, Waldrop MP, Wallenstein MD, Zak DR, Zeglin LH (2008) Stoichiometry of soil enzyme activity at global scale. Ecol Lett 11:1252–1264. https://doi.org/10.1111/j.1461-0248.2008.01245.x
Song Y, Song C, Hou A, Ren J, Wang X, Cui Q, Wang M (2018) Effects of temperature and root additions on soil carbon and nitrogen mineralization in a predominantly permafrost peatland. CATENA 165:381–389. https://doi.org/10.1016/j.catena.2018.02.026
Sonneveld MPW, Lantinga EA (2011) The contribution of mineralization to grassland N uptake on peatland soils with anthropogenic A horizons. Plant Soil 340:357–368. https://doi.org/10.1007/s11104-010-0608-7
Steffens D, Pfanschilling R, Feigenbaum S (1996) Extractability of 15N-labeled corn-shoot tissue in a sandy and a clay soil by 0.01 M CaCl2 method in laboratory incubation experiments. Biol Fertil Soils 22:109–115. https://doi.org/10.1007/BF00384441
Stevens WB, Hoeft RG, Mulvaney RL (2005) Fate of nitrogen-15 in a long-term nitrogen rate study: II. Nitrogen Uptake Efficiency Agron J 97:1046–1053. https://doi.org/10.2134/agronj2003.0313
Tateno R, Takeda H (2010) Nitrogen uptake and nitrogen use efficiency above and below ground along a topographic gradient of soil nitrogen availability. Oecologia 163:793–804. https://doi.org/10.1007/s00442-009-1561-0
Templer P, Mack M, Iii FC, Christenson L, Compton J, Crook H, Currie W, Curtis C, Dail D, D’antonio C (2012) Sinks for nitrogen inputs in terrestrial ecosystems: a meta-analysis of 15N tracer field studies. Ecology 93:1816–1829. https://doi.org/10.1890/11-1146.1
Tiemeyer B, AlbiacBorraz E, Augustin J, Bechtold M, Beetz S, Beyer C, Drosler M, Ebli M, Eickenscheidt T, Fiedler S, Forster C, Freibauer A, Giebels M, Glatzel S, Heinichen J, Hoffmann M, Hoper H, Jurasinski G, Leiber-Sauheitl K, Peichl-Brak M, Rosskopf N, Sommer M, Zeitz J (2016) High emissions of greenhouse gases from grasslands on peat and other organic soils. Glob Change Biol 22:4134–4149. https://doi.org/10.1111/gcb.13303
Wang Y, Paul SM, Jocher M, Espic C, Alewell C, Szidat S, Leifeld J (2021) Soil carbon loss from drained agricultural peatland after coverage with mineral soil. Sci Total Environ 800:149498. https://doi.org/10.1016/j.scitotenv.2021.149498
Wang Y, Paul SM, Jocher M, Alewell C, Leifeld J (2022) Reduced nitrous oxide emissions from drained temperate agricultural peatland after coverage with mineral soil. Front Environ Sci 10:856599. https://doi.org/10.3389/fenvs.2022.856599
Wardle DA (1998) Controls of temporal variability of the soil microbial biomass: a global-scale synthesis. Soil Biol Biochem 30:1627–1637. https://doi.org/10.1016/S0038-0717(97)00201-0
Wesselsperelo L, Jimenez M, Munch J (2006) Microbial immobilisation and turnover of 15N labelled substrates in two arable soils under field and laboratory conditions. Soil Biol Biochem 38:912–922. https://doi.org/10.1016/j.soilbio.2005.07.013
Yang S, Liu W, Guo L, Wang C, Deng M, Peng Z, Liu L (2022) The changes in plant and soil C pools and their C: N stoichiometry control grassland N retention under elevated N inputs. Ecol Appl 32:e2517. https://doi.org/10.1002/eap.2517
Yu Z, Loisel J, Brosseau DP, Beilman DW, Hunt SJ (2010) Global peatland dynamics since the last glacial maximum. Geophys Res Lett 37:L13402. https://doi.org/10.1029/2010gl043584
Zhang M, Alves RJE, Zhang D, Han L, He J, Zhang L (2017) Time-dependent shifts in populations and activity of bacterial and archaeal ammonia oxidizers in response to liming in acidic soils. Soil Biol Biochem 112:77–89. https://doi.org/10.1016/j.soilbio.2017.05.001
Zhang X, Zhu B, Yu F, Cheng W (2021) Plant inputs mediate the linkage between soil carbon and net nitrogen mineralization. Sci Total Environ 790:148208. https://doi.org/10.1016/j.scitotenv.2021.148208
Zistl-Schlingmann M, KwatchoKengdo S, Kiese R, Dannenmann M (2020) Management intensity controls nitrogen-use-efficiency and flows in grasslands—a 15N tracing experiment. Agronomy 10:606. https://doi.org/10.3390/agronomy10040606
We appreciate the help from Markus Jocher, Robin Giger, Steven Nagel, Martin Zuber, and Shiva Ghiasi at Agroscope during field sampling and lab analysis. We are thankful for many useful discussions with Christof Ammann and Chloé Wüst, Agroscope. We are grateful to Bernhard Schneider for the opportunity to collaborate on his farm.
Open access funding provided by Agroscope The work was supported by funding received from the Swiss Federal Office for the Environment (contract number 06.0091.PZ/R261-2425) and the China Scholarship Council (NSCIS 201806350221).
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Wang, Y., Paul, S.M., Alewell, C. et al. Reduced nitrogen losses from drained temperate agricultural peatland after mineral soil coverage. Biol Fertil Soils 59, 153–165 (2023). https://doi.org/10.1007/s00374-022-01689-y