Non-canonical reductive nitrous oxide production pathways in a seasonally stratified lake basin

Non-canonical reductive nitrous oxide production pathways in a seasonally stratified lake basin | bioRxiv /* */ /* */ <!-- <!-- /*! * yepnope1.5.4 * (c) WTFPL, GPLv2 */ (function(a,b,c){function d(a){return"[object Function]"==o.call(a)}function e(a){return"string"==typeof a}function f(){}function g(a){return!a||"loaded"==a||"complete"==a||"uninitialized"==a}function h(){var a=p.shift();q=1,a?a.t?m(function(){("c"==a.t?B.injectCss:B.injectJs)(a.s,0,a.a,a.x,a.e,1)},0):(a(),h()):q=0}function i(a,c,d,e,f,i,j){function k(b){if(!o&&g(l.readyState)&&(u.r=o=1,!q&&h(),l.onload=l.onreadystatechange=null,b)){"img"!=a&&m(function(){t.removeChild(l)},50);for(var d in y[c])y[c].hasOwnProperty(d)&&y[c][d].onload()}}var j=j||B.errorTimeout,l=b.createElement(a),o=0,r=0,u={t:d,s:c,e:f,a:i,x:j};1===y[c]&&(r=1,y[c]=[]),"object"==a?l.data=c:(l.src=c,l.type=a),l.width=l.height="0",l.onerror=l.onload=l.onreadystatechange=function(){k.call(this,r)},p.splice(e,0,u),"img"!=a&&(r||2===y[c]?(t.insertBefore(l,s?null:n),m(k,j)):y[c].push(l))}function j(a,b,c,d,f){return q=0,b=b||"j",e(a)?i("c"==b?v:u,a,b,this.i++,c,d,f):(p.splice(this.i++,0,a),1==p.length&&h()),this}function k(){var a=B;return a.loader={load:j,i:0},a}var l=b.documentElement,m=a.setTimeout,n=b.getElementsByTagName("script")[0],o={}.toString,p=[],q=0,r="MozAppearance"in l.style,s=r&&!!b.createRange().compareNode,t=s?l:n.parentNode,l=a.opera&&"[object Opera]"==o.call(a.opera),l=!!b.attachEvent&&!l,u=r?"object":l?"script":"img",v=l?"script":u,w=Array.isArray||function(a){return"[object Array]"==o.call(a)},x=[],y={},z={timeout:function(a,b){return b.length&&(a.timeout=b[0]),a}},A,B;B=function(a){function b(a){var a=a.split("!"),b=x.length,c=a.pop(),d=a.length,c={url:c,origUrl:c,prefixes:a},e,f,g;for(f=0;f<d;f++)g=a[f].split("="),(e=z[g.shift()])&&(c=e(c,g));for(f=0;f<b;f++)c=x[f](c);return c}function g(a,e,f,g,h){var i=b(a),j=i.autoCallback;i.url.split(".").pop().split("?").shift(),i.bypass||(e&&(e=d(e)?e:e[a]||e[g]||e[a.split("/").pop().split("?")[0]]),i.instead?i.instead(a,e,f,g,h):(y[i.url]?i.noexec=!0:y[i.url]=1,f.load(i.url,i.forceCSS||!i.forceJS&&"css"==i.url.split(".").pop().split("?").shift()?"c":c,i.noexec,i.attrs,i.timeout),(d(e)||d(j))&&f.load(function(){k(),e&&e(i.origUrl,h,g),j&&j(i.origUrl,h,g),y[i.url]=2})))}function h(a,b){function c(a,c){if(a){if(e(a))c||(j=function(){var a=[].slice.call(arguments);k.apply(this,a),l()}),g(a,j,b,0,h);else if(Object(a)===a)for(n in m=function(){var b=0,c;for(c in a)a.hasOwnProperty(c)&&b++;return b}(),a)a.hasOwnProperty(n)&&(!c&&!--m&&(d(j)?j=function(){var a=[].slice.call(arguments);k.apply(this,a),l()}:j[n]=function(a){return function(){var b=[].slice.call(arguments);a&&a.apply(this,b),l()}}(k[n])),g(a[n],j,b,n,h))}else!c&&l()}var h=!!a.test,i=a.load||a.both,j=a.callback||f,k=j,l=a.complete||f,m,n;c(h?a.yep:a.nope,!!i),i&&c(i)}var i,j,l=this.yepnope.loader;if(e(a))g(a,0,l,0);else if(w(a))for(i=0;i (function(w,d,s,l,i){w[l]=w[l]||[];w[l].push({'gtm.start':new Date().getTime(),event:'gtm.js'});var f=d.getElementsByTagName(s)[0];var j=d.createElement(s);var dl=l!='dataLayer'?'&l='+l:'';j.src='//www.googletagmanager.com/gtm.js?id='+i+dl;j.type='text/javascript';j.async=true;f.parentNode.insertBefore(j,f);})(window,document,'script','dataLayer','GTM-M677548'); Skip to main content Home About Submit ALERTS / RSS Search for this keyword Advanced Search New Results Non-canonical reductive nitrous oxide production pathways in a seasonally stratified lake basin View ORCID Profile Teresa Einzmann , View ORCID Profile Moritz F. Lehmann , View ORCID Profile Jakob Zopfi , View ORCID Profile Claudia Frey doi: https://doi.org/10.1101/2025.08.15.670496 Teresa Einzmann † Department of Environmental Sciences, University of Basel , 4056 Basel, Switzerland Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Teresa Einzmann For correspondence: teresa.einzmann{at}unibas.ch Moritz F. Lehmann † Department of Environmental Sciences, University of Basel , 4056 Basel, Switzerland Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Moritz F. Lehmann Jakob Zopfi † Department of Environmental Sciences, University of Basel , 4056 Basel, Switzerland Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Jakob Zopfi Claudia Frey † Department of Environmental Sciences, University of Basel , 4056 Basel, Switzerland Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Claudia Frey Abstract Full Text Info/History Metrics Supplementary material Preview PDF ABSTRACT Nitrous oxide (N 2 O), a potent greenhouse gas and ozone-depleting agent, was detected at high concentrations in the anoxic bottom-waters of a monomictic eutrophic lake basin in Switzerland (Lake Lugano). The observed high site-specific nitrogen (N) isotope preference (SP), was inconsistent with bacterial denitrification, which typically exhibits low SP, thereby challenging its role as the primary N 2 O source. This pointed to chemo-denitrification and/or fungal denitrification, both characterized by high SP, as possible alternative pathways. We conducted incubation experiments with sediment and bottom-water samples to assess N 2 O production and reduction dynamics and associated natural-abundance stable isotope signatures. We demonstrate that N 2 O accumulation predominantly originated from sedimentary production, and that elevated SP values in the bottom water reflected fractional bacterial N 2 O reduction. Using an isotope mass balance mixing model, we identified bacterial denitrification as the dominant sedimentary process (∼75%), followed by chemo-denitrification (∼20%), and fungal denitrification (∼5%). Additional 15 N tracer incubation experiments, combined with selective inhibitors to quantify isolated pathways, confirmed model-estimated contributions. These findings validate the use of literature-based SP values in mixing models, and provide evidence for non-canonical N 2 O production via chemo- and fungal denitrification, highlighting the need to broaden our understanding of N 2 O cycling in lakes beyond classical bacterial pathways. SYNOPSIS Eutrophic lake conditions significantly enhance nitrous oxide production through diverse microbial processes, including non-traditional pathways beyond classical nitrification and denitrification. This study highlights the importance of incorporating chemo- and fungal denitrification when investigating aquatic nitrous oxide cycling. Download figure Open in new tab 1 INTRODUCTION Nitrous oxide (N 2 O) is a potent greenhouse gas with a warming potential up to 300 times greater than that of carbon dioxide (CO 2 ) over a 100-year timescale, and it is the primary agent responsible for ozone depletion. 1 Since the pre-industrial era, atmospheric N 2 O concentrations have increased from 270 ppb in 1750 to 336 ppb in 2022, with a record growth rate of 1.38 ppb per year in 2021. 2 A major driver of this trend is the intensification of agricultural fertilization leading to increased nitrogen (N) loading into ecosystems, and results in elevated N 2 O emissions from terrestrial and aquatic environments. 3 , 4 Despite evidence of rising N 2 O emissions from inland waters, 5 some freshwater environments act as sinks for atmospheric N 2 O. 6 , 7 Understanding the dual role of lacustrine systems in both N 2 O production and removal is crucial for predicting their response to anthropogenic nutrient inputs. In aquatic systems, N 2 O is primarily produced through three key microbial pathways: nitrification, nitrifier-denitrification and denitrification. Nitrification, an aerobic process, involves the oxidation of ammonia (NH₃) or ammonium (NH 4 + ), the latter being the dominant species of NH 4 + /NH 3 in the pH range of most natural aquatic system, to nitrate (NO 3 - ) via nitrite (NO₂⁻), with N 2 O released as a by-product. Under low-oxygen conditions, nitrifiers may simultaneously reduce NO 2 - to N 2 O and dinitrogen (N 2 ), a process known as nitrifier- denitrification. In contrast, denitrification is an anaerobic process where bacteria reduce NO 3 - to N 2 , with N 2 O formed as an intermediate. Dissimilatory nitrate reduction to ammonium (DNRA) is another anaerobic pathway, which converts NO 3 - to NH + via NO 2 - , and can also lead to N 2 O production. However the underlying enzymatic mechanisms are not fully understood. 8 , 9 Of these processes, denitrification is the only known biological sink for N 2 O, especially under oxygen-depleted conditions. However, many denitrifying bacteria lack the complete gene set necessary to reduce N 2 O to N 2 , resulting in a truncated denitrification pathway. 10 , 11 Moreover, the enzymatic reactions involved in denitrification vary in their sensitivities to oxygen, with nitrous oxide reductases (NOS) being particularly oxygen-sensitive. 12 As a result, N 2 O often accumulates under suboxic conditions while complete denitrification dominates in the anoxic part of the water column. Thus, oceanic oxygen-deficient zones (ODZ), for example, enhance N 2 O consumption, whereas conditions along suboxic-anoxic interfaces favour N 2 O production, leading to higher N 2 O yields due to incomplete denitrification. 13 – 15 Recently, alternative N 2 O production pathways have been identified, including denitrification by fungi and chemo-denitrification, an abiotic process involving reactions between ferrous iron (Fe(II)) and nitrite or nitrate in coastal sediments and ferruginous waters. 16 – 18 Denitrifying fungi, which lack the gene encoding NOS, produce N 2 O as a terminal product, making their role in N 2 O production ecologically relevant. 19 Fungal denitrification is a recognised source of N 2 O in terrestrial and coastal environments, 16 , 17 , 20 , 21 and has also been implicated in the open ocean, where 18-22% of N 2 O at the oxic-anoxic interface of an ODZ may be of fungal origin. 22 Despite growing evidence for both fungal and abiotic N 2 O production, the role of these alternative N 2 O production pathways in freshwater lakes remains underexplored. This, in turn, suggests that microbial denitrification may have been overestimated as the dominant source of N 2 O in lacustrine environments. Stable N and O isotopic analysis of N 2 O is widely used to differentiate among N 2 O production pathways. Isotopic signatures depend the kinetic isotope effects (ε), which reflect the preferential use of the lighter isotopes (e.g., 14 N versus 15 N) during enzymatic or abiotic reactions, and on the isotopic composition of precursor substrates. Moreover, the intramolecular distribution of 15 N between the central (α) and outer (β) N atoms, expressed as site preference (SP = δ¹⁵Nα – δ¹⁵Nβ), provides insight into the formation mechanisms. 23 , 24 Characteristic SP ranges have been reported for different N 2 O production pathways: ∼-8 to 4‰ for incomplete bacterial denitrification, ∼30 to 39‰ for fungal denitrification, and ∼10 to 27‰ for chemo-denitrification. 25 – 27 However, the use of SP for differentiating N 2 O production pathways is not always straightforward. Hybrid N 2 O formation, where N atoms originate from different substrates, can yield variable SP values (Kelly et al. 2024). Additionally, flavohemoglobin (fhp), a nitric-oxide (NO) reducing, detoxifying enzyme common in bacterial denitrifiers, has been shown to produce N 2 O with elevated SP values (∼11‰) compared to respiratory nitric oxide reductases (NOR). 28 Furthermore, N 2 O reduction increases SP, 29 potentially obscuring the SP signature of the production process. To address this, combining SP with δ 18 O-N 2 O measurements has proven effective in disentangling N 2 O production and reduction processes. 16 , 30 In this study, we investigate reductive N 2 O production pathways in the bottom water and surface sediment in the South Basin (SB) of Lake Lugano (Switzerland), where N 2 O concentrations as high as 900 nM have been observed during summer stratification. 31 High sediment-to-water N 2 O fluxes have been observed year-round, 32 but previous studies have attributed N 2 O production in the sediment solely to bacterial denitrification, potentially overlooking alternative pathways. We performed incubation experiments with bottom water and surface sediments to determine the relative contributions of bacterial, fungal and chemo-denitrification to N 2 O production/ accumulation in deep-hypolimnetic waters of Lake Lugano during summer stratification. Incubation experiments with 15 N-NO 3 - , combined with pathway-specific inhibitors, enabled quantification of N 2 O production and consumption rates. Parallel natural-abundance (NA) incubations, conducted under the same inhibitory conditions, were used for isotopic signature identification of N 2 O. Dual-isotope mapping and a Bayesian isotope mixing model were applied to the NA-incubation data to apportion contributions from individual N 2 O formation pathways. The 15 N-incubations further served to validate the reliability of using literature- derived isotope endmembers for N 2 O source attribution in freshwater systems. 2 MATERIALS AND METHODS 2.1 Study site and sampling The study was conducted in the monomictic, eutrophic South Basin (SB) of Lake Lugano, located at the Swiss/Italian border. The basin stratifies seasonally from June to January and has a maximum depth of 95 m. 32 , 33 Sampling was carried out in June and August 2023 at the basin’s deepest point (8°53’37”E, 45°57’31”N). A conductivity-temperature-depth profiler (CTD) (SBE 19plus V2 SeaCAT Profiler CTD, Seabird) was used to measure temperature, oxygen concentrations and turbidity in the water column. The first sampling campaign in June focussed on assessing thermal stratification and bottom-water N 2 O accumulation, while the August sampling targeted the collection of bottom-water and surface sediments for incubation experiments. Duplicate samples for N 2 O concentration and isotopic analysis were collected at 5-10 m depth intervals using 5 L Niskin bottles. After overflow, samples were transferred into acid-washed, muffled (6 h at 450 °C) 160 mL glass serum bottles via gas-tight Tygon tubing. A 10 mL headspace was created, and 3 mL of 10 M NaOH was added as a preservative. Bottles were sealed with a grey butyl rubber septum (hollow stoppers grey, 20 mm, Chromatographie Service GmbH), crimped, and stored upside down at room temperature in the dark until isotope-ratio mass spectrometry (IRMS) analysis. For NO 3 - and NO 2 - analysis, samples were filtered (0.2 µm pore-size, Filtropur S, Sarstedt) and frozen at -20 °C. Samples for Fe(II) analysis were not filtered and were fixed with sulfamic acid (40 mM final concentration). For incubation experiments with unfiltered bottom water, water was collected at 93 m depth directly into muffled 2 L Schott bottles and sealed with a rubber stopper to avoid oxygen contamination. For filtered bottom-water incubations, water was passed through a Sterivex filter (Millipore, membrane material, 0.2 µm) using a peristaltic pump. For sediment slurry incubations, three sediment cores (diameter: 6 cm) were collected using a custom-made gravity corer, transported back to the laboratory and stored in the cold room until further processing within 24 hours. For slurry incubations, the top 5 cm of each core were combined and homogenised and bottom water was filtered as described above into acid-washed (10% HCl) and DI water-rinsed (5x times) plastic canisters. All samples were transported to the laboratory at 4 °C in the dark. 2.2 Chemical and stable isotope analyses Concentrations of NO 3 - were quantified by ion chromatography (940 Professional IC Vario, Methrom), while NO 2 - was determined using standard spectrophotometric methods involving sulfanilamide and the Griess reagent (naphtal-ethylenediamine dihydrochloride). 34 Concentrations of Fe 2+ in sulfamic-acid-preserved samples were determined spectrophotometrically using the ferrozine assay. 35 For N 2 O concentration and isotopic analysis, serum bottles were purged with helium (He) as carrier gas for 38 min. The N 2 O was trapped in liquid N 2 following H 2 O removal using an ethanol trap (-60 °C) and a Mg(ClO 4 ) 2 trap; CO 2 was removed with Ascarite. Remaining traces of CO 2 and sample N 2 O were separated using a gas chromatograph (GC). Masses 44, 45, 46 (and their respective isotope ratios 45/44, 46/44) of the purified N 2 O, as well as the NO + fragment ions at masses 30 and 31 (and isotope ratio 31/30), were analyzed by GC-IRMS (Delta V Plus, Thermo), with reference gas injection and flow control handled via a Conflo IV interface. The N 2 O isotope ratios were calibrated against reference N 2 O gas (≥ 99.9986%, Messer), aligned to the Tokyo Institute of Technology scale (Mohn et al. 2012) for both bulk and site- specific isotopic composition. Ratios of m/z 45/44, 46/44, and 31/30 signals were converted to δ 15 N-N 2 O (referenced to AIR), δ 18 O-N 2 O (referenced to Vienna Standard Mean Ocean Water, V-SMOW), and site-specific δ 15 N α and δ 15 N β -N 2 O using three calibration standards (mixtures of N 2 O in synthetic air; CA06261, 53504, CA08214) and two scrambling factors (gamma and kappa) according to Frame and Casciotti (2010). N 2 O isotopocule corrections were performed using the pyisotopomer Python package (Kelly et al. 2023). N 2 O concentrations were calculated by calibrating the total peak areas against standards of known concentrations, prepared by concerting 20 nmol NO 3 - standards to N 2 O using the denitrifier method (see below). The isotope ratios of NO 3 - (N and O) were measured using the denitrifier method, 36 , 37 in which NO 3 - was converted to N 2 O by Pseudomonas chlororaphis subsp. aureofaciens (ATCC 13985) and subsequently analysed by GC-IRMS as above. Nitrate isotope measurements were calibrated using the international NO 3 - standards IAEA-N3 (δ 15 N = 4.7‰, δ 18 O = 25.6) and USGS34 (δ 15 N = -1.8‰, δ 18 O = -27.9‰), and an in-house standard (UBN-1; δ 15 N = 14.15‰, δ 18 O = 25.7‰). Isotope ratios of N 2 were determined with a GasBench II coupled to an IRMS (Delta V Plus, Thermo). Air N 2 was used as both calibration and drift-correction standard for δ 15 N-N 2 . 2.3 Incubation experiments Two sets of experiments were conducted: one with 15 N-labelled nitrate to measure N 2 O production and consumption rates, and another with NA nitrate additions to examine shifts in δ 18 O and SP of the produced N 2 O in association with specific processes. In both experimental sets, four treatments with specific inhibitors were applied to selectively suppress different N 2 O production pathways (Table S1 and S2). The four treatments were as follows: 1. Control - with no inhibitors added; 2. Bacterial - fungal inhibitor added, allowing N 2 O accumulation only from incomplete bacterial or chemo- denitrification; 3. Fungal - bacterial inhibitor added to allow N 2 O production only via fungal or chemo-denitrification; 4. Chemo - both fungal and bacterial inhibitors added to supress biological N 2 O production completely, leaving only chemo-denitrification. Although chemo-denitrification may occur in all treatments to some degree, treatment 2 and 3 are referred to as bacterial and fungal for clarity. The occurrence of DNRA and its potential contribution to N 2 O production was not considered in our incubation experiments; however the importance of DNRA in the SB of Lake Lugano was investigated in previous studies, reporting varying contributions below 12% 38 and up to 50% 39 to NO 3 - reduction compared to bacterial denitrification. Streptomycin (1.29 mM final concentration) and cycloheximide (5.33 mM) were used as bacterial and fungal inhibitors, respectively, following concentrations previously shown to achieve up to 90% inhibition. 16 , 40 A 24 h pre-incubation period allowed the inhibitors to take effect. Afterward, all vials were He-purged for 30 minutes to remove O 2 and N 2 O. To ensure sufficient Fe(II) availability for chemo-denitrification (in case in situ Fe(II) had already been oxidised before the start of the incubations), ferrous iron was added as anoxic FeSO 4 solution to a final concentration of 2 µM, reflecting in situ lake conditions. No Fe(II) was added to filtered bottom-water samples, which served as abiotic controls. In sediment incubations, Fe(II) was assumed to be naturally available based on previously determined porewater and solid-phase profiles. 41 Oxygen concentrations were monitored in separate incubation bottles per treatment using non-invasive optical trace-range O 2 sensor spots (Pyrocience, Firesting, detection limit 0.005% O 2 ). 15N-incubation experiments 15 N-incubations were conducted to determine potential N 2 O production rates following the addition of 15 N-labelled KNO 3 (≥98 atom % 15 N, Cambridge Isotopes). For bottom-water incubations ( 15N BW), 18 ml of inhibitor-treated bottom water was filled into muffled 20 mL glass serum vials, leaving a 2 mL headspace. For sediment incubations ( 15N Sed), 2 g of mixed sediment (top 5 cm) was transferred into 20 ml vials, and 8 ml of treated, filtered bottom water was added. All vials were crimp-sealed and shaken vigorously. After 24 h-pre-incubation and subsequent He-purging (30 min.), natural-abundance N 2 O was injected into each vial to achieve a background concentration of ∼50 nM. This ensured sufficient total N 2 O for reliable isotopic measurements by IRMS and allowed detection of 15 N-labelled N 2 O as it accumulated during the incubation. Experiments comprised a total of eight treatments: 1) 15N BW-Control, 2) 15N BW-Bacterial, 3) 15N BW-Fungal, 4) 15N BW-Chemo, 5) 15N Sed-Control, 6) 15N Sed- Bacterial, 7) 15N Sed-Fungal, 8) 15N Sed-Chemo (Table S1). A filtered bottom-water treatment without Fe(II) served as negative abiotic control. Each vial received 15 N-NO 3 - to a final concentration of 10 µM, and was incubated in the dark at 7 °C. For each treatment, 15 replicate vials were prepared. At designated time points (0, 6, 12, 16, 24 h for 15N BW; 0, 1, 2, 4, 6 h for 15N Sed) three vials per time point were sacrificed by adding 200 µL of 10 M NaOH. For N 2 O reduction rate determination, 6 ml of liquid was subsampled from each vial prior to NaOH addition, transferred to He-purged exetainers, and subsequently analysed for δ 15 N 2 . Two assumptions were made for calculating N 2 O production: 1) chemo-denitrification occurs across all treatments and must be subtracted from total N 2 O production in fungal and bacterial treatments (for both 15N BW and 15N Sed), and 2) N 2 O reduction by bacterial denitrifiers occurs in the 15N Control and 15N Bacterial treatments and must be added back to production estimates to correct for fractional N 2 O loss due to reduction to N 2 . These corrections ensured accurate quantification of both N 2 O production and consumption in the absence of acetylene (C 2 H 2 ), which was not used in 15 N incubations to block N 2 O reduction (Table S1). Natural-abundance isotope incubation experiments NA-incubations were performed to determine the isotopic composition (δ 15 N, δ 18 O, SP) of newly produced N 2 O over a 6-day period. After the pre-incubation, 145 mL of treated unfiltered bottom water was transferred into acid-washed, muffled 160 mL serum bottles, and crimp-sealed with grey butyl rubber septa and aluminium caps. NA-sediment incubations were prepared in the same way as described for the 15 N-sediment incubations. Eight treatments were prepared in quadruplicate: 1) NA BW-Control, 2) NA BW-Bacterial, 3) NA BW-Fungal, 4) NA BW-Chemo, 5) NA Sed-Bacterial, 6) NA Sed-Chemo, 7) NA Sed-Fungal, 8) NA Sed-Chemo. Again, a filtered bottom-water incubation without Fe(II) was included as an abiotic negative control. To inhibit N 2 O reduction to N 2 in a subset of samples, 1.5 ml of 100% acetylene gas (Carbagas) was injected into duplicate vials per treatment, reaching 10% v/v in the headspace (+C 2 H 2 ). In the other subset of duplicate vials, no C 2 H 2 was added to allow the potential occurrence of N 2 O reduction (Table S2). After 30 min of He-purging, NA KNO 3 - (δ 15 N = -29.6‰) was added to 10 µM final concentration. All samples were then incubated at 7 °C in the dark for 6 days. After incubation, duplicate samples per treatment were fixed by adding 3 ml of 10 M NaOH to 160 mL bottom- water bottles and 200 µL of 10 M NaOH to 20 mL sediment vials. In cases where N 2 O accumulation exceeded 15 nmol per vial, samples were split into two separate vials to allow analysis of technical replicates for NA BW and NA Sed incubations. 2.4 Rate calculations Rates of N 2 O production (R N2O in nM N/d) were calculated from the linear increase in mass 45 and 46 isotopologues of N 2 O over time in 15 N incubation experiments. 42 The total N 2 O production rate (R) was calculated according to Eq. 1 . 43 where F refers to the fraction of 15 N in the NO 3 - substrate pool (𝐹𝐹 = 15 𝑁𝑁/ ( 15 𝑁𝑁 + 14 𝑁𝑁)), which is assumed to be constant over the duration of the incubation. The probability of 46 N 2 O production is proportional to 1/F 2 , therefore the term includes an extra factor of 1/F for 46 N 2 O relative to 45 N 2 O production. F was 0.48 in the 15N BW-incubations and 0.78 in 15N Sed-incubations. Rates were considered significant at p < 0.05. N 2 O reduction rates (R N2O_red in nM N/d) were calculated from the increase in excess 30 N 2 over time, using the same time points as for N 2 O production rate determination ( Eq. 2 ). 44 No 29 N 2 formation was observed. The reduction rated was calculated as: 2.5 Modelling the relative contributions of denitrification pathways using an isotope mass balance approach To quantify the contributions of bacterial denitrification, fungal denitrification and chemo-denitrification to N 2 O production and reduction, we applied the Fractionation and Mixing Evaluation (FRAME) model. FRAME is a Bayesian isotope mixing model with a user-friendly graphical interface (malewick.github.io/frame), recently developed by Lewicki et al 45 . The model enables simultaneous N 2 O source apportionment (and associated probability distributions) and estimation of isotope fractionation due to N 2 O reduction, using a Markov-Chain Monte Carlo approach. 45 The model has been successfully applied in recent N 2 O source apportionment studies. 46 – 48 Here, FRAME was applied to NA-control incubations of both bottom water and sediment (i.e. treatments permissive to all denitrification pathways), and an in situ bottom-water sample taken at 92.5 m depth in August. The model used prior knowledge on isotope endmember values (SP, δ 15 N, δ 18 O) from the literature for bacterial, fungal, and chemo-denitrification as potential N 2 O sources (Table S3 and Supplemental Material). Nitrification was excluded as a N 2 O source, as all incubations and in situ water samples were anoxic. To test the model’s ability to identify N 2 O reduction, FRAME was applied to incubation samples both with and without acetylene (C 2 H 2 ). While C 2 H 2 inhibits N 2 O reduction it may not fully suppress it. 49 – 51 Including C 2 H 2 -amended samples in the modelling efforts allowed us to estimate how much residual unreacted N 2 O the model would estimate. Notably, model outputs generally exhibited less uncertainty for incubation samples with C 2 H 2 , likely due to a cleaner N 2 O isotopic source signal (less overprinting by N 2 O reduction; Fig. S6a, d). Therefore, in the results section, we focus on results from C 2 H 2 -amended incubations. Model outputs from non-C 2 H 2 incubations are provided in the Supplementary Material (Fig. S4a, d, Fig. S7a, f). The model can run in 2D-isotopic space (SP and δ 18 O) or 3D (SP, δ 18 O and δ 15 N bulk ). In this study, 3D modelling provided the most conclusive and robust source-separation constraints (see Supplementary Material for details). 3 RESULTS 3.1 Vertical distribution of nitrogen species and isotopic signatures in the South Basin During the sampling months of June and August, water column stratification was evident, marked by a steep thermocline and oxycline ( Fig. 1a, e ). Surface-water temperatures ranged from 25-26 °C and decreased to 7 °C at depths of 25 m and below. The oxic-anoxic interface, defined here as the depth where O 2 concentrations dropped below 2 µM, shifted upward from 85 m in June to 76 m in August ( Fig. 1a , 1e). Download figure Open in new tab Figure 1. Depth profiles of physical and hydrochemical parameters in the Lake Lugano South Basin in June 2023 (upper panels) and August 2023 (lower panels). Shown are temperature, dissolved oxygen concentrations, and turbidity ( a, e ); concentrations of NO 2 - , NO 3 - , N 2 O and Fe(II), as well as calculated N 2 O concentrations for atmospheric equilibrium conditions, ( b, f ); isotopic values for N 2 O ( c, g ); and isotopic values for NO 3 - ( d, h ). Concentrations of N 2 O in surface waters reached 11 nM, slightly exceeding the atmospheric equilibrium levels (8 nM) ( Fig. 1b ). Concentrations increased to 34 nM at the onset of the oxycline (defined here as the water-column layer where O 2 decreases from 40 µM to 2 µM; i.e. 80-85m in June and 70-76m in August; Fig. 1a, e ), with a δ 15 N bulk -N 2 O value of approximately 3‰ ( Fig. 1c, g ). SP increased from 24‰ at the surface to 34 ‰ at the start of the oxycline in both months. δ 18 O-N 2 O mirrored SP, with values around 51‰ in the upper water column, rising to 60‰ at greater depths ( Fig. 1c, g ). Within the oxycline N 2 O concentrations decreased, reaching a minimum of 12 nM in June and 13 nM in August just below the oxic-anoxic interface ( Fig. 1b, f ). At these depths, δ 15 N bulk -N 2 O values were lowest (June: -14‰ and August: -7‰, Fig. 1c, g ), while SP (June: 43‰, August: 51‰) and δ 18 O (June: 60‰ August: 70‰) reached their highest values. These trends intensified, and shifted upward in August, corresponding to the upward migration of the oxic- anoxic interface. Toward the sediment, N 2 O concentrations sharply increased, reaching 920 nM in June and 3032 nM in August in the near-bottom-water ( Fig. 1b, f ), where increased turbidity indicates the development of a benthic nepheloid layer 33 ( Fig. 1a, e ). In August, this N 2 O accumulation was accompanied by an increase in δ 15 N bulk -N 2 O to 12‰, and decreases in SP and δ 18 O-N 2 O to 34‰ and 60‰, respectively ( Fig. 1g ). Nitrate concentrations declined throughout the anoxic water column toward the sediment, while δ 15 N-NO 3 - and δ 18 O-NO 3 - values increased ( Fig. 1d, h ). Nitrite concentrations were below the detection limit in the oxic water column but reached a maximum of 20 µM in the lower anoxic water layer. Similarly, Fe(II) was only detected in the anoxic bottom water, with concentrations reaching 9 µM ( Fig. 1f ). 3.2 N 2 O production and N 2 O reduction rates from 15 N-incubations In bottom-water (BW) incubations, the highest N 2 O production rates were observed in the control treatment followed, by the bacterial treatment, which accounted for 58 ± 15% of total production ( Table 1 ). Chemo and fungal treatments showed similar N 2 O production, representing 18 ± 2% and 24 ± 9% of the total N 2 O production rate. No N 2 O production was observed in the abiotic control (data not shown). Within error margins, the sum of chemo, fungal and bacterial-derived N 2 O production aligned with the rates observed in the control treatment, supporting the validity of the selective inhibition approach. View this table: View inline View popup Download powerpoint Table 1. Rates of N 2 O production and reduction from 15 N-incubation experiments. All rates are based on triplicate incubations and were significant (p <0.05). Sediment incubations yielded substantially higher volumetric N 2 O production rates overall ( Table 1 ). The control and bacterial treatments displayed the highest N 2 O production rates with the bacterial contribution, representing 75 ± 12% of total production. Chemo- and fungal denitrification showed significantly lower rates, accounting for 20 ± 2% and 6 ± 2% of the total N 2 O production rate, respectively. N 2 O reduction was detected only in sediment incubations, specifically in the control and bacterial treatment ( Table 1 ). No N 2 O reduction activity was observed in the fungal and chemo treatments. 3.3 Natural-abundance incubations – δ 18 O and site preference of accumulated N 2 O After six days of incubation, N 2 O accumulation was much higher in sediment incubations with inhibited N 2 O reduction (+C 2 H 2 ; up to 20’000 nM N 2 O) compared to BW incubations (+C 2 H 2 ; up to 2’000 nM N 2 O) (Fig. S1). In the BW+C 2 H 2 treatments, SP and δ 18 O values were similar across all inhibitor conditions ( Fig. 2a , Table S5). The bacterial+C 2 H 2 treatment yielded values within the expected range for bacterial denitrification. Unexpectedly, the chemo+C 2 H 2 and fungal+C 2 H 2 treatments produced comparable isotopic signatures (SP -1.8 ± 0.7‰ and 0.0 ± 0.1‰; δ 18 O 20.4 ± 0.7‰ and 22.1 ± 0.2‰, respectively). The control+C 2 H 2 treatment showed slightly lower SP (-2.5 ± 1.2‰) and δ 18 O (17.0 ± 0.2‰) values. Download figure Open in new tab Figure 2. N 2 O site preference (SP) plotted against δ 18 O of N 2 O in situ bottom-water (92.5 m) and for the various treatments after six days of incubation: a) with C 2 H 2 addition to inhibit N 2 O reduction, and b) without C 2 H 2 . The δ 18 O values are corrected for the δ 18 O of ambient water (Δδ 18 O(N 2 O – H 2 O); δ 18 O-H 2 O = -7‰ 52 ). Black dashed lines with arrows indicate the expected isotopic enrichment from bacterial N 2 O reduction of εSP/ε 18 O slopes ranging from of 0.23 to 0.45. 53 The abiotic control (filtered bottom-water without Fe(II) addition) is excluded, as no N 2 O accumulation was observed. Standard errors are shown for samples where technical replicates were analysed. In BW incubations without C 2 H 2 (i.e., allowing N 2 O reduction), only the bacterial treatment exhibited a significant shift in the N 2 O isotopic signature, with SP increasing by 6.4‰ and δ 18 O by 47.6‰ ( Fig. 2b ). Other treatments exhibited isotopic signatures similar to their +C 2 H 2 counterparts. No N 2 O accumulation occurred in the abiotic control (filtered bottom-water without Fe(II)). In sediment incubations with C 2 H 2 , the bacterial and control treatments showed low SP values (8.2 ± 0.6‰ and 8.5 ± 1.4‰) and elevated δ 18 O values (51.6 ± 1.2‰ and 51.4 ± 1.9‰), consistent with N₂O production via bacterial denitrification. In contrast, the chemo and fungal treatments exhibited higher SP values (26.8 ± 2.5‰ and 27.5 ± 1.9‰) and slightly lower δ 18 O (45.5 ± 5.6‰ and 45.0 ± 3.2‰) ( Fig. 2a ), suggesting distinct production pathways from the bacterial treatment. Without C 2 H 2 , the isotopic composition of N 2 O shifted toward higher SP and δ 18 O values in all sediment treatments ( Fig. 2b ). The SP and δ 18 O values of the chemo and fungal treatments showed relatively modest increases with respect to the corresponding +C 2 H 2 treatments (∼3‰ for SP and ∼14‰ for d 18 O), while the control and bacterial treatments exhibited larger shifts (∼6‰ for SP and ∼42‰ for δ 18 O). 3.4 N 2 O source contributions determined by FRAME model N 2 O source partitioning using the FRAME model, based on three isotopic parameters (SP, δ 18 O and δ 15 N bulk ), aligned well with the results from the 15 N-incubation experiments ( Table 1 and 2). The model identified incomplete bacterial denitrification as the dominant N 2 O production pathway in both control treatments ( Table 2 ). Smaller contributions were attributed to chemo- and fungal denitrification both in the bottom-water and sediment ( Table 2 ). View this table: View inline View popup Download powerpoint Table 2. Contribution of different denitrification pathways to N 2 O accumulation in NA BW/Sed-Control+C 2 H 2 treatments and in situ bottom-water, as estimated by the FRAME model. We also assessed N₂O accumulation in bottom-water in situ (August sampling). FRAME- model estimates indicated similar contributions from all three pathways, though with high associated uncertainties due to model limitations (discussed in section 4.2 ) ( Table 2 ). 4 DISCUSSION 4.1 N 2 O accumulation in bottom-waters originates mostly from sediments Extremely high N 2 O accumulation (up to 3 µM) was observed in the anoxic bottom-waters of the SB, where high concentrations have been measured previously during summer stratification (up to 900 nM 31 ). These N 2 O levels far exceed peak N 2 O concentrations typically observed in other lacustrine 18 , 54 , 55 or marine systems. 14 , 56 , 57 Sediment incubation experiments revealed that volumetric N 2 O production rates and accumulation over six days were more than an order of magnitude higher than those in bottom- water incubations (Fig. S1). This strongly suggests that most of the N 2 O in the bottom-waters originates from the sediments rather than being produced in situ within the water column. It should be noted that NO 3 - in the in situ sediments is depleted within the first few millimetres. Hence, the sediment volume capable of sustaining reductive N 2 O production via denitrification is much smaller compared to the anoxic water volume. Nonetheless, high benthic N 2 O fluxes, around 1115 nmol N 2 O h -1 m -2 , reported in previous research at the same site during summer stratification, support a sedimentary origin for the N 2 O accumulating in the bottom water. 32 Moreover, benthic N 2 O fluxes were found to increase as stratification progresses (up to 2605 nmol N 2 O h -1 m -2 during late stage of the stratification period), whereas fluxes were much lower during the mixing period (140 to 293 nmol N 2 O h -1 m -2 ). 32 The high benthic N 2 O fluxes over the stratification period result in an increase in bottom-water N 2 O concentrations from June (920 nM) to August (3032 nM). Despite the highly elevated N 2 O concentrations in near-bottom-waters, N 2 O remains undersaturated immediately below the oxic-anoxic interface, indicating that active, complete denitrification prevents N 2 O from evading into the atmosphere during the stratification. Thus, the anoxic waters below the oxic- anoxic interface function as a biogeochemical filter, impeding upward diffusion of N 2 O from sediments into the upper water column. Yet, at least at this time during the seasonal cycle, N 2 O reduction rates in the anoxic water layer are insufficient to fully counterbalance benthic N 2 O fluxes, leading to persistent N 2 O accumulation. At the end of the stratification period, this stored N 2 O could be transported into the upper water column during winter mixing, potentially rendering the SB at least transiently a significant source of N 2 O to the atmosphere. The fate of accumulated bottom-water N 2 O however remains to be further investigated. Nonetheless, these findings highlight the potential for lakes to act as pulsed sources of greenhouse gases, with implications for regional and global N 2 O budgets under changing stratification regimes. 4.2 Multi-isotope constraints on N 2 O source attribution Source attribution for bottom-water and sediment incubations using the FRAME model showed strong agreement with results from 15 N-incubation experiments, consistently identifying incomplete bacterial denitrification as the dominant N 2 O production pathway, followed by chemo-denitrification and fungal denitrification. While FRAME effectively distinguished among these processes in the controlled incubations, its application to in situ bottom-water samples led to less conclusive insight. More specifically, in the in situ samples, FRAME inferred roughly equal contributions from the three reductive N 2 O production pathways, but with high uncertainties ( Table 2 , Fig. S8a, b). In fact, when initially only SP and δ 18 O were used as isotopic input parameters, FRAME misleadingly suggested fungal denitrification as the dominant source of N 2 O in the in situ bottom-water (Fig. S5a, b). However, results from both 15 N- and NA-incubations clearly indicated bacterial denitrification as the primary source, with fungal denitrification playing only a minor role. This obvious discrepancy prompted the refinement of the FRAME model through the inclusion of a third parameter (δ 15 N bulk ), which "improved" the source attribution output by reducing the overestimation of fungal denitrification due to the extension of the mixing space. However, including δ 15 N bulk as a parameter led to source apportionment with large standard deviations, indicating that the model failed to converge on a probable solution. This is evident from the pronounced oscillations of the Markov chain for the in situ bottom-water sample, reflecting that maximum likelihood was not achieved (Fig. S8). The high uncertainties in the in situ sample source-attribution results expose limitations of the FRAME model, particularly when isotopic signatures fall outside the mixing space, or when N 2 O reduction isotope fractionation effects are not well constrained, and strong isotopic overprinting due to N 2 O reduction is expected (see Supplementary Material, Figs. S3–S8). This highlights the need to consider multiple isotopic parameters and contextual information to assess and improve the reliability and accuracy of process identification and N 2 O source attribution (e.g., information about the study site can help exclude certain pathways, such as nitrification as a source of N 2 O due to anoxic conditions; or, data on substrate isotopic composition can better constrain the range of isotopic endmembers, and independent 15 N rate measurements can be used to verify model outputs) . Nonetheless, more precise baseline data on the isotope effects associated with different N 2 O production/consumption processes are still needed. For instance, recent work suggests that the N 2 O isotope signature of bacterial denitrification may extend to higher SP values (∼10‰). 28 Furthermore, N 2 O production via DNRA performed by bacteria belonging to the family of Geobacteraceae was reported to exhibit SP values >43‰. 8 This would clearly impact future model calibrations. Despite these challenges, the use of canonical literature values for isotopic endmembers used in the FRAME model remains a valuable tool for disentangling N 2 O production, particularly when combined with targeted incubation experiments and careful site-specific validation. 4.3 Incomplete bacterial denitrification dominates N 2 O production in bottom water and sediments Both 15 N- and NA-incubation experiments identified incomplete bacterial denitrification as the dominant N 2 O production mechanism. Bacterial N 2 O production accounted for 58 ± 15% in the bottom water and 68 ± 3% in sediments, based on direct rate measurements, and 88 ± 21% and 79 ± 9%, respectively, based on the isotope mass balance approach. In bottom-water treatments, the addition of a bacterial inhibitor (STP) was expected to suppress bacterial N 2 O production. However, the isotopic signatures (SP and δ 18 O) of the N 2 O produced remained characteristic of bacterial denitrification ( Fig. 2a ). This suggests incomplete inhibition of bacterial activity (i.e., in the NA BW-Chemo+C 2 H 2 and the NA BW-Fungal+C 2 H 2 treatments). Given that fungal and chemo-denitrification contributions were minimal in the NA BW incubations, even low residual bacterial activity could dominate the N 2 O isotopic signal. As with many inhibitor-based approaches, full inhibition is often difficult to ensure and may be accompanied by non-target effects by the inhibitors. 58 – 60 Further supporting bacterial dominance, the isotopic data for all bottom-water treatments (including the bacteria-inhibition treatments targeting fungal and chemo-denitrification) plotted within the bacterial endmember box in the SP-vs.-δ 15 N bulk space (Fig. S2a, b). Nevertheless, a closer examination reveals that chemo and fungal treatments exhibited slightly elevated SP values compared to the bacterial control (zoom-in of Fig. 2a ), suggesting limited but detectable activity of these alternative N 2 O production processes. In sediment incubations, the control and bacterial treatment had overlapping SP and δ 18 O signatures, again pointing to bacterial denitrification as the primary N 2 O source ( Fig. 2a ). Notably, SP and δ 18 O values in the bacterial treatment were higher than typically reported for bacterial denitrification. Several factors may explain this: Firstly, the range of SP values for bacterial denitrification was recently extended following the identification of a NO-detoxifying enzyme (fhp), widely present in marine and terrestrial denitrifying bacteria, which produces N 2 O with a SP of 10‰. 28 This suggests that bacterial denitrification can produce N 2 O with SP values above the traditionally accepted range of -8 to 4‰. Additionally, DNRA has been shown to produce N 2 O with SP values exceeding +43‰. 8 The contribution of DNRA to total benthic NO 3 - reduction in Lake Lugano’s South Basin was reported to be less than 12% in a study by Wenk et al. 38 . However more recent research indicated that DNRA may contribute up to ∼50% at the same study site. 39 Despite this, a high contribution of N 2 O production from DNRA is unlikely in our study, as the SP value of ∼8‰ in the control and bacterial treatments (i.e., the only treatments that permitted the occurrence of DNRA) was much lower than the SP values typically associated with DNRA. This suggests that DNRA played only a minor role, and that bacterial denitrification was the dominant process. Nonetheless, a relatively small contribution of DNRA could explain the shift in SP values to ∼8‰. in the bacterial and control treatments Elevated SP and δ 18 O values may also result from incomplete inhibition of either N 2 O reduction to N 2 or fungal denitrification. Indeed, the effectiveness of C 2 H 2 as an inhibitor of N 2 O reduction is known to vary. 49 – 51 Lastly, in the bacterial treatment, chemo-denitrification was not specifically inhibited, and its contribution could have shifted the isotopic signatures towards SP and δ 18 O values that are higher than typically associated with bacterial denitrification. Despite these complexities, the consistently high benthic N 2 O reduction rates observed in both control and bacterial treatments further support bacterial denitrification as the dominant N 2 O source in the sediment and bottom water Lake Lugano’s South Basin. 4.4 Isotopic signatures of N 2 O in the bottom-waters reflect N 2 O reduction in sediments The elevated SP values of N 2 O observed in the bottom water cannot be explained by N 2 O production via incomplete bacterial denitrification in the water column alone (i.e., they are uncharacteristic for N 2 O production by bacterial denitrification, which typically yields lower isotopic signatures). Instead, these "enriched values" are best explained by microbial N 2 O reduction occurring within the sediments, where the majority of the bottom-water N 2 O originates from, and where we demonstrated N 2 O reduction to occur concurrently ( Table 1 ). While N 2 O reduction rates in bottom-water incubations were below detection limit ( Table 1 ), sediment incubations clearly demonstrated the effect of N 2 O reduction on the N 2 O isotopic composition. In NA Sed incubations without C 2 H 2 (i.e., a N 2 O reduction inhibitor), only the control and bacterial treatments showed a strong shift toward elevated SP and δ 18 O values ( Fig. 2b ), along with significantly lower N 2 O concentrations compared to C 2 H 2 -amended treatments (Fig. S1). This directly links the observed isotopic enrichment to active N 2 O reduction. As expected, no N 2 O reduction was measured in the chemo- and fungal denitrification treatments (i.e., bacteria-inhibited conditions; Table 1 ), consistent with the understanding that bacteria are the primary N 2 O reducers. 19 , 61 Together, these results demonstrate that the enriched SP and δ 18 O values in bottom-water N 2 O are primarily driven by microbial reduction processes in the sediment, rather than reflecting distinct N 2 O production pathways alone. This highlights the importance of considering reduction dynamics when interpreting isotopic data from stratified aquatic systems. 4.5 Secondary but non-negligible role of chemo-denitrification and fungal denitrification Chemo-denitrification and fungal denitrification contributed much less to N 2 O production than bacterial denitrification in the Lake Lugano southern basin, but their roles, especially that of chemo-denitrification, were nonetheless substantial. Together, they represent non-canonical but significant pathways of N 2 O formation, which are often overlooked in biogeochemical models and source attribution studies. Chemo-denitrification accounted for approximately 20 ± 2% in sediment and 18 ± 2% in bottom water, based on 15 N tracer experiments. Even when considering the possibility of incomplete inhibition in some experiments, these values indicate a non-negligible contribution, particularly for a pathway that is not traditionally emphasized in aquatic N 2 O budgets. Previous studies have shown that chemo-denitrification can contribute 13 to 28% of total N 2 O production in estuarine sediments, 16 and up to 70% under elevated NO 3 - conditions in coastal sediments. 17 The activity of chemo-denitrification has been reported to be enhanced by the presence of Fe(II)-bearing minerals. 62 , 63 Under natural, pH-neutral conditions, the majority of aqueous Fe(II) binds to mineral surfaces or ligands, forming Fe(II)-bound phases that act both as Fe 2+ sources as well as a reactive surfaces for NO 2 - reduction. 62 In the Lake Lugano South Basin, dissolved Fe 2+ concentrations were higher than those of particulate Fe(II), and Fe concentrations increased significantly in the sediment. 41 This likely explains the higher chemo- denitrification rates observed in sediments compared to the water column ( Table 1 ). Fungal denitrification contributed modestly to N 2 O production in the sediments, with both NA- and 15 N-incubation approaches estimating a similar contribution of ∼6%. However, results for the bottom-water incubations were inconsistent. The 15 N approach resulted in a fungal N 2 O production rate of 5 ± 2 nM N/d, contributing 24 ± 9% of total N₂O production, while the isotope mass balance approach suggested a much smaller contribution (5 ± 14%). This discrepancy may stem from the overall low N₂O production in bottom-waters, where even small absolute contributions/differences can appear large in relative terms, and can therefore result in larger errors. Moreover, despite the use of a bacterial inhibitor in the 15 N-label incubations, the N 2 O produced in the fungal treatment may still reflect residual bacterial activity (i.e., incomplete inhibition). This hypothesis is supported by bottom-water NA- incubations, where, as mentioned above, SP values in both chemo- and fungal denitrification treatments were consistent with bacterial denitrification ( Fig. 2 ). A slight shift toward higher SP values in the fungal treatment compared to the bacterial treatment may indicate a somewhat greater relative (but still minor) contribution to total N 2 O production from fungal denitrifiers. Further investigations are needed to fully understand and confirm the role of fungi in N 2 O cycling in Lake Lugano, and other aquatic systems. Overall, while fungal denitrification appears to play a minor role in Lake Lugano, and is difficult to isolate with certainty, both chemo-denitrification and fungal pathways contribute to N 2 O production in the lake’s southern basin. These findings highlight the importance of accounting for non-canonical reductive N 2 O sources beyond bacterial denitrification, both in Lake Lugano and in other lake systems. Their relevance may increase under elevated NO 3 - conditions or shifting redox regimes, 16 , 17 making them increasingly significant components of N 2 O cycling in eutrophic aquatic environments. ASSOCIATED CONTENT Supporting information The Supporting information is available free of charge on the ACS Publications website at DOI: Literature values for N 2 O production of different denitrification pathways (Table S3) and for N 2 O reduction (Table S4), Isotopic values for NA-incubation experiments (Table S5), N 2 O concentrations in NA-incubation experiments (Fig. S1), dual isotope plot (SP vs δ 15 N) for NA-incubation samples, FRAME model approach description and outputs (Fig. S3-S8). Author Contributions T.E. and C.F. designed research and planned experiments; T.E. performed experiments and analysed data; all authors discussed the results; T.E. wrote the paper with contributions from all authors. Funding Sources This work was funded by the Swiss National Science Foundation (Grant number 201027). Notes The authors declare no competing financial interest ACKNOWLEDGMENTS We thank Camilla Capelli and Fabio Lepori from the University of Applied Sciences and Arts of Southern Switzerland (SUPSI) for support during sampling. We thank Thomas Kuhn and Franz Conen for help with N 2 O concentration and isotope analysis. Funder Information Declared Swiss National Science Foundation, https://ror.org/00yjd3n13 , 201027 References (1). ↵ Ravishankara , A. R. ; Daniel , J. S. ; Portmann , R. W . Nitrous Oxide (N 2 O): The Dominant Ozone-Depleting Substance Emitted in the 21st Century . Science . 2009 , 326 (5949), 123 – 125 . doi: 10.1126/science.1176985 . OpenUrl Abstract / FREE Full Text (2). ↵ Tian , H. ; Pan , N. ; Thompson , R. L. ; Canadell , J. G. ; Suntharalingam , P. ; Regnier , P. ; Davidson , E. 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