Denitrification is the major nitrous acid production pathway in boreal agricultural soils

Nitrous acid (HONO) photolysis produces hydroxyl radicals—a key atmospheric oxidant. Soils are strong HONO emitters, yet HONO production pathways in soils and their relative contributions are poorly constrained. Here, we conduct 15N tracer experiments and isotope pool dilution assays on two types of agricultural soils in Finland to determine HONO emission fluxes and pathways. We show that microbial processes are more important than abiotic processes for HONO emissions. Microbial nitrate reduction (denitrification) considerably exceeded ammonium oxidation as a source of nitrite—a central nitrogen pool connected with HONO emissions. Denitrification contributed 97% and 62% of total HONO fluxes in low and high organic matter soil, respectively. Microbial ammonium oxidation only produced HONO in high organic matter soil (10%). Our findings indicate that microbial nitrate reduction is an important HONO production pathway in aerobic soils, suggesting that terrestrial ecosystems favouring it could be HONO emission hotspots, thereby influencing atmospheric chemistry. Microbial processes, particularly denitrification, are more important in driving nitrous acid production and emissions in aerobic soils than abiotic processes, according to 15N tracer and isotope pool dilution experiments in boreal agricultural soils.

U p to 60% of primary production of atmospheric hydroxyl (OH) radicals can be attributed to photolysis of nitrous acid (HONO) 1 . Hydroxyl radicals are vital for the removal of pollutants and reactive gases, such as carbon monoxide and methane 2,3 from the atmosphere, and they are also an essential precursor of cloud formation by transforming volatile organic compounds into secondary organic aerosols 4,5 . The current global HONO budget is not balanced when all known sources and sinks are included, which suggests that there is a missing source of HONO [6][7][8] . Although soils have only recently been identified as important HONO emitters [9][10][11][12][13][14][15][16][17] , soils have been proposed as a key component of this missing HONO source. Soil nitrite (NO 2 − ) may be the key HONO precursor 9 but this has not yet been directly demonstrated using 15 N tracer approaches and accounting for primary microbial pathways producing it. Nitrite is mainly produced during microbial ammonia (NH 3 ) oxidation (autotrophic and heterotrophic nitrification) and nitrate (NO 3 − ) reduction (denitrification and dissimilatory nitrate reduction to ammonium (DNRA)) pathways [18][19][20] . In addition, direct release of HONO by NH 3 oxidizers (autotrophic nitrifiers) has also been demonstrated 10,21 . Although the mechanisms that underpin HONO production are not yet fully understood, several studies have demonstrated the importance of microbial pathways in soil HONO production [9][10][11][12][13][14][15][16][17] . However, abiotic pathways (e.g., hydro-xylamine+water = HONO) have been proposed but have not been investigated in soils to date 21 .
The first evidence on a potential link between soil HONO emissions and underlying microbial processes was provided conducting 15 N tracer experiments, showing that nitrifiers outpaced denitrifiers in soil HONO emissions 22 . Another study 23 , proving the concept, also provided evidence on the participation of microbial processes in soil HONO emissions, especially the involvement of NH 3 oxidizers (nitrifiers) by using 15 N-labeled urea. To date, only one systematic study has addressed the pathways of soil HONO formation using a tracer approach (labeled ammonium ( 15 NH 4 + )), highlighting the sole importance of NH 3 oxidizers in HONO production and excluding the role of NO 3 − reduction in aerobic soils 11 . However, microbial NO 3 − reduction is a widespread phenomenon in soils of many ecosystems 24 , with NO 2 − produced as an intermediate of the process. Recently, the participation of denitrification processes in water-saturated soils (~100% water-holding capacity (WHC)) was demonstrated for soil HONO formation, suggesting that NO 3 − reduction to NO 2 − under anoxic conditions could be crucial for soil HONO emissions 25 . Nevertheless, besides watersaturated conditions, soil aggregates can also generate anoxic microsites that can favor microbial NO 3 − reduction in waterunsaturated soils, e.g., in agricultural soils 26,27 . In addition, the relative contribution of different microbial pathways to HONO production may vary in soils, as-beyond soil moisture and oxygen saturation-microbial activity is also influenced by soil pH, organic matter (OM) content, soil carbon to nitrogen (C:N) ratio, and temperature 19,28,29 . Thus, HONO production could occur at the microscale level and by contrasting pathways that need to be further investigated. Moreover, the proposed abiotic pathway 21 needs to be addressed in soils. Thus, in order to get a better understanding of the processes that contribute to HONO emissions from soils and their impact on the global HONO budget, the expected link between the soil inorganic N cycle and HONO formation requires further investigation 9,30,31 . Agricultural soils, which cover 50% of the global habitable land area 32,33 , possess a strong potential to release HONO because of enhanced N fertilizer use, which is expected to increase by~2-4-fold by 2050 34 . Therefore, it is crucial to understand HONO production pathways in agricultural ecosystems.
This study focused on the understanding of HONO production and emissions from agricultural soils of the boreal region, which are typically acidic (pH < 7) 9 . To date, only three studies 12,13,15 have investigated HONO emissions from boreal agricultural soils. Yet, the processes and pathways behind those HONO emissions have not been demonstrated. In this study, we performed a series of 15 N tracer experiments to understand HONO production pathways in two distinct boreal agricultural soils in Finland; one with a low organic matter (LOM) content and the other with a high organic matter (HOM) content (Supplementary Table 1). First, we aimed to distinguish between biotic and abiotic HONO production pathways by adding 15 N-labeled NO 2 − (LNi) to both, live, and sterile soils. Second, to determine the contribution of the two main biotic HONO production pathways (microbial NH 3 oxidation and NO 3 − reduction), we performed separate experiments using 15 N-labeled ammonium (NH 4 + , LA) and NO 3 − (LNa) in live (non-sterilized) soils. The HONO flux rates were then compared with data compiled in a global HONO flux rate synthesis, which was generated for the first time here. In addition, gross NO 2 − production rates across soil pH levels, and the contribution of nitrification and denitrification to the soil NO 2 − pool were also compared with published data.

Results
Time kinetics of added 15 NO 2 − and HO 15 NO production in live and sterile soils. We sterilized the soil samples by autoclaving to ensure that the biotic pathways that produce HONO via NO 2 − or NH 3 -oxidizing microbes were completely blocked. Autoclavation is the most effective sterilization method in soils, though autoclavation can alter the structure of OM content and cause disaggregation, for example, increasing the availability of dissolved organic matter 35 . We examined the 15 N atom percent excess (APE) of emitted HONO (HO 15 NO APE) and of soil NO 2 − ( 15 NO 2 − APE) in live and sterile soil samples from time zero to 240 minutes, after the addition of labeled 15 NO 2 − (LNi) and unlabeled NO 2 − (control) (Fig. 1).
In the LNi treatment under live conditions, labeled 15 NO 2 − gradually declined (Fig. 1a, b), indicating the presence of active microbial pathways that produce and consume NO 2 − . Further, 15 NO 2 − consumption was not equally accompanied by 15 NO 3 − production (by nitrifiers) under live conditions (Supplementary Figure 1). Microbial NO 2 − production pathways led to a 50% decrease in 15 NO 2 − APE by the end of the experiments in both soils (Fig. 1a, (Fig. 1c, d). Despite stable 15 NO 2 − APE, HO 15 NO APE decreased towards the end of the incubation period under sterile conditions (Fig. 1c, d), although the 15 N depletion of HO 15 NO occurred later in time and was less pronounced (at least for LOM) than under live conditions.

Relationship between biotic pathways and HONO production.
To further disentangle the biotic HONO production pathways (in addition to 15 NO 2 − (LNi; see experiment above)), we added 15 NH 4 + (LA) and 15 NO 3 − (LNa) to the live soil samples and measured the subsequent change in 15 NO 2 − and emissions of HO 15 NO. We also quantified gross production and gross consumption rates of NH 4 + , NO 2 − , and NO 3 − in the live soil samples, based on 15 N pool dilution approaches, to better explain the relationship between the microbial pathways that use 15 NH 4 + and 15 NO 3 − to produce HONO. Production (i.e., N mineralization) and consumption rates of NH 4 + were relatively low and did not differ between soils. Production (i.e., nitrification) and consumption rates of NO 3 − were 6-24 times higher than gross NH 4 + turnover rates and were significantly higher in the HOM soil than the LOM soil (Fig. 2a). Gross NO 2 − production and consumption rates ranged between the values observed for nitrification and mineralization, and were not significantly different between the two studied soils. Production of 15 NO 2 − increased gradually in the LNa treatment in both soils, which exhibited high NO 3 − consumption rates ( Fig. 2a-c, Supplementary Table 4), indicating that the reduction of NO 3 − to NO 2 − is a crucial biotic pathway in both water-unsaturated soils. Despite high nitrification rates (22 µg N g dw −1 d −1 ) in the HOM soil, the magnitude of 15 NO 2 − and HO 15 NO production during nitrification was smaller compared with NO 3 − reduction (Fig. 2b, treatments LA vs. LNa), which suggests that the capability to nitrify by the NH 3 oxidizers, which reside in the HOM soil, was lower compared with the NO 3 − reducers that shared the same environment.
The contribution of the microbial pathways that use NH 4 + and NO 3 − to total production and emissions of HONO was evident. In general, reduction of 15 Figure 2).

Quantification of HONO emissions by biotic and abiotic processes.
In the presence of microbial activity in the live soil samples, a two-source isotope mixing model 36 showed that NO 2 − contributed 100% to HONO production in the LOM soil and 81% in the HOM soil. In the latter, (in addition to NO 2 − ) an additional biotic HONO source contributed 19% to HONO production towards the end of the incubation period (Fig. 3a). Despite the presence of a consistent HONO precursor (i.e., NO 2 − ), the dynamics of biotic HONO emissions in the LOM soil differed from the HOM soil (Fig. 3b). Biotic HONO emissions in the LOM soil peaked at 180 min and then decreased slightly until the end of the incubation, whereas biotic HONO emissions increased  15 NO declined, which indicates the presence of an unknown abiotic HONO source (Fig. 3c, d). In the absence of microbial activities in the sterile soils, the unknown HONO production source increased with time (Fig. 3c). Furthermore, the unknown HONO source was more pronounced in the HOM soil than in the LOM soil (Fig. 3c). In the latter, the contribution of the unknown source to total HONO flux peaked at the end of the incubation, reaching a maximum of 22%. In the HOM soil, the contribution from the unknown source increased gradually, exceeding HONO production from soil NO 2 − at the end of the incubation, and contributed up to 52%. In both soils, the fractional contributions of the unknown source and abiotic HONO emissions showed similar increasing trends (Fig. 3c, d). In comparison with the LOM soil, the HOM soil had a greater contribution of an unknown source and also greater abiotic HONO emission rates (Fig. 3d). Biotic pathways of soil HONO formation and significance. A consistent decrease in the 15 N enrichment of HONO, in parallel with that in NO 2 − in the live soil samples, and much higher rates of HONO production in the live soil samples compared with the sterile soil samples confirmed that microbial processes are the most important HONO sources in soils. Although soil HONO emissions have been previously linked to soil NO 2 − based on physicochemical reasoning (acid-base equilibria between H + and NO 2 − -producing HONO) 9 , here for the first time we unequivocally and causally link the soil NO 2 − pool with HONO emissions, based on 15 N tracing and liquid chromatography-high resolution mass spectrometry. The findings further demonstrate that microbial NO 3 − reduction dominated HONO production and by far exceeded the concurrently accepted major HONO production pathway, i.e., NH 3 oxidation by autotrophic nitrifiers, in aerobic, water-unsaturated soils, here boreal agricultural soils (Figs. 2a-c, 3b). Among the three soil NO 2 − -producing microbial pathways, only NO 3 − reduction triggered significant 15 NO 2 − and HO 15 NO production in both soils. Soil NO 3 − reduction was solely linked to the presence of the periplasmic NO 3 − reductase (napA) gene, as the cytoplasmic NO 3 − reductase (narG) gene 38 was absent in both soils (Supplementary Figure 7). The lack of 15 NH 4 + production from 15 NO 3 − in the LNa treatment (Supplementary Figure 1) suggests that DNRA (dissimilatory nitrate reduction to ammonium) was absent and that denitrification dominated microbial NO 3 − reduction and soil HONO emissions. This is further confirmed by the absence of the key DNRA gene, i.e., nitrite reductase (nrfA) 18,39 , whereas the genes of key denitrifying genes were abundant, e.g. napA, nirK, nirS, and nosZ (Supplementary Figure 7). Nitrification only contributed to a small extent to soil HONO production (Fig. 4), and was only found in the HOM soil (Fig. 2b). Nitrification could be attributed to the presence of bacterial and archaeal NH 3 oxidizers 10,21 , as suggested by higher gene copy numbers of bacterial (and archaeal) ammonia monooxygenase (amoA) in the HOM soil ( Supplementary Figure 7), and these are capable of directly producing HONO 10 before accumulating NO 2 − (Supplementary Figure 2). Decoupling of microbial NO 3 − reduction (high 15 NO 2 − output, Fig. 2a) and biotic HONO production (Fig. 4) in the HOM soil could be associated with the combined effect of microbial NO 2 − consumption and abiotic NO 2 − reactions with soil OM via nitration and nitrosation 40 . In these soils, we observed a curvilinear relationship between soil NO 2 − concentrations and HONO emission rates ( Supplementary Figure 6a), which indicates a gradual conversion of NO 2 − to NO 3 − (Supplementary Figure 1) and concurrently occurring chemical reactions between NO 2 − and soil OM 40 , the latter of which surpass HONO formation via NO 2 − after reaching an optimum.  Soil moisture is an important environmental driver of soil HONO emissions. Maximum soil HONO emissions were identified in a wide range of ecosystems under low soil moisture contents ("dry" peak at 20-30% WHC) 9,10,25 . Such HONO emission peaks occurred during progressive reduction in soil moisture in the experiments and were linked to nitrifiers as NH 3 oxidation is an aerobic pathway and requires oxygen. Therefore, it was assumed that with soil moisture reduction and progressively developing the aerobic condition, NH 3 oxidation 10,21 dominates soil HONO production and emissions. However, the 15 NH 4 + supplemented LOM soil did not emit HO 15 NO, even until the end of the experimental period, despite a slight reduction in soil moisture (from 60% start to 51% end WHC; Supplementary Figure 8). In contrast, 15 NH 4 + supplemented HOM started to emit HO 15 NO, presumably by NH 3 oxidizers 10,21 , from the middle of the experiment (Fig. 2b, Supplementary Table 2-3), despite a smaller reduction in soil moisture (from 60% start to 54% end WHC) compared with the LOM soil (Supplementary Figure 8). In our study, we did not measure HONO emissions for longer periods and therefore did not reach lower soil moisture regimes, for example, 20-30% WHC, at which the contribution of nitrifier pathways could increase 10,21 . For example, Wu et al. 23 showed that soil HONO emissions (at 20-30% WHC) were driven by NH 3 oxidation in a pH neutral soil from the temperate region. On the other hand, Kubota and Asami 22 estimated that~69% of the soil emitted HONO originated from NH 3 oxidation at 60% WHC from sub-tropical acidic soils (pH = 4.8-5.8) incubated at 30°C for 30 days. Our estimation shows that NH 3 oxidation contributed only 0-10% to the total biotic HONO emissions (at 60-54% WHC range) from slightly acidic (pH = 6.1) boreal agricultural soils incubated at 21°C for 48 hours. The likely reason for higher contributions of nitrifiers to soil HONO emissions in the two earlier studies 22,23 could be differences in soil moisture and temperature between studies, which are known to control the activities of NH 3 oxidizers 29 and thereby of soil HONO emissions 10,23,25 . Nevertheless, our results obtained from the LA treatments in two boreal agricultural soils suggest that the capabilities of NH 3 oxidizers to produce HONO 10,21 vary according to soil type and/or climatic zone, which requires further studies.

Discussion
Besides soil moisture, soil pH also has a significant role in HONO emissions. Although nitrification (especially by bacterial ammonia oxidizers) decreases towards low soil pH levels 41 , at the same time, regardless of its source, rapid protonation of NO 2 − in acidic soils 9 can promote soil HONO formation and release. The reported HONO release from soils by acidophilic and high NH 3 affinity archaea 11,41,42 was lower than HONO emissions via microbial NO 3 − reduction (Supplementary Figure 9), further supporting our finding that soil NO 2 − via microbial NO 3 − reduction is crucial for HONO production, especially in soils with pH < 7. In addition, denitrification demands the presence of protons (2NO 3 − + 10e − + 12H + →N 2 + 6H 2 O) and therefore shows a pH optimum under acidic conditions unlike nitrification, which does the opposite, i.e., it releases H + (NH 3 + 2O 2 → NO 3 − + H + + H 2 O) into the soil and therefore shows a pH optimum at neutral to alkaline conditions. Therefore, acidic soils are likely to favor NO 3 − reduction via denitrification under suitable conditions, such as ample NO 3 − availability and anoxic microsites. Moreover, our data synthesis clearly shows that gross NO 2 − production rates increase in acidic soils (Fig. 5b, Supplementary Data 2). In published studies that have examined the contribution of nitrifiers and denitrifiers to soil NO 2 − production, an average 48% and 57% of the NO 2 − source was contributed by denitrifiers (n = 82, Fig. 5c, Supplementary Data 2), when organic N oxidation to NO 2 − by heterotrophic nitrifiers was included or excluded. The relationship between soil NO 3 − reduction by denitrifiers and soil HONO emissions as found in our study is in agreement with Wu et al. 25 who recently demonstrated high HONO emissions from soils under water-saturated, anaerobic conditions ("wet" peak caused by microbial NO 3 − reduction at 100% WHC). However, here we show the importance of microbial NO 3 − reduction for HONO production under waterunsaturated, aerobic soil conditions (between 51% and 60% WHC). This suggests that not only water-saturated, anoxic soils, but also anoxic microsites in aerobic soils that are commonly generated in the interior of soil aggregates 26 are key to the promotion of microbial NO 3 − reduction and soil HONO formation. Therefore, soil HONO production via microbial NO 3 − reduction in aggregates is likely a more widespread phenomenon because of the anoxic soil microsites and the ubiquity and diversity of NO 3 − reducers (archaea, fungi, bacteria) in a wide range of ecosystems 24,43 , and across a wide range of soil pH and moisture contents.
Abiotic pathway of soil HONO formation. Decreasing 15 N enrichment in HONO in sterile soils, despite the constant 15 N enrichment in NO 2 − , suggests the presence of an abiotic HONO production pathway. With consistent 15 N enrichment in NO 2 − and zero microbial production and consumption of NO 2 − (Fig. 1), we can only assign the observed HONO flux to an unknown abiotic source, not related to the soil NO 2 − pool. As our HONO measurements were conducted in Teflon coated (inner wall) chambers in the dark, heterogeneous reactions that require nitrogen dioxide (NO 2 ) and light conditions to form HONO 44,45 are expected to be insignificant. Also, the absence of microbial NO 2 − production in sterile conditions precludes the possibility of HONO production via self-decomposing HNO 2 , as its products (NO 2 and nitric oxide (NO) gases) must pass through the soil NO 2 − pool 46 . Recently, a reaction between soil hydroxylamine and H 2 O has been suggested to produce HONO abiotically on glass beads 21 , although there is little evidence for this phenomenon in soils. Nevertheless, abiotic production of HONO was twofold higher (Fig. 4) in the HOM soil (1.3 ng N m −2 s −1 ) compared with the LOM soil (0.6 ng N m −2 s −1 ), which indicates that soil OM may play a role in the abiotic formation of HONO. This unidentified abiotic HONO pathway, which does not require passage through the soil NO 2 − /HNO 2 pool, clearly needs further investigation.

Conclusion
We confirmed that microbial processes are essential for HONO emissions from agricultural soils, as they contributed 8-42 times more to soil HONO emissions than abiotic processes (Fig. 4). Using 15 NO 2 − , 15 NO 3 − , and 15 NH 4 + tracers, we showed, for the Pathways that are yet to be confirmed (e.g., NH 2 OH + H 2 O and DNRA) or having potential but were not studied yet, i.e., chemo-denitrification concerning HONO production in soils, are indicated by dotted colored arrows and with question marks (?). Chemo-denitrification is a process associated with the abiotic reaction of nitrite or nitrate in the presence of amines, reduced metals (e.g., Fe 2+ ), and high soil organic carbon [47][48][49] to gaseous nitrogen forms. Arrows denoting denitrification are thickest, indicating denitrification to be the most significant pathway of HONO production via the soil nitrite (NO 2 − ) pool in soils with pH <7. This conceptual model is based on findings here, and literatures published earlier.
first time, that microbial NO 3 − reduction (denitrifiers) is driving soil HONO emissions in aerobic soils, by fueling the soil NO 2 − . Other microbial pathways contributed little or not to HONO production, i.e., nitrification (NH 3 and organic N oxidation) and DNRA (Figs. 4, 6). We conclude that microbial processes are essential for soil HONO emissions, and that the microbial NO 3 − reduction pathway could be a significant contributor in aerobic soils of many ecosystems. Moreover, abiotic HONO production pathways that have remained elusive could exist in soils, thereby contributing to notable emissions of HONO to the atmosphere, although microbial pathways are dominant. Therefore, in order to better understand the impact of soil emitted HONO on the seasonally changing atmospheric HONO budget, atmospheric OH radical production and associated atmospheric chemistry 30,31,50 (at present and in the future under a changing climate), we suggest that future studies should assess the relative contribution of microbial pathways, in tandem with concurrent abiotic processes (Fig. 6), to soil HONO production, by applying appropriate 15 N tracer approaches.

Methods
Soil sampling and sample preparation. Two soils (HOM and LOM) with distinct edaphic properties (Supplementary Table 1) were sampled from two separate agricultural fields maintained by the Natural Resources Institute Finland (LUKE) in eastern Finland, Maaninka (63°09ʹ N, 27°20ʹ E). According to the World Reference Base of soils 51 , the LOM soil is classified as a Dystric Regosol, which covers 2% of the global land surface and the HOM soil is classified as a Histosol, typical of the northern latitudes, and covers 2.5% of the global land area, though 30% in Finland. The sampled bulk soils (0-20 cm soil depth) were immediately transported to the laboratory where the roots and remaining plant parts were removed manually. Then the soils were homogenized, sieved (4 mm), and stored at 4°C until the start of the experiments. 15 N tracer experiments were conducted with live (non-sterile) and sterile (autoclaved) soil conditions. More details are described in the Supplementary Methods. 15 N tracer experiments. In all, 100 g of soil (live or sterilized) were transferred to a sterile petri dish (Ø 0.137 m, h = 0.017 m) with a sterilized spatula, and 15 N tracer solution was added evenly to the soil surface with a pipette, immediately prior to the HONO measurements. We used three different tracers, 15  HONO flux measurements and flux calculation. The 15 N-amended samples were immediately transferred inside an opaque dynamic chamber, and HONO concentrations were measured with a Long Path Absorption Photometer instrument (LOPAP) 52 . The HONO flux was measured as described in Bhattarai et al. 15 . More details are provided in the Supplementary Methods.
Isotope pool dilution assays and determination of isotope ratios. The Isotope pool dilution (IPD) assays 53 were performed for all three N pools in live and sterile soil samples. For IPD assays, 4 g fresh soil (live or sterile) were weighed into 50 ml sterile polypropylene tubes (CELLSTAR®, Greiner BIO-ONE) and supplemented with 300 µl 15 N tracer mix. Here, concentrations and atom% of tracer solutions and the soil moisture content were the same as used in the HONO flux experiments. After tracer addition, the tubes were vortexed for 2 minutes to allow the tracers to mix sufficiently in the soil matrix. The incubations were stopped by extraction with 30 ml cold (4°C) 1 M KCl at zero, 30, 60, 90, 120, 180, and 240 minutes. The isotope ratios ( 15 N/ 14 N) in NH 4 + and NO 3 − were determined using the microdiffusion method 53 in the soil extracts collected at zero, 90, 180, and 240 during the IPD experiment. More details are provided in the Supplementary Methods.
Collection and purification of HONO and NO 2 − azo dyes. The collection and purification of HONO and NO 2 − azo dyes were performed following Wu et al. 23 , who established a method to analyze 15 N APE in gaseous HONO. In the LOPAP instrument, HONO is reacted to an azo dye (C 18 H 19 O 2 N 5 S). Gaseous HONO is scrubbed in two sequential reactions, first with acidic sulfanilamide solution (R 1 ) in a two-channel striping coil, and second with N-(1-Naphthyl)ethylenediamine solution (R 2 ), which was collected separately from channel 1 and channel 2 after photometric detection in the LOPAP 52 . Each 5 ml of HONO azo dye solution was collected in 15 ml sterile polypropylene tubes (CELLSTAR®, Greiner BIO-ONE) at six sampling points that ended at 30, 60, 90, 120, 180, and 240 minutes after connecting the soil samples to the LOPAP. In parallel, NO 2 − azo dyes were produced from the soil extracts collected during the IPD experiment and standards using an identical approach for preparing azo dye 23 . Immediately upon extract collection at each time point, 5 ml azo dye was generated by reacting 2.5 ml R 1 + R 2 mix (mixed at 1:1 (v:v) ratio) with 2.5 ml soil extract or standards. The R 1 and R 2 solutions used in the LOPAP instrument and for generating NO 2 − azo dyes in the soil extracts were of the same concentrations and the same chemical brand. The generation of NO 2 − azo dye from the soil extracts in the IPD assays was done for the first time in this study to understand the dynamics in the isotope ratios ( 15 N/ 14 N) of NO 2 − produced by different microbial N cycling pathways and to relate them to the isotope dynamics of HONO emitted from soils. HONO and NO 2 − azo dyes were purified by a reversed-phase extraction, according to Wu et al. 23 , except for the volume of acetonitrile and milli-Q H 2 O used for the preconditioning and final washing step of SPE (Solid Phase Extraction) columns. We used 1 ml acetonitrile and 2 ml milli-Q H 2 O more than used by Wu et al. 23 , to ensure the best cleaning of SPE columns (during preconditioning) and to maximize the removal of inorganic ions (in the final washing step). In brief, the pH of the azo dye solutions was adjusted to~5 with 2 M sodium hydroxide and loaded onto 6 ml pre-conditioned SPE columns (Discovery® DSC-18 6 ML/500MG SPE, Sigma-Aldrich). The SPE columns were pre-conditioned by washing with 3 ml acetonitrile (HPLC grade, ≥99.8%, Thermo fisher Scientific) followed by 4 ml milli-Q H 2 O (18.2 MΩ). The SPE columns loaded with azo dye were then washed with 4 ml milli-Q H 2 O and stored at −20°C for further analyses.
Elution and analyses of HONO and NO 2 − azo dyes. The frozen SPE columns were transported to the Terrestrial Ecosystem Research laboratory, University of Vienna, Austria, within one month after collection, where elution and analyses were performed. To avoid temperature effects (if any) during transportation, the frozen samples were transported in a sealed cooling box that contained ice bags. Immediately upon arrival, they were transferred to a −20°C freezer until elution. The elution and analyses of HONO and NO 2 − azo dyes were different from those in Wu et al. 23 . The fundamental difference was the instrument used to analyze the isotope ratios ( 15 N/ 14 N) in HONO and NO 2 − azo dyes. In our study, we analyzed HONO and NO 2 − azo dyes with an UPLC system (Ultimate 3000, Thermo Fisher Scientific, Bremen, Germany) coupled to an Orbitrap Exactive HCD MS (Thermo Fisher Scientific) with a mass resolution of 50,000. The choice of the device was made to optimize the detection of the 15 N signal in azo dyes, as suggested by Wu et al. 23 . The high resolution, high mass accuracy, and sensitivity of the Orbitrap-MS allowed us to separate the 15 N and 13 C isotopologue peaks that have a very small mass difference (Δ mass of 0.00632 Da) 54,55 . This and similar instruments are widely used to study metabolites, including those in soil extracts 56 . The choice of instrument (with electrospray ionization) eventually also led us to use eluents and reagents different from Wu et al. 23 for elution and chromatographic separation. Prior to elution, the SPE columns were transferred to room temperature for 20 minutes. Azo dyes were eluted with 5 ml eluent [80% methanol (HPLC grade, Sigma-Aldrich, ≥99.9%)+1% formic acid (98-100%, Merck Pro analysis)] using SPE vacuum manifolds. The eluates were diluted at 1:1 (v:v) ratio with milli-Q H 2 O and analyzed on the same day of elution. More details are provided in the Supplementary Methods.
Calculation of 15 N enrichment in HONO and NO 2 − azo dyes. The 14 N and 15 N peak areas in the samples and calibration standards were integrated manually with Xcalibur software and exported into Microsoft Excel sheets for further calculations. After blank correction of the 14 N and 15 N peaks of the samples and standards, we then calculated the concentrations (µM) of the samples using the peak area sum (area of 14 N plus area of 15 N peaks), based on the slope and intercept build from the unlabeled concentration calibration standards using equation 1. Next, we calculated the percentage (%) of isotopically heavy (i.e., 15 N) molecules of total molecules by dividing the peak area of the heavy isotopologues by the peak area sum of heavy and light isotopologues, as shown in Eq. 2. The atom% 15 N of the azo dyes of HONO and NO 2 − were quantified from a series of polynomial functions generated from % of the heavy atoms (theoretical 15 N in NO 2 − azo dye of isotope calibration standards) and atom% 15 N measured in NO 2 − azo dyes by LCMS at two concentration levels of nitrite, 1.56 µM and 25 µM, and their average. These polynomial functions were applied depending upon the concentration and % of heavy isotopologues in the azo dye samples, which adapted within and between samples over time (e.g., LA vs. LNa vs. LNi treatments). The calibration curves for concentration from natural abundance calibration standards and theoretical 15  Source contribution and determination of abiotic and biotic HONO fluxes. The contributions (%) of NH 4 + oxidation, NO 3 − reduction, and organic nitrogen oxidation (the inset in Fig. 4) were calculated using the slopes (Fig. 2b-c) generated between 15 NO 2 − and HO 15 NO from the live soil samples of the LNi, LNa, and LA (only in the HOM soil) treatments. The contribution of NH 4 + oxidation and NO 3 − reduction was obtained by dividing the respective slopes by the slope between 15 NO 2 − and HO 15 NO in the LNi treatment. The contribution of organic N oxidation was obtained by subtracting the sum of NH 4 + oxidation and NO 3 − reduction from 100%. The contribution of a potentially unknown abiotic HONO source was quantified from the LNi treatment using Eqs. 3 and 4.
In Eqs. 3 and 4, atom% represents the values of the mixture and the sources (i.e., source 1 and source 2), and the fractional contributions of source 1 and source 2 are represented by f 1 and f 2 , respectively. Here, the mixture was HONO, source 2 was NO 2 − and source 1 assumed to be at natural 15 N abundance, i.e., 0.3663 atom % 15 N. The fractional contribution of an unknown source (i.e., f 1 ) was determined using Eq. 4, with the assumption that the sum of fractional source contributions is 1, and that the contribution of NO 2 − (f 2 ) is determined by 1−f 1 . The abiotic and biotic HONO emissions were quantified using the HONO emissions from live and sterile LNi treatment. To obtain an estimate of the abiotic HONO flux, we quantified the HONO flux, which solely originated from NO 2 − and was subtracted from the total HONO emitted there, in the LNi treatment that used fractional contributions of NO 2 − to HONO emissions (ng N m −2 s −1 ) in the sterile soil samples. For the biotic HONO flux, HONO that originated from the NO 2 − added to the sterile soils was subtracted from the HONO measured from the live soil samples.
Total HONO. Total abiotic and biotic HONO represents the emission rates of soil HONO production averaged over 240 minutes of HONO flux measurements. To do that, first, total HONO in ng N m −2 was obtained for all replicates separately. Here, the HONO emission rates (in µg N m −2 s −1 ) at each sampling point were multiplied by the time in seconds between the two consecutive sampling points, and their sum was calculated to obtain the total HONO. The total HONO (ng N m −2 ) thus obtained was divided by 14,400 (240 × 60) for all replicates separately to get the HONO emissions rate averaged over 240 minutes in ng N m −2 s −1 . Finally, mean values and standard deviations were calculated for the abiotic (sterile soil samples) and biotic (live soil samples) sources. Table 1 Statistics. Statistical analysis and graphical presentations were performed in R statistical software (R version 3.6.1), unless otherwise specified. Before performing a statistical test, the distribution of the data was assessed using the Shapiro-Wilk test, Q-Q plots, and histograms, and were transformed (log or square root) when required. More details are provided in the Supplementary Methods.