Rapid Permafrost Thaw Removes Nitrogen Limitation and Rises the Potential for N 2 O Emissions

: Ice–rich Pleistocene permafrost deposits (Yedoma) store large amounts of nitrogen (N) and are susceptible to rapid thaw. In this study, we assess whether eroding Yedoma deposits are potential sources of N and gaseous carbon (C) losses. Therefore, we determined aerobic net ammoniﬁcation and nitriﬁcation, as well as anaerobic production of nitrous oxide (N 2 O), carbon dioxide (CO 2 ), and methane (CH 4 ) in laboratory incubations. Samples were collected from non-vegetated and revegetated slump ﬂoor (SF) and thaw mound (TM) soils of a retrogressive thaw slump in the Lena River Delta of Eastern Siberia. We found high nitrate concentrations (up to 110 µ g N (g DW) − 1 ) within the growing season, a faster transformation of organic N to nitrate, and high N 2 O production (up to 217 ng N 2 O-N (g DW) − 1 day − 1 ) in revegetated thaw mounds. The slump ﬂoor was low in nitrate and did not produce N 2 O under anaerobic conditions, but produced the most CO 2 (up to 7 µ g CO 2 -C (g DW) − 1 day − 1 ) and CH 4 (up to 65 ng CH 4 -C (g DW) − 1 day − 1 ). Nitrate additions showed that denitriﬁcation was substrate limited in the slump ﬂoor. Nitrate limitation was rather caused by ﬁeld conditions (moisture, pH) than by microbial functional limitation since nitriﬁcation rates were positive under laboratory conditions. Our results emphasize the relevance of considering landscape processes, geomorphology, and soil origin in order to identify hotspots of high N availability, as well as C and N losses. High N availability is likely to have an impact on carbon cycling, but to what extent needs further investigation.


Introduction
Recent estimates suggest that if global warming continues, about 10% of the Northern Hemisphere permafrost area will be affected by abrupt permafrost thaw by the end of the century, causing an additional release of 624 Tg carbon dioxide equivalents (CO 2 eq) per year between 2000-2100 [1]. The release of CO 2 eq by rapid permafrost thaw might be even higher, as nitrous oxide (N 2 O) emissions are not yet considered. Studies of collapsed permafrost proved that these sites are not only significant sources of carbon dioxide (CO 2 ) and methane (CH 4 ), but also of N 2 O [2][3][4][5][6]. However, intensive research has been conducted on permafrost carbon-climate feedback, but comparatively less research focused on N biogeochemistry and N 2 O release from degrading permafrost soils. Voigt et al. [7] reviewed in 2020 that there were only 40 published studies on N 2 O across permafrost regions worldwide, thereof only three studies were conducted in the Russian Arctic. This major knowledge gap leads to uncertainties in earth system models [8]. For decades, it was assumed that Arctic ecosystems are generally N limited and that N is directly immobilized by microorganisms and plants and hence are not available for microbial processing [7,[9][10][11]. Nitrous oxide, primarily formed aerobically via nitrification but also anaerobically via denitrification [12], has 298 times higher global warming potential (GWP) than CO 2 [13] and further leads to ozone depletion in the stratosphere [13,14]. In addition to gaseous losses, elevated concentrations of dissolved nitrogen have been found in eroded permafrost [5, [15][16][17][18][19]. In particular deposits from the Late Pleistocene-called Yedoma-are susceptible to thermoerosion due to their high ice content (up to 80 vol-%) [20,21]. The Yedoma domain (Yedoma deposits covered by sediment layers from the Pleistocene and Holocene), is suggested to store about 6 to 8 times (15)(16)(17)(18)(19)(20)(21)(22)(23) kg N m −2 , total: 40-60 Pg N) [22] more N than permafrost-affected peatlands (4.2 kg N m −2 , total: 7.1 Pg N) [23], for an area only 1.5 times larger. However, permafrost-affected peatlands are more commonly discussed as N hotspots [7], but N release from thawing Yedoma coasts represents an important source of N losses into aquatic systems [6,24]. About 78% of the coastal regions of the Laptev Sea are already affected by thermoerosion [25], and retrogressive thaw slumps (RTS)-typical features formed by abrupt permafrost thaw-are frequently observed along riverbanks and coastlines [26][27][28]. Fuchs et al. [24] reported that 0.4 × 10 6 kg of N were exported annually from Sobe-Sise Island (Siberia) into the Lena River between 2015 and 2018. Nitrogen in form of nitrate is particularly mobilized into the Lena River in winter months [29].
Thermoerosion could favor N mineralization (sum of ammonification and nitrification) and in this way N losses by affecting moisture, temperature, texture, pH, as well as organic matter distribution in soils [3,30]. N mineralization in Yedoma sediments was found to be functionally limited by a low abundance of ammonia oxidizers [6,31]. As ammonia-oxidizing bacteria and archaea are most abundant in the active layer [32], N mineralization in Yedoma sediments might be favored by being mixed with younger sediments through erosion. Furthermore, mixing organic-rich topsoils with carbon (C) limited mineral subsoils, could stimulate soil organic matter (SOM) decomposition through priming [33,34]. Specifically, reduced competition for N between plants and microbes due to the abrasion of the vegetation cover could enhance the microbial transformation of N and thus N losses. To date, it has not been elucidated how the stabilization of RTS by revegetation affects N availability and greenhouse gas (GHG) emissions [6,26].
In this study, we investigate potential N and C release from non-vegetated and revegetated soils with different origins from a retrogressive thaw slump (RTS) formed at a coastal Yedoma cliff on Kurungnakh island in the Lena River Delta (Siberia). We hypothesize (I) that revegetated soils have a lower N availability and thus lower N 2 O production due to plant N uptake. We further suggest (II) that anaerobic N 2 O production is predominately limited by substrate availability and that nitrate additions would stimulate N 2 O production via denitrification. Due to different soil origin and landscape processes, we further expect (III) enhanced N availability, N 2 O production, and (IV) total release of CO 2 eq from soils where erosion mixed the Holocene cover with Pleistocene aged Yedoma deposits (slump floor) than from soils consisting only of Yedoma (thaw mounds).

Study Sites
The studied thaw slump is located in the southeast of the island Kurungnakh within the Lena River Delta in North Siberia (72.339 • N, 126.292 • E). The study area is underlain by continuous permafrost and is characterized by a continental Arctic climate. The average annual air temperature and average annual rainfall measured on the neighboring island Samoylov are −12.3 • C and 169 mm, respectively [35]. Soil temperatures at 6 cm depth vary between −28 and 10 • C [36]. On Kurungnakh, ice-rich Yedoma sediments with large ice wedges, which formed during the Late Pleistocene, are covered by Holocene deposits [37]. Combined rapid permafrost thaw and thermoerosion of the massive icecomplex led to the formation of a retrogressive thaw slump (RTS). By erosion of the headwall, surface vegetation and deposits from the Holocene and Pleistocene (Yedoma) got mixed and transported down the slope, forming the slump floor (SF). The former polygon centrers-baydjarakhs (russ.) or thaw mounds (TM)-are exposed within the slump floor ( Figure 1). They are not affected by relocation and consist only of Yedoma deposits. All soils were classified as Protic Cryosols due to the presence of permafrost and missing soil horizon development [26]. The plant cover at the thaw slump was disturbed, but increasing revegetation could be observed between 2013 and 2019. Revegetated sites are covered with mosses, grasses like Puccinellia neglecta, and herbs such as Tephroseris palustris and Descurainia sophioides. According to Laschinskiy et al. [38], the thaw slump can be classified into the first and second phases of plant succession after thermoerosion.
to the formation of a retrogressive thaw slump (RTS). By erosion of the headwall, surface vegetation and deposits from the Holocene and Pleistocene (Yedoma) got mixed and transported down the slope, forming the slump floor (SF). The former polygon centrersbaydjarakhs (russ.) or thaw mounds (TM)-are exposed within the slump floor ( Figure  1). They are not affected by relocation and consist only of Yedoma deposits. All soils were classified as Protic Cryosols due to the presence of permafrost and missing soil horizon development [26]. The plant cover at the thaw slump was disturbed, but increasing revegetation could be observed between 2013 and 2019. Revegetated sites are covered with mosses, grasses like Puccinellia neglecta, and herbs such as Tephroseris palustris and Descurainia sophioides. According to Laschinskiy et al. [38], the thaw slump can be classified into the first and second phases of plant succession after thermoerosion. . The ice-rich headwall can be seen on the right site in the picture. Vegetation density increases with distance from the thaw front and hence time since thawing. Accordingly, sites SF6 and TM2 were more recently thawed than sites SF3 and TM1.

Soil Sampling and Sample Preparation
Soil samples were taken in July 2019 from revegetated (V) and non-vegetated (NV) sites from two different spots from the slump floor (sites SF3 and SF6) and two different thaw mounds (sites TM1 and TM2) (Figures 1 and S1, Table S1). Due to the retrogression of the headwall, the sites have been thawing for different lengths of time. Soils from sites SF6 and TM2 were more recently thawed than soils from sites SF3 and TM2 ( Figure 1). Unlike all other revegetated sites, soils from site SF6, which were closest to the headwall, were only covered by mosses and not by grasses ( Figure S1). In 2019, the thaw mound at site TM1 was already partly eroded and samples were taken at a lower part while TM2 was still intact and samples were taken from the upper part of the thaw mound. Surface waters accumulated at the sampled slump floor in July 2019, while the sampled thaw . The ice-rich headwall can be seen on the right site in the picture. Vegetation density increases with distance from the thaw front and hence time since thawing. Accordingly, sites SF6 and TM2 were more recently thawed than sites SF3 and TM1.

Soil Sampling and Sample Preparation
Soil samples were taken in July 2019 from revegetated (V) and non-vegetated (NV) sites from two different spots from the slump floor (sites SF3 and SF6) and two different thaw mounds (sites TM1 and TM2) (Figures 1 and S1, Table S1). Due to the retrogression of the headwall, the sites have been thawing for different lengths of time. Soils from sites SF6 and TM2 were more recently thawed than soils from sites SF3 and TM2 ( Figure 1). Unlike all other revegetated sites, soils from site SF6, which were closest to the headwall, were only covered by mosses and not by grasses ( Figure S1). In 2019, the thaw mound at site TM1 was already partly eroded and samples were taken at a lower part while TM2 was still intact and samples were taken from the upper part of the thaw mound. Surface waters accumulated at the sampled slump floor in July 2019, while the sampled thaw mounds were well drained and had large cracks penetrating up to 40 cm into the soil. From the surface down to the permafrost (Table 1) (30 to 110 cm), two to five soil samples (depending on sampleable material) as independent replicates were taken every 10 cm to account for small-scale variability. Samples were transported frozen (−18 • C) to Hamburg (Germany), and composite samples between 0-10 and 10-20 cm ( Figure S2) were created in a N 2 (purity: 5.0) flushed glove box under near anoxic conditions. We also analyzed subsoil samples from site TM2 NV (20-40 cm and 40-70cm, Figure S2), which are presented in the Supplementary Material (Table S2). Table 1. Active layer (AL) depth and soil properties (water content (WC), water-filled pore space (WFPS), soil pH (pH), total carbon (TC), total carbon-nitrogen ratio (C/N), delta 15 N in bulk soil (δ 15 N), water-extractable organic carbon (WEOC), and phosphorus (P)) of sampled revegetated (V) and non-vegetated (NV) slump floor (SF) and thaw mounds (TM) soils (n = number of replicates). Values (mean ± standard deviation), * n = 4, ** n = 2.

Soil Analysis
After creating composite samples (see above), soils were sieved (<2 mm) and soil water content (WC) was determined by the weight difference between fresh and samples which were dried for 24 h at 105 • C. Bulk density (BD) was determined for each depth interval (0-10 and 10-20 cm) by subtracting the water content from the mass of the volumetric fresh samples. For composite samples, the mean BD from 0-10 and 10-20 cm was calculated. Particle density (PD) was determined by a gas (He) pycnometer (AccuPyc II 1340, Micromeritics Instrument Corporation, Norcross, GA, USA). The water-filled pore space (WFPS) was calculated according to Haney and Haney (2010) [39] on basis of WC, mean BD, and PD. Dried samples were analyzed for total carbon (TC) and nitrogen with an elemental analyzer (VarioMAX, Elementar Analysensysteme GmbH, Hanau, Germany). Delta δ 15 N in bulk soil was determined using IRMS (Thermo Delta V Plus, Thermo Fisher Scientific, Bremen). Soil pH was measured in a suspension of 2 g fresh soil in 5 mL deionized water with a potentiometric pH-meter (Typ CG820, Xylem Analytics Germany Sales GmbH & Co. KG., Weilheim, Germany). Dissolved inorganic nitrogen (DIN) was extracted from 5 g fresh soil in 20 mL 0.0125 M CaCl 2 . Ammonium was analyzed photometrically at 655 nm [40] (DR 5000, Hach Lange GmbH, Düsseldorf, Germany) while nitrite and nitrate were analyzed by high-performance liquid chromatography (HPLC) (1200 Series, Agilent Technologies, Santa Clara, CA, USA). Total dissolved nitrogen (TDN) was determined by extracting 12.5 g fresh soil in 50 mL 0.5 M K 2 SO 4 and analyzed by a TOC/TON analyzer (Shimadzu TNM-L, Shimadzu Corporation, Kyoto, Japan) [41]. Dissolved organic nitrogen (DON) was calculated by subtracting DIN from TDN. Water-extractable organic carbon (WEOC) was determined according to Surey et al. [42] by suspending 10 g dry soil in 100 mL deionized water and a subsequent analysis by a TOC/TON analyzer (Shimadzu TNM-L, Shimadzu Corporation, Kyoto, Japan). Labile phosphorus (P) was extracted from 2.5 g air-dried soil in 0.02 M 125 mL calcium lactate solution with a pH of 3.6 [43] and subsequently analyzed photometrically at 882 nm [44] (Thermo Spectromic Genesys 10-S, Thermo Fisher Scientific, MA, USA).

Nitrogen Transformation
Net N transformation rates were determined by analyzing ammonium and nitrate concentrations from aerobic and continuously horizontally shaken (125 rpm) slurries which were incubated at 4 • C in a dark environment. Since a lag phase of microbial activity was expected during the first weeks [29], an incubation period of 40 days for nitrogen transformation and GHG production (see Section 2.5) was chosen in order to observe a maximal increase. On a clean bench, approximately 25 g of fresh and sieved soil and 100 mL phosphate buffer was weighed into 300 mL Erlenmeyer flasks with loose lids, enabling air exchange. The phosphate buffer was prepared by mixing 28 mL 0.2 M KH 2 PO 4 with 72 mL 0.2 M K 2 HPO 4 and 100 mL deionized water, followed by 1:100 dilution with deionized water and adjustment to pH 7.2 with 0.5 M HCl. Twice per week, samples from the slurry were taken with sterile pipette tips and centrifuged at 16,000 g for 10 to 20 min (20 min for organic-rich samples), and ammonium and nitrate concentrations were measured in the supernatants (as described above). Nitrate was analyzed directly after sampling while one part of the supernatant was frozen at −20 • C and analyzed later for ammonium. Since there was no constant increase in ammonium and nitrate over time, maximum net ammonification and maximum net nitrification rates were determined by calculating the maximal difference between the two measurements for ammonium and nitrate, respectively. On average, the maximum rates were observed after 20 days. Maximum rates of net ammonification and net nitrification were not necessarily observed between the same measurement points, therefore net N mineralization as the sum of ammonification and nitrification is not provided.

Greenhouse Gas Production in Anaerobic Laboratory Incubations
For the anaerobic incubation, 25 g of fresh unsieved soil was weighed into 100 mL glass flasks flushed with N 2 under nearly anoxic conditions in a glove box. Subsequently, the headspace was flushed again with pure N 2 , and samples were incubated for 40 days at 4 • C in a dark environment. Headspace gas concentrations were analyzed with a gas chromatograph (7890 A, Agilent Technologies, Santa Clara, CA, USA) using a FID with methanizer for CO 2 and CH 4 and an ECD for N 2 O. Fluxes were calculated from cumulative gas production and gas dissolution in porewater calculated according to Carroll et al. [45] and Millero et al. [46] for CO 2 , Yamamoto et al. [47] for CH 4 and Weiss & Price [48] for N 2 O. Carbon dioxide production was calculated based on a slope of a linear regression fitted to four measurement points. The mean of all slopes calculated over time is considered with the CO 2 production rate. In contrast, for anaerobic CH 4 and N 2 O production, there was no constant production observable, therefore maximal rates were calculated by the maximal difference between the two measurements. On average, the maximum rates were observed after 10 and 1 days for CH 4 and N 2 O, respectively. CO 2 equivalents (CO 2 eq) were calculated by multiplying production rates with the respective global warming potential (GWP) of a 100-year horizon (CH 4 × 34; N 2 O × 298) [13]. CO 2 eq was given as the sum of all three GHG.

Nitrate Addition
After four weeks (28 days), 1 mL of 152 mM sterilized nitrate (KNO 3 ) was added through a septum into the anaerobically incubated flasks, after they were flushed with pure N 2 . The added nitrate concentration was based on the highest concentration measured in samples from site TM1 V (110 µg N gDW −1 ). GHG production rates were determined before and after the substrate addition and were treated separately. By adding 1 mL KNO 3 to the soil, the WFPS increased by about 10% ( Figure S3). At the time nitrate was added, N 2 O concentrations in the headspace were declining or zero.

In-Situ N 2 O Fluxes
In-situ N 2 O fluxes from the soil surface were quantified only from non-vegetated sites with triplicates for each site on three different days in July 2019. At site TM2 NV only two measurement days were possible. Fluxes were measured by static closed chamber measurements and subsequent GC-ECD analysis of headspace gas samples. For each measurement, an opaque gas chamber (volume: 0.01 m 3 ) was placed on top of the preinstalled collars (diameter: 24.2 cm) and the headspace was mixed by an internal fan. Collars were installed 24 h before the first measurement. Headspace gas concentrations increased differently with time, therefore five gas samples were taken in varying time intervals between 10 and 25 min from a three-way-valve installed on top of the chamber using a syringe (Braun Melsungen AG, Melsungen, Germany). The samples were directly transferred to pre-flushed (N 2 ) and subsequently evacuated 12 mL glass vials (Exetainer ® , Labco Limited, Lampeter, UK) and stored under overpressure. Nitrous oxide concentrations of the gas samples were measured after six months using a GC-ECD (7890 A, Agilent Technologies, Santa Clara, CA, USA). Nitrous oxide emissions were calculated from the linear increase of N 2 O concentrations in the chamber over 10 to 20 min, temperature, chamber volume, and collar area. Average flux rates from 2 to 4 sampling dates (Table S6) are presented.

Statistics
Statistical analyses were conducted with IBM SPSS STATISTICS 27. Normal distribution was tested by applying the Shapiro-Wilk test (p > 0.05) and visual consideration of Q-Q plots and histograms. Correlations for non-parametric data sets were determined by using the Spearman test while parametric data sets were correlated by the application of the Pearson test. Correlations were calculated based on topsoil and subsoil samples including 30 samples. Due to the limited amount of sample material, correlations with WEOC were only based on 22 topsoil samples. Statistically, significant differences were tested between slump floor and thaw mound topsoils, revegetated (V) and non-vegetated (NV) slump floor topsoils, as well as between revegetated (V) and non-vegetated (NV) thaw mound topsoils ( Table 2). The Mann-Whitney-U test (p < 0.05) was applied to test significant differences for non-parametric datasets. Parametric data sets were first tested on variance homogeneity with the Levene test (p > 0.05), followed by a t-test (p < 0.05) for equal variances or Welch-test (p < 0.05) for unequal variances. Table 2. Significance of differences between sites: thaw mound (TM) and slump floor (SF) sites, revegetated (TM V) and non-vegetated (TM NV) thaw mounds, as well as revegetated (SF V) and non-vegetated slump floor (SF NV).

Soil Properties
Soil properties of slump floor and thaw mound soils differed in most of their physical and chemical properties (Table 1). Slump floor soils were generally wetter with higher WC and WFPS as they were more strongly affected by meltwaters from the headwall while thaw mound soils were drier with lower WC and WFPS due to the increased drainage by cracks and their exposed position ( Figure S3). Since site TM1 was already partly eroded in 2019, samples were taken at a lower location which likely resulted in a higher water content relative to site TM2 (sampled at the top of the thaw mound). Slump floor soils had a moderately acidic pH (5-6), while thaw mound soils had a moderately alkaline pH (7)(8). The average total carbon (TC) (t-test, p < 0.01) and C/N ratio (Mann-Whitney-U test, p < 0.01) were significantly higher in slump floor (TC: 3.4-6.5%; C/N: 15.3 to 19.2) than in thaw mound soils (TC: 2.0-4.4%; C/N: 12. 8-14.4). No significant contrasts were observed for WEOC and P between slump floor (WEOC: 1302-2083 µg C (g DW) −1 ; P: 7.0 to 18.6 µg P (g DW) −1 ) and thaw mounds (WEOC: 753-1919 µg C (g DW) −1 ; P: 9.0 to 22.9 µg P (g DW) −1 ). However, most of the parameters were enhanced in the revegetated soils, but this observation was only significant for soil pH (Welch, p < 0.01), C/N ratio (Mann-Whitney-U test, p < 0.05), and WEOC (Welch, p < 0.05) in slump floor soils.

Nitrogen Pool Composition
Slump floor soils had significantly higher DON (19.05-37.54 µg N (g DW) −1 ) and ammonium (0.06-16.73 µg N (g DW) −1 ) concentrations (Mann-Whitney-U test, p < 0.01) than thaw mound soils with low DON (3.06-9.30 µg N (g DW) −1 ) and ammonium concentrations close to the detection limit. In contrast, nitrate concentrations were significantly higher in the thaw mounds (7.05 to 81.64 µg N (g DW) −1 ) (Mann-Whitney-U test, p < 0.01) and near zero in the slump floor, except at site SF3 NV where the second highest nitrate concentration was measured (mean ± standard deviation: 31.23 ± 6.20 µg N (g DW) −1 ) ( Figure 2, Table S3). Soils from site SF3 NV contained relatively high amounts of DON (19.05 ± 4.22 µg N (g DW) −1 ) and ammonium (6.31 ± 2.36 µg N (g DW) −1 ) in comparison to thaw mound soils, where high nitrate concentrations were observed together with relatively low DON and ammonium availability. While in slump floor soils DON exceeded DIN (not at site SF3 NV), in thaw mound soils DIN exceeded DON due to high nitrate availability. There was 1.6 times more DIN and 3.2 times less DON in thaw mounds than in the slump floor. Nitrogen availability was generally enhanced in non-vegetated slump floor soils in comparison to revegetated slump floor soils, but only for nitrate significant (Mann-Whitney-U test, p < 0.05). However, this statistical significance was caused by high nitrate con-centrations in samples from site SF3 NV. Revegetation had no significant effect on N availability in thaw mounds, but nitrate concentrations were relatively enhanced in revegetated thaw mounds and the highest nitrate concentration was measured in soils from site TM1 V (81.64 ± 24.37 µg N (g DW) −1 ). Dissolved organic nitrogen correlated positively with C/N ratio (R = 0.49, p < 0.01) soil pH (R = 0.55, p < 0.01), and ammonium (R = 0.78, p < 0.01) while ammonium correlated positively with C/N ratio (R = 0.63, p < 0.01) and δ 15 N in bulk soil (R = 0.51, p < 0.01) but also negatively with WFPS (R = −0.38, p < 0.05). Delta δ 15 N in bulk soil correlated negatively with WC (R = −0.73, p < 0.01) and WFPS (R = −0.73, p < 0.01). There was no similar correlation with nitrate. None of the samples contained nitrite.

Nitrogen Transformation
Under optimal laboratory conditions (oxygen supply and neutral pH), net ammo fication, as well as net nitrification were positive (uptake < production) in all soils, a only in soils from site SF6 NV no nitrification was observed (Table 3). Soils from site NV had the highest net ammonification rates (69.51 ± 5.26 ng N (g DW) −1 day −1 ). Howe in all other samples, net nitrification rates exceeded net ammonification rates and were average about 22 times higher. Net ammonification rates were significantly higher in slump floor than in the thaw mounds (Welch, p < 0.01), varying between 23.25 to 69.   Table S3 in the supplements.

Nitrogen Transformation
Under optimal laboratory conditions (oxygen supply and neutral pH), net ammonification, as well as net nitrification were positive (uptake < production) in all soils, and only in soils from site SF6 NV no nitrification was observed (Table 3). Soils from site SF6 NV had the highest net ammonification rates (69.51 ± 5.26 ng N (g DW) −1 day −1 ). However, in all other samples, net nitrification rates exceeded net ammonification rates and were on average about 22 times higher. Net ammonification rates were significantly higher in the slump floor than in the thaw mounds (Welch, p < 0.01), varying between 23.25 to 69.51 ng N (g DW) −1 day −1 in slump floor and 1.97 to 7.60 ng N (g DW) −1 day −1 in thaw mound soils. However, there was no statistically significant difference in net nitrification between slump floor and thaw mounds (Mann-Whitney-U test, p = 0.406) (SF: 0.00-1088.40 ng N (g DW) −1 day −1 ; TM: 197.13-1437.41 ng N (g DW) −1 day −1 ). Revegetation had no significant impact on N transformation rates. Net ammonification and net nitrification were not correlated. Additionally, net ammonification correlated positively with TC (R = 0.54, p < 0.01), WC (R = 0.30, p < 0.40), anaerobic CO 2 (R = 0.63, p < 0.01), CH 4 (R = 0.68, p < 0.01), and N 2 O (R = 0.49, p < 0.01) production, and additionally negatively with soil pH (R = −0.52, p < 0.01), as well as with nitrate (R = −0.54, p < 0.01). None of these parameters correlated with net nitrification. Table 3. Net N transformation rates of incubated soil slurries (4 • C for 40 days) from revegetated (V) and non-vegetated (NV) slump floor (SF) and thaw mounds (TM) soils under oxygen supply and pH 7.2 (n = number of replicates).

Anaerobic N 2 O Production and In-Situ N 2 O Fluxes
Anaerobic N 2 O production was most strongly dependent on initial nitrate availability (R = 0.74, p < 0.01) thus high N 2 O production rates were observed particularly from thaw mound soils (0.00 to 216.63 ng N 2 O-N (g DW) −1 day −1 ) while N 2 O production from slump floor soils was near or below the detection limit, except for site SF3 NV where the second highest N 2 O production was observed with 95.43 ± 43.21 ng N 2 O-N (g DW) −1 day −1 (Figure 3, Table S4). Revegetation significantly decreased N 2 O production in the slump floor (Mann-Whitney-U test, p < 0.05), but significance was caused by high N 2 O production from site SF3 NV soils. Nitrous oxide production was elevated in revegetated thaw mounds, but not statistically significant (Mann-Whitney-U test, p > 0.05). The highest N 2 O production was determined from soils from site TM1 (216.63 ± 158.25 ng N 2 O-N (g DW) −1 day −1 ). In addition, we noticed that soils from site TM1 NV did not produce N 2 O, although having higher nitrate concentrations than slump floor soils (7.05 ± 6.26 µg N (g DW) −1 ). Nitrous oxide production correlated moderately with the WC (R = −0.46, p < 0.05), but not with the WFPS. Anaerobic N 2 O production correlated positively with WEOC (R = 0.52, p < 0.05) and δ 15 N in bulk soil (R = 0.49, p < 0.01), but not with P.  Values (mean ± standard deviation).

Anaerobic N2O Production and In-Situ N2O Fluxes
Anaerobic N2O production was most strongly dependent on initial nitrate availability (R = 0.74, p < 0.01) thus high N2O production rates were observed particularly from thaw mound soils (0.00 to 216.63 ng N2O-N (g DW) −1 day −1 ) while N2O production from slump floor soils was near or below the detection limit, except for site SF3 NV where the second highest N2O production was observed with 95.43 ± 43.21 ng N2O-N (g DW) −1 day −1 (Figure 3, Table S4). Revegetation significantly decreased N2O production in the slump floor (Mann-Whitney-U test, p < 0.05), but significance was caused by high N2O production from site SF3 NV soils. Nitrous oxide production was elevated in revegetated thaw mounds, but not statistically significant (Mann-Whitney-U test, p > 0.05). The highest N2O production was determined from soils from site TM1 (216.63 ± 158.25 ng N2O-N (g DW) −1 day −1 ). In addition, we noticed that soils from site TM1 NV did not produce N2O, although having higher nitrate concentrations than slump floor soils (7.05 ± 6.26 µg N (g DW) −1 ). Nitrous oxide production correlated moderately with the WC (R = −0.46, p < 0.05), but not with the WFPS. Anaerobic N2O production correlated positively with WEOC (R = 0.52, p < 0.05) and δ 15 N in bulk soil (R = 0.49, p < 0.01), but not with P. After 28 days, nitrate was added to the incubation flasks to evaluate whether denitrification in soils with missing N 2 O production was limited by substrate availability (Figure 3, Table S4). Nitrous oxide production increased significantly (Mann-Whitney-U test, p < 0.01) after nitrate addition in slump floor soils from zero up to 3312.74 ± 426.27 ng N 2 O-N (g DW) −1 day −1 (SF6 NV), but increased only weakly in thaw mound soils and in soils from site SF3 NV. Nitrous oxide production rates after nitrate addition correlated positively with carbon components as TC (R = 0.56, p < 0.01) and C/N (R = 0.47, p < 0.01) but not with WEOC or P. The headspace gas concentration was resampled after six months and N 2 O concentrations declined to near zero in all samples except in replicates from site SF3 V and SF3 NV, as well as from site TM1 V and TM1 NV where N 2 O concentrations declined slower and were still between 16 and 1867 ppm.
In-situ N 2 O fluxes were only measured from non-vegetated sites on three different days in July 2019 (Table 4). However, in-situ fluxes varied strongly between different days and ranged on average between 0.01 and 2.03 mg N 2 O-N m −2 day −1 . Comparable to N 2 O production from the incubation, site SF3 NV, which had a high nitrate availability and an optimal WFPS, also produced the most N 2 O in-situ (2.03 ± 1.39 mg N 2 O-N m −2 day −1 ), but otherwise in-situ measurements did not follow the pattern of the incubation experiment.  (Table S6). Anaerobic CH 4 (ng CH 4 -C (g DW) −

Anaerobic CO 2 and CH 4 Production
Besides N 2 O, also CO 2 and CH 4 production from anaerobic laboratory incubations were measured (Table 4). Anaerobic CO 2 production was on average 164 times higher than CH 4 production and both rates were positively correlated (R = 0.64, p < 0.01). Carbon dioxide and CH 4 production were both significantly higher (Mann-Whitney-U test, p < 0.01 for CO 2 and p < 0.05 for CH 4 ) from slump floor (CO 2 : 0.99-7.07 µg CO 2 -C (g DW) −1 day −1 ; CH 4 : 0.00 to 65.46 ng CH 4 -C (g DW) −1 day −1 ) than from thaw mound soils (CO 2 : 0.24-2.34 µg CO 2 -C (g DW) −1 day −1 ; CH 4 : 0.11-7.80 ng CH 4 -C (g DW) −1 day −1 ). Revegetation had only a positive statistically significant impact on CO 2 (Mann-Whitney-U test, p < 0.01) and CH 4 (Mann-Whitney-U test, p < 0.05) production from slump floor soils. Both, CO 2 and CH 4 production correlated positively with TC (R = 0.52-0.64, p < 0.01). We found a negative correlation for CO 2 and CH 4 with nitrate and N 2 O (R = −0.65 to −0.67, p < 0.01). Consistent with that, replicates from site SF3 NV, which contained the second highest nitrate concentration, had a 62 to 85% lower CO 2 production (0.99 ± 0.33 µg CO 2 -C (g DW) −1 day −1 ) than other slump floor samples although WC and TC were in a similar range. As long as there was N 2 O in the headspace, we observed no CH 4 production and after the N 2 O concentration declined, CH 4 production started on the same measurement day even though no reliable strong production was indicatable within the observation period ( Figure S4). In addition, samples, which did not produce N 2 O in the beginning but CH 4 , stopped producing CH 4 after they were flushed with N 2 , and artificial nitrate was added ( Figure S4). Carbon dioxide and CH 4 production correlated also negatively with soil pH (CO 2 : R = −0.65, p < 0.01, CH 4 : R = −0.47, p < 0.01). Carbon dioxide production correlated additionally positively with DON (R = 0.38, p < 0.05) and WEOC (R = 0.64, p < 0.01).

Anerobic Production of CO 2 Equivalents
As long as no N 2 O production was observed in the anaerobic incubation, the production of CO 2 equivalents (CO 2 eq) was dominated by CO 2 (76-92%), otherwise, N 2 O was the major contributor (94-98%) (Figure 4, Table S5). Methane production had only a minor impact on CO 2 eq production (0-24%). No significant impact of revegetation was observed when GHG production was summarised as CO 2 eq. In general, CO 2 eq from the slump floor was dominated by CO 2 (83%) (except site SF3 NV, high N 2 O production) while CO 2 eq from the thaw mounds mainly came from N 2 O (95%). Consequently, under anaerobic conditions, thaw mounds had the highest production of CO 2 eq (SF: 12.37 ± 11.63 µg CO 2 eq-C (g DW) −1 day −1 ; TM: 21.21 ± 35.16 µg CO 2 eq-C (g DW) −1 day −1 ). However, after nitrate was added, the production rates of CO 2 and CH 4 decreased from slump floor soils, but since N 2 O production increased, mean CO 2 eq were about 45 times higher than before nitrate addition (559.60 ± 35.73 µg CO 2 eq-C (g DW) −1 day −1 ). Since N 2 O production from thaw mound soils was less affected by nitrate addition, CO 2 eq slightly decreased (33%) if compared to the CO 2 eq from before.
there was N2O in the headspace, we observed no CH4 production and after the N2O co centration declined, CH4 production started on the same measurement day even thoug no reliable strong production was indicatable within the observation period ( Figure S In addition, samples, which did not produce N2O in the beginning but CH4, stopped pr ducing CH4 after they were flushed with N2, and artificial nitrate was added (Figure S Carbon dioxide and CH4 production correlated also negatively with soil pH (CO2: R −0.65, p < 0.01, CH4: R = −0.47, p < 0.01). Carbon dioxide production correlated additional positively with DON (R = 0.38, p < 0.05) and WEOC (R = 0.64, p < 0.01).

Discussion
This study differentiated between slump floor (SF) and thaw mound (TM) soils within a RTS and found differences in soil properties, N availability, and GHG release. Differences can be attributed to different soil origins and landscape processes. Slump floor soils contain mixed Pleistocene and Holocene material and are actively affected by persistent erosion and meltwaters while thaw mound soils consist solely of Pleistocene Yedoma deposits and are less affected by disturbance due to their exposed position. Contrary to our hypotheses, revegetated thaw mounds had the highest N availability, produced the most N 2 O, and released more CO 2 eq than non-vegetated soils from the slump floor.

Nitrogen Availability
Most studies on Arctic permafrost soils describe higher N availability in the absence of vegetation [2,5,49,50] and a domination of DON over DIN, with higher ammonium than nitrate concentrations [3,[51][52][53][54]. In general, Arctic ecosystems are N-limited, and available N is directly consumed by plants and microbes, often leading to a tight N cycle and negative N transformation rates [6,10,[55][56][57][58]. Despite our expectations (I and III), we found the highest N availability in revegetated thaw mounds and not in soils from nonvegetated parts of the slump floor, suggesting that competition for N between plants and microbes is low and that N availability was affected by soil origin and landscape processes rather than by revegetation. Different from previous studies [3,[51][52][53][54], we report that DIN, dominated by nitrate, exceeded DON in thaw mound soils. Nitrate concentrations in thaw mounds exceeded most of those reported from other (disturbed) permafrost-affected soils [15,16,19,51,52], but were in a comparable range to those reported from soils from Arctic lowlands of Canada [59], non-vegetated peat plateaus in the subarctic tundra of Russia [2], upland thermokarst soils on the Tibetan Plateau [5], and thaw mounds soils from a Siberian RTS [6]. Significantly smaller net ammonification rates in thaw mound than in slump floor soils, as well as lower DON and ammonium concentrations with simultaneously elevated nitrate concentrations, indicate a faster turnover of DON to nitrate by immobilization and nitrification in thaw mounds. Positive N transformation rates and higher net nitrification rates than the mean nitrification reported from permafrost-affected soils (−700 to −100 ng N (g DW) −1 day −1 ), [10], confirm that microbes at the thaw slump of Kurungnakh are not generally N limited. However, higher nitrate concentrations in thaw mounds might be not only due to an enhanced microbial N turnover, but also due to the high N content in Pleistocene Yedoma deposits [22].
Nitrification is particularly favored in thaw mounds as their exposed position and deep cracks lead to drainage and aeration ( Figure 5A). Similarly, Marushchak et al. [6] found a comparable high concentration (116 µg N (g DW) −1 ) from a moss-covered thaw mound at Duvanny Yar exposure near the Kolyma River (Siberia) (max. at site TM1 V: 110 µg µg N (g DW) −1 ) and Tanski et al. [19] found elevated nitrate concentrations in cracked and well-aerated surfaces at a slump floor at a thaw slump in Alaska. Strong differences in N pools between the slump floor and thaw mounds were most probably not linked to microbial functional limitation (III), as suggested in Monteux et al. [31] and Marushchak et al. [6], since our N transformation experiment showed a continuous nitrate production (except site SF6 NV) under optimal conditions with oxygen supply and at a neutral pH. It is more likely that the high ammonium concentrations in slump floor soils were caused by an inhibited nitrification under field conditions due to high soil moisture [60][61][62] and acidic pH ( Figure 5A), despite the fact that there were only negative correlations between ammonium and WFPS, between δ 15 N and WC and WFPS, as well as between DON and soil pH. In particular, ammonia oxidizers (archaea and bacteria) prefer oxic conditions and react sensitively to elevated soil moisture [61]. The impact of soil pH on nitrification is contentious and depends strongly on the microbial community [63][64][65], although at a neutral pH less ammonia is converted into ammonium, being generally optimal for nitrification. The SF is affected by meltwaters increasing soil moisture. High soil moisture and low pH of SF soils slow down or inhibit nitrification leading to an accumulation of DON and ammonium (NH4 + ). N2O via nitrification and CO2 losses from TM soils are relatively small [6,26] and more CO2 equivalents are produced from the SF. (B): When the TMs become anaerobic in the end or beginning of the next growing season after rain events or snow melt, accumulated nitrate can be reduced via denitrification to N2O and less CO2 is produced due to the slowdown of SOM decomposition. When the SF becomes predominately anaerobic, it produces in addition to CO2, also CH4, but no N2O since denitrification is nitrate limited at the SF. Due to the GWP of N2O, more CO2 equivalents are released from TMs, CH4 emissions play a minor role. (C): When nitrate is leached from the TMs towards the SF by meltwaters or soil erosion, it removes the nitrate limitation for the denitrification at the SF and leads to high N2O losses. Nitrate leaching might also affect CO2 and CH4 production or adjacent aquatic ecosystems (here Lena River).
At the beginning of the growing season, the slump floor is more affected by meltwaters, which increase soil moisture and might slow down nitrification ( Figure 5A). Accordingly, the mobilization of nitrate by meltwaters could be an additional reason why nitrate concentrations were lower in the slump floor than in the thaw mounds [19]. Contrary, thaw mound topsoils are presumably water saturated only after rain events or overlain snow melt but are otherwise well aerated. However, the second-highest nitrate concentration was measured at site SF3 NV. Samples from site SF3 NV were taken from the steep middle part of the slope and were rich in DON and ammonium compared to thaw mound soils, suggesting that nitrate might not be produced in-situ at site SF3 NV, but could be enriched by meltwaters. Soils from site SF6 were only recently thawed in 2016 and sampled from the upper slope close to the headwall, where they are likely less affected by N leaching.

Anaerobic N2O Production
Under anaerobic conditions, N2O is released as an intermediate during denitrification, but often denitrification is limited in continuously water-saturated environments as the required substrate (nitrate) is produced via nitrification under aerobic conditions [7,12]. We did not measure N2O production via nitrification, but Marushchak et al. [6] presented that less N2O is produced under oxygen supply than under anoxic conditions  [6,26] and more CO 2 equivalents are produced from the SF. (B): When the TMs become anaerobic in the end or beginning of the next growing season after rain events or snow melt, accumulated nitrate can be reduced via denitrification to N 2 O and less CO 2 is produced due to the slowdown of SOM decomposition. When the SF becomes predominately anaerobic, it produces in addition to CO 2 , also CH 4 , but no N 2 O since denitrification is nitrate limited at the SF. Due to the GWP of N 2 O, more CO 2 equivalents are released from TMs, CH 4 emissions play a minor role. (C): When nitrate is leached from the TMs towards the SF by meltwaters or soil erosion, it removes the nitrate limitation for the denitrification at the SF and leads to high N 2 O losses. Nitrate leaching might also affect CO 2 and CH 4 production or adjacent aquatic ecosystems (here Lena River).
At the beginning of the growing season, the slump floor is more affected by meltwaters, which increase soil moisture and might slow down nitrification ( Figure 5A). Accordingly, the mobilization of nitrate by meltwaters could be an additional reason why nitrate concentrations were lower in the slump floor than in the thaw mounds [19]. Contrary, thaw mound topsoils are presumably water saturated only after rain events or overlain snow melt but are otherwise well aerated. However, the second-highest nitrate concentration was measured at site SF3 NV. Samples from site SF3 NV were taken from the steep middle part of the slope and were rich in DON and ammonium compared to thaw mound soils, suggesting that nitrate might not be produced in-situ at site SF3 NV, but could be enriched by meltwaters. Soils from site SF6 were only recently thawed in 2016 and sampled from the upper slope close to the headwall, where they are likely less affected by N leaching.

Anaerobic N 2 O Production
Under anaerobic conditions, N 2 O is released as an intermediate during denitrification, but often denitrification is limited in continuously water-saturated environments as the required substrate (nitrate) is produced via nitrification under aerobic conditions [7,12]. We did not measure N 2 O production via nitrification, but Marushchak et al. [6] presented that less N 2 O is produced under oxygen supply than under anoxic conditions from the same thaw slump emphasizing that denitrification is the major pathway that releases N 2 O in the study area. Denitrification was also estimated to be the main process releasing N 2 O in permafrost-affected peat soils when nitrate is available [2,66]. As already known, N 2 O release via denitrification strongly depends on nitrate availability and soil moisture [66,67]. Consequently, the highest N 2 O fluxes were observed in nitrate-enriched thaw mounds while there was no N 2 O production in the nitrate-poor slump floor ( Figure 5B). Confirming our hypothesis II, nitrate addition stimulated N 2 O release via denitrification in soils that did not produce N 2 O. This indicates that N 2 O production is most likely limited by nitrate availability. Contrary to our hypotheses I and III and previous studies [2,4,5,49,66,68], we found higher N 2 O production rates in revegetated thaw mounds [6] and not in non-vegetated soils of the slump floor as revegetated thaw mounds had higher nitrate concentrations and hence provided more substrate for denitrification. We found only a correlation between N 2 O and the water content, but not with N 2 O and the WFPS. Nevertheless, soils with the highest N 2 O production, site TM1 V and SF3 NV, had in addition to the highest nitrate concentration an optimal WFPS (66-67%) for N 2 O production via denitrification [7] (Table 1, Figure S3). Soils from site TM1 NV did not produce N 2 O, although having a similar nitrate concentration to site TM2 NV, but since samples from site TM1 NV had the highest water content (29 ± 0%) and WFPS (74 ± 1%) (Table 1, Figure S3), N 2 O was most likely directly reduced to N 2 and therefore not detectable. Differences in N 2 O production and reduction could also be attributed to differences in microbial community composition which were not tested here. N 2 O production rates were in a similar range to Marushack et al. [6] from the same study area. N 2 O production from the incubation experiment overestimated the N 2 O fluxes measured in the field and was 3 to 73 times higher. Variations between N 2 O fluxes from the incubations and in-situ measurements can be best explained by a higher oxygen supply in the field limiting denitrification. In-situ fluxes are in a comparable range to those reviewed by Voigt et al. [7] for permafrost-affected soils. Based on our data, we hypothesize that due to low soil moisture in thaw mounds, less N is lost via denitrification within the growing season enabling nitrate accumulation ( Figure 5A). When soil conditions become anaerobic (e.g., after rain events and overlain snow melt at the beginning of the growing season), there is a high potential that thaw mounds produce considerable amounts of N 2 O via denitrification ( Figure 5B). Under anaerobic conditions, N 2 O was the major GHG contributing to the total release of CO 2 eq when nitrate was available. Consequently, in contradiction to our expectations (IV), thaw mounds released more CO 2 eq than slump floor soils ( Figure 5B).

Effects of Nitrate on Greenhouse Gas Production
As expected, (II) N 2 O production could be stimulated by the addition of nitrate in soils that did not produce N 2 O at the beginning of the incubation. This indicates that denitrification, mainly in slump floor soils, was limited by substrate availability. Consequently, the slump floor could become a considerable N 2 O source if nitrate is leached by erosion or meltwaters from nitrate-rich thaw mounds towards the slump floor ( Figure 5C). However, it remains unclear whether nitrate is leached from thaw mounds, for how long, and in which concentrations. Studies of abrupt permafrost thaw on the Tibetan Plateau and in Alaska have shown a decline in N availability with time after erosion [5, 16,62,69,70]. Nitrous oxide production after nitrate addition increased only weakly in thaw mound soils as most probably the reduction of N 2 O to N 2 exceeded already N 2 O production. Maximum N 2 O production rates from slump floor soils were observed relatively late after 7 to 12 days after nitrate addition, supporting the assumption that microbes were not adapted to these conditions in the field, resulting in a lag phase of N 2 O production in the incubation.
So far, less attention has been paid to the effects of N on the C cycle at N hotspots in permafrost regions. Most studies in permafrost regions discuss GHG production separately for C and N. As the paradigm of an N-limited Arctic is changing [7,10], and more studies are proving elevated nitrate export and N 2 O production from disturbed permafrost [2][3][4][5][6][15][16][17][18][19]50], the effects of N on the C cycle need to be urgently considered in the future, in particular in hotspots of high N availability [58]. In our study, nitrate and N 2 O, in addition to total carbon (TC), had major control over anaerobic CH 4 production. Apart from the fact that nitrate is a preferred electron acceptor for microorganisms, it was proven in other soils and ecosystems [71][72][73], that denitrification intermediates (even at low concentrations) can have negative effects on methanogenesis and methanogen abundance. As a consequence, short-term incubations measuring CH 4 , might give biased estimations when N 2 O is not considered. Yang et al. [5,30] measured in-situ CH 4 and N 2 O fluxes, as well as nitrate concentration from a thermoerosion gully (published in two different studies) on the Tibetan Plateau. They show that CH 4 production, as well as methanogen abundance, increases with time after erosion (20 years) while at the same time nitrate concentration and hence N 2 O production have declined. As demonstrated in our study, Yang et al. [74] found a negative correlation between N 2 O (R = −0.40, p < 0.05) and nitrate (R = −0.47, p < 0.01) with CH 4 , as well as between nitrate and methanogen gene abundance (R = −0.48, p < 0.01) (personal communication). In addition, Marushchak et al. [75] studied soils across a thawing Yedoma exposure, and found a lack of CH 4 production in anaerobic incubations when substantial N 2 O production or nitrate content was observed (personal communication). They also reported that substantial N 2 O fluxes were mainly associated with zero CH 4 emissions in the field [75]. Furthermore, Abbott et al. [18] described a 90% reduction of dissolved CH 4 due to high inorganic N concentration in thermokarst outflows in Alaska. Those observations might indicate that N leaching from N-rich soils could negatively affect methanogenesis ( Figure 5C). However, since nitrate is rapidly leached or immobilized, it needs to be examined whether the impact of N on CH 4 has an impact on a regional or global scale. High nitrogen availability might also slow down CO 2 production by satisfying microbial N demand [70,76] (Figure 5C), but the effects of N on SOM decomposition are controversial [33,34,70,76]. However, decomposition depends on more factors such as SOM age [77], quality [78], or microbial accessibility [79]. Disturbance and methanogen abundance could have affected our observations. Knoblauch et al. [26,80] and Holm et al. [81] reported a generally lower abundance or even a lack of methanogens in Pleistocene-aged Yedoma sediments.

Role of Revegetation on Greenhouse Gas Production
The impact of plants could not be generalized due to a larger heterogeneity of selected slump floor sites. We hypothesize three major aspects that could have enhanced N turnover in revegetated soils: The control of soil moisture on N mineralization and denitrification are well described [82]. Plant water uptake, transpiration, soil loosening via rooting, as well as oxygen supply via roots into the soil, potentially affect soil aeration and in this way favor N mineralization. However, in our study, revegetated soils tended to be wetter but there was no statistically significant difference in WC or WFPS between revegetated and non-vegetated soils. Plants also have a direct impact on N mineralization by stimulating SOM decomposition through root exudation (priming effect) [83][84][85]. We found that WEOC was generally 36-38% higher in revegetated than in non-vegetated soils, however, it did not correlate with dissolved N, but with N 2 O production [42]. Dissolved carbon is an important electron donor for heterotrophic denitrification and potentially could have favored N 2 O production in revegetated soils. With increasing stabilization and plant succession, plant N demand might exceed microbial supply. At this point, the thaw slump could return to an N-limited system, lowering the potential for N losses as shown for other thermokarstaffected soils [5, 16,62,69,70,78]. So far it is not clear how long eroded Yedoma coasts remain as a source for N and C losses after stabilization.

Conclusions
At the studied retrogressive thaw slump, due to thermoerosion Yedoma deposits were exposed and mixed with the Holocene topsoil cover forming the two studied thermokarst features, thaw mounds, and slump floor. The different origins and geomorphology of thaw mound and slump floor soils led to differences in soil properties, N availability, N transformation, as well as GHG production. Despite our expectation, nitrification was not limited in exposed Yedoma deposits due to microbial functional limitation or plant N uptake and high amounts of nitrate could accumulate in revegetated thaw mounds during the growing season. In particular, thaw mounds have a high potential for N losses in form of nitrate and N 2 O, when soils become water saturated at the beginning or end of the growing season. At the slump floor, nitrification was likely inhibited by wet and acidic field conditions causing substrate limitation for the identification. However, when substrate limitation is removed, e.g., by nitrate leaching of nitrate-rich thaw mound soils, the slump floor has a high potential to become a N 2 O source. We showed here that Arctic soils can be potential sources of significant N losses. Nitrous oxide and CO 2 were the main components of produced CO 2 equivalents, while CH 4 played a minor role. The role of revegetation remains unclear and requires more research, but we assume that an increasing plant demand for N will determine the fate of N and GHG emissions with further stabilization and plant succession. This study highlights the relevance of considering landscape processes, such as thermoerosion, to estimate N and C losses at a landscape scale. More studies are necessary in order to characterize the seasonality and extent of N losses and to further evaluate the importance of C-N interactions in hotspots of high N availability.
Supplementary Materials: The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/nitrogen3040040/s1, Figure S1: Studied field sites; Figure S2: Soil profiles with indication of analyzed depths of studied field sites; Figure S3: Water-filled pore space; Figure S4: Cumulative CH 4 and N 2 O headspace gas concentration; Table S1: Coordinates of the sampled soil profiles in July 2019; Table S2: Subsoil samples from site TM2 NV; Table S3: Dissolved nitrogen components; Table S4: Anaerobic N 2 O production rates without and with nitrate addition; Table S5: CO 2 equivalents based on the anaerobic incubation; Table S6: Dates of in-situ N 2 O measurements in 2019.