Accurate and precise quantification of atmospheric nitrate in streams draining land of various uses by using triple oxygen isotopes as tracers

Land use in a catchment area has significant impacts on nitrate eluted from the catchment, including atmospheric nitrate deposited onto the catchment area and remineralised nitrate produced within the catchment area. Although the stable isotopic compositions of nitrate eluted from a catchment can be a useful tracer to quantify the land use influences on the sources and behaviour of the nitrate, it is best to determine these for the remineralised portion of the nitrate separately from the unprocessed atmospheric nitrate to obtain a more accurate and precise quantification of the land use influences. In this study, we determined the spatial distribution and seasonal variation of stable isotopic compositions of nitrate for more than 30 streams within the same watershed, the Lake Biwa watershed in Japan, in order to use 17O excess (117O) of nitrate as an additional tracer to quantify the mole fraction of atmospheric nitrate accurately and precisely. The stable isotopic compositions, including 117O of nitrate, in precipitation (wet deposition; n= 196) sampled at the Sadoseki monitoring station were also determined for 3 years. The deposited nitrate showed large 17O excesses similar to those already reported for midlatitudes: 117O values ranged from +18.6 to +32.4 ‰ with a 3-year average of +26.3 ‰. However, nitrate in each inflow stream showed small annual average 117O values ranging from +0.5 to +3.1 ‰, which corresponds to mole fractions of unprocessed atmospheric nitrate to total nitrate from (1.8± 0.3) to (11.8± 1.8) % respectively, with an average for all inflow streams of (5.1± 0.5) %. Although the annual average 117O values tended to be smaller in accordance with the increase in annual average stream nitrate concentration from 12.7 to 106.2 μmol L−1, the absolute concentrations of unprocessed atmospheric nitrate were almost stable at (2.3± 1.1) μmol L−1 irrespective of the changes in population density and land use in each catchment area. We conclude that changes in population density and land use between each catchment area had little impact on the concentration of atmospheric nitrate and that the total nitrate concentration originated primarily from additional contributions of remineralised nitrate. By using the average stable isotopic compositions of atmospheric nitrate, we excluded the contribution of atmospheric nitrate from the determined δ15N and δ18O values of total nitrate and estimated the δ15N and δ18O values of the remineralised portion of nitrate in each stream to clarify the sources. We found that the remineralised portion of the nitrate in the streams could be explained by mixing between a natural source with values of (+4.4± 1.8) and (−2.3± 0.9) ‰ for δ15N and δ18O respectively and an anthropogenic source with values of (+9.2± 1.3) and (−2.2± 1.1) ‰ for δ15N and δ18O respecPublished by Copernicus Publications on behalf of the European Geosciences Union. 3442 U. Tsunogai et al.: Accurate and precise quantification of atmospheric nitrate tively. In addition, both the uniform absolute concentration of atmospheric nitrate and the low and uniform δ18O values of the remineralised portion of nitrate in the streams imply that in-stream removal of nitrate through assimilation or denitrification had little impact on the concentrations and stable isotopic compositions of nitrate in the streams, except for a few streams in summer with catchments of urban/suburban land uses.

Abstract.Land use in a catchment area has significant impacts on nitrate eluted from the catchment, including atmospheric nitrate deposited onto the catchment area and remineralised nitrate produced within the catchment area.Although the stable isotopic compositions of nitrate eluted from a catchment can be a useful tracer to quantify the land use influences on the sources and behaviour of the nitrate, it is best to determine these for the remineralised portion of the nitrate separately from the unprocessed atmospheric nitrate to obtain a more accurate and precise quantification of the land use influences.In this study, we determined the spatial distribution and seasonal variation of stable isotopic compositions of nitrate for more than 30 streams within the same watershed, the Lake Biwa watershed in Japan, in order to use 17 O excess ( 17 O) of nitrate as an additional tracer to quantify the mole fraction of atmospheric nitrate accurately and precisely.The stable isotopic compositions, including 17 O of nitrate, in precipitation (wet deposition; n = 196) sampled at the Sadoseki monitoring station were also determined for 3 years.The deposited nitrate showed large 17 O excesses similar to those already reported for midlatitudes: 17 O values ranged from +18.6 to +32.4 ‰ with a 3-year average of +26.3 ‰.However, nitrate in each inflow stream showed small annual average 17 O values ranging from +0.5 to +3.1 ‰, which cor-responds to mole fractions of unprocessed atmospheric nitrate to total nitrate from (1.8 ± 0.3) to (11.8 ± 1.8) % respectively, with an average for all inflow streams of (5.1 ± 0.5) %.Although the annual average 17 O values tended to be smaller in accordance with the increase in annual average stream nitrate concentration from 12.7 to 106.2 µmol L −1 , the absolute concentrations of unprocessed atmospheric nitrate were almost stable at (2.3 ± 1.1) µmol L −1 irrespective of the changes in population density and land use in each catchment area.We conclude that changes in population density and land use between each catchment area had little impact on the concentration of atmospheric nitrate and that the total nitrate concentration originated primarily from additional contributions of remineralised nitrate.By using the average stable isotopic compositions of atmospheric nitrate, we excluded the contribution of atmospheric nitrate from the determined δ 15 N and δ 18 O values of total nitrate and estimated the δ 15 N and δ 18 O values of the remineralised portion of nitrate in each stream to clarify the sources.We found that the remineralised portion of the nitrate in the streams could be explained by mixing between a natural source with values of (+4.4 ± 1.8) and (−2.3 ± 0.9) ‰ for δ 15 N and δ 18 O respectively and an anthropogenic source with values of (+9.2 ± 1.3) and (−2.2 ± 1.1) ‰ for δ 15 N and δ 18 O respec-Published by Copernicus Publications on behalf of the European Geosciences Union.
U. Tsunogai et al.: Accurate and precise quantification of atmospheric nitrate tively.In addition, both the uniform absolute concentration of atmospheric nitrate and the low and uniform δ 18 O values of the remineralised portion of nitrate in the streams imply that in-stream removal of nitrate through assimilation or denitrification had little impact on the concentrations and stable isotopic compositions of nitrate in the streams, except for a few streams in summer with catchments of urban/suburban land uses.

Nitrate (NO −
3 ) in stream water can be an important source of information for understanding the biogeochemical cycles within the catchment area of the stream (Likens et al., 1970;Durka et al., 1994;Swank et al., 2001).In addition, the nitrate concentration in stream water is important to primary production, and an excess of nitrate can lead to eutrophication in downstream areas, including receiving lakes, estuaries, and oceans (McIsaac et al., 2001;Paerl, 2009).However, nitrate concentrations in stream water are determined through a complicated interplay of several processes within the catchment area including (1) the addition of atmospheric nitrate (NO − 3 (atm)) through deposition, (2) the production of remineralised nitrate (NO − 3 (re)) through microbial nitrification, (3) the removal of nitrate through assimilation by plants and microbes, and (4) the removal of nitrate through denitrification by microbes.In addition to natural processes, anthropogenic processes can have a significant impact on the sources and dynamics of nitrate within each catchment area, particularly those with urban or agricultural catchment zones.Therefore, interpretation of the processes regulating nitrate concentration in stream water is not always straightforward.
The 15 N / 14 N and 18 O / 16 O ratios of nitrate have been widely applied worldwide in the determination of the sources and behaviours of nitrate in stream water (Durka et al., 1994;Campbell et al., 2002;Silva et al., 2002;Barnes and Raymond, 2010;Nestler et al., 2011;Lohse et al., 2013).By combining the two isotopic ratios, the relative mole fractions among various nitrate sources such as atmospheric (unprocessed), fertiliser, manure, and sewage plants can be quantified through a simple isotope mass balance approach.Partial removal of nitrate through either assimilation or denitrification, however, results in residual nitrate being enriched with 15 N and 18 O (Böttcher et al., 1990;Granger et al., 2010), which complicates the interpretation of the ratios beyond that of the simple isotope mass balance approach.In addition, trace contributions of unprocessed NO − 3 (atm) can have a significant impact on the 18 O / 16 O ratios of the total nitrate in stream water (Durka et al., 1994;Kendall, 1998;Mayer et al., 2001;Michalski et al., 2004;Tsunogai et al., 2010).Therefore, 18 O / 16 O ratios are used as tracers based on assumptions such as (1) the 18 O / 16 O ratios of nitrate in stream water simply reflect the mole fraction of unprocessed NO − 3 (atm) within total nitrate (Durka et al., 1994;Williard et al., 2001;Ohte et al., 2004;Campbell et al., 2006;Barnes et al., 2008;Burns et al., 2009;Ohte et al., 2010;Tobari et al., 2010;Thibodeau et al., 2013;Zeng and Wu, 2015), (2) the mole fractions of unprocessed NO − 3 (atm) within total nitrate are minimum for specific samples (such as soil solution samples) studied (Hales et al., 2007), and (3) the mole fractions of unprocessed NO − 3 (atm) within total nitrate are uniform in the entire samples studied (Wankel et al., 2006;Johannsen et al., 2008).To verify the reliability of these assumptions and to utilise the 18 O / 16 O ratios for quantification of the mole fractions among various nitrate sources based on the isotope mass balance approach, the mole fraction of NO − 3 (atm) within the total nitrate in stream water must be better understood based on more accurate and more precise quantification rather than on traditional quantification using the 15 N / 14 N and 18 O / 16 O ratios of nitrate.
To overcome the limitation in using the 15 N / 14 N and 18 O / 16 O ratios, the 17 O / 16 O ratios of nitrate have been used as an additional tracer of NO − 3 (atm) in stream water in recent studies (Michalski et al., 2004;Tsunogai et al., 2010;Costa et al., 2011;Dejwakh et al., 2012;Riha et al., 2014;Tsunogai et al., 2014;Rose et al., 2015).Because the oxygen atoms of NO − 3 (re) are derived from either terrestrial O 2 or H 2 O through usual chemical reactions such as nitrification, NO − 3 (re) shows mass-dependent relative variations between 17 O / 16 O and 18 O / 16 O ratios.However, only unprocessed NO − 3 (atm) displays an anomalous enrichment in 17 O from the mass-dependent relative variations, reflecting oxygen atom transfers from ozone anomalously enriched in 17 O during the conversion of NO x to NO − 3 (atm) (Michalski et al., 2003;Morin et al., 2008).By using the 17 O signature (the magnitude of 17 O excess) defined by the following equation (Miller, 2002;Kaiser et al., 2007), we can distinguish unprocessed NO − 3 (atm) ( 17 O > 0) from NO − 3 (re) ( 17 O = 0): where the constant β is 0.5279 (Miller, 2002;Kaiser et al., 2007), δ 18 O = R sample /R standard − 1, and R is the 18 O / 16 O ratio of the sample (or the 17 O / 16 O ratio in the case of δ 17 O or the 15 N / 14 N ratio in the case of δ 15 N) and each standard reference material.Please note that all the nitrate other than the unprocessed NO − 3 (atm) is classified into NO − 3 (re), including the nitrate produced through natural/anthropogenic processes in the biosphere/hydrosphere/geosphere and that stored in soil, fertiliser, manure, sewage, etc.
In addition, 17 O is stable during the mass-dependent isotope fractionation processes within surface ecosystems.Therefore, although the atmospheric δ 15 N or δ 18 O signature can be overprinted by biogeochemical processes subsequent to deposition, 17 O can be used as a robust tracer of unprocessed NO − 3 (atm) to reflect the accurate mole fraction of un- processed NO − 3 (atm) within total NO − 3 regardless of biogeochemical partial removal processes subsequent to deposition by using the following equation: where C atm and C total denote the concentrations of NO − 3 (atm) and NO − 3 in each water sample respectively, and 17 O atm and 17 O denote the 17 O values of NO − 3 (atm) and nitrate (total) in each water sample respectively.This is the primary merit of using the 17 O / 16 O ratio as an additional tracer of NO − 3 (atm).Moreover, additional measurements of the 17 O values of nitrate together with δ 15 N and δ 18 O enable us to exclude the contribution of NO − 3 (atm) in the determined δ 15 N and δ 18 O values and to estimate the corrected δ 15 N and δ 18 O values (δ 15 N re and δ 18 O re respectively) for accurate evaluation of the source and behaviour of NO − 3 (re) (Tsunogai et al., 2010(Tsunogai et al., , 2011(Tsunogai et al., , 2014;;Dejwakh et al., 2012;Liu et al., 2013;Riha et al., 2014), including NO − 3 (re) produced through anthropogenic processes.The details of the calculation are presented in Sec.2.5.
Previous studies have successfully applied the 17 O tracer to nitrate eluted from arid/semi-arid watersheds (Michalski et al., 2004;Dejwakh et al., 2012;Riha et al., 2014), forested watersheds (Tsunogai et al., 2010(Tsunogai et al., , 2014;;Rose et al., 2015), and a large river basin (Liu et al., 2013) to determine mole fractions of unprocessed NO − 3 (atm) in total nitrate more accurately and precisely than ever before, in addition to the fate of the NO − 3 (atm) that had been deposited into each watershed.However, relative changes in the fate of NO − 3 (atm) in accordance with the changes in land use of catchments have not been studied thus far by using the 17 O tracer of nitrate.
In this study, we measured the concentrations and the stable isotopic compositions of nitrate including 17 O values for more than 30 streams flowing into a lake in Japan with catchments of widely varying land uses, which includes urban, suburban, agricultural (mostly rice paddies), and forested catchments.By using the 17 O tracer, we quantified both spatial and temporal variations in the concentrations of both NO − 3 (atm) and NO − 3 (re) in streams across the land use settings accurately and precisely to gain insight into the processes controlling the sources, transport, and fate of NO − 3 (atm) and NO − 3 (re) (Fig. 1).Although NO − 3 (re) increases during nitrification within each catchment area, NO − 3 (atm) is stable during nitrification, so we were able to evaluate the progress of nitrification within each catchment area by using the changes in the concentrations of both NO − 3 (atm) and NO − 3 (re).In addition to those from the streams, we determined the stable isotopic compositions of nitrate including 17 O values in precipitation (wet deposition) for comparison to obtain accurate and precise mole fractions of both NO − 3 (atm) and NO − 3 (re) within nitrate (total) in each stream.Moreover, by using both the estimated mole fractions of NO − 3 (atm) in nitrate (total) and the δ 15 N and δ 18 O values of NO − 3 (atm), we estimated the δ 15 N re and δ 18 O re values for accurate evaluation of the source and behaviour of NO − 3 (re) in streams.Furthermore, we determined the concentrations and the stable isotopic compositions of nitrate including 17 O values in an outflow river of the same lake to evaluate the influences of flow stagnation in the lake on the concentrations of both NO − 3 (atm) and NO − 3 (re) by using the differences between inflows and outflows (Fig. 1).The results presented herein increase our understanding of the fate of NO − 3 (atm) deposited onto land, particularly the fate of that deposited on urban/suburban and forested catchments (Fig. 1).
2 Experimental section 2.1 Steam water samples Lake Biwa, located in the central part of the Japanese islands, is the largest freshwater lake in Japan (Fig. 2).It has a surface area of 670.4 km 2 , a total catchment area of 3174 km 2 , and annual precipitation of around 2000 mm.More than 120 streams flow into the lake, but the Seta River (no. 33 in Fig. 2b) at the southern end of the lake, also known as the Yodo River, is the only natural outflow.The average residence time of water in the lake is 5.5 years.
Similar to many lakes throughout the world, Lake Biwa has experienced eutrophication in the past.Urbanisation near the lake, beginning in the 1960s, particularly on the southern and eastern shores, likely caused an increase in nutrient loading.Blooms of Uroglena americana and cyanobacteria have occurred since 1977 and 1983 respectively (Hsieh et al., 2011).To clarify the pathways and sources of nitrate that www.biogeosciences.net/13/3441/2016/Biogeosciences, 13, 3441-3459, 2016  were fed into the lake, the stable isotopic compositions (δ 15 N and δ 18 O) of dissolved nitrate were determined in the major streams flowing into the lake (Ohte et al., 2010).Based on the δ 15 N values of nitrate showing positive correlation with the population densities of each catchment area, it was concluded that sewage effluent was the dominant source contributing to the increase in the δ 15 N values of nitrate.
In this study, stream water samples were collected near the mouths of 33 inflow streams and 1 outflow river (Seta River) of Lake Biwa (Table 1; Fig. 2b) during base flow periods four times in 2013: on 15 March, 17 June, 5 August, and 21 October, except for stream nos. 3 and 28 in June, which became dry arroyos at that time.The catchments of the studied inflow streams occupied 70 % of the entire Lake Biwa basin area.The streams were selected to cover those in which the concentrations and stable isotope compositions of nitrate, δ 15 N and δ 18 O, had already been determined in 2004-2006 (Ohte et al., 2010).The categories of locations classified by Ohte et al. (2010) were also used in this study to classify the location of each stream (Table 1).Either a bucket or dipper was used to collect samples as far from the bank as possible.Each sample was transferred into a dark polyethylene bottle that was pre-rinsed at least twice with the sample itself and subsequently stored in a refrigerator.Then, the samples were filtered through a pre-combusted Whatman GF/F filter with a 0.7 µm pore size within a few hours after collection, and the filtrate was stored in a different dark polyethylene bottle at 4 • C until analysis.
In this study, we defined the sampling number n, where n = 1, 2, 3, and 4, which represent the sampling in March, June, August, and October respectively.In addition, we de-fined one more hypothetical sampling number (n = 5) set just 1 year later than the n = 1 date.Please note that there are no data for sampling n = 5.Furthermore, we rated the intervals between n = 1 and n = 2, n = 2 and n = 3, n = 3 and n = 4, and n = 4 and n = 5 as spring, summer, autumn, and winter respectively for the streams in this study.

Wet deposition samples
The Sado-seki National Acid Rain Monitoring Station (38 • 14 59 N, 138 • 24 00 E) was established on Sado Island (Fig. 2a), at 110 m above sea level, as a monitoring observatory of the Acid Deposition Monitoring Network in East Asia (EANET) representing the central Japan area (EANET, 2014).Samples of wet deposition were taken at the station by using standard methods for evaluating acid deposition in Japan for the 3 Japanese financial years (FYs) from April 2009 to March 2012.An automatic wet deposition sampler (US-420, Ogasawara) was used in the collection.All of the deposition samples were stored in 1 L polyethylene bottles under refrigeration until daily recovery.After measuring the volume (i.e.precipitation rate), conductivity, and pH, the recovered samples were filtered through a 0.2 µm pore-size membrane filter (Dismic-25CS, ADVANTEC) and stored in a refrigerator until analysis.

Analysis
The concentrations of nitrate (NO − 3 ) and nitrite (NO − 2 ) in each filtrate sample were measured by ion chromatography (Prominence HIC-SP, Shimadzu, Japan) within a few days (stream water samples) and within 2 weeks (wet deposition samples) after each sampling.The error (standard error of the mean) in the determined concentrations of nitrate was ±3 %.The δ 18 O values of H 2 O in the samples were analysed using the cavity ring-down spectroscopy method by employing an L2120-i instrument (Picarro Inc., Santa Clara, CA, USA) equipped with an A0211 vaporiser and auto sampler; the error (standard error of the mean) in this method was ±0.1 ‰.Both Vienna Standard Mean Ocean Water (VSMOW) and Standard Light Antarctic Precipitation (SLAP) were used to calibrate the values to the international scale.
To determine the stable isotopic compositions, nitrate in each filtrate sample was chemically converted to N 2 O by using a method originally developed to determine the 15 N / 14 N and 18 O / 16 O ratios of seawater and freshwater nitrate (McIlvin and Altabet, 2005) and that was later modified (Tsunogai et al., 2008;Konno et al., 2010;Yamazaki et al., 2011).In brief, the procedures were as follows.Approximately 10 mL of each sample solution was pipetted into a vial with a septum cap.Then, 0.5 g of spongy cadmium was added, followed by 150 µL of a 1 M NaHCO 3 solution.The sample was then shaken for 18-24 h at a rate of 2 cycles s −1 .Then, the sample solution was decanted into a different vial with a septum cap.After purging the solution using high purity helium, 0.4 mL of the azide/acetic acid buffer was added.After 45 min, the solution was made basic by adding 0.2 mL of 6 M NaOH.
Then, the stable isotopic compositions (δ 15 N, δ 18 O, and 17 O) of N 2 O in each vial were determined by using a continuous-flow isotope ratio mass spectrometry (CF-IRMS) system at Nagoya University.The analytical procedures using the CF-IRMS system were the same as those detailed in previous research (Komatsu et al., 2008;Hirota et al., 2010).The obtained values of δ 15 N, δ 18 O, and 17 O for N 2 O derived from the nitrate in each sample were compared with those derived from our local laboratory nitrate standards that had been calibrated using the internationally distributed isotope reference materials to calibrate the values of the sample nitrate to an international scale and to correct for both the isotope fractionation during the chemical conversion to N 2 O and the progress of oxygen isotope exchange between the nitrate-derived reaction intermediate and water (ca.20 %).In this study, we adopted the internal standard method (Nakagawa et al., 2013;Tsunogai et al., 2014) for the calibrations of sample nitrate.All values in this paper are expressed relative to air (for nitrogen) and VSMOW (for oxygen).
To determine whether samples were deteriorated or contaminated during storage and whether the conversion rate from nitrate to N 2 O was sufficient, concentrations of nitrate in the samples were determined each time we analysed isotopic compositions using CF-IRMS based on the N 2 O + or O + 2 outputs.We adopted the δ 15 N, δ 18 O, or 17 O values only when concentrations measured by CF-IRMS correlated with those measured by ion chromatography just after the sampling within a difference of 10 %.About 10 % of the whole isotope analyses showed conversion efficiencies lower than this criterion.Nitrate in these samples was converted to N 2 O again and re-analysed for the stable isotopic compositions.None of the samples showed significant nitrate deterioration or nitrate contamination during storage.
We repeated the analyses of the δ 15 N, δ 18 O, and 17 O values of nitrate for each sample at least three times to attain high precision.Most of the samples had a nitrate concentration of more than 5.0 µmol L −1 , which corresponded to a nitrate quantity greater than 50 nmol in a 10 mL sample.This amount was sufficient for determining the δ 15 N, δ 18 O, and 17 O values with high precision.For cases of nitrate concentration less than 5.0 µmol L −1 , the sample volume was increased to 30 mL and the number of analyses was also increased.Thus, all isotopic data presented in this study have an error (standard error of the mean) better than ±0.2 ‰ for δ 15 N, ±0.3 ‰ for δ 18 O, and ±0.1 ‰ for 17 O.
Nitrite (NO − 2 ) in the samples interferes with the final N 2 O produced from nitrate (NO − 3 ) because the chemical method also converts NO − 2 to N 2 O (McIlvin and Altabet, 2005).Therefore, it is sometimes necessary to correct for the contribution of NO − 2 -derived N 2 O to determine the stable isotopic compositions of the sample nitrate accurately.However, in this study, more than 90 % of the samples analysed for stable isotopic compositions had NO − 2 concentrations lower than the detection limit (0.05 µ mol L −1 ).Even for the samples with NO − 2 concentrations higher than the detection limit, the NO − 2 / NO − 3 ratio was less than 1 %.Thus, in this study, the results were used with no correction.

Calculating average concentration and isotopic compositions in each stream
To clarify the chemical and isotopic characteristics of each stream, we determined both the flow-weighted annual average concentration (C total ) and the flow-weighted annual average δ 15 N, δ 18 O, and 17 O values (δ) of nitrate for each stream assuming the same flow rate, the same nitrate concentration, and the same isotopic compositions for the interval until the next observation by using Eqs.( 3), (4), and ( 5): where C n and δ n denote the concentration (C total in Eq. 2) and isotopic values (δ 15 N, δ 18 O, or 17 O) of nitrate in each stream during each observation n respectively; f n denotes the flow rate of each stream during each observation n; and t n denotes the time interval between the observation n and the next observation n + 1.When possible, we used the flow rate of each stream that was determined monthly by the Shiga Prefecture (Shiga Prefecture, 2015) for f n .For small streams with no data for flow rate (n = 13), we used a small and stable flow rate of 0.1 m 3 s −1 for f n .

Calculating δ 15 N and δ 18 O of remineralised nitrate
To exclude the contribution of NO − 3 (atm) from the δ 15 N and δ 18 O values of nitrate (total) and to clarify the sources and behaviour of NO − 3 (re) by using both δ 15 N and δ 18 O as tracers, we estimated the end-member δ 15 N and δ 18 O values of the remineralised nitrate portion, δ 15 N re and δ 18 O re , by excluding the contribution of NO − 3 (atm) in nitrate (total; Tsunogai et al., 2010Tsunogai et al., , 2011Tsunogai et al., , 2014;;Dejwakh et al., 2012;Liu et al., 2013;Riha et al., 2014) by using Eqs.( 6) and ( 7): where C atm and C total denote the concentrations of NO − 3 (atm) and nitrate (total) in each water sample respectively, and δ 15 N atm , δ 18 O atm , and 17 O atm denote the δ 15 N, δ 18 O, and 17 O values of NO − 3 (atm) in each sample respectively.The actual values of δ 15 N atm , δ 18 O atm , and 17 O atm used in this study were determined from the δ 15 N, δ 18 O, and 17 O values of nitrate in the wet deposition samples and are reported in Sect.3.1 along with the ranges of errors.Please note that the errors in the estimated values of δ 15 N re and δ 18 O re become larger in accordance with an increase in the C atm / C total ratio due to the propagation law of errors, even if the errors in the values of δ 15 N atm and δ 18 O atm are the same.

Possible variations in 17 O during the progress of partial removal and mixing
Because we used the power law shown in Eq. ( 1) for the definition of 17 O, the 17 O values are different from those based on the linear definition (Michalski et al., 2002).Please note that our 17 O values would be (0.1 ± 0.1) ‰ higher for the stream water nitrate and (0.9 ± 0.3) ‰ higher for the atmospheric nitrate if we had used the linear definition for calculation.
Compared with 17 O values based on the linear definition, 17 O values based on the power law definition are more stable during mass-dependent isotope fractionation processes, so we rated the 17 O values of nitrate as always stable irrespective of any biogeochemical partial removal processes subsequent to deposition, such as assimilation or denitrification.However, 17 O values based on the power law definition are not conserved during mixing processes between fractions with different 17 O values, so the C atm / C total ratios estimated using Eq. ( 2) are somewhat deviated from the actual C atm / C total ratios in the samples.However, in this study, the extent of the deviations of the C atm / C total ratios of the stream nitrate was less than 0.15 %, so we disregard this effect in the discussion.taken is also shown in Fig. 3c.The daily depositional flux of NO − 3 (atm) was calculated from the nitrate concentration and the daily precipitation (Table S2 in the Supplement).
The atmospheric nitrate at the Sado-seki monitoring station showed large 17 O excesses with 17 O values from +18.6 to +32.4 ‰.Moreover, a clear normal correlation between 17 O and δ 18 O was shown (r 2 = 0.878; Fig. 4).A similar trend was reported for atmospheric nitrate aerosols collected for a 1-year period in La Jolla, California (32.7 • N, 117.2 • W; Michalski et al., 2003), and similar results also have been obtained in other areas of the world (Kaiser et al., 2007;Morin et al., 2009).Michalski et al. (2003) interpreted that the linear correlation corresponds to the mixing line between tropospheric ozone and tropospheric H 2 O and thus tropospheric OH radicals, with 17 O = 0 ‰ and δ 18 O = −5 ‰.However, the NO − 3 (atm) data obtained at the Sado-seki monitoring station showed a somewhat different trend in the 17 O-δ 18 O plot between summer, from May to October, and winter, from November to April (Fig. 4).Although the line fitted to the summer data showed a slope of 2.21 ± 0.22 and an intercept of (+19.7 ± 5.1) ‰ in the 17 O-δ 18 O plot, that of the winter data showed a statistically significant larger slope of 2.89 ± 0.38 and a smaller intercept of (+3.0 ± 9.2) ‰; all errors were in the 2σ range.Although the winter data included an intercept of −5 ‰ reported by Michalski et al. (2003) as the end-member δ 18 O value of the tropospheric OH radical within the possible error range, the intercept of summer data deviated strongly from the value.Because the monitoring station is located in the Asian monsoon area, the major air mass that arrived at the station was different seasonally: Pacific air originated from south-east was dominant in summer, whereas continental air originated from the north-west was dominant in winter.The present results imply seasonal and regional changes in the δ 18 O / 17 O ratios of tropospheric ozone and the OH radical.
To apply the 17 O avg values for the 3 years obtained at the Sado-seki monitoring station as 17 O of NO − 3 (atm) deposited on the studied watershed (i.e. 17 O atm in Eq. 2), additional corrections could be needed because the 17 O value of NO − 3 (atm) is a function of the NO x oxidation channels in the atmosphere, which shift depending on the intensity of sunlight, temperature, and oxidant levels (e.g.Michalski et al., 2003;Morin et al., 2008;Kunasek et al., 2008;Alexander et al., 2009;Morin et al., 2012;Savarino et al., 2013).The latitudinal difference between the Sado-seki monitoring station (38 • 15 N, 138 • 24 E; Fig. 2) and the watershed studied (35 • 15 N, 136 • 5 E; Fig. 2) could change the intensity of sunlight and thus the NO x oxidation channel.Moreover, Tsunogai et al. (2010) reported that nitrate in polluted air masses derived directly from megacities in winter showed slightly larger 17 O values than nitrate in background air masses in the same seasons owing likely to the relative increase in the reaction via NO 3 radicals within the entire NO − 3 (atm) production channel to produce NO − 3 (atm) in the polluted air mass.The annual average 17 O value determined in this study was lowest in FY2009 when the deposition rate of nitrate was the smallest, at 19.3 mmol m −2 yr −1 (EANET, 2014.), whereas the annual average 17 O value was highest in FY2010 when the deposition rate was the largest, at 28.0 mmol m −2 yr −1 , within the 3 years of observation.These results also imply that the difference in the deposition rate of nitrate must be also corrected to apply the 17 O avg value to 17 O atm .Nevertheless, both the annual average and the seasonal variation range of 17 O correlated strongly with those determined at the Rishiri monitoring station (45 • 07 11 N, 141 • 12 33 E; Fig. 2) at +26.2 ‰ in FY2008 (Tsunogai et al., 2010), where the wet deposition rate of NO − 3 (atm) was an average of 40 % smaller than that at the Sado-seki monitoring station from 2000 to 2013 (EANET, 2014).Moreover, the values also coincided with those reported for midlatitudes, such as at La Jolla, at 33 • N (Michalski et al., 2003), and at Princeton, at 40 • N (Kaiser et al., 2007).We concluded that by allowing 1 ‰ of error, the standard deviation of the 17 O avg values determined at the four different monitoring stations located within the same midlatitude range in the past, the obtained 17 O avg value of NO − 3 (atm) can be considered representative for middle latitudes worldwide, including the Lake Biwa watershed basin.
In addition, the actual 17 O atm values of NO − 3 (atm) in each stream water sample can differ from the 17 O avg owing to the seasonal variation in the 17 O values of NO − 3 (atm).In correcting for the seasonal variation, however, it is not adequate to use the 17 O values determined for the seasons of sampling as 17 O atm in Eq. (2), because the residence time of water is longer than a few months for most of the catchments in Japan with a humid temperate climate (Takimoto et al., 1994;Kabeya et al., 2007).That is, the nitrate in base flow stream water had been stored previously in subsurface runoff and groundwater, for which seasonal 17 O changes have not been found thus far (Tsunogai et al., 2010;Nakagawa et al., 2013).We concluded that by allowing an additional 1.8 ‰ of error, the standard deviation of the 6-month moving average of atmospheric nitrate determined at the Sado monitoring station in this study, the obtained 17 O avg value of NO − 3 (atm) represented those eluted from the Lake Biwa watershed basin.
In summary, we used the 3-year average 17 O value of NO − 3 (atm) obtained at the Sado-seki monitoring station in this study ( 17 O avg = +26.3‰) as the 17 O atm in Eq. ( 2) to estimate C atm in the streams of the Lake Biwa watershed basin by allowing an error range of 3 ‰, considering the factor changes of 17 O atm from 17 O avg described above.About 65 % of all of the 17 O data of NO − 3 (atm) obtained at the Sado-seki monitoring station were included in this range of (+26.3 ± 3.0) ‰.
In the case of the δ 15 N and δ 18 O of NO − 3 (atm) in each stream water sample (i.e.δ 15 N atm and δ 18 O atm in Eqs.6 and 7), the values differed further from δ 15 N avg and δ 18 O avg owing to isotopic fractionation during partial removal subsequent to deposition.As a result, while using the 3-year average values of δ 15 N (δ 15 N avg = −4.2‰) and δ 18 O (δ 18 O avg = +79.8‰) as δ 15 N atm and δ 18 O atm , we assumed much a larger error range in the values, i.e. ±10 ‰ for both δ 15 N and δ 18 O, twice the enrichment factor during assimilation of nitrate.Because the concentration of atmospheric nitrate would be reduced to e −2 of the original value if δ 15 N and δ 18 O values increased +10 ‰ from their original values through assimilation, it might be difficult to detect atmospheric nitrate within total nitrate.Of course, this estimation is less reliable, and we further discuss the appropriateness of these error ranges in Sect.3.3.Because of the small C atm / C total ratios of stream water of generally less than 7 % (Sect.3.2), the error propagated to δ 15 N re and δ 18 O re was generally small, less than 1 and 2 ‰ respectively for most of the data presented in this study (Table S5).

Stream nitrate overview
The concentrations (C total ) and the δ 15 N, δ 18 O, and 17 O values of nitrate in the stream water samples determined for each observation (n = 1 2, 3, and 4) are presented in Fig. 5a-d.The annual average concentration (C total ) and the In this figure, each stream was plotted on the x axis in the order of location, beginning from stream no.31, which lies south-west of all of the streams (Fig. 2), and proceeding in a clockwise direction.The errors are comparable to the sizes of Although the 17 O values presented significant spatial and temporal variation from +0.0 to +6.8 ‰, the range of the 17 O values from +0.5 to +3.1 ‰ was typical for nitrate in natural stream water (Michalski et al., 2004;Tsunogai et al., 2010Tsunogai et al., , 2014;;Liu et al., 2013;Rose et al., 2015).These results correspond to mole fractions of unprocessed NO − 3 (atm) to total nitrate from (1.8 ± 0.3) to (11.8 ± 1.3) %, obtained by using Eq.(2).
By using the concentration (C total ) and the 17 O values of nitrate, the NO − 3 (atm) concentration (C atm ) was calculated using Eq. ( 2) and is plotted in Fig. 5e.In addition, the annual average concentration of NO − 3 (atm) (C atm ) in stream nitrate was calculated and is also presented in Fig. 5e as black bars.Please note that the errors in C atm and C atm are not presented in Fig. 5e (see Table S4 for the respective ranges of errors).
To verify possible secular changes (i.e.long-term nonperiodic variation), the estimated C total , δ 15 N, and δ 18 O for each stream were compared with those determined by Ohte et al. (2010), in which annual average concentration and annual average δ 15 N and δ 18 O values of nitrate (total) were determined for the same streams in 2004 to 2006.Although both concentrations and δ 15 N and δ 18 O values in the streams showed significant spatial and temporal variations during 2013, as presented in Fig. 5, the annual average values al-most correlated with the values determined in 2004 to 2006 (Fig. S1 in the Supplement).The average differences from the values determined in the streams in 2004 to 2006 were +5.3 µmol L −1 for C total , +0.6 ‰ for δ 15 N, and +1.6 ‰ for δ 18 O, whereas the standard deviation ranges of the differences were 14.9 µmol L −1 for C total , 1.6 ‰ for δ 15 N, and 2.1 ‰ for δ 18 O.That is, the differences from the values determined in 2004 to 2006 were smaller than their standard deviation ranges, so the differences were not significant.We concluded that secular changes were minimal for nitrate in the streams, at least for the most recent 10-year period of observations.

Relationship between 17 O and δ 18 O
One of the features in the spatial variation shown in Fig. 5 is the positive correlation between 17 O and δ 18 O.As is clearly represented by the relationship between 17 O and δ 18 O (Fig. 6), these values showed linear correlation with an r 2 value of 0.88.Because NO − 3 (atm) is characterised by highly elevated values of both 17 O and δ 18 O (Figs. 3 and 4), changes in the mole fraction of unprocessed NO − 3 (atm) within the total nitrate pool must be responsible for the positive correlation between 17 O and δ 18 O for nitrate in the streams.
The slope value of the least-squares-fitted line between 17 O and δ 18 O (Fig. 6) also supports this hypothesis.By extrapolating the least-squares-fitted line to the region of NO − 3 (atm) with a 17 O value of +26.3 ‰, we obtained δ 18 O = (+86 ± 7) ‰, which also corresponds with the average δ 18 O value of NO − 3 (atm) of +79.8 ‰ obtained in Sect.3.1.We concluded that the δ 18 O values also primarily reflect the mole fraction of NO − 3 (atm) within nitrate.Without 17 O data and without some assumptions, it was difficult to decide the major factor controlling the δ 18 O values of the stream nitrate.However, by adding 17 O data, as presented above, it became apparent that the changes in the mole fraction of unprocessed NO − 3 (atm) within the total nitrate pool were primarily responsible for the δ 18 O variation between the streams.These results further support our hypothesis, presented in Sect. 1, that interpretations on the stable isotopic compositions of nitrate (δ 15 N and δ 18 O) made without 17 O values can often be misleading.When using stable isotopic compositions, particularly the δ 18 O value, of nitrate in freshwater environments to trace its sources and fate, the determination of 17 O values is essential.
By extrapolating the linear correlation between 17 O and δ 18 O to 17 O = 0 ‰, we obtained the δ 18 O value of (−2.9 ± 1.2) ‰ as the average δ 18 O value of the remineralised portion of nitrate (NO − 3 (re)) in the streams.Although the δ 18 O value was substantially 18 O depleted compared with that produced through microbial nitrification in soil during in vitro incubation experiments in past studies (Mayer et al., 2001;Burns and Kendall, 2002;Spoelstra et al., 2007), it correlated strongly with the δ 18 O values of NO − 3 (re) determined recently by using the linear relationship between 17 O and δ 18 O of nitrate eluted from forested watersheds, such as NO − 3 (re) in the groundwater of cooltemperate forested watersheds at (−4.2 ± 2.4) ‰, where the δ 18 O(H 2 O) was around −13 ‰ (Tsunogai et al., 2010), and NO − 3 (re) in stream water in a cool-temperate forested watershed at (−3.6 ± 0.7) ‰, where the δ 18 O(H 2 O) was around −11 ‰ (Tsunogai et al., 2014).Moreover, the δ 18 O value of NO − 3 (re) obtained in this study, (−2.9 ± 1.2) ‰, is close to the possible lowermost δ 18 O value of NO − 3 (re) (Buchwald et al., 2012) produced through microbial nitrification under H 2 O of (−7.8 ± 1.0) ‰ (the average and the standard deviation of the δ 18 O values of H 2 O of the streams; Table S3).Furthermore, the δ 18 O value of NO − 3 (re) correlates strongly with that obtained through in vitro incubation experiments in recent studies that simulated temperate forest soils (Fang et al., 2012).We concluded that the δ 18 O value of NO − 3 (re) produced through nitrification in the temperate watershed with a δ 18 O(H 2 O) value of (−7.8 ± 1.0) ‰ was (−2.9 ± 1.2) ‰ and that we should use such a low δ 18 O value for the NO − 3 (re) produced through nitrification in the watershed.Understanding the relationship between 17 O and δ 18 O of nitrate shown in Fig. 6 is highly useful for determining the δ 18 O value of NO − 3 (re) in each watershed (Tsunogai et al., 2010).
Although the 17 O values of nitrate were stable during partial biogeochemical processing, such as partial removal through assimilation or denitrification, the δ 18 O values of nitrate could vary through the isotopic fractionation processes within each catchment area.Nevertheless, the δ 18 O values of nitrate in the streams plotted on the mixing line between the NO − 3 (atm) that had been deposited in the watershed and NO − 3 (re) with δ 18 O and 17 O values close to those produced through nitrification in the catchments.Thus, we concluded that the range of isotopic fractionations owing to partial removal through assimilation or denitrification subsequent to deposition of NO − 3 (atm) or production of NO − 3 (re) within each catchment area was generally small for the major portion of nitrate eluted from the watershed.If such isotopic fractionations were significant for the portion of NO − 3 (atm) in total nitrate, the data should plot on the 18 O-enriched side of Fig. 6, especially for those data enriched in NO − 3 (atm) (i.e.those showing high 17 O values).This result also supports our assumption in Sect.3.1 that the actual δ 15 N and δ 18 O values of NO − 3 (atm) in each stream water sample (δ 15 N atm and δ 18 O atm in Eqs.6 and 7) correlate with the δ 15 N avg and δ 18 O avg estimated at the Sado-seki monitoring station within an error of ±10 ‰.

δ 15 N values of remineralised nitrate in streams
To trace the source of the 18 O-depleted NO − 3 (re) eluted from the watershed into the lake, the annual average δ 15 N and δ 18 O values of the remineralised portion of nitrate (δ 15 N re and δ 18 O re ) in each inflow stream were estimated using Eqs.( 6) and ( 7) and are plotted as a function of population density in Fig. 7c and d respectively.The original δ 18 O is also in Fig. 7d.Because of the large δ 18 O differences of about 80 ‰ between the nitrate in streams and the NO − 3 (atm), the δ 18 O re values were a few ‰ lower than each original δ 18 O value in total nitrate (Fig. 7d).On the contrary, because of the small δ 15 N differences of less than 15 ‰ between the total nitrate in streams and NO − 3 (atm), as well as the small C atm / C total ratios in the streams, most of the δ 15 N re values showed small deviations of less than 1 ‰ from each corresponding original δ 15 N value in the total nitrate in most of the streams, so the original δ 15 N values are not presented in Fig. 7c.Although the annual average δ 18 O re values were low and almost uniform, from −4.0 to −0.1 ‰, as implied in the linear correlation between 17 O and δ 18 O in Fig. 6, δ 15 N re showed larger variation from +1.7 to +10.9 ‰.
Moreover, δ 15 N re showed positive linear correlation with the population density in logarithmic scale (r 2 = 0.64, p < 0.001; Fig. 7c).A similar trend was reported for the δ 15 N values of total nitrate (= NO − 3 (atm) + NO − 3 (re)) in past studies in this watershed (Ohte et al., 2010) and others (Mayer et al., 2002).We further verified that the remineralised portion of nitrate (NO − 3 (re)) was responsible for the positive correlation between the δ 15 N values of total nitrate and population density that has been found often in various streams in the world.
Both the concentrations and the isotopic compositions shown in Fig. 7 clearly demonstrate that most portions of the nitrate eluted from the catchments with lower population densities of less than 100 km −2 , showing δ 15 N values of (+4.4 ± 1.8) ‰ and δ 18 O values of about (−2.3 ± 0.9) ‰, were produced through nitrification in naturally occurring soil organic matter (Kendall et al., 1995;Ohte et al., 2010).In the latter half of this section, we discuss the source of the 15 N-enriched NO − 3 (re) eluted from the catchments with higher population densities of more than 1000 km −2 , showing δ 15 N values of (+9.2 ± 1.3) ‰ or more and δ 18 O values of about (−2.2 ± 1.1) ‰.
Denitrification in riverbed sediments adjacent to riparian zones or groundwater bodies (McMahon and Böhlke, 1996) can increase the δ 15 N value of stream nitrate.However, if such post-production alternation were responsible for the 15 N enrichment of NO − 3 (re) and thus the total nitrate, the values of δ 18 O re in addition to those of δ 15 N re would be increased (Granger et al., 2008).Moreover, the absolute concentration of NO − 3 (atm) (C atm ) would decrease in accordance with the progress of denitrification, but the observed C atm was almost uniform irrespective of population density.Of course, the initial C atm could vary between the streams.However, to explain the observed uniform C atm , unrealistic assumptions are needed for C atm , such as the initial C atm being higher in accordance with higher population density.The low and uniform δ 18 O re values (Fig. 7d) as well as the uniform C atm irrespective of population density (Fig. 7b) imply that denitrification in riverbed sediments was minor for the nitrate in the streams.Rather, the NO − 3 (re) must be enriched in 15 N from its initial production through nitrification within the catchments with high population densities.In addition, the small differences in δ 18 O values of NO − 3 (re) between those values irrespective of the population densities in the catchment area (Fig. 7d) imply that the essential parameters www.biogeosciences.net/13/3441/2016/Biogeosciences, 13, 3441-3459, 2016 for determining the δ 18 O values of nitrate during nitrification, such as the δ 18 O values of H 2 O and the pH of soils (Buchwald et al., 2012;Fang et al., 2012), should be similar among them.
Based on the δ 15 N values of total nitrate eluted from catchments with high population densities, as well as the positive correlation between the δ 15 N values of total nitrate and the population densities, Ohte et al. (2010) proposed sewage effluent as the dominant source contributing to the increase in the δ 15 N values of total nitrate eluted from such catchments.The δ 15 N and δ 18 O values of NO − 3 (re) newly estimated in this study, (+9.2 ± 1.3) ‰ or more and (−2.2 ± 1.1) ‰ respectively, also imply that the dominant source contributing to the increase in the δ 15 N values of total nitrate had been produced through nitrification in which the source N of the nitrate had already been enriched in 15 N.Although the δ 15 N and δ 18 O values of total nitrate in sewage effluent determined in past studies (Aravena et al., 1993;Widory et al., 2005;Wankel et al., 2006;Xue et al., 2009) were a few per mille higher than the δ 15 N and δ 18 O values of NO − 3 (re) eluted from the high population density catchments, δ 15 N re = (+9.2± 1.3) ‰ and δ 18 O re = (−2.2± 1.1) ‰, the slight deviations in the reported δ 15 N and δ 18 O values from our results can be explained by several factors, such as (1) a slight contribution of NO − 3 (atm) and (2) the progress of denitrification subsequent to production.We concluded that sewage effluent was the most probable pollution source of nitrate to explain the observed concentrations and isotopic compositions of nitrate eluted from the catchments with high population densities, particularly for those of more than 1000 km −2 .

Seasonal variation
Although the annual average values of 17 O and δ 18 O in each stream, 17 O and δ 18 O respectively, showed linear correlation, as presented in Fig. 6, the same results were not always attained for those in each season.Particularly for those values obtained during June and August (i.e.summer), some of the streams showed significant deviations in δ 18 O of more than a few per mille from the hypothetical mixing line between NO − 3 (atm) ( 17 O = +26.3and δ 18 O = +79.8‰) and NO − 3 (re) ( 17 O = 0 and δ 18 O = −2.9‰; Fig. 8).Even though the values of 17 O and δ 18 O of NO − 3 (atm) showed seasonal variation, as presented in Figs. 3 and 4, the large deviations from the mixing line could not be explained based on the seasonal changes in NO − 3 (atm).Rather, we must assume some seasonal changes in the biogeochemical nitrogen cycles within each catchment area to explain the relationship because, unlike the 17 O values, the δ 18 O values of nitrate could vary during biogeochemical processing within each catchment area.As a result, we can evaluate the seasonal changes in the biogeochemical processing within each catchment area by using the seasonal changes in the relationship between 17 O and δ 18 O shown in Fig. 8.
The increases in the number of data deviated from the hypothetical mixing line by more than a few per mille from the hypothetical mixing line in June and August, especially those plotted on the 18 O-enriched region (i.e.vertically upward direction in the figures), imply that partial nitrate removal through assimilation or denitrification was active within each catchment area in these months.The spatial differences in the 18 O enrichment also support this hypothesis.As presented in Fig. 8 by the orange squares, 18 O enrichment was common in samples obtained from the southern streams with high population densities in their catchment areas.We can assume elevated loading of both nutrients and organic matter of anthropogenic origin in these catchments, both of which naturally enhance assimilation and denitrification.
In contrast, most samples obtained during March and October were distributed on the hypothetical mixing line between NO − 3 (atm) and NO − 3 (re), as presented in Fig. 8.We concluded that the range of isotopic fractionation subsequent to production, such as partial removal through assimilation or denitrification, in winter was generally small for the major portion of nitrate eluted from the watershed and fed into the lake.Therefore, the annual average values (i.e.δ 18 O and 17 O) of the streams distributed on the hypothetical mixing line, as shown in Fig. 6, because the nitrate influx in winter occupied a major portion of the annual nitrate influx.Active removal of nitrate from the streams through denitrifi-U.Tsunogai et al.: Accurate and precise quantification of atmospheric nitrate into the stream via storm sewers without penetrating the ground.Thus, no interaction with soils occurred, as presented in Fig. 1.Because biogeochemical interactions within soils are the major sink for NO − 3 (atm) and thus for 17 O excess of nitrate (Nakagawa et al., 2013;Tsunogai et al., 2014), the development of such drainage systems in urban/suburban areas is largely responsible for the high C atm in stream no. 21.Similar bypassing effects of NO − 3 (atm) from soil contact by paved surfaces have been suggested in urban/suburban watersheds by using δ 18 O values of nitrate as a tracer (Burns et al., 2009;Kaushal et al., 2011).We further verified that the drainage systems in urban/suburban catchments changed the fate of the NO − 3 (atm) deposited onto land to some extent.The observed uniform C atm irrespective of population density and headwater elevation shown in Fig. 7 implies that the influences of snowpacks and paved surfaces were still minor in determining the C atm values in the streams.Rather, the observed stable C atm implies that most of the NO − 3 (atm) in the streams had been stored in groundwater/subsurface runoff in the watershed, which had similar C atm concentrations, and then gushed to the surface at the respective headwater zones with various elevations and land uses, as presented in Fig. 1.
When using the δ 18 O tracer, it was difficult to determine the precise absolute concentration of NO − 3 (atm) (C atm ) in each stream water, as presented in this study, and to determine whether the absolute concentration of NO − (atm) was stable among the streams.However, by using the 17 O values, we can determine the precise C atm in each stream for each season and thus clarify the fate of NO − 3 (atm).

Differences in outflows from inflows
The concentrations and δ 15 N, δ 18 O, and 17 O values of nitrate in the outflow river (Seta River; no.33) are also presented in Fig. 5.In a manner similar to the inflow streams (i.e. by using Eqs. 3 to 5), we estimated the annual average concentration of total nitrate in the outflow river (C total ) to be 13.3 µmol L −1 , the annual average δ 15 N value (δ 15 N) to be +13.1 ‰, the annual average δ 18 O value (δ 18 O) to be +1.5 ‰, and the annual average 17 O value ( 17 O) to be +0.9 ‰, as presented in Fig. 5.Moreover, in a manner similar to that used for the inflow streams (i.e. by using Eqs.2, 6, and 7), we estimated the annual average concentration of NO − 3 (atm) in the outflow river (C atm ) to be (0.4 ± 0.1) µmol L −1 (Fig. 5e), the annual average δ 15 N re value (δ 15 N re ) to be (+13.7 ± 0.6) ‰, and the annual average δ 18 O re value (δ 18 O re ) to be (−1.2± 0.9) ‰ (Table S5).Similar to those for inflows, the 17 O values were typical for nitrate in natural stream waters.
The striking features of the outflow in comparison with the inflows were the depletions in the outflow of both C total and C atm as well as the enrichment in 15 N (Fig. 5).Because the denitrification/assimilation processes remove both nitrate and NO − 3 (atm) and preferentially consume 14 N during the re-moval, the process of denitrification/assimilation in the lake water column can be strongly responsible for the removal of both nitrate and NO − 3 (atm) and for the 15 N enrichment of nitrate in the outflow compared with the inflow.If this were the case in Lake Biwa, the total nitrate efflux would have to be smaller than the total nitrate influx.To verify this hypothesis quantitatively and to evaluate the influences of the stagnant flow in the lake on nitrate, we estimated the total influx through all of the inflow streams for nitrate and NO − 3 (atm), N in , and A in respectively and the total efflux for nitrate and NO − 3 (atm), N out , and A out respectively, as well as the flow-weighted average δ 15 N, δ 18 O, and 17 O values of all inflows and outflows to discuss their changes in the lake.
The N in and A in in each interval between the observation n and the next observation n+1 (i.e. each season) and the flow-weighted average δ 15 N, δ 18 O, and 17 O values of the inflows (δ(n)) during each interval between the observation n and the next observation n + 1 were determined by using the following equations, assuming the same flow rate, the same nitrate concentration, and the same isotopic compositions for the interval until the next observation: where Q in denotes the annual gross influx of water into the lake; C i and δ i denote the concentration and isotopic values (δ 15 N, δ 18 O, or 17 O) of nitrate in each stream i during each observation n respectively; f i denotes the flow rate of each stream i during each observation n; and t n denotes the time interval between the observation n and the next observation n + 1.
For Q in , we used the annual influx of water estimated by Kunimatsu (1995), in which influx via streams and via groundwater were included.To include the influx of nitrate via groundwater and the other minor streams not measured  al., 1996).As implied by the significant 15 N enrichment in the remineralised portion of nitrate (δ 15 N re ) in the outflow, (+13.7 ± 0.6) ‰, compared to the inflow, (+5.6 ± 0.3) ‰, partial removal of nitrate through either assimilation or denitrification is strongly responsible for the (8.1 ± 1.1) ‰ increase in δ 15 N re as well as the net removal of both nitrate and NO − 3 (atm) from the lake.In contrast, the δ 18 O differences in the remineralised portion of nitrate (δ 18 O re ) between the inflows and outflows were significantly smaller than those of δ 15 N re , at an annual average of only (1.6 ± 1.8) ‰ (Table 2).When the nitrate in the outflows is the residual nitrate of assimilation/denitrification in the lake, δ 18 O re should also increase (Granger et al., 2004(Granger et al., , 2008)).The much smaller δ 18 O re difference implies that nitrate supplied directly from inflows occupied a small portion of the nitrate in the outflows and that most of nitrate with high δ 15 N values in the outflows was produced through nitrification in the lake water column in which the fixed N was enriched in 15 N. Isotopic fractionations during fixed N cycling in the lake, such as during denitrification or assimilation, and the subsequent removal of 15 N-depleted organic N during sedimentation (Fig. 1) are likely responsible for the 15 N enrichment of the total fixed N. That is, most of nitrate fed into the lake via the inflows was removed at least once from the lake water column and was involved in the total fixed N cycling in the lake, in which the 15 Nenriched nitrate in the outflow was produced (Fig. 1).The stagnation of flow in the lake (around 5 years) encouraged primary production and thus the net removal of total fixed N through either denitrification or sedimentation, which resulted in 15 N enrichment of the total fixed N pool compared with that in the inflows.Further studies on N cycling in the lake are needed to verify these results.

Concluding remarks
In this study, we applied the 17 O tracer of nitrate to determine accurate and precise mole fractions of unprocessed NO − 3 (atm) within the total nitrate value for more than 30 streams in the Lake Biwa watershed basin.Although the nitrate concentration varied from 12.7 to 106.2 µmol L −1 among the inflow streams and the mole fraction of NO − 3 (atm) within the total nitrate also varied from 1.8 to 11.8 %, the absolute concentration of NO − 3 (atm) (C atm ) in each stream water was almost stable at (2.3 ± 1.1) µmol L −1 irrespective of the changes in population density and land use among the catchment areas.We concluded that changes in population density and land use among the catchment areas had little impact on C atm and that the total nitrate concentration was determined primarily by the extent of the additional NO − 3 (re) contribution, which was mostly from anthropogenic sources.When relying on only the δ 15 N and δ 18 O tracers of nitrate, it was difficult to determine the precise C atm in the stream water and whether C atm was uniform among the streams.By using the 17 O values, we were able to estimate accurate and precise C atm in each stream for each season; thus, we could clarify the fate of the NO − 3 (atm) deposited into the catchments.
Moreover, additional measurements of the 17 O values of nitrate together with δ 15 N and δ 18 O enabled us to exclude the contribution of NO − 3 (atm) from the determined δ 15 N and δ 18 O values and to use the corrected δ 15 N and δ 18 O values, δ 15 N re and δ 18 O re , to evaluate the source and behaviour of NO − 3 (re) in each stream.Based on the correction, we successfully estimated the δ 15 N and δ 18 O values of NO − 3 (re) in the streams to be (+4.4± 1.8) ‰ and (−2.3 ± 0.9) ‰ respectively for NO − 3 produced through nitrification in naturally occurring soil organic matter and (+9.2 ± 1.3) ‰ and (−2.2 ± 1.1) ‰ respectively for NO − 3 supplied from anthropogenic sources, most of which were sewage effluent.In addition, the low and uniform annual average δ 18 O re values of NO − 3 (re) in the streams implied that denitrification in the riverbed sediments was minor in the streams.
Furthermore, we clarified the seasonal changes in the range of isotopic fractionation through partial nitrate removal via assimilation or denitrification by using the relationship between 17 O and δ 18 O of nitrate in the streams.The changes were small in winter in all of the catchment areas but large in summer in some catchments.Therefore, the relationship between 17 O and δ 18 O of nitrate eluted from a catchment area is a powerful indicator for evaluating the biogeochemical nitrogen cycles within a catchment area, including the seasonal changes.
Based on the annual influx and efflux of both nitrate and NO − 3 (atm) in Lake Biwa newly estimated in this study, we found that Lake Biwa is a net sink for both nitrate and NO − 3 (atm).Additionally, we found significant 15 N enrichment ((+8.1 ± 1.1) ‰) in the remineralised portion of nitrate in the outflow compared with those in the inflows, whereas the 18 O enrichment was only (+1.6 ± 1.8) ‰.We concluded that most of the nitrate fed into the lake via the inflows was removed at least once from the lake water column and was involved in the total fixed N cycling in the lake, by which the 15 N-enriched nitrate in the outflow was produced (Fig. 1).The stagnation of flow in the lake encouraged primary production and thus net removal of total fixed N through either denitrification or sedimentation, which resulted in 15 N enrichment of the total fixed N pool compared with that of the inflows.
The Supplement related to this article is available online at doi:10.5194/bg-13-3441-2016-supplement.

Figure 1 .
Figure 1.Schematic diagram showing the biological processing of atmospheric nitrate (NO − 3 (atm)) and remineralised nitrate (NO − 3 (re)) in the watershed with catchments of various land uses and in the lake water column.

Figure 2 .
Figure 2. (a) Map showing the location of the Lake Biwa watershed basin in Japan and Sado Island, where the Sado-seki National Acid Rain Monitoring Station is located.(b) Map showing the boundary of the Lake Biwa watershed basin (dashed line) and the locations of the inflows (represented by numbers) and the outflow (Seta River, no.33) studied in this paper (modified from Ohte et al., 2010).

3
Results and discussion 3.1 Atmospheric nitrate The δ 15 N, δ 18 O, and 17 O of atmospheric nitrate NO − 3 (atm) are shown in Fig. 3a-c respectively as a function of sampling day (local time, UT + 09:00), and the daily depositional flux of NO − 3 (atm) when each of the wet deposition samples was

Figure 3 .
Figure 3. Temporal variations in the values of (a) δ 15 N, (b) δ 18 O, and (c) 17 O of nitrate in wet deposition recorded at the Sado-seki National Acid Rain Monitoring Station.The errors were smaller than the sizes of the symbols.The daily depositional flux of nitrate when each of the wet deposition samples was taken is also presented (c).

Figure 4 .
Figure 4. Relationship between 17 O and δ 18 O for atmospheric nitrate in wet deposition.The errors were smaller than the size of the symbols.The blue circles represent data for winter, from November to March, whereas the orange squares represent data for summer, from April to October.The blue solid and orange dashed lines represent the least-squares fitted lines to the data of winter and summer respectively.
, and the 17 O value, we estimated the monthly average 17 O value of NO − 3 (atm) ( 17 O(m)) deposited at the Sado-seki monitoring station for each month (m) from April 2009 to March 2012 by using 17

Figure 5 .
Figure 5. Distribution of (a) total concentrations, (b) δ 15 N, (c) δ 18 O, and (d) 17 O for nitrate and (e) atmospheric nitrate concentrations in inflow streams by the various station numbers and the outflow station no.33 in the Lake Biwa watershed in March (blue diamonds), June (green squares), August (red triangles), and October (orange circles) 2013 together with the annual averages for each river (black bars).The errors were smaller than the sizes of the symbols except for (e).

Figure 6 .
Figure 6.Relationship of the annual average values of 17 O and δ 18 O of NO − 3 in the inflow streams.The symbols represent the location of each river (west: yellowish green diamonds; north: green circles; east: red triangles; south: orange squares).A hypothetical mixing line between atmospheric nitrate (NO − 3 (atm)) and remineralised nitrate (NO − 3 (re)) is shown together with the end-member value of NO − 3 (re) (large white square) and the 10 and 15 % mole fractions of NO − 3 (atm) on the line.The errors were smaller than the sizes of the symbols.

Figure 7 .
Figure 7. Annual average concentration of (a) nitrate (C total ) and (b) NO − 3 (atm) (C atm ) in each inflow stream plotted as a function of the population density in each catchment, together with (c) the annual average value of δ 15 N and (d) that of δ 18 O for remineralised NO − 3 (NO − 3 (re)).The annual average δ 18 O values of nitrate (total) are also presented in (d).The symbols represent the location of each river (west: diamonds; north: circles; east: triangles; south: squares).The errors larger than the sizes of the symbols are presented by error bars.Some of the streams in (b) are indicated by number.

Figure 8 .
Figure 8. Temporal changes in the relationship between the values of 17 O and δ 18 O of the total nitrate in the stream water.The symbols are the same as those in Fig. 6.A hypothetical mixing line between atmospheric nitrate (NO − 3 (atm)) and remineralised nitrate (NO − 3 (re)) is also shown together with the end-member value of NO − 3 (re) (large white square).The errors were smaller than the sizes of the symbols.

Table 1 .
List of studied streams.Dominant land use in the respective catchment of agricultural land (Agr), forest, or residential (Res).See TableS1for the specific contents.

Table 2 .
Estimated gross influx/efflux of total nitrate ( N) and atmospheric nitrate ( A) via inflows/outflows during each observation interval, together with the average δ 15 N, δ 18 O, and 17 O values of total nitrate and remineralised portions of nitrate (δ 15 N re and δ 18 O re ) in the inflows/outflows during each interval.