Direct O 2 control on the partitioning between denitrification and dissimilatory nitrate reduction to ammonium in lake sediments

. Lacustrine sediments are important sites of fixed nitrogen (N) elimination through the reduction of nitrate to N 2 by 10 denitrifying bacteria, and are thus critical for the mitigation of anthropogenic loading of fixed N in lakes. In contrast, dissimilatory nitrate reduction to ammonium (DNRA) retains bioavailable N within the system, promoting internal eutrophication. Both processes are thought to occur under oxygen-depleted conditions, but the exact O 2 thresholds particularly of DNRA inhibition are uncertain. In O 2 -manipulation laboratory experiments with dilute sediment slurries and 15 NO 3-additions at low-to sub-micromolar O 2 levels, we investigated how, and to what extent, oxygen controls the balance between 15 DNRA and denitrification in lake sediments. In all O 2 -amended treatments, oxygen significantly inhibited both denitrification and DNRA compared to anoxic controls, but even at relatively high O 2 concentrations (≥ 70 µmol L -1 ), nitrate reduction by both denitrification and DNRA was observed, suggesting a relatively high O 2 tolerance. Nevertheless, differential O 2 control and inhibition effects were observed for denitrification versus DNRA in the sediment slurries. Below 1 µmol L -1 O 2 , denitrification was favored over DNRA, while DNRA was systematically more important than denitrification at higher O 2 20 levels. Our results thus demonstrate that O 2 is an important regulator of the partitioning between N-loss and N-recycling in sediments. In natural environments, where O 2 concentrations change in near bottom waters on an annual scale (e

While oxygen inhibition/tolerance of denitrification and anammox has been studied previously in the ocean water column 65 (Jensen et al. 2008, Kalvelage et al. 2011, Babbin et al. 2014, Dalsgaard et al. 2014, investigations into the O2 control on benthic N-reduction are rather rare, and limited to sandy and low organic matter marine sediments (Gao et al., 2010;Jäntti and Hietanen, 2012;Rao et al., 2007). Despite intensified research, the exact O2 thresholds with regards to the direct inhibition of benthic N reduction are still poorly constrained. This is particularly true for DNRA. Recent work has highlighted the significance of DNRA even in the presence of relatively high O2 concentrations (i.e., at hypoxic levels (i.e., 10-62 µmol L -1 ), 70 or concentrations even greater than 62 µmol L -1 ) in estuarine sediments (Roberts et al., 2012(Roberts et al., , 2014 and marine sediments (Jäntti and Hietanen, 2012), but a systematic investigation of how DNRA is impacted at low micromolar O2 levels in aquatic sediments (and how in turn the balance between denitrification and DNRA is affected), does to our knowledge not exist.
In this study, we provide first experimental evidence for direct O2 control on the fate of reactive N in lacustrine sediments with 75 high organic matter content. Through slurry incubation experiments with sediment from a eutrophic lake in Switzerland (Lake Lugano), 15 N-labelled substrates and manipulated O2 concentrations, we investigated the functional response of benthic Nreducing processes to changing O2 levels. We demonstrate that denitrification and DNRA are differentially sensitive towards O2, which has important implications for fixed N removal in environments that undergo short-and longer-term O2 changes, such as seasonally stratified (anoxic) lakes or other aquatic environments with expanding volumes of hypoxia/anoxia. 80 4 containing bottom waters. Ammonium concentrations in bottom water were relatively high (~ 30-140 µmol L -1 ) during anoxia and close to the detection limit during months when the water column was mixed.

Porewater sampling
Porewater oxygen microprofiles were generated using an O2 microsensor (Unisense) with a tip diameter of 100 µm in duplicate cores from both sites. The overlying water was gently stirred (without disturbing the sediment-water interface) and aerated to 100 determine the O2 penetration depth under oxygenated conditions at room temperature (~20 °C). Porewater samples for the analysis of dissolved inorganic nitrogen concentrations were obtained by sectioning of a separate set of cores from the same sites at 1 cm-interval under normal atmosphere and at room temperature (~20 °C), and centrifuging of the samples (4700 rpm, 10 min).

N-transformation incubation experiments 105
In a first step, incubations to measure potential denitrification and DNRA rates under control (i.e., anoxic) conditions were performed in an anaerobic chamber (N2 atmosphere). At each site, fresh surface sediments (upper 2 cm) from duplicate sediment cores were homogenized to prepare dilute sediment slurries. Aliquots of 1 g sediment and 70 mL anoxic artificial lake water (NO3 -, NO2 -, NH4 + -free; Smith et al., 2002) were transferred into 120 mL serum bottles. The use of dissolved-NOxfree artificial water is important to avoid any potential underestimation of N-transformation process rates due to 28 N2 production 110 from ambient NO3or NO2 --present in bottom waters. Serum bottles were sealed and crimped using blue butyl rubber stoppers and aluminum caps. The sediment slurries (generally in triplicates, Table 1) were He-flushed for 10 min to lower the atmospheric N2 and O2 backgrounds, and placed overnight on a shaker (80 rpm) at 8 °C in the dark to remove any residual O2.
It needs to be noted that this He-flushing step, although crucial in our experimental set-up, may have interfered with in situ conditions by altering microbe-particle interactions through disruption of larger aggregates in the sediments or by slightly 115 changing the pH in the sediment slurries. Labeled 15 N substrate (i.e., Na 15 NO3 -, 99% 15 N, Cambridge Isotopes Laboratories, 120 ± 2 µmol L -1 final conc.) was added in order to quantify potential rates of denitrification and DNRA. During the incubation period (ca. 10 hours), anoxic sediment slurries were kept in an incubator on an orbital shaker (80 rpm; 8 °C). Preliminary tests were performed in order to assess the minimal incubation time required to obtain clear and robust 15 N-N2 production trends, and during which it was feasible to maintain a more or less constant O2 concentration in parallel slurry experiments. For 120 subsampling of gas and liquid samples, the incubation vials were transferred to an anaerobic chamber with N2-atmosphere.
There, 2-mL gas samples were taken at four successive time points (T0, T1, T2, T3) for N2 isotope measurement, in exchange with 2 mL He (T0) or anoxic Milli-Q water (T1 to T3) in order to compensate for any pressure decrease inside the vials. Gas samples were then transferred into 3 mL exetainers (Labco), prefilled with anoxic water, and stored upside down until isotope analysis. Liquid samples (6 mL) were taken at T0 and T3 for the quantification of DNRA rates through N-NH4 + isotope analysis 125 (see below) and for the assessment of nutrient (NO3 -, NO2 -, NH4 + ) concentrations. Samples were filtered (0.2 µm) inside the anaerobic chamber prior to freezing.

O2 manipulation experiments
For the O2 manipulation experiments, serum bottles were equipped with TRACE Oxygen Sensor Spots (TROXSP5, detection limit = 6 nmol L -1 O2, Pyroscience, Germany), allowing non-invasive, contactless monitoring of dissolved O2 concentrations 130 in the dilute slurry. The sensor spots were fixed at the inner side of the glass wall with silicone glue and the sensor signal was read out from outside using a Piccolo2 fiber-optic oxygen meter (PyroScience). Different volumes of pure O2 (99,995%) were injected into the headspace of pre-conditioned and 15 NO3 --amended slurries using a glass syringe (Hamilton). For each treatment, the gas volume required to reach the targeted O2 equilibrium concentration (0.8, 1.2, 2, …, 78.6 µmol L -1 ) was calculated based on the headspace versus slurry volumes, salinity, and temperature (Garcia and Gordon, 1992). Measured O2 135 concentrations in slurries after injection of the respective O2 gas volumes were always close to the ones calculated (the first measurement was performed 30 minutes after injection to ensure gas equilibration between the gas and the liquid phase).
Oxygen concentrations in the slurry incubations were monitored with the fiber-optic oxygen meter every 30 minutes and, in case of a decline in dissolved O2 due to microbial consumption, O2 was added in order to return to the initial target O2 concentrations (Supp. shaken by hand every 30 minutes to avoid the formation of anoxic microniches, which may act to increase rates of anaerobic N-transformation processes (Kalvelage et al., 2011).

Hydrochemical analyses
Nitrite concentrations were determined colorimetrically using sulfanilamide and Griess-reagent, according to Hansen and Koroleff (1999). Total NOx (i.e., NO3 -+ NO2 -) concentrations were measured using a NOx-analyser (Antek Model 745, 145 detection limit = 0.02 µmol L -1 ), by reduction to nitric oxide (NO) in an acidic V 3+ solution, and quantification of the generated NO by chemiluminescence detection (Braman and Hendrix, 1989). Nitrate concentrations were then calculated from the difference between NOx and NO2concentrations. Ammonium was measured by suppression-ion chromatography with conductivity detection (940 Professional IC Vario, Metrohm, Switzerland).

15 N-based rate measurements 150
For the determination of denitrification rates, gas samples from the 15 N-isotope enrichment experiments were analysed for 15 N/ 14 N isotope ratios in the N2 using a Delta V Advantage isotope ratio mass spectrometer (IRMS; Thermo Fisher Scientific) with a ConFlo IV interface (Thermo Fisher Scientific). Denitrification (and negligible anammox) rates were calculated based on the quantification of 15 N in the N2 gas in excess of the natural abundance, i.e. by calculating the linear regression of 30 N2 concentrations (and to a minor extent 29 N2) over incubation time (Nielsen, 1992;Thamdrup and Dalsgaard, 2002;Supp. Fig. 155 SI.1). DNRA rates were quantified using the isotope-pairing method described by Risgaard-Petersen et al. (1995). Briefly, 2 mL liquid samples were transferred into 6 mL exetainers (Labco) and closed with plastic screw septum caps. The headspace was flushed with He for 2 min to reduce the 28 N2 background, and 25 µL mL -1 of alkaline (16 mol NaOH) hypobromite iodine solution (3.3 mol L -1 ) were added through the septum to convert NH4 + to N2 (Robertson et al., 2016). Exetainers were stored upside down and placed on a shaker (100 rpm) for 24 h at room temperature. The produced N2 was then analysed by IRMS as 160 described above. DNRA rates were determined based on the 15 NH4 + production with time, as calculated from the total 15 N-N2 in the hypobromite-treated samples (i.e., calculated from the excess 29 N2 and 30 N2 signals). The recovery of 15 N-label from identically treated standards was >95%.

Statistics
Results are presented as the mean and standard error of n replicate experiments (Table 1). Correlation analyses were performed 165 using Microsoft Excel software, and significant differences between results were verified using a Student's t-test (P < 0.05).

Porewater hydrochemistry
The O2 microsensor profiles revealed that the O2 penetration at the two sites under aerated conditions ranged between 2.4 mm (Melide) and 3.7 mm (Figino, Fig. 2). The relatively low oxygen penetration depth is consistent with a high organic carbon 170 content (~8%, data not shown). According to the observed O2 concentration gradients at the two stations, the potential O2 flux into sediments was greater at Melide suggesting a higher reactivity of the sedimentary organic matter. In contrast to the microsensor profiling, the sectioning-centrifuging technique was not sufficient to resolve the exact porewater nitrate concentration gradient, yet the observed nitrate concentration data across the sediment-water interface ( Fig. 2) clearly indicate that the sediments at both sites represent a sink for water-column nitrate, and that nitrate is consumed to completion already 175 within the top centimeter of the sediments. In contrast, ammonium concentrations just below the sediment-water interface at Figino and Melide increased steeply from 830 and 600 µmol L -1 NH4 + to 1.7 and 1.2 mmol L -1 , respectively.

N-transformations in control experiments
Potential rates of denitrification and DNRA under true anoxic conditions were quantified at both sampling sites in October 2017. Anammox rates were measured in a previous study at different times of the year, and its contribution to the total fixed-180 N removal was always less than 1%, thus negligible with respect to other processes (Cojean et al., in prep.). Indeed, in all experiments, denitrification and DNRA were the main benthic N-transformation processes with an essentially equal contribution to the total nitrate reduction (»0.1 µmol N g -1 wet sed. d -1 ; Table 1 caption). We ensured that measured DNRA rates were not underestimated due to 15 NH4 + loss through adsorption on mineral surfaces. Previous results (Cojean et al., in prep.) demonstrate that adsorption of ambient or tracer ammonium does not occur at detectable levels in the dilute sediment Indeed, oxic slurry incubation experiments (≥ 73 µmol L -1 O2) revealed that at least at high O2 concentrations net NO3production occurs (≤ 1 µmol N g -1 wet sed. d -1 ).

Impact of O2 on NO3reduction in sediments
The O2 sensitivity of denitrification and DNRA and inhibition kinetics were investigated through slurry incubation experiments 190 under O2-controlled conditions. At both sites, potential denitrification and DNRA rates consistently decreased with increasing O2 concentration (Fig. 3). While the general pattern was systematic for both processes (i.e., an exponential attenuation of both denitrification and DNRA rates with increasing O2), the response of denitrifiers versus nitrate ammonifiers towards manipulated O2 differed across sites and treatments. We compared O2-addition experiments to the anoxic controls to estimate the inhibition of nitrate reduction by O2. At the lowest O2 concentration (~ 1 µmol L -1 O2), denitrification was less inhibited 195 than DNRA at Figino (29 ± 20 % and 51 ± 7 % inhibition, respectively) while the suppression was almost equivalent at Melide (43 ± 8 % and 37 ± 9 % inhibition of denitrification and DNRA respectively, Table 1). At O2 concentrations around 2 ± 0.2 µmol L -1 , both denitrification and DNRA rates were more than 50% inhibited compared to the anoxic control (Table 1, Fig. 3). At O2 concentration ≥ 2 µmol L -1 , DNRA rates were generally higher than those of denitrification (with one exception, i.e., 16 µmol L -1 O2 at Figino; Fig. 3), indicating that denitrification was more sensitive than DNRA to elevated O2 levels. Oxygen 200 concentrations higher than 73 µmol L -1 resulted in almost complete inhibition of denitrification at both sites (96 ± 1 % and 93 ± 2 % at Figino and Melide, respectively, Table 1). Oxygen inhibition thresholds for DNRA were even higher, as DNRA rates were significantly less impaired compared to denitrification at these elevated O2 levels (79 ± 5 % and 75 ± 4 % inhibition compared to the anoxic controls at Figino and Melide, respectively; Table 1). A correlation analysis between the relative contribution of DNRA to the total NO3reduction (%) and the increase of O2 concentration displayed a positive correlation 205 coefficient of 0.57 and 0.91 for Figino and Melide, respectively (Supp. Fig. SI.2). Hence, the relative contribution of the two processes to total nitrate reduction was significantly affected by changing O2 concentrations. At anoxic and submicromolar levels of O2 (≤ 1 ± 0.2 µmol L -1 O2), denitrification rates were higher than those of DNRA, while at higher O2 concentration the ratio between denitrification and DNRA was shifted in favour of the nitrate ammonifiers (Fig. 4).

210
Consistent with the observed decline in denitrification and DNRA rates based on the 15 N-N2 and 15 NH4 + measurements in the 15 N-label incubations, nitrate consumption in slurries decreased with increasing O2 concentration at both stations (Table 1).
Similarly, maximum ammonium accumulation was observed in the anoxic controls, whereas at higher O2 levels ammonium underwent net consumption, indicating the concomitant decrease of DNRA and the increasing importance of nitrification under more oxic conditions, particularly at Melide. In incubations where nitrate concentrations decreased, the ratio of (NO3 -)consumed 215 discuss this puzzling discrepancy further, but we speculate that excess NO3consumption may be linked to bacterial and algal 220 uptake (Bowles et al., 2012). Biotic immobilization of NO3in marine sediments has been attributed previously to the intracellular storage of nitrate by filamentous bacteria (Prokopenko et al., 2013;Zopfi et al., 2001) and/or diatoms (Kamp et al., 2011), but we do not know yet whether such nitrate sinks are important also in Lake Lugano sediments.

Anaerobic N-cycling in the South Basin of Lake Lugano 225
Benthic denitrification and DNRA were the predominant anaerobic N-transformation processes at the two studied stations.
Interestingly, the contribution of DNRA was systematically higher than observed in flow-through whole-core incubations performed with sediment from the same basin. Wenk et al. (2014) reported a maximum DNRA contribution to NO3reduction of not more than 12%, but also argued that their DNRA rate measurements must be considered conservative, because they did not account for the production of 14 NH4 + from ambient natural-abundance nitrate. The reason for such a discrepancy is unclear, 230 but there seems to be a tendency for slurry incubations to yield higher DNRA rates compared to denitrification (Kaspar, 1983), implying biasing methodological effects. The observed discrepancies may also be related to natural sediment heterogeneity and/or seasonal/interannual fluctuations in benthic N transformation rates. As for the latter, in 2016, the annual water overturn and bottom-water ventilation was exceptionally suspended and sediments remained anoxic for more than a year. In contrast, in 2017, the water column mixed in January and surface sediments were oxygenated throughout June. Our O2 manipulation 235 experiments revealed that redox conditions have a marked impact on the partitioning between the two nitrate reduction pathways, and consistent with the slurry incubation data, the extended O2 exposure of microbes at the sediment-water interface in 2017 compared to the preceding year may have favoured nitrate ammonifiers over denitrifiers. Independent of any possible spatio-temporal variability, in this study, DNRA rates were equal, or even higher, than denitrification. Such a partitioning of the two nitrate reducing processes is not implausible and was similarly observed in a wide range of environments, particularly 240 in more reduced sediments with high organic matter content and comparatively low nitrate levels (Brunet and Garcia-Gil, 1996;Dong et al., 2011;Papaspyrou et al., 2014). More generally, substrate-availability changes induced by O2 fluctuations may be important drivers of the partitioning between denitrification and DNRA (Cojean et al., in prep.), and environmental conditions that favour DNRA over denitrification may be quite common. However, to our knowledge, experimental evidence for the direct O2 control on the balance between these two nitrate-reducing processes is still lacking. 245

O2 inhibition thresholds of benthic nitrate reduction
Our study shows that submicromolar O2 levels significantly lowered both, denitrification and DNRA rates. Denitrification and DNRA were inhibited by about 30-50% at 1 µmol L -1 O2, while in previous studies that investigated O2 effects on fixed-N elimination in the water column, denitrification was almost completely suppressed at this O2 level already. For example, by conducting incubation experiments using samples from oxygen minimum zones in the Eastern Tropical Pacific, a 50% 250 inhibition of denitrification was noticed already at 0.2 µmol L -1 O2, and complete suppression at 1.5-3 µmol L -1 O2 (Dalsgaard et al., 2014, Babbin et al., 2014. Similarly, incubation experiments with samples from a Danish fjord exhibited full inhibition of denitrification at 8-15 µmol L -1 O2 (Jensen et al., 2009). In marine sediments, in contrast, denitrification was occurring even at O2 concentrations greater than 60 µmol L -1 (Gao et al., 2010, Rao et al., 2007. This is in agreement with our results showing that at higher O2 levels (≥73 µmol L -1 ) denitrification was still active although at very low rates compared to the anoxic control 255 (≥ 93% inhibition). Similarly, DNRA was still occurring, and was less impaired by the elevated O2 concentration compared to denitrification (≥ 75% inhibition relative to the anoxic control). An increase of DNRA relative to denitrification rates under oxic conditions (> 100 µmol L -1 O2) was also observed in estuarine sediments, though N-removal remained predominant (Roberts et al., 2012(Roberts et al., , 2014. In brackish sediments in the Gulf of Finland in the Baltic Sea, at elevated O2 concentrations (from 50 to 110 µmol L -1 in bottom waters), benthic DNRA rates were generally higher than denitrification rates (Jäntti and Hietanen,260 2012), further supporting our findings. Yet, in contrast to our study, their observations suggest a higher O2 sensitivity (i.e., greater inhibition) of DNRA compared to denitrification in sediments with higher bottom water O2 concentrations (> 110 µmol L -1 ). Given the paucity and discrepancy of existing data in this context, it is premature to conclude that DNRA microbes are generally less or more oxygen-tolerant than denitrifiers. A direct comparison of DNRA O2 inhibition thresholds in this study and in the study of Jäntti and Hietanen (2012) is difficult because of the differing methodological approaches. There, nitrate 265 reduction rates were determined in whole-core incubations, without manipulating (and measuring) the O2 concentrations at the sediment depth where nitrate is actually reduced. And although the O2 penetration depth and porewater O2 concentrations will respond to a certain degree to the O2 content in the bottom water, deducing the actual O2 concentrations for the active nitrate reduction zone within the sediment from O2 concentrations in the overlying water is problematic. Here, we tested the oxygen sensitivity of a microbial community in suspension, directly exposed to defined O2 conditions. These incubation data indicate 270 that DNRA is less inhibited than denitrification at O2 concentrations ≥ 73 µmol L -1 and, at the same time, imply that anoxia per se is not a strict requirement for DNRA, as previous ecosystem-scale work has also suggested (Burgin and Hamilton, 2007). Our results also are consistent with observations made in soil microcosms showing that DNRA is less sensitive to increasing O2 partial pressures than denitrification within the range of 0-2% O2 v/v (Fazzolari et al., 1998;Morley and Baggs, 2010). 275 The observed O2 inhibition thresholds for nitrate reduction are significantly higher than reported from most incubation studies with water column samples (Dalsgaard et al., 2014, Babbin et al., 2014, Jensen et al., 2008. Elevated O2 tolerance in prior studies was often attributed to the formation of anoxic microniches that may foster anaerobic N-reduction (Kalvelage et al., 2011). It is unlikely that such microniches formed during our incubation experiments since slurries were heavily diluted (1 g 280 sediment in 70 mL water) and vigorously shaken by hand every 30 min, in addition to the continuous agitation on a shaking table during the incubation. Also, experiments were replicated 2-3 times for some O2-amended treatments, and measured rates were very similar between replicates. If anoxic microniches had formed, we would have expected that their formation is more variable, resulting in a lower reproducibility of the determined rates.
The existence of aerobic denitrifiers (e.g. microbes that reduce NO3 -/NO2to N2 in presence of O2) in soils and sediments has been confirmed through isolation of bacterial strains (e.g. Robertson et al., 1995), and it was suggested that they contribute to the total fixed N loss in marine sediments (Carter et al., 1995;Patureau et al., 2000;Zehr and Ward, 2002). Recent studies of permeable marine sediments (Gao et al., 2010) and soils (Bateman and Baggs, 2005;Morley et al., 2008) also observed significant N2 production in the presence of O2 and attributed it to aerobic denitrification 290

DNRA favoured under less reducing conditions
It is generally assumed that strongly reducing conditions favour DNRA over denitrification, yet in our study, particularly at elevated O2 concentrations, DNRA rates were higher than those of denitrification. That DNRA often seems to be more important under true anoxic conditions may therefore not be linked directly to the absence of O2 and differential O2 inhibition levels of the two nitrate-reducing processes. Indirect mechanisms are likely to be important. For instance, H2S accumulation, 295 which often accompanies prolonged anoxia, can inhibit denitrification and simultaneously enhance DNRA (An and Gardner, 2002;Rysgaard et al., 1996). Another indirect, redox-dependent factor may be the availability of nitrate. Higher DNRA rates were observed under more NO3 --limiting conditions induced by prolonged anoxia, probably because nitrate ammonifiers are able to gain more energy per NO3reduced than denitrifiers (Dong et al., 2011). As nitrate concentrations are generally much lower under oxygen-free conditions, it appears plausible that anoxia-associated nitrate and nitrite depletion is conducive to 300 higher DNRA/denitrification rates. While these examples seem to support that DNRA is favoured under true anoxic conditions, results of other studies are more consistent with our observation of higher DNRA than denitrification rates at elevated O2 concentrations. For example, in estuarine sediments, DNRA was stimulated relative to denitrification under more oxidizing conditions (Roberts et al., 2014(Roberts et al., , 2012. The authors argued that DNRA is enhanced by increasing Fe 2+ availability at the oxicanoxic sediment layer during more oxygenated conditions. These studies highlight the importance of redox conditions in 305 regulating the balance between dentrification and DNRA, however, to what extent O2 directly controls the partitioning between the two nitrate-reducing processes at the enzyme levels remains, to our knowledge, still unknown. Apparent contradictions with regards to how changing O2 levels may impact nitrate reduction may simply be due to the counteracting and variable influence of direct versus indirect effects of the variable O2 concentrations.

310
We cannot fully exclude that through O2 manipulation in this study, we partly affected nitrate-reduction indirectly through its control of H2S or Fe 2+ . Yet, we set up the experiments in a way that indirect effects should be minimized (e.g., no free sulfide measured in any of the incubations, same organic matter content, same excess NO3concentrations), and this study can thus be considered an investigation into the direct O2 effect on the partitioning between N-loss by denitrification and N-recycling by DNRA in aquatic sediments. The fact that in our experiments we can essentially exclude the effects of redox-related 315 parameter changes (i.e., H2S, NO3 -, and Fe 2+ ) leads us to the conclusion that in the studied sediments from Lake Lugano, O2 likely controls the balance between denitrification and DNRA at the organism-level, and that denitrification is in fact more sensitive towards increasing O2 concentrations than DNRA.

Direct O2 control on benthic NO3reduction
It has been previously reported that O2 can either suppress the synthesis of enzymes involved (Baumann et al., 1996) or the 320 enzyme activity itself (Dalsgaard et al., 2014). The observed DIN concentration trends (i.e. decreasing nitrate consumption) with increasing O2 concentrations suggest that the overall activity is modulated mainly at the nitrate reduction step. Without conclusive information on enzyme activities in hand, we can only speculate at this point about any real difference in O2dependent response of the enzymes involved in denitrification versus DNRA. The differential response of denitrifiers and nitrate ammonifiers may, however, suggest a distinct O2 sensitivity of the nitrate reductase enzymes involved. Denitrifiers and 325 nitrate ammonifiers utilize the same nitrate reductase enzymes (Nar, Nap), and while a differential O2 sensitivity of the same type of enzyme is difficult to explain, it is certainly possible for different enzymes. Indeed, the membrane-bound (Nar) and the periplasmic (Nap) nitrate reductases have distinct affinities towards NO3and O2 tolerance (Mohan and Cole, 2007).
Periplasmic nitrate reduction is almost exclusively found in the Proteobacteria and many of the organisms possess both Nar and Nap systems, whose production is regulated in response to ambient NO3and O2 concentrations (Simon and Klotz, 2013). 330 lake basins such as the south basin of Lake Lugano plays an important role in regulating the contribution of N-removal and Nrecycling in the water column (Lehmann et al. 2004;Wenk et al, 2014). To which extent O2 fluctuations affect N transformation 350 reactions within the sediments remains uncertain. Winter water column turnover ventilates the bottom waters and reoxygenates surface sediments that were anoxic for several months. Hence, at least in the top millimeters of the sediment column, we can expect changes in the benthic N cycling. Based on our incubation experiments, the O2 inhibition threshold was lower for denitrification than for DNRA, possibly reflecting differential adaption of the in situ microbial community of denitrifiers and nitrate ammonifiers to fluctuating O2 conditions of bottom waters. Indeed, many nitrate ammonifiers possess 355 both nitrate reductase enzymes (Nap and Nar) and can switch between the two respiratory systems providing them with an ecological advantage over denitrifiers when substrates become limiting (i.e., with regards to the primary reductant used in energy metabolism; Mohan and Cole, 2007). During oxygenated bottom-water conditions, within the benthic redox transition zone, nitrate-reducing microbes at the sediment-water interface will be exposed to elevated O2 concentrations, similar to the ones tested here. Our experimental data imply that then, at least in the uppermost sediments, DNRA is favoured over 360 denitrification. We may even expect an O2-regulated zonation of DNRA and denitrification. As a consequence, when denitrification-driven nitrate-reduction is pushed down, it is possible that NO3will be partially consumed through DNRA before it gets to the "denitrification layer", as nitrate ammonifiers are less O2 sensitive than denitrifiers. In contrast, denitrification is likely to be a more important nitrate-reducing process compared to DNRA during water column stratification (suboxia/anoxia of bottom waters), when the sediments are fully anoxic. 365 In the discussion thus far, we implicitly assume that the main control O2 exerts on the absolute and relative rates of denitrification and DNRA is due to its inhibitory effects at the organism-level, yet the effect of O2 on the coupling of nitrification and nitrate reduction by either denitrification or DNRA remained unaddressed. Oxygen fluctuations in the natural environment will affect nitrate regeneration by nitrification, and hence determine how much nitrate is available for microbial 370 reduction. It has been shown previously that through oxygenation events (e.g., the increase in bottom water O2 concentrations during episodic mixing/ventilation), the overall benthic N elimination in lakes may be enhanced through coupled nitrificationdenitrification, at least transiently (Hietanen and Lukkari, 2007;Lehmann et al., 2015). So, while the direct effect of elevated O2 would be to hamper fixed N elimination by denitrification at the organism-level, the oxygenation of previously ammoniumladen but nitrate free (pore-) waters would help to better exploit the benthic nitrate-reduction potential by increasing the nitrate 375 availability for nitrate-reducing microbes within the sediments, so that the overall nitrate reduction may be stimulated (Lehmann et al. 2015). Yet, as shown in the present study, oxygenation of the water column and the upper surface sediments may also act to shift the balance between denitrification and DNRA towards DNRA, thus promoting N-recycling rather than fixed-N elimination through denitrification. Total nitrification rates were not measured in this study, but nitrate concentration changes in sediment slurries suggest that at elevated O2 levels there is at least some production of nitrate. There is no obvious 380 reason to assume that O2-stimulation of the coupling of nitrification and denitrification on the one hand, and of nitrification and DNRA on the other would per se be different. Yet, as demonstrated here, DNRA appears to be less O2 sensitive compared to denitrification. It is thus reasonable to expect a higher coupling of nitrification with DNRA than with denitrification during oxygenated bottom-water conditions. Indeed, there is putative evidence for such an indirect link between O2 and elevated coupled nitrification-DNRA. In a recent study with estuarine sediments, stronger stimulation of DNRA compared to 385 denitrification was observed during oxygenation of bottom waters, in parts attributed to the coupling to nitrification (Roberts et al., 2012). Additional experimental work is required to better understand the role of nitrification in regulating the balance between benthic denitrification and DNRA during oxygenation of bottom waters.
It is important to understand that in the natural environment, O2 will not be the only regulator of the balance between 390 denitrification and DNRA. As previously mentioned, the partitioning of the two nitrate-reducing processes can also be modulated by the substrate (e.g., NO3 -, NO2 -, TOC, H2S, Fe 2+ ) availability. The latter may be redox controlled or not. Such regulation may be linked to the differential substrate affinity of the two processes when competing for the same electron acceptor (e.g., nitrate/nitrite) providing selective pressure that can drive communities either towards denitrification or DNRA (Kraft et al. 2014), or simply due to differing substrate requirements in the case of chemolithotrophic versus organotrophic 395 nitrate reduction.
For example, nitrate concentrations in the water column of the lake sampled in this study (Lake Lugano) varied significantly over the year, with very low NO3concentrations during the stagnation period (during anoxia; Fig. 1). As a consequence, it is reasonable to assume that the relative partitioning between denitrification and DNRA in a natural environment is affected by 400 the fluctuating nitrate concentrations (e.g., Tiedje et al., 1988, Dong et al., 2011. Similarly, Fe 2+ levels in near-bottom waters and sediment porewaters in Lake Lugano are greater during the anoxia/stratification period (Lazzaretti et al., 1992). At least in environments where chemolithotrophic processes contribute to the overall nitrate reduction, such redox-dependent Fe 2+ concentration changes (or changes of other electron donors such as HS -) may affect the balance between DNRA and denitrification (e.g., Robertson et al. 2015). Hence, in addition to the direct regulating effects of O2 on the partitioning between 405 denitrification and DNRA, which we have demonstrated here experimentally, O2 can act as indirect regulator of fixed N elimination versus regeneration. The ultimate ecosystem-scale DNRA/denitrification ratio in environments that are subject to O2 fluctuating conditions is difficult to predict, because direct and indirect O2 regulation may act concomitantly and in opposite ways.

Conclusion 410
The presented results broaden the range of O2 inhibition thresholds of benthic denitrification at micromolar O2 levels, demonstrating that benthic denitrification may resist full inhibition up to almost 80 µM O2. Similarly, sedimentary DNRA does not necessarily require true anoxia, and was even less sensitive than denitrification to higher O2 levels. Our data suggest that the balance between DNRA and denitrification is modulated by O2 at the nitrate-reducing enzyme level. However, more in-depth investigations on the exact role of oxygen in regulating other denitrification and/or nitrate-ammonification enzymes 415 in microbial pure culture experiments are needed. The differential tolerance of denitrifiers versus nitrate ammonifiers towards O2 has important implications for natural environments with fluctuating O2 conditions. Based on our results, one might argue that DNRA may be more important during phases of bottom-water oxygenation, while anoxic conditions during the stratification period may favour full denitrification to dinitrogen. Whether and when fixed nitrogen is preserved in a lake or eliminated by denitrification is, however, difficult to predict, as this will depend also on multiple indirect effects of changing 420 O2 levels. For example, nitrification and the redox-dependent modulation of substrates that may be relevant for denitrification or DNRA (such as nitrite, the substrate at the branching point between the two processes, and/or sulfide as potential inhibitor of denitrification and stimulator of chemolithotrophic DNRA) will play an important role both with regards to the overall nitrate reduction rate, as well as the balance between different nitrate reducing processes. Internal eutrophication from N in high-productivity lakes is generally less of a concern than from P. Nevertheless, it needs to be considered that oxygenation 425 may reduce the overall fixed N-elimination capacity of the bottom sediments by impairing denitrification more than DNRA, partially counteracting the generally positive effects of hypolimnetic ventilation in the context of benthic nutrient retention/elimination, and with implications on the nutrient status in the water column.

Data availability
Data can be accessed upon request to the corresponding author. 430 Akademische Gesellschaft (FAG) Basel that also financially supported the study. correspond to net NO3or NH4 + consumption and production rates over incubation time, respectively. Standard errors are indicated in bracket for n replicates. Average denitrification and DNRA rates (µmol N g -1 wet sed. d -1 ) in anoxic control experiments were: 0.11 ± 0.01 and 0.12 ± 0.04, respectively, at Figino; 0.12 ± 0.01 and 0.11 ± 0.01, respectively, at Melide.