Field Application of Organic Fertilizers Triggers N2O Emissions From the Soil N Pool as Indicated by 15N-Labeled Digestates

Anaerobic digestion (AD) can generate biogas while simultaneously producing digestate which can be used as fertilizer. Feedstocks used for AD influence digestate composition, which in turn may affect carbon (C) and nitrogen (N) turn-over in soils and subsequently influence nitrous oxide (N2O) emissions after soil application. Assessment of greenhouse gas emissions from digestates can help to evaluate the overall sustainability of an agricultural production system. The objective of this study was therefore to evaluate and understand the effect of differences in digestate composition on in situ N2O emissions within the 1st weeks after application of seven digestates. The digestates were derived from different feedstocks and 15N-labeled, either in total N or only in ammonium-N. Therefore, the experimental design enabled us to differentiate between potential N2O-N sources (i.e., digestate N or soil N). Furthermore, it allowed to distinguish to some extent between organic-N and ammonium-N as potential N sources for denitrification. Digestates were homogeneously incorporated into the upper 5 cm of microplots in an arable Haplic Luvisol in South Germany at a rate of 170 kg N ha−1. After application, N2O fluxes were measured for ~60 days (May-July) using the closed chamber method in 2 experimental years. Mainly due to higher precipitations in the 1st year, cumulative N2O emissions were higher (312–1,580 g N2O-N ha−1) compared to the emissions (133–690 g N2O-N ha−1) in the 2nd year. Between 16–33% (1st year) and 17–38% (2nd year) of N2O emissions originated from digestate N, indicating that digestate application triggered N2O production and release mainly from soil N. This effect was strongest immediately after digestate application. It was concluded that the first (short term) peak in N2O emissions after digestate application is largely related to denitrification of soil-N. However, the experimental setup does not allow to differentiate between the different denitrification pathways. Weather conditions showed a substantial effect on N2O emissions, where the correlation between N2O and CO2 flux rates hinted on denitrification as main N2O source. The effect of digestate composition, particularly organic N from the digestate, on soil N2O emissions seems to be of minor relevance.

Anaerobic digestion (AD) can generate biogas while simultaneously producing digestate which can be used as fertilizer. Feedstocks used for AD influence digestate composition, which in turn may affect carbon (C) and nitrogen (N) turn-over in soils and subsequently influence nitrous oxide (N 2 O) emissions after soil application. Assessment of greenhouse gas emissions from digestates can help to evaluate the overall sustainability of an agricultural production system. The objective of this study was therefore to evaluate and understand the effect of differences in digestate composition on in situ N 2 O emissions within the 1st weeks after application of seven digestates. The digestates were derived from different feedstocks and 15 N-labeled, either in total N or only in ammonium-N. Therefore, the experimental design enabled us to differentiate between potential N 2 O-N sources (i.e., digestate N or soil N). Furthermore, it allowed to distinguish to some extent between organic-N and ammonium-N as potential N sources for denitrification. Digestates were homogeneously incorporated into the upper 5 cm of microplots in an arable Haplic Luvisol in South Germany at a rate of 170 kg N ha −1 . After application, N 2 O fluxes were measured for ∼60 days (May-July) using the closed chamber method in 2 experimental years. Mainly due to higher precipitations in the 1st year, cumulative N 2 O emissions were higher (312-1,580 g N 2 O-N ha −1 ) compared to the emissions (133-690 g N 2 O-N ha −1 ) in the 2nd year. Between 16-33% (1st year) and 17-38% (2nd year) of N 2 O emissions originated from digestate N, indicating that digestate application triggered N 2 O production and release mainly from soil N. This effect was strongest immediately after digestate application. It was concluded that the first (short term) peak in N 2 O emissions after digestate application is largely related to denitrification of soil-N. However, the experimental setup does not allow to differentiate between the different denitrification pathways. Weather conditions showed a substantial effect on N 2 O emissions, where the correlation between N 2 O and CO 2 flux rates hinted on denitrification as main N 2 O source. The effect of digestate composition, particularly organic N from the digestate, on soil N 2 O emissions seems to be of minor relevance.

INTRODUCTION
In the EU, about 180 million tons of anaerobic digestate are estimated to be produced per year, most of which is used as organic fertilizer (Corden et al., 2019). Digestates have been shown to have the potential to substitute mineral fertilizers and contribute to a sustainable soil management (Gutser et al., 2005;Cavalli et al., 2016;Verdi et al., 2019). However, application of organic as well as mineral nitrogen (N) fertilizers is also known to increase greenhouse gas (GHG) emissions from soils. Globally, agriculture contributes up to 20% to carbon dioxide equivalents (CO 2 -eq.) from all human activities (2010)(2011)(2012)(2013)(2014)(2015)(2016)(2017), with nitrous oxide (N 2 O) and methane (CH 4 ) as main GHGs (FAO, 2020). About 60% of anthropogenic N 2 O emissions are emitted by agricultural soils (Ciais et al., 2013), thus it is of high relevance to assess N 2 O in relation to fertilizer application.
Studies have shown that digestates might lead to a higher risk of N 2 O formation than manures, which is related to the higher share of ammonium (NH + 4 -N) after AD (Möller and Stinner, 2009). Ammonium is quickly nitrified to nitrate (NO − 3 ), which can further be denitrified to dinitrogen gas (N 2 ). Both processes, as well as nitrifier denitrification, bear the risk of producing N 2 O and are considered as main N 2 O source from soils (Granli and Bøckman, 1994;Bremner, 1997;Koola et al., 2010). Application of liquid manures like slurry or digestates provides available N and carbon (C), which in turn promotes heterotrophic activity (oxidation of C, N, S, etc.), depleting oxygen (O 2 ) availability in soil, and thus favors creation of anaerobic microsites that ultimately trigger N 2 O production and release via denitrification (Chadwick et al., 2000;Petersen et al., 2003). Hence, N 2 O emissions largely depend on the availability of labile organic C (C org ), mineral N, O 2 and water in the soil and their subsequent effect on soil microbial processes (Flessa and Beese, 2000;Ruser et al., 2001). However, AD has also been reported to reduce the N 2 O potential compared to the initial feedstock e.g., by decreasing slurry viscosity or increasing the recalcitrance of organic matter (OM) (Petersen, 1999;Möller, 2015).
The different organic substrates that are used as feedstock for anaerobic digestion (AD) affect the physico-chemical characteristics of the digestate (Fouda et al., 2013;Zirkler et al., 2014). For example, comparing food wastes and maize silage, food wastes are already processed goods with a high degradability and high protein content. Thereby, food waste-based digestates tend to have a higher OM degradability and a higher share of NH + 4 -N than maize silage, that could enhance soil microbial activity (Möller and Müller, 2012;Guilayn et al., 2020).
Based on compositional differences, such as N content, C/N ratio and OM degradability [C org /organic N (N org )], it can be assumed that digestates from different feedstocks will show differences in N 2 O emissions after field application. However, a differentiated consideration of the GHG emission potential for digestates from different feedstocks is currently scarce, and therefore will be the main research focus of this study.
The largest share of N 2 O release during the growing season usually occurs shortly after field application, with further peaks correlated to rainfall-events (Guzman-Bustamante et al., 2019;Herr et al., 2019) or freeze-thaw periods Rochette et al., 2008). For this reason, the following experiment was conducted to evaluate digestates regarding short-term N 2 O emissions on fallow land. To calculate the amount of N 2 O derived from the digestate, 15 N-stable isotope labeling was used. The following hypotheses were tested: (1) Digestates with varying physical and chemical properties will show different temporal N 2 O and 15 N-N 2 O flux patterns.
(2) Application of these digestates will also result in different cumulative N 2 O emissions and N 2 O emission factors. (3) The amount of N 2 O-N directly derived from the digestate will differ among the digestate types.

MATERIALS AND METHODS 15 N Labeling and Digestate Production
Labeled anaerobic digestates were prepared by cultivation of 15 N-enriched plants in a comparable approach as applied by Schouten et al. (2012). Maize (Zea mays L. cv. Ronaldinio), ryegrass (Lolium perenne L. cv. Kentaur), and sugar beet (Beta vulgaris subsp. vulgaris, Altissima Group) were 15 N labeled, by addition of 15 N ammonium sulfate ((NH 4 ) 2 SO 4 ) as fertilizing solution. Ryegrass was cultivated in sand culture in 12 kg boxes of 10 cm height. For fertilization, 96.2 mg N kg −1 as (NH 4 ) 2 SO 4 (30 atom% 15 N) solution was applied before sowing ryegrass. We cut the ryegrass three times in 30-days intervals. Sugar beet was grown in Mitscherlich pots with 12 kg sand. After pregrowing sugar beet seedlings in peat, two plants were set for each pot. Four rates of (NH 4 ) 2 SO 4 (50 atom% 15 N) solution were applied during growth (in total 1.5 g N per pot). Maize was grown in a hydroponic system with two plants per 10liter pot. Nutrient solution adapted after Engels (1999) with modified N concentration was exchanged twice a week. Within the first 5 weeks of growth, NH + 4 -N concentration was gradually increased, while NO − 3 -N supply was decreased to acclimate maize plants to primary NH + 4 -N nutrition. After this adaption phase for the plants, the N concentration was kept stable at 0.5 mM NO − 3 -N and 3 mM NH + 4 -N, in the form of calcium nitrate and ammonium sulfate. For 15 N labeling, four additions of NH + 4 -N were substituted by 50 atom% 15 N-NH + 4 and applied at BBCH stages: 16-19, 30-33, 51-55, and 71. As commonly done for maize, as energy crop, it was harvested at the dough-ripe stage. After harvest, ryegrass, maize sugar beet, as well as sugar beet leaves were immediately cut and homogenized by short blending (Thermomix TM31, Wuppertal, Gemany). The 15 N enrichment of crops and harvest residues was determined by IRMS with previous freeze-drying, leading to 19.3 atom% 15 N in maize, 26.1 atom% in ryegrass, 43.8 atom% in sugar beet, and 45.3 atom% in sugar beet leaves. After weighing the 15 N-plant biomass into small portions, they were frozen at −20 • C until anaerobic digestion in a batch reactor as previously described by Brulé (2014), Mönch-Tegeder et al. (2014).
Anaerobic digestion of the 15 N-labeled plants and plant residues was carried out at the State Institute of Agricultural Engineering and Bioenergy, at the University of Hohenheim. Before, digestates from maize, grass silage and sugar beet were collected from biogas plants in southern Germany to be used as inoculum for AD of the 15 N-feedstocks (maize, ryegrass, sugar beet, and leaves). Digestates were "starved" for 10 days according to the German standard VDI 4630 guideline (2016) to minimize residual gas production. During this starvation phase, the vessels were kept open and stirred to volatilize ammonia (NH 3 ) from the inoculum. By decreasing NH + 4 -N in the digestate, hence total N concentration, a high 15 N-signature could be assured, with only marginal N dilution of the added 15 N-feedstock. Prior to AD, the inoculum was sieved to produce a homogeneous slurry. 15 N-labeled ryegrass, maize, sugar beet, and sugar beet leaves were separately added to the substrate-specific inoculum in a ratio of 1:2.5 organic total solids (oTS) (VDI 4630, 2016). Anaerobic digestion was carried out in 2 liter fed-batch systems under mesophilic temperature at 37.5 ± 1 • C for 60 days. During digestion, three feeding portions of 15 N enriched plant substrates were added: at the start of the experiment, after 20, and after 40 days, respectively. Due to the amount of oTS added by the digestates, the 15 N amount of the feedstocks was diluted by N contained in the inoculum, leading to a lower labeling of the 15 N-digestates compared to the initial plant feedstock (Table 1).
Additionally, digestates from existing biogas plants were included and the mineral NH + 4 -N fraction was labeled: organic waste digestate, food waste digestate, and cattle slurry digestate. The digestates were analyzed for total N and NH + 4 -N. Each digestate was filled into a glass beaker and put into a rotating water bath for 12 h at 37 • C, to volatilize a small amount of NH 3 . Afterwards the digestates were analyzed again for total N and NH + 4 -N to assess the amount of N that was emitted. The lost N was substituted by addition of 15 N-enriched (NH 4 ) 2 SO 4 solution to 5 atom% 15 N excess. If more N was lost than resupplied by 15 N-NH + 4 , ammonium chloride solution was added.

Experimental Design
The experiment was performed at the research station "Heidfeldhof " at the University of Hohenheim, 13 km south of Stuttgart, in South-Germany. The research station has a mean annual precipitation of 686 mm and a mean annual temperature of 8.8 • C, monitored by a local meteorological station. The soil type of the arable field was a Haplic Luvisol (IUSS Working Group, 2015) with a silty loam soil texture (2% sand, 68% silt, and 30% clay), a bulk density of 1.24 g cm −3 in the upper 30 cm. Soil analytical results are presented in Table 2. The micro-plot field experiment (1 × 1 m plot size) was conducted as randomized block design with four replicates per treatment in 2 years 2016 (1st) and 2017 (2nd year) from May to July. The treatments consisted of one unfertilized control and seven 15 N-labeled digestates based on maize (M), grass (G), sugar beet (SB), sugar beet leaves (SBL), organic waste (OW), food waste (FW), and cattle slurry (CS) ( Table 1).

Gas Measurements and Analysis
The closed chamber system was used to monitor N 2 O, CO 2 , and methane (CH 4 ) soil fluxes (Hutchinson and Mosier, 1981). The system consisted of a polyvinyl chloride (PVC) base ring (30 cm inner diameter) and a corresponding chamber . Within the center of the 1 m 2 micro plot, the PVC base ring was embedded 10 cm deep in the soil. The 15 N-labeled digestates ( Table 1) were applied at a rate of 170 kg N ha −1 , meaning 1.2 g N per base ring and quickly incorporated into the upper 5-10 cm of the soil. In order to do so, a 10 cm deep furrow was dug across the ring, using a spade. Digestate was filled into the furrow, covered by soil and mixed. The same procedure was done for the unfertilized control using water. The amount of water (290 ml) corresponded to the average volume of digestate application. Directly after application, the first gas measurement was performed. For gas sampling between 8.00 and 12.00 am, the base ring was covered with the dark, vented PVC chamber, sealed by a rubber ring to collect the trace gas. The chambers were closed for 45-60 min. The first gas sample was directly taken after closure, followed by additional sampling every 15-20 min using a syringe and transferred into evacuated 20 ml gas vials. At the same time, two additional gas samples were collected into 100 ml vials at the start and end of each measurement for 15 N-N 2 O determination. Soil and chamber temperature were recorded within each block from two random plots at beginning and end of sampling. Within the 1st month, gas samples were taken 3-4 times a week. Afterwards the sampling frequency was reduced to once or twice per week, with additional samplings after strong rainfall events. In both years 20 gas samplings were performed and measured for N 2 O, CO 2 , and CH 4 , whereas 13-15 out of 20 15 N-N 2 O gas samples could be measured due to cost and time reasons in the 1st and 2nd year, respectively. Gas samples were measured with a gas chromatograph (GC 450 Greenhouse Gas Analyzer, Bruker Daltonic, Bremen, Germany) equipped with electron capture detector (ECD) and flame ionization detector (FID) and an automatic sampler (GX-281, Gilson, Limburg, Germany). During GC measurements, concentrations of N 2 O and CO 2 were analyzed with a 63 Ni ECD and CH 4 concentrations were determined with the FID. Fluxes of N 2 O, CO 2 , and CH 4 were calculated by an extended version of the R (R Core Team, 2016) package "gasfluxes" (Fuß and Asger, 2014).

Analysis of Digestate and Soil
Digestates were dried at 105 • C for dry matter (DM) analysis. Total C (C t ) was measured by Dumas combustion via elemental analysis (Elementar vario MAX CN, Analysensysteme GmbH, Hanau, Germany). Carbonate content was determined volumetrically using the Scheibler method according to DIN 10693 (2014). Thus, organic C can be calculated as the difference between C t and Carbonate-C. Total N (N t ) and NH + 4 of fresh matter (FM) digestate sample was determined by Kjeldahl method. Organic N was derived by the difference of NH + 4 from N t . The pH value was measured in FM digestate using 0.01 mol L −1 calcium chloride solution (1:10 w/w). Soil mineral N (N min )was determined by extraction with 0.5 M potassium sulfate solution (1:4) and measured colorimetrically with a photometer (Flow-injection-analyzer 3 QUAAtro, SEAL Analytical, UK). Bulk density of the top soil was determined using 100 ml stainless steel cylinders in the field.
Total N and C, and 15 N-signature of 15 N-labeled plant substrates, soil and digestate was measured with a CN-elemental analyzer (EuroVector, HEKAtech, Wegberg, Germany) with Isotope Ratio Mass Spectrometer (IRMS) (Delta plus Advantage, Thermo Finnigan, Bremen, Germany). For the determination of the 15 N abundance in N 2 O we used an IRMS delta plus (Finnigan MAT, Bremen, Germany) coupled with an automated PreCon-Interface (Brand, 1995).

Statistics and Calculations
Trapezoidal linear interpolation of daily gas fluxes (N 2 O and CH 4 ) was used to calculate total cumulative emissions for the 55-58 days of the experiment in the 1st and 2nd year, respectively. The percentage of N 2 O-N originating from digestate N (Nd) was calculated by equation (1), with digestate i at sampling time t. Atom% 15 N excess was calculated by subtraction of the natural abundance of N 2 O in the atmosphere (0.369 atom%) from the measured 15 N. The daily N 2 O flux rate (µg N 2 O-N m −2 h −1 ) was multiplied with Nd in equation (2) to determine the amount of N 2 O derived from digestate ( 15 N-N 2 O) as reported by Senbayram et al. (2009). We calculated the recovery (%) of 15 N applied by summing up 15 N content of the soil at the end of the experiment and cumulative 15 N-N 2 O loss. This sum was then divided by the amount of 15 N applied by the digestate as described by Pfab (2011).
Digestate derived fluxes ( 15 N-N 2 O) were also linearly interpolated to calculate cumulative (cum) 15 N-N 2 O emissions. The total share of N derived from digestate (total Nd) in cumulative N 2 O was calculated by Equation (3). As suggested by Schleusner et al. (2018), the amount of primed N 2 O-N lost by fertilizer application was calculated by a simplified approach accounting for cumulative N 2 O-N emissions of the unfertilized control treatment (Equation 4), without considering other gaseous losses via NH 3 or N 2 .  (Hafner et al., 2019). The model is used to predict NH 3 -N losses within the first 72 h from animal slurry and therefore holds a higher uncertainty for digestates. Digestate NH 3 -N losses mainly served as an indicator for the amount of indirect N 2 O emissions. Indirect emissions from NO − 3 leaching were not accounted for. According to IPCC (2019) 1% of NH 3 -N losses was assumed to be re-deposited as N 2 O-N. Soil organic C stocks were presumed to be stable over the experimental period, thus, CO 2 fluxes were not considered for the calculation of total GWP (Herr et al., 2020). Water filled pore space (WFPS) was calculated by Equation (6) using the measured volumetric water content (WC vol ) and porosity (P), where P is depicted as soil bulk density (ρd) and solid particle density (ρs) ( Equation 7). For ρs the density of quartz (2.65 g cm −3 ) was assumed.
For each year, a regression analysis of N 2 O fluxes was calculated, using a stepwise forward selection in a multiple linear regression approach. Air temperature (2 m height), WFPS and CO 2 fluxes were included as independent variables within the model (8). Only significant variables remained in the model (α = 0.05) and the square root of the partial R² was determined. Same regression procedure was applied for cumulative N 2 O and 15 N-N 2 O emissions within each year separately. For this approach the effect of digestate composition was determined, using NH + 4 -N share, and the ratios C/N and C org /N org in model (9).
where y i is the observation of the i th digestate treatment, µ represents the average response, β n are the parameters of fixed effects, b i is the complete block effect and e i is the error of y i .
Significant differences among treatments for cumulative N 2 O, CH 4 , CO 2 , as well as 15 N-N 2 O and total Nd (%) were determined by the Proc MIXED procedure and the Tukey test (α = 0.05). The MIXED procedure can fit various mixed linear models to data and produces the appropriate statistics (SAS Institute Inc, 2015). All statistical analyses were performed with SAS 9.4 (SAS Institute, Cary NC, USA). Graphics were produced with SigmaPlot 11.0 (Systat Software GmbH, Erkrath, Germany).

Meteorological Conditions
Weather conditions showed distinct differences in precipitation between the 2 experimental years. Over the 1st and 2nd year, 183 and 178 mm of precipitation were measured during the 55 and 58 days when the experiment lasted, respectively (Figure 1). Within  (Figures 2, 3). Climate data with mean daily precipitation on the left y-axis and mean daily temperature on the right y-axis.

Temporal N 2 O Fluxes
Nitrous oxide fluxes measured in the 2 experimental years showed distinct differences in peak number and flux magnitude. In both years N 2 O pulses occurred directly after digestate fertilization and after strong rainfall events (Figure 1). Three major peaks were detected in the 1st year: one directly after digestate application, the second and third peak after 13 and 24 days, and a minor peak after 1 week, following strong rainfall events on day 12 and 23. Highest N 2 O flux rate in the 1st year was measured with the SBL treatment on day 13 (1,260 µg N 2 O-N m −2 h −1 ). In the 2nd year, the N 2 O pulse developing directly after N fertilization did not reach the same magnitude as in the 1st year and appeared 1 day later. Highest N 2 O flux in the 2nd year followed a rainfall event 1 week after digestate application reaching up to 424 µg N 2 O-N m −2 h −1 with SB (Figure 1). The peak decreased sharply in case of SB, and gradually until day 14 for the other digestates. After another strong rainfall event on day 22 (2nd year), only FW showed a slight N 2 O rise (38 µg N 2 O-N m −2 h −1 ). Approaching the end of the experiment, 50 days after digestate application, a small peak (5.22-21.9 µg N 2 O N m −2 h −1 ) was noted within 4 days of continuous rainfall (Figure 1). In both years, WFPS showed a significant positive linear correlation with N 2 O flux rates, r = 0.400 (p < 0.001) and r = 0.454 (p < 0.001) in the 1st and 2nd year, respectively. Similarly, CO 2 fluxes correlated with N 2 O fluxes, exhibiting r = 0.233 (p < 0.001) in the 1st year, and a weaker coefficient of correlation in the 2nd year (r = 0.144, p < 0.001). Soil temperature showed a negative correlation with N 2 O fluxes (r = −0.340; p < 0.001) in the 2nd year, but no significant correlation in the 1st year. All parameters (WFPS, soil temperature and CO 2 ) combined in a linear regression model (model 8) could account for 28.1 to 20.9 % of the prediction of N 2 O fluxes in the 1st and 2nd year, respectively ( Table 3).

Temporal 15 N-N 2 O Fluxes
Total N 2 O and digestate derived 15 N-N 2 O fluxes in the 1st year are shown in Figure 2 and the 2nd year data are shown in Figure 3. A comparable trend was observed for both 15 N t -and 15 NH + 4 -N-labeled digestates in each year, with variations in flux magnitude of 15 N-N 2 O among digestates (Supplementary Tables 2, 3).
In the 1st year, the emerging peak directly after digestate application showed a low 15 N signature, indicating that 92.4-96.5% of N 2 O was derived from soil internal N sources (Supplementary Table 2). Within the first 10 days after digestate application, 15 N-N 2 O fluxes showed no significant differences among treatments (Supplementary Table 2 Table 2). The last major peak appeared on day 24 (10-114 µg 15 N-N 2 O-N m −2 h −1 ), with 10-27% N 2 O-N originating from digestates. Flux rates peaking on day 24 were comparable among most treatments, and only OW exhibited significantly higher flux rates than SB. Prior to peaks of day 24, OW already indicated a rising flux rate on day 20, being significantly higher than all digestates, except G and FW. The flux rate further increased on day 22, where OW significantly exceeded all other treatments (Figure 2 and Supplementary Table 2). Within the first 3 weeks of measurements, flux pattern of CS digestate significantly differed from the other treatments, where 15 N-N 2 O gradually increased after 7 days and reached its maximum on 13 (Figure 2). At both peaks, on day 13 and 24, 15 N-N 2 O flux rates of CS were in a comparable range (48-40 µg 15 N-N 2 O-N m −2 h −1 ). The 2nd year showed a similar temporal pattern in 15 N-N 2 O abundance over the duration of the experiment. The first pulse after digestate application was observed 2 days after application with more than 80% soilborne N 2 O-N. Only CS showed lower soil-borne N 2 O-N, thus highest digestate derived N 2 O-N (45%) among digestates (Supplementary Table 3). The major peak in total N 2 O and 15 N-N 2 O appeared after 1 week (Figure 3)    Unfertilized soil as control and application of Nt -labeled or NH + 4 -N -labeled digestates. Mean values ± standard error (n = 4). Different letters indicate significant differences at p < 0.05 (Tukey Test), ns = no significant differences. Small letters represent statistical differences among all treatments. For cumulative 15 N-N 2 O-N, large letters refer to significant differences only among Nt-labeled or NH + 4 -N-labeled digestates. § no significant differences among Nt -labeled labeled digestates, when excluding NH + 4 -N-labeled digestates. § § no significant differences among NH + 4 -N-labeled digestates, when excluding Nt -labeled digestates. treatment continued to show higher 15 N fluxes compared with the other digestates, even though these emission rates were quite low (from 1.1 to 2.7 µg 15 N-N 2 O-N m −2 h −1 ).

Cumulative N 2 O and 15 N-N 2 O Evolution and Emission Factors
Total cumulative N 2 O emissions in the 1st year (302-1,345 g N 2 O-N ha −1 ) were more than twice as high as in the 2nd year (124-613 g N 2 O ha −1 ) (Figure 4 and Table 4). In both years, digestates lead to significantly higher N 2 O emissions than the unfertilized control. Differences among digestates were observed only in the 1st year, with significantly higher N 2 O emissions for SBL compared to CS and FW (Figure 4). Compared to total N 2 O emissions, digestate-based 15 N-N 2 O emissions indicated larger differences between the different treatments in both years (Figure 4). In the 1st year, G and SBL emitted significantly more 15 N-N 2 O than CS, while all other treatments did not differ significantly. In the 2nd year, highest 15 N-N 2 O emission was  (Table 4). In the 2nd year, there were no significant differences among digestates. Total Nd emitted by the digestates was 16-33% in the 1st, and 17-38% in the 2nd year ( Table 5). Grass digestate tended to show the highest share of digestate derived Nd in both years.
There was no correlation of digestate properties (C/N, C org /N org , NH + 4 -N/N t ) with N 2 O emissions. The respective digestate characteristics did not help to predict cumulative N 2 O or 15 N-N 2 O emissions in a multiple (stepwise forward) linear regression model (model 9).
According to IPCC guidelines, ∼0.6% of the annual amount of total N of organic amendments applied as fertilizer is lost as N 2 O-N in wet climates (IPCC, 2019). In the 1st year, most digestates approached this IPCC EF within only 55 days and SBL even exceeded it with 0.75% (Table 5). Only FW and CS indicated lower EFs than the IPCC default value in the 1st year, with 0.44 and 0.30%, respectively. Related to the overall lower cumulative N 2 O emissions of the 2nd year, mean N 2 O EFs were below 0.33% and in a comparable range for all digestates.
Total 15 N recovery within cumulative N 2 O and soil N at the end of the experiment was 10-57% and 27-64% in the 1st and 2nd year, respectively (Supplementary Table 4). The largest share of digestate 15 N remained in the soil.

Total Global Warming Potential
Cumulative CH 4 -C emissions were significantly higher in the 1st year compared to the 2nd. In both years, unfertilized soil served as CH 4 sink (−147 to −184 g CH 4 -C ha −1 ) ( Table 5).
Within the 1st year, emissions among digestates ranged between 0.26 and 1.82 kg CH 4 -C ha −1 and decreased in the following order SBL ≥ CS ≥ FW, OW, SB ≥ M, G ≥ control. In the 2nd year, digestates as well as unfertilized soil were comparable and digestates indicated CH 4 -C emissions close to zero ( Table 6).
In both years, the release of CO 2 -eq after digestate application was significantly higher than in the control ( Table 7). Significant differences among digestates were only noted in the 1st year, where SBL caused significantly higher total CO 2 -eq. than M, FW, and CS. In both years, N 2 O emissions made up the largest share in total GHG emissions, based on CO 2 -eq, above 85.6% in the 1st and almost 100% in the 2nd year.

Temporal N 2 O and 15 N-N 2 O Fluxes
The high temporal variability of N 2 O fluxes in this study, with increased flux rates after application of crop residues or organic fertilizers and after rainfall events, was similarly documented in 7 | Carbon dioxide equivalents (CO 2 -eq.) of unfertilized soil (control) and soil after application of different digestates, originating from maize, grass, sugar beet, sugar beet leaves, organic waste, food waste, and cattle slurry, based on cumulative N 2 O and CH 4 emissions (kg ha −1 ) after 55 (1st) and 58 days (2nd year); and indirect N 2 O emission as NH 3 -N volatilization over 72 h after application.

Year Treatment
Share of total CO 2− eq. Total CO 2 -eq. other experiments (Pfab et al., 2012;Herr et al., 2019). Ultimately, N 2 O fluxes leveled off 30 days after digestate application in both years with drying of the soil during warm periods with low rainfall. Dry conditions with low WFPS have often been reported to result in low N 2 O emissions from arable soils even if these soils were well-provided with microbial easily degradable C and available N (Möller and Stinner, 2009;Pezzolla et al., 2012). However, in the 2nd year strong rainfall events were recorded 50 days after digestate application and only caused a minor increase in N 2 O fluxes (Figure 1). Hence, digestate-related effects were short-term and had the highest impact on N 2 O release within the first 30 days after application.

First Peak After Digestate Application
Peaks evolving shortly after organic N fertilizer application, such as digestates or manures, have been reported by several studies (Wulf et al., 2002;Johansen et al., 2013;Holly et al., 2017). As indicated by 15 N measurements in both years, the first N 2 O peak after digestate application showed a low 15 N abundance, demonstrating that more than 90% of N 2 O-N was derived from soil N (Figures 2, 3). However, the experimental setup does not allow for a differentiation between nitrification and denitrification. Therefore, we can only conclude that the first N 2 O peak was mainly derived from soil N. The high share of soil-borne N suggests that the addition of OM positively affected microbial activity which further enhanced the turnover of native soil-N, as also stated by Schleusner et al. (2018).
Furthermore, digestate or slurry application moistened the soil close to the applied fertilizer, another factor that has been shown to promote denitrification of NO − 3 -N (Comfort et al., 1990). Moreover, CO 2 flux rates were elevated directly after digestate fertilization (Supplementary Figure 1), supporting the assumption of increased microbial activity which further stimulated denitrification of NO − 3 by O 2 depletion (Buchen-Tschiskale et al., 2020). However, these are only speculations, as soil N min and its 15 N-signature was not measured during the experiment. It should also be considered that digestates contain carbonate-C (HCO − 3 and CO 2− 3 ): the higher the total N content, the higher the carbonate-C content (Georgacakis et al., 1982). Carbonate-C in the digestates can also contribute to soil CO 2 release within the 1st days after application (Chen et al., 2011). For example, carbonate-C release from digestates can occur after application to acidic soils (Chen et al., 2011), which is not the case in present study, or due to microbial turnover processes (Tamir et al., 2013). Therefore, the immediate effects of digestate application on soil microbial activity and the related CO 2 release might be masked by decomposition of carbonate-C to CO 2 . In order to elucidate the driving processes related to the N turnover processes in the soil shortly after digestate application, a more detailed measurement of the pathways of the different fractions of soil and digestate N (NH + 4 , N org ), as well as digestate C (C org , carbonate-C), is necessary.

Rainfall-Induced Peaks
The emission pattern found in present study strongly coincided with the precipitation pattern, providing a major indication that the environmental conditions are the main driving factor for soil N 2 O fluxes. Also the unfertilized control showed a significant increase in N 2 O flux rates after rainfall, whereas almost no fluxes were observed in dry periods. The occurrence of increased N 2 O fluxes in conjunction with heavy rainfall events, hence a high soil WFPS, is typical for arable fields and has extensively been described in the literature Senbayram et al., 2014;Ruser et al., 2017).

Contribution of Digestate N and Soil N Pool to N 2 O Emissions
The largest rainfall-induced N 2 O peaks in both years, had also the highest 15 N abundance, with up to 56-66% of N 2 O-N derived from the digestate (Figures 3, 4 and Supplementary Tables 2, 3). Although it was shown that even at a high soil moisture of 70% WFPS nitrification may also contribute to the N 2 O-release from soils (Ruser et al., 2006), the positive correlations between N 2 O flux rates and CO 2 flux rates as well as between N 2 O fluxes and soil moisture ( Table 3) indicate that denitrification is the driving process releasing N 2 O after rainfall. The contribution of denitrification to the N 2 O release generally increases with increasing soil moisture (Davidson, 1991). When compared to soil air, the ∼10 −4 lower diffusion coefficient for atmospheric O 2 in soil water (Heincke and Kaupenjohann, 1999) restricts O 2 delivery, the creation of anaerobic conditions is favored. Similarly, the turn-over of fresh OM, as indicated by the increased CO 2 fluxes, further depletes O 2 availability and thus fuels anaerobiosis .
The largest peaks evolved 13 (1st year) and 7 days (2nd year) after digestate application, where presumably digestate NH + 4 -N was already nitrified (Johansen et al., 2013) and available for denitrification, thus, explaining the high share of digestate-based N 2 O-N. Senbayram et al. (2009) observed that nitrification of 15 N-labeled digestate rapidly increased 1 week after application in a pot experiment initiating a rise in N 2 O flux rates, as also noted in our study. It cannot be excluded that beside denitrification also nitrification contributed a share of the measured N 2 O. Yet, Köster et al. (2011) measured the intramolecular 15 N distribution in N 2 O within a 43-days incubation experiment, showing that bacterial denitrification was the main process emitting N 2 O after application of food waste digestate, driven by C availability. This is in line with other studies, reporting that the largest N 2 O-N contribution of digestates was caused by denitrification, even at 65% WHC (Senbayram et al., 2009). Later N 2 O peaks (1st year, day 24) showed lower 15 N-N 2 O fluxes, hence indicating an increasing share of N 2 O from soil-internal N. This shift in 15 N-N 2 O abundance over the measuring period indicates increased effects of the soil microbial processes, affecting N availability and N 2 O emissions, and might result from mineralization of digestate 15 N org -N and subsequent processes. A comparable shift was observed by Senbayram et al. (2014).
For both, NH + 4 -N or N t -labeled digestates, the low shares of fertilizer-derived N 2 O-N supported the notion that the largest source of N 2 O was native soil N (> 62%, Table 5). The open hypothesis of an "enhanced soil-derived N 2 O" stated by Senbayram et al. (2014), regarding the low share of emitted digestate-N, can therefore be confirmed. This triggering effect on N 2 O emissions due to digestate application was accounted for by a simplified calculation via equation 4. The amount of triggered N 2 O-N reflects the high share of soil-derived N 2 O, and was approximately half of total cumulative N 2 O emissions ( Table 4). Significant differences in primed N 2 O-N among digestates followed the same trend as N 2 O emissions. For NH + 4 -N-labeled digestates, N 2 O-N losses which might originate from digestate N org , were not accounted for. Therefore, the amount of primed N 2 O-N might be overestimated.
In general, the rather comparable share of total digestatederived N 2 O-N losses among the digestates with different labeling approaches indicates that digestate-N org plays only a minor role in short-term N 2 O formation. Senbayram et al. (2014) labeled only the mineral N fraction of a digestate and found 31% of N 2 O-N was derived from the digestate mineral fraction. The share of digestate-derived N 2 O-N losses among NH + 4 -N-labeled digestates FW, OW and CS ranged from 15.7 to 24.2% over the 2 years. For these digestates as well as for the digestates in the study of Senbayram et al. (2014) it cannot be excluded that nonlabeled organic N was mineralized and emitted as N 2 O. However, the N t -labeled digestates, M, G, SB, and SBL showed a rather comparable range with 18.2-37.8% digestate-N being emitted as N 2 O over the 2 years. Similar to our findings, also other studies reported a higher share of N 2 O-N originating from the soil N pool than from fertilizer N. For instance, only 22% of N 2 O-N was derived from 15 N-labeled manure after 22 days (Ingold et al., 2018) or 40.4% from 15 N-urea after 35 days (Roman-Perez and Hernandez-Ramirez, 2020) in incubation experiments. In a field study, NH + 4 -N-labeled cattle slurry was applied, which produced higher fertilizer-derived N 2 O emissions within the first 10 days, but higher soil-derived N 2 O 11-22 days after application (Dittert et al., 2001). However, the study was carried out on grassland and using the injection technique (Dittert et al., 2001), which has been reported to increase N 2 O emissions compared to trail hose application with immediate incorporation (Herr et al., 2019).

N 2 O Emissions Affected by Fertilizer Type
As previously described, N 2 O fluxes were shaped and influenced by weather conditions and soil microbial processes. Environmental conditions and soil type may play a more important role than the fertilizer type, as previously suggested by Senbayram et al. (2014): the authors noted no significant differences in N 2 O emissions between mineral and organic N fertilization. However, in both years, significant differences among digestates were noted on several sampling dates, for N 2 O as well as 15 N-N 2 O fluxes (e.g., flux rates from M digestate vs. fluxes from SB digestate in Figure 2 and Supplementary Tables 1-3), indicating that digestate composition affects N 2 O emissions. This supports hypothesis (1), that digestates from different feedstocks will differ in N 2 O flux rates.
However, regarding cumulative effects, there was no clear indication that the digestate type influenced total N 2 O emissions. This was supported by the lack of a significant correlation between digestate composition (NH + 4 /N, C/N, C org /N org ) and cumulative N 2 O or 15 N 2 O. Only measurements of the 1st year showed significant differences among digestates. Therefore, hypothesis (2) had to be rejected for the 2nd year and could be partly accepted for the 1st year. Yet, when separating cumulative 15 N-N 2 O data into N t -labeled and NH + 4 -N-labeled digestates, there was a significant effect of C/N ratio in the 1st year, predicting 22.1% of 15 N-N 2 O emissions of 15 N t -labeled digestates (R 2 = 0.36, F-value = 3.81, p-value = 0.077). For the 2nd year, C org /N org accounted for 39.2% (R 2 = 0.56, F-value = 9.85, p-value = 0.0094) of 15 N-N 2 O emissions among 15 N tdigestates (M, G, SB, and SBL). Also Abubaker et al. (2013) noted significantly different cumulative N 2 O emissions after 24 days between two types of urban waste digestates, which were low or high organic C. For NH + 4 -N-labeled digestates, there was no significant relation of digestate properties to 15 N-N 2 O emissions. Hence, the correlation between digestate properties and N 2 O emissions seems more strongly related to the total N and N org content of digestates, than NH + 4 -N. Regarding the total share of digestate-derived N 2 O-N (Nd), significant differences among digestates (Table 5) could support hypothesis (3).
Ultimately, the results of the present study suggest that the different digestate types influenced cumulative N 2 O, flux rates and digestate derived N 2 O-N only marginally. Hence, N 2 O emissions were more strongly affected by environmental conditions ( Table 3). The effect of digestate properties on total N 2 O emissions was overlaid to some extent by the high amount of N 2 O from the native soil N pool. Abubaker et al. (2013) incubated two digestates in three soil textures and noted considerable differences regarding emission peaks and cumulative N 2 O-N emissions among digestates, particularly in the sandy soil. In loam, digestates showed comparable total N 2 O emissions (Abubaker et al., 2013). Therefore, N 2 O emissions discussed in this study might differ on soils with different soil textures or amendment history (Rosace et al., 2020). In this context, soil texture, soil amendment history and fertility status, especially OM content, plays a crucial role, exceeding the effect of digestate properties.

Digestate Mission Factors and Practical Consequences
Digestate EFs determined in this study (0.21-0.75%) were all within the range of the IPCC default value, except for SBL in the 1st year (Table 5). However, these EFs will not cover the whole year and might underestimate the total EF of the digestates. Shang et al. (2020) determined 10-30% lower EFs when only the growing season and not the whole year N 2 O emissions were considered. Moreover, the authors found that the differences between EFs of the whole year and growing season were higher with higher precipitation (Shang et al., 2020). The experimental design of the present study used bare soil, hence there were no crops removing the applied digestate N. Crop N uptake could have decreased available N from the soil as well as soil moisture, which could have lowered digestate-derived N 2 O emissions and EFs. Thorman et al. (2020) determined annual N 2 O EFs from different organic amendments, topdressed to a cereal crop (0.15-0.73% in 2011 and 0.27-0.51% in 2012), which were in a comparable range with our EFs. Most digestate EFs did not show significant differences, except SBL compared with CS in the 1st year, thus hypothesis (2) cannot be fully confirmed.
Soil derived N 2 O-N contributed to a large extent to digestate EFs. As a consequence of the high share of N 2 O-N from the native soil pool within the first 30 days after digestate application, crop cultivation should be synchronized with available soil N. In particular, mineral N from the soil pool should be taken up by the crops, before digestates are applied. Thereby, the triggering effect of short-term soil-enhanced N 2 O emissions by digestates could be decreased. For example, N min supply in the present study would be sufficient for maize cultivation in the early growth stage. Digestates could then be top-dressed ∼1 month after emergence when most of soil M was already taken up by the crop. Also de Neve (2017) emphasized that in ideal cropping systems fertilizer availability and soil mineral N should be synchronized with crop demand, which could mitigate potential N losses.

Experimental Limitations
Determination of N 2 O isotopomers in the present study, including the δ 18 O and site preference of 15 N in the N 2 O molecule, could have helped to understand the underlying soil microbial processes, differentiating between denitrification and nitrification (Köster et al., 2015). Yet, distinguishing nitrifier denitrification from nitrification is not possible using site preference (Köster et al., 2011). A dual isotope labeling approach of 15 N and 18 O-labeled water would be required (Koola et al., 2010), which is not feasible in field studies (Baggs, 2008). Also the N 2 O/N 2 O+ N 2 product ratio could have provided a better indication of denitrification in the study (Buchen-Tschiskale et al., 2020). However, measuring N 2 in the field is rather difficult due to the high N 2 background level in the atmosphere, as well as its spatial and temporal heterogeneity (Groffman et al., 2009). Instead N 2 is often studied in incubation experiments using an artificial helium-oxygen atmosphere (Scholefield et al., 1997). Regular soil N min and 15 N min analysis at the sampling dates could have given a hint for respective microbial processes, but would not have completely identified them. Thus, allocating the specific N 2 O pathways after digestate application in the field is still challenging and needs further research and suitable methods to provide accurate measurements (Well et al., 2019).

CONCLUSION
The major finding of this study was the large share of N 2 O-N from the soil pool, showing that digestate application triggers "enhanced soil-derived N 2 O." The major driving forces of the emission pattern are the weather conditions, the specific chemical composition of digestates do have only minor effects on the denitrification. The different 15 N-labeling approaches of the digestates indicate that contribution of the organic fraction seems to be of very low significance for short-term N 2 O emissions. The 15 N labeling approach helped to determine the source of N 2 O emissions, but not the underlying processes (nitrifier denitrification or heterotrophic denitrification). Analysis of isotopomers and N 2 is needed to further identify the N 2 O-releasing microbial processes in the soil. Emission factors were comparable for most digestates, but reached and even exceeded the default IPPC EF (0.6%) within only 60 days in the 1st year.

DATA AVAILABILITY STATEMENT
The original contributions presented in the study are included in the article/Supplementary Materials, further inquiries can be directed to the corresponding author/s.

AUTHOR CONTRIBUTIONS
FH prepared original draft, including statistical analysis, and graph production. RR supported calculation of global warming potential, provided guidance, and new input for determination of indirect N 2 O emissions. The basis for the experimental idea was based on KM, with further contribution by RR regarding design of the field experiment, sampling frequency, and measurement technique. IC-M supported calculation of 15 N abundance in N 2 O from IRMS data. FH and IC-M conducted the experiment. All authors reviewed and proofread the whole manuscript and gave critical feed-back to all sections.