Sources and cycling of nitrogen in a New England river discerned from nitrate isotope ratios

Coastal waters globally are increasingly impacted due to the anthropogenic loading of nitrogen (N) from the watershed. In order to assess dominant sources of N contributing to the eutrophication of the Little Narragansett Bay estuary in New England, we carried out an annual study of N loading from the Pawcatuck River. We conducted weekly monitoring of nutrients and nitrate (NO 3- ) isotope ratios 15 ( 15 N/ 14 N, 18 O/ 16 O and 17 O/ 16 O) at the mouth of the river and from the larger of two Waste Water Treatment Facilities (WWTFs) along the estuary, as well as seasonal along-river surveys. Our observations reveal a direct relationship between N loading and the magnitude of river discharge, and a consequent seasonality to N loading into the estuary – rendering loading from the WWTFs and from an industrial site upriver more important at lower river flows during warmer months, comprising 20 ~ 23 % and ~ 18 % of N loading, respectively. Riverine nutrients derived predominantly from deeper groundwater and the industrial point source upriver during low base flow in summer, and from shallower groundwater and surface flow at higher river flows during colder months. Loading of dissolved organic nitrogen appeared to increase with river discharge, ostensibly delivered by surface water. The NO 3- associated with deeper groundwater had higher 15 N/ 14 N ratios than shallower 25 groundwater, consistent with the expectation fractionation due to partial denitrification. Along-river, NO 3- 15 N/ 14 N ratios showed a correspondence to regional land use, increasing from agricultural and forested catchments to the more urbanized watershed downriver, with the agricultural and urbanized portions of the watershed contributing disproportionately to total N loading. Corresponding NO 3- 18 O/ 16 O ratios were lower during the warm season, a dynamic that we ascribe to 30 increased biological cycling in-river. The 18 O/ 16 O isotope ratios along-river were consistent with the notion of nutrient spiraling, reflecting NO 3- input from the watershed and in-river nitrification and its coincident removal by biological consumption. Uncycled atmospheric NO 3- , detected from its unique mass-independent NO 3- 17 O/ 16 O vs. 18 O/ 16 O fractionation, accounted for < 3 % of riverine NO 3- , even at elevated discharge. We explore the implications of our findings for the mitigation of eutrophication 35 in Little Narragansett Bay. watershed is partially attenuated through biological cycling in soils and aquifers. Specifically, organic N is degraded to reduced N species that are oxidized (nitrified) to nitrate (NO 3- ) in oxygenated zones of groundwater. NO 3- is otherwise removed from anoxic groundwater by denitrification, reduced to inert N 2 . Reactive N is further cycled and attenuated in-river: The 65 hyporheic zone, where groundwater interchanges with stream and river water, creates a complex the seasonal and flow-dependent nature of N cycling within a riverine system transitioning to an estuarine system. This last finding has direct relevance to water quality modeling efforts in 120 temperate estuaries.


Introduction
Human activities have resulted in a substantial increase in the delivery of nutrients from terrestrial to aquatic and marine systems (Gruber and Galloway, 2008). In marine systems, increased loading of reactive nitrogen (N) has resulted in coastal eutrophication, engendering the loss of 40 valuable nearshore habitat such as seagrass beds and oyster reefs, depletion of dissolved oxygen (creating so-called "dead zones"), and increased frequency and severity of algal blooms -including toxic brown and red tides causing fish kills (Heisler et al., 2008). In densely populated areas like the northeast United States, excess anthropogenic nitrogen loads originate from Waste Water Treatment Facilities (WWTFs), septic systems, industrial discharge, fertilizer applied to turf and agricultural lands, 45 and atmospheric sources from industry and fossil fuel use McClelland et al., 2003, Latimer andCharpentier, 2010). The pervasive degradation of coastal marine ecosystems is alarming and of significant concern to coastal communities worldwide.
The transfer of nutrients from land to the coast is facilitated by rivers, which constitute an effective pipeline that collects nutrients from the watershed, ultimately discharging these to the 50 coast. The mitigation of estuarine eutrophication thus relies on identifying primary sources of nutrients to riverine systems. Nutrients are fundamentally delivered to rivers from non-point sources: from waters entering the river via surface runoff, sub-surface groundwater in the unsaturated zone, and groundwater within the water table. Nutrients also enter rivers from point sources, including WWTFs as well as industrial discharge, which can dominate N loading in urbanized watersheds 55 (Howarth et al., 1996). The nutrient loads contained in surface and deeper groundwater entering rivers differ markedly depending on land use. In temperate pristine systems, soil and groundwater concentrations are generally low, with reactive N originating from atmospheric deposition, biological N2 fixation in soils, and from N in rocks and minerals (Hendry et al. 1984;Holloway et al. 1998;Morford et al., 2016). Higher concentrations of reactive N are found in waters draining agricultural 60 and urbanized areas (Dubrovsky et al., 2010;Baron et al., 2013).
The N loaded to the watershed is partially attenuated through biological cycling in soils and aquifers. Specifically, organic N is degraded to reduced N species that are oxidized (nitrified) to nitrate (NO3 -) in oxygenated zones of groundwater. NO3is otherwise removed from anoxic groundwater by denitrification, reduced to inert N2. Reactive N is further cycled and attenuated in-river: . Riverine δ 18 ONO3, in turn, integrates across values of exogenous NO3delivered to the river from the watershed and from atmospheric deposition, those of NO3produced in-river by nitrification, minus the NO3lost concurrently to denitrification and assimilation (see Sigman et al. 2019). Interpreted in tandem, NO3 -N and O isotopologue ratios thus offer complementary constraints 100 to identify important source terms and characterize inherent cycling.
Here we present a study of annual N loading from the Pawcatuck River to the Little Narragansett Bay in southern New England (U.S.A.), wherein we exploit measurements of the N and O isotope ratios of riverine NO3to draw inferences on dominant N sources from the watershed and on riverine N cycling. The site is heavily impacted by nitrogen loading as evidenced by the history of the habitat: 105 Vast seagrass beds of Zostera marina (eelgrass) historically established in Little Narragansett Bay were overtaken in the early 1990's by extensive mats of filamentous macroalgae dominated by the Cladophoraceae clade, whose substantial biomass has been linked to frequent events of night-time hypoxia in the bay's shallow-water coves (Dodds and Gudder. 1992;Dillingham et al., 1993;D'avanzo and Kremer 1994;Tiner et al., 2003;Berezina and Golubkov, 2008;National Water Quality Monitoring 110 Council, 2020). The evident eutrophication of the estuary has raised questions regarding the magnitude of N loading from the Pawcatuck River and from the local WWTFs, whose respective contributions must be assessed in order to devise targets for mitigation. To this end, we conducted weekly monitoring of nutrients and NO3isotopologue ratios at the mouth of the Pawcatuck River and of nutrients discharged from the larger of two WWTFs along the estuary, as well as parallel 115 measurements of samples collected from seasonal along-river surveys. Utilizing NO3isotopologue ratios to identify N sources has immediate local implications for management of the watershed, allows for extrapolation to similar watersheds throughout the temperate zone, and most importantly, isolates the seasonal and flow-dependent nature of N cycling within a riverine system transitioning to an estuarine system. This last finding has direct relevance to water quality modeling efforts in 120 temperate estuaries.
Wakefield, RI, and extends 47 km southwest to Westerly, RI. It is joined by the Wood River, which originates in northern RI and runs 29 km south to Wood River Junction. The drainage basin is mostly in RI from development (Dillingham et al. 1992;U.S. Geological Survey, 2011). Agricultural areas comprise 8 % of land use (U.S. Geological Survey, 2011)

Sample collection
We conducted four distinct sampling regimens: (a) weekly river samplings at the mouth of the river, (b) weekly WWTF effluent samplings, (c) seasonal along-river surveys, and (d) rainwater 145 samplings. We collected weekly river samples (a) from January 10, 2018 through to January 12, 2019 at two sites: The Stillman Bridge near the mouth of the freshwater portion of the river, and ~1 km downstream at the Westerly Bridge, which marks the limit of seawater intrusion (Figure 1 (approximately 18 km west of the Pawcatuck River), from September 6, 2018 to December 2, 2018, in order to define regional NO3isotopic endmembers.
Weekly samplings at the Stillman and Westerly bridges occurred around sunrise, before the onset of photosynthetic activity, whereas along-river samples were collected sequentially from sunrise to mid-day. During each sample collection, river temperature and dissolved oxygen concentrations were 160 measured in situ with a Thermo Orion Star A123 portable dissolved oxygen meter. At each site, river water was collected at ~0.5 m depth with a Van Dorn bottle and transferred into a 5 L carboy for transport, on ice, back to the laboratory for processing. In the laboratory, the conductivity of each sample was measured with an Oakton CON 450 conductivity meter. Sub-samples for analyses of dissolved nutrient and NO3isotope ratios were filtered through pre-combusted 25 mm GF/F glass 165 fiber filters and collected in acid washed polypropylene bottles, then stored at -20˚C pending analysis.
The filters were placed in pre-combusted aluminum foil and frozen at -20˚C in preparation for particulate nitrogen isotope ratio analyses. Samples for chlorophyll-a analysis were similarly collected onto 25 mm GF/F filters.
The weekly effluent samples at the Westerly WWTF were collected by facility personnel into 0.5 170 L acid-washed polypropylene bottles and frozen pending monthly pick-ups by our team. Two types of samples were collected on a weekly basis: grab and composite samples. Grab samples correspond to treated effluent collected prior to its release to the river, while composite samples are effluent collected continually over a 24-hour period, thus providing a concentration-weighted daily average.
In the laboratory, samples for nutrient analysis were thawed and filtered through a 25 mm GF/F filter 175 and frozen at -20 ˚C pending analysis. Samples for particulate N analysis were not collected from the WWTF.
Rainwater samples were collected into trace-metal-clean 1-L Teflon bottles outfitted with a glass funnel to create a vapor lock preventing evaporation. These samples were stored unfiltered at -20˚C pending nutrient and NO3isotope ratio analyses. 180
The concentration of total dissolved nitrogen, [TDN], in filtered river and WWTF samples was 195 measured by persulfate oxidation to NO3 -, then measured via chemiluminescent NOx analyzer as described above (Sólorzano and Sharp, 1980;Knapp et al, 2005). The persulfate reagent was first recrystallized following protocol by Grasshoff et al. (1999). A ratio of sample to reagent of 5 to 10 was used in the oxidations. Reagent blanks accounted for ≤ 0.

Chlorophyll-a analyses
Chlorophyll-a was extracted from duplicate 25 mm GF/F filter samples in 5 mL of 90 % acetone, incubated overnight at -20˚C and quantified by florescence detection on a Turner Trilogy Laboratory Fluorometer (Arar and Collins, 1997 (Sigman et al. 2001;Casciotti et al, 2002;Kaiser et al. 2007). Briefly, NO3was converted quantitatively to a nitrous oxide (N2O) analyte by denitrifying bacteria that lack a terminal reductase (Pseudomonas chlororaphis f. sp. aureofaciens; 210 ATCC® 13985™), followed by analysis of the N2O product at the University of Connecticut on a Thermo Delta V GC-IRMS prefaced with a custom-modified Gas Bench II device with two cold traps and a PAL autosampler (Casciotti et al., 2002). The NO3 -17 O/ 16 O in rainwater (as well as 18 O/ 16 O) was similarly analyzed by bacterial conversion to N2O, followed by pyrolysis in a gold tube to N2 and O2 and analysis on a Thermo Delta V GC-IRMS at Brown University (Kaiser et al. 2007 Thiemens (1999): The analytical reproducibility for ∆ 17 ONO3 averaged 0.3‰ based upon the pooled standard deviation 225 of repeated measures of reference materials. The fraction (%) of atmospheric NO3in river water was derived from a two-end-member mixing equation of river water NO3 -(∆ 17 O = 0) with the corresponding atmospheric NO3 -∆ 17 O value (19.7 to 27.2 ‰; Section S1), with an associated uncertainty of ~1 % based on the pooled standard deviations of Monte Carlo error propagations.

Particulate nitrogen analyses 230
Particulate nitrogen (PN) in river samples was collected on pre-combusted 25 mm GF/F glass fiber filters that were freeze dried then compacted into tin capsules for analysis on a Costech Elemental Analyzer connected to a Thermo Delta V isotope ratio mass spectrometer via a Conflow IV interface.

Nutrient flux estimates
Instantaneous nutrient fluxes were estimated from the product of the nutrient concentration and the corresponding mean daily river discharge recorded by USGS gauges, or discharge reported by the Westerly WWTF. Given the large number of available river flow data from the USGS gauge at the Stillman Bridge and effluent discharge from the WWTF in relation to comparatively fewer 240 concentration data, annual fluxes of respective constituents were calculated using Beale's ratio estimator (Beale 1962;Quiblé et al. 2006), which accounts for the covariance between load and river The term µq is the mean of all river discharge measurements, Ci is the concentration on day i, Qi the average river discharge on day i, n the total number of days for the period of load estimation and nd is the number of observations of Ci. Overbars denote sample arithmetic means, and L is the resulting load. 250

Weekly river samplings
The concentration of NO3measured in samples collected weekly at the Stillman Bridge was lowest in winter and highest in the summer months, ranging from to 9.7 µM to as high as 73.  Table 1). The [NO3 -] at the Stillman Bridge, upstream of potential 260 seawater intrusion, also correlated directly with conductivity ( Figure 3a). Values of δ 15 NNO3 were lowest in winter and increased in summer, ranging from 5.3 ‰ to 9.4 ‰ -thus decreasing with increasing river discharge (Figure 2c-d; Table 1). Values of δ 18 ONO3 followed a contrasting trend, being lower during the summer months and increasing in winter months, with values ranging from 1.6 ‰ to 6.8 ‰, barring a single an outlying value of 8.1 ‰ (Figure 2e). Values of δ 18 ONO3 at the bridges 265 increased directly with discharge ( Figure 2f; Table 1). Measurements of ∆ 17 ONO3 at the Stillman Bridge ranged from -0.5 to 1.9 ‰. Uncycled atmospheric NO3was not detected in the majority of the river samples analyzed, with only 10 of 41 samples showing values above our lower limit of detection of ~1 % atmospheric NO3 -. The fraction of atmospheric NO3was otherwise < 3%, notwithstanding a single sample in which atmospheric NO3accounting for ~7 % of total riverine NO3 - (Figure 2g, S2; 270 Section S1). Values of ∆ 17 ONO3 nevertheless correlated with river discharge (Figure 2h; Table 1).
https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. Figure 2. Weekly measurements of solute concentrations and NO3isotopologue ratios at the Stillman and Westerly bridges vs. the sampling date (superposed onto river discharge), and vs. the mean daily river discharge recorded at the Stillman Bridge. The secondary axis on left-hand panels is the river discharge (x 10 6 m 3 d -1  Concentrations of PO4 3appeared to correlate inversely with discharge, yet only at the Westerly Bridge but not the Stillman Bridge ( Figure 2l; Table 1).
The concentration of DON at the bridge sites ranged from 9 to 56 µM, appeared similar among 280 seasons, and did not show a statistically significant relationship to river discharge (Figure 2m-n; Table   1). Nevertheless, [DON] and coincident [DIN] were inversely correlated, albeit weakly so, and significantly so only at the Stillman Bridge ( Figure 3b; Table 1 Table 1). 285 Concentrations of chlorophyll-a, which we measured only from June through December, ranged from 0.5 µg L -1 to 12.1 µg L -1 , with higher values occurring in late summer to early fall. Chlorophyll-a showed no correlation with discharge (Figure S3e-f; Table 1).
The daily riverine flux of dissolved inorganic nitrogen (DIN) delivered to the estuary from the Pawcatuck River, computed from the product of river discharge and the sum of [NO3 -] and [NH4 + ] 290 https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. recorded at the bridges, varied ~10-fold over the annual sampling period, ranging from 0.1 to 1.1 (x 10 5 ) moles of NDIN per day -omitting a single outlier of 1.8 x 10 5 moles of NDIN per day ( Figure 4a).
The riverine DIN flux increased directly with river discharge, such that it was lowest in summer, averaging 0.2 ± 0.1 (x 10 5 ) moles of NDIN per day from May through October ( Figure 4b; Table 1). The riverine DON flux, in turn, ranged from < 0.1 to 2.0 (x 10 5 ) moles of NDON per day, and also increased 295 directly with discharge (Figure 4c-d; Table 1). The total riverine N flux (TN flux), which is the sum of https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. respective DIN, DON and PN fluxes, ranged from 0.2 to 3.0 (x 10 5 ) moles of NTN per day and correlated directly with discharge (Figure 4e-f; Table 1).   (Table 1). There was an apparent 310 increase in [DON] with discharge, albeit, with high variability during high flow in winter months, whereas facility-reported [TON] did not correlate with discharge ( Figure 5f; Table 1). Our limited [PO4 3-] measurements were not significantly correlated with facility-reported discharge ( Figure 5hj; Table 1).

WWTF samples
In contrast to the riverine N fluxes, which increased with river discharge, the DIN and TON fluxes 315 from the Westerly WWTF were remarkably constant, and were substantially lower than corresponding riverine fluxes, averaging 3.2 x 10 3 moles of NDIN per day, 1.0 x 10 3 moles of NTON per day, and 4.1 x 10 3 moles of NTN per day in 2018 (Figure 4 a-f; Table S1). The daily TN loading at the Westerly WWTF was notably lower than the permitted allowable daily discharge from May through November of 13.5 x 10 4 moles of NTN per day. 320 https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. Table 1. Correlation coefficients, corresponding intercepts, coefficients of determination (r 2 ) and statistical probability of least-squared regression analyses from property-property plots of riverine solutes and fluxes. Statistically significant relationships are signaled by an asterisks (p-value ≤ 0.05*; ≤ 0.01**).

Along-river samplings
Samples collected at stations along the length of the river showed both spatial and seasonal patterns in nutrients and NO3isotope ratios ( Figure 6, S5 May 2018 and March 2019 samplings at all river sites (Tukey HSD, both p < 0.05; Figure 6b; Table S2). . Values during the March 2019 sampling ranged from 3.4 ‰ at Station 3 to 6.7 ‰ at the 345 bridges (river section 4). Values in November 2018, which were similar to those in May 2018, ranged from 5.8 ‰ at Station 3 to 8.5 ‰ at the bridges. NO3delivered by the Wood River (Station 6) had δ 15 NNO3 values similar to or greater than those of NO3originated upstream in the Pawcatuck River.
In contrast to δ 15 NNO3, δ 18 ONO3 values tended to decrease downriver (F3,9 = 8.6, p < 0.01; Figure   6d), despite relatively large variability. Relative maxima between 3.2 and 5.0 ‰ were apparent at 350 Stations 3 and 4 (river section 2), decreasing to values to values oscillating between 2.7 and 4.8 ‰ toward the bridges (F3,9 = 8.6, p < 0.01; Figure 6d). The δ 18 ONO3 values upriver were generally higher in November (in contrast to δ 15 NNO3) but otherwise occupied comparable ranges among sampling dates. Values contributed by the Wood River were similar to or marginally greater than those upstream in the Pawcatuck River on corresponding dates. 355 The concentration of NH4 + did not vary systematically across river sections (F3,9 = 3.2, p = 0.08), The term Qi is mean river discharge on day i in units of L d -1 , and [DIN]i is the corresponding 390 concentration in units of moles L -1 . Implicit in Eq. 4 is the assumption of negligible in-river N consumption, a notion supported by the low incident [PN]; we return to this dynamic further below.
The mixing relationship can serve to approximate the DIN flux from the Pawcatuck River into Little Narragansett Bay based on the river discharge recorded continually by the USGS at the Stillman Bridge. 395 https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License.
The inverse correlation of [DON] with [DIN], in turn, suggests that [DON] is transported into the river by shallow ground water and surface flow from the catchment. Shallow flow, which increases with increased precipitation, is apt to transport organic material from soils and surface plant materials (Elwood and Turner, 1989;Mulholland et al. 1990;Pabich et al. 2001). The import of DON by shallow flow is consistent with the visibly elevated concentrations of riverine tannins. In this regard, 400 the lack of direct correlation of [DON] to discharge is surprising, but may be masked by the relatively high variability of the [DON] measurements, even between replicate water samples.
Nutrient loading from the Pawcatuck River into Little Narragansett Bay was investigated previously by Fulweiler and Nixon (2005). As discerned herein, they observed an inverse relationship of [DIN] to discharge from biweekly measurements at the Stillman Bridge over an annual cycle. 405 Contrary to our interpretations, however, they argued that the decline in [DIN] with discharge was due to seasonal uptake by vegetation within the catchment, specifically during spring. They observed the lowest [DIN] in spring, corresponding to the highest discharge during their annual study period.
Here, we otherwise argue that increased water discharge dilutes the low base-flow nutrients derived from groundwater and point source discharge, such that concentrations are most elevated at low 410 base flow. While the concentration is lower during periods of high river flow, the riverine DIN flux nevertheless increases with discharge, carrying nutrients imported by shallow flow.

Corroborating insights from NO3isotope ratios 435
We turn to the N and O isotope composition of NO3to further investigate relationships of nutrients with river discharge and to characterize N sources and cycling in the river. Like [NO3 -], the isotope ratios of NO3co-varied with discharge. Values of δ 15 NNO3 decreased with discharge, suggesting that (a) NO3added by shallow flow had lower δ 15 NNO3 values than low base flow NO3 -, and/or (b) δ 15 NNO3 values at low base flow increased during warmer months compared to their 440 groundwater end-member due to biological cycling in-river. Concurrently, δ 18 ONO3 values increased with discharge, suggesting that (c) NO3added by shallow flow had higher δ 18 ONO3 values than low https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License.
base flow NO3 -, and/or (d) δ 18 ONO3 values decreased in summer due to biological cycling. We consider these hypotheses in turn.

Sources of DIN in shallow flow evidenced from δ 15 NNO3 values 445
In order to evaluate whether the lower δ 15 N DIN values observed at higher discharge can be explained by the addition of relatively low δ 15 N DIN by shallow flow, we plotted the δ 15 NNO3 values Plot; Keeling, 1958Keeling, , 1961 Figure 8a). Because we lack measurements of the δ 15 N values of the 450 incident NH4 + pool (which we could not assess due to an analytical interference from dissolved organic material; see Zhang et al. 2007), we assume that the N isotope composition of NO3captures that of bulk DIN, on the basis that NH4 + imported from the catchment was largely nitrified in-river, wherein NH4 + accounted for only a small fraction of the DIN reservoir. The riverine δ 15 NNO3 data conform to a linear relationship expected for the addition of DIN with a lower mean δ 15 N to the low 455 base flow reservoir (Table 1). The intercept of the resulting linear regression suggests that the NO3associated with increased discharge had a mean δ 15 N value of ~6.7 ‰ (Table 1), compared to a low base flow value of ~8 ‰ observed at the bridges. The average δ 15 NNO3 value of atmospheric NO3in rainwater was -2.5 ± 2.1 ‰ (Section S1; Figure S2), indicating that NO3added by shallow flow did not originate predominantly from direct atmospheric deposition as uncycled atmospheric NO3 -. While the 460 δ 15 NNO3 of atmospheric NO3could conceivably be fractionated by biological cycling in-river following its import by shallow flow, increased discharge occurred largely during the cold season, at which time biological cycling in-river was presumably curtailed. Thus, we surmise that the DIN added by shallow flow did not originate from direct atmospheric deposition as uncycled atmospheric NO3 -, but rather derived from catchment soils and shallow groundwater. The δ 15 NNO3 end-member value of ~6.7 ‰ is 465 in the upper range observed for soil NO3in temperate forested catchments (Mayer et al. 2002;Barnes and Raymond, 2009)

Negligible fraction of uncycled atmospheric NO3confirmed by O isotope ratios 480
The inference that uncycled atmospheric NO3did not contribute substantially to the increased NO3flux at higher discharge is corroborated by the ∆ 17 ONO3 measurements at the Stillman Bridge.
The low values observed evidenced only a slight contribution of < 3% uncycled atmospheric NO3to total riverine NO3in a few samples, suggesting efficient processing of atmospheric NO3in soils shallow groundwater (Mengis et al, 2001;Barnes et al. 2008). This observation is further echoed in a 485 recent metanalysis of North American rivers, wherein the contribution of uncycled atmospheric NO3to base flow was inferred to be generally modest (Sebestyen et al. 2019). The NO3delivered to the Pawcatuck River by shallow flow evidently originated from a reservoir that was biologically cycled within catchment soils -and potentially in-river -thus losing its atmospheric ∆ 17 O signature.
A Keeling plot of δ 18 ONO3 values vs. the inverse of the NO3flux at the bridges suggests that NO3 -490 added by surface flow had a mean δ 18 ONO3 value of ~4.5 ‰ (Figure 8b; Table 1), compared to a mean low base flow value of 2.8 ± 0.2 ‰. Although the contribution of uncycled atmospheric NO3to the riverine reservoir was modest, we nevertheless consider that the increase in δ 18 ONO3 values with discharge may derive in part from uncycled atmospheric NO3 -, given the direct relationship of ∆ 17 ONO3 to discharge, and considering the characteristically elevated δ 18 ONO3 values of 60 -80 ‰ observed in 495 the local rainwater NO3 -. Indeed, when the weighted contribution of atmospheric NO3is subtracted from individual δ 18 ONO3 values (attributed from corresponding ∆ 17 O measurements, accounting for precipitation-dependent differences in the mean ∆ 17 O and δ 18 ONO3 values of rainwater), the intercept of the Keeling plot decreases slightly to ~3.8 ‰, nevertheless remaining greater than the δ 18 ONO3 of low base flow NO3 -( Table 1). The higher δ 18 ONO3 with higher discharge is thus partially explained by 500 the small component of uncycled atmospheric [NO3 -] with elevated δ 18 ONO3 values.
water is -7 ‰ and the δ 18 OO2 of atmospheric oxygen is ~23.5 ‰ (Kroopnick and Craig, 1972), the nitrification δ 18 ONO3 value thus expected is on the fortuitous order of 3.2 ‰. This empirical metric, however, demonstrably overlooks substantive isotope effects associated with O-atom incorporation 510 into the NO3molecule during nitrification and isotopic exchange of the nitrite intermediate with water, which otherwise give way to nitrified NO3whose δ 18 ONO3 value is close to that of ambient water (Sigman et al., 2009;Casciotti et al., 2008;Buchwald and Casciotti, 2010;Snider et al., 2010;Boshers et al., 2019). This consideration explains frequent observations of relatively low δ 18 ONO3 in some soils and saturated systems, which are not explained by simple fractional contribution of 515 reactants (Hinkle et al. 2008;Xue et al., 2009;Fang et al. 2012;Veale et al., 2019). Thus, we posit that the O isotope composition of the NO3imported into the river with increased discharge, which is typical of that in soils and shallow groundwater, does not strictly indicate that shallow flow NO3originated from proximate nitrification therein, as generally presumed, but also signals that NO3underwent partial denitrification in soils and shallow groundwater, resulting in a coupled increase in 520 its δ 15 N and δ 18 O relative to source values (Houlton et al. 2006;Granger and Wankel 2016;Boshers et al. 2019). Although increased discharge occurred largely in winter, some in-river biological cycling during colder months could additionally influence the shallow flow δ 18 ONO3 end-member, specifically reducing it from its soil value due to the nitrification of incident NH4 + . Thus, δ 18 ONO3 values imported by shallow flow, once adjusted for modest contributions of uncycled atmospheric NO3 -, fall within the 525 range typically observed in soils, potentially modified by nitrification in-river.

Values of δ 15 NNO3 at low base flow reflect groundwater DIN
The higher δ 15 NNO3 values at low base flow compared to shallow flow may derive directly from those of the ground-water end-members and point source(s); The δ 15 NNO3 values in deeper groundwater are generally higher than in shallower groundwater above, fractionated by 530 denitrification in the saturated zone. Alternatively, the higher NO3isotope ratios at low base flow may result from increased biological cycling in summer -modifying the isotope composition of low base flow NO3relative to its groundwater and/or point source values. The expectation of increased biological activity in summer months is consistent with the incident decrease in [NH4 + ] with lower discharge, which can be explained by a seasonal increase in algal assimilation and nitrification. 535 Fulweiler and Nixon (2005) similarly observed lower [NH4 + ] in the summer, but saw no correlation to https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. river discharge, further supporting our contention that increased seasonal biological cycling underlies the [NH4 + ] dynamics, rather than river discharge.
The extent to which the coincident NO3pool is also assimilated during summer months -and isotopically fractionated -is unclear. The fraction of the NO3pool assimilated by algae may be 540 modest, even in summer, on the basis that the phytoplankton biomass was relatively small due to the high tannin content of the river water, which limits light penetration. Median chlorophyll-a concentrations in summer were ~1.3 µg L -1 at the Stillman and Westerly bridges -save for late summer where higher concentrations were detected -while the median [PN] was ~2.5 µM, and no greater than 7 µM. There are, however, populations of emergent plants along some shallow reaches 545 of the river, which may assimilate NO3as well as reduced N substrates. Nevertheless, the inference that the riverine NO3pool is minimally assimilated, even in summer, appears consistent with alongriver distribution of NO3isotope ratios. If a sizeable fraction of the incident NO3pool were assimilated into biomass during summer months, both the δ 15 NNO3 and δ 18 ONO3 values of low base flow NO3would expectedly increase in proportion to the fraction of NO3assimilated (Granger et al., 550 2004;Johannsen et al., 2008). However, the δ 15 NNO3 increase along-river observed during the seasonal surveys, which could be construed as signaling partial assimilation of riverine NO3 -, was not matched by coincident along-river increases in δ 18 ONO3 values. Similarly, [PN] and chlorophyll-a did not increase along-river, as would otherwise be expected for the progressive and sizeable conversion of the NO3pool into the particulate pools ( Figure S6c-d). Thus, we rule out a dominant influence of 555 algal assimilation in fractionating the riverine NO3isotope ratios.
A more nuanced framework from which to interpret the NO3isotope ratios is afforded by the concept of riverine nutrient spiraling, namely, the continual assimilation of nutrients in the water column, the remineralization of organic material in sediments, and the return of remineralized nutrients to the water column where they can undergo assimilation into new biomass (reviewed by 560 Ensign and Doyle, 2006;Harvey et al., 2013). A small fraction of the NO3pool is likely assimilated during the growing season, resulting in the production of PN with a lower δ 15 N than coincident NO3due to N isotope fractionation during assimilation (Needoba et al. 2003; Figure 9). Considering the small summertime pools of PN and NH4 + relative to the NO3pool, δ 15 NNO3 values will be minimally fractionated by assimilation. Moreover, the concomitant recycling of PN and its subsequent 565 nitrification will ostensibly regenerate NO3with a δ 15 NNO3 value roughly equivalent to that assimilated into organic material then ammonified -given an approximate steady state between NO3 -https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. assimilation and nitrification -such that δ 15 NNO3 values will not incur a progressive increase from continual assimilation along-river. These dynamics will result in little net change in riverine δ 15 NNO3 values relative to the mean catchment end-member. 570 The NO3isotope ratios could, however, be influenced by denitrification in-river (Kellman and Hillaire-Marcel 1998; Figure 9). While direct benthic denitrification does not communicate an isotope enrichment to NO3in the overlying water column due to a reservoir effect (Brandes and Devol, 1997;Sebilo et al., 2003;Lehmann et al., 2005), δ 15 N-and δ 18 O-enriched NO3from the sediment depth of denitrification can be entrained into the water column by hyporheic flows in the riparian zone (Sebilo 575 et al., 2003). Moreover, coupled nitrification-denitrification can fractionate the N isotopologues of NH4 + in surface sediments in proportion to the corresponding fraction of nitrified NO3lost concurrently to denitrification, thus contributing to an increase in d 15 N of the water column reactive N reservoir (Brandes and Devol 1997;Granger et al, 2011).
The along-river increase in δ 15 NNO3 values could then result from isotopic fractionation by 580 sedimentary denitrification. Yet a downstream increase in δ 15 NNO3 was notably apparent in all seasons, not only in summer. On the presumption that water-column and benthic N cycling were substantially reduced during the March 2019 sampling when river waters were colder (average temperature of 5.9 ˚C), we surmise that the increase in δ 15 NNO3 values along-river arises principally from differences in the δ 15 N of DIN input from respective reaches of the catchment -although some influence of benthic 585 denitrification on riverine δ 15 NNO3 values cannot be ruled out. We thus interpret the riverine δ 15 NNO3 values to predominantly reflect the N isotope composition of DIN input from the catchment. We return to this insight in a subsequent section, to identify N sources along the catchment.
Observations of decreasing along-river values are then consistent with the notion of higher 605 catchment δ 18 ONO3 end-member values converging onto lower values determined by the ratio of nitrification to consumption in-river -and associated isotopic fractionation. Within this framework, δ 18 ONO3 values in winter, when biological cycling is dampened, would expectedly increase to values closer to the catchment sources, a prediction that appears to be borne out in our observations. Barnes et al. (2008)

Regional N sources to the Pawcatuck River
Observations from the along-river surveys provide insights into the contribution of different 615 reaches of the catchment to the riverine N reservoir. Areas of disproportionate loading can be identified from distinct concentration increases, and areas of lesser loading and/or net attenuation from concentration decreases. Reaches of the river that exhibit disproportionate loading present potential targets for mitigation. As detailed above, we interpret changes in δ 15 NNO3 values along-river to reflect differences in the δ 15 N of DIN inputs from respective reaches of the catchment, thus serving 620 to identify dominant regional N sources. Figure 9. Conceptual illustration of the influence of nutrient spiraling on the N and O isotope ratios of riverine NO3 -. Nutrient spiraling describes the cycling of nutrients as they are assimilated from the water column into biomass that is temporarily retained on the benthos, then mineralized and released back into the water column or denitrified. (a) The δ 15 N of the riverine NO3reservoir integrates the NO3and NH4 + delivered continually from groundwater (δ 15 NNO3-GW and δ 15 NNH4-GW), minus the NO3removed concurrently by sedimentary denitrificationthe δ 15 N of which depends on the sedimentary isotope fractionation communicated to the water column reservoir, 15 eD. Given the small size of the respective PN and NH4 + pools relative to NO3 -, ammonification and subsequent nitrification produce NO3with a δ 15 NNO3-N value approximating that lost concurrently to assimilation (δ 15 NNO3 -15 eA), notwithstanding the NH4 + input from groundwater. The input of groundwater NH4 + (δ 15 NNH4-GW) implicitly subsumes the input of reactive allochthonous PN and DON. (b) The riverine δ 18 ONO3 integrates the NO3input from groundwater and precipitation (δ 18 ONO3-GW) and from in-river nitrification (δ 18 ONO3-N), minus NO3lost to algal assimilation and sedimentary denitrification -whose respective values depend on the net isotope effects associated with assimilation and denitrification, 18 eA and 18 eD.  (Kreitler et al., 1978 ;Katz et al., 1999;Townsend et al., 2002;Deutsch et al, 2005). Input of uncycled atmospheric NO3by surface flow due to reginal snow melt, which could also explain lower δ 15 NNO3 values, is not supported by the corresponding  (Kendall et al. 2007;Böhlke et al., 2009;Kasper et al, 2015). Thus, changes in land use along the catchment best explain the δ 15 NNO3 increase in the lower portion of the river.
In all, the substantial difference in [DIN] between Stations 2 and 5 signals disproportionate input 660 from this section of the watershed, likely owing to the proximity of turf farms and discharge from Kenyon Industries. Indeed, the riverine DIN flux at Wood River Junction amounted to 28 ± 11 % of the DIN flux recorded at the Stillman bridge among the 3 sampling dates, while accounting for only 11 ± 2 % of the riverine discharge. A fraction of the N loaded in this portion of the river may arguably be partially attenuated by denitrification along-river; nevertheless, this regional input remains 665 substantial even assuming some biological attenuation. This portion of the river also contributed disproportionately to the riverine PO4 3burden, although we do not explicitly consider this contribution in relation to the total discharge into the estuary, given the complex geochemistry of PO4 3that involves adsorption and release from authigenic particles in sediments (Froelich, 1988).
The increase in [DIN] and δ 15 NNO3 values in the lower portion of the river, in light of the large 670 coincident river discharge, also signals a disproportionate contribution from the urbanized portion of the catchment. However, lacking estimates of river discharge at Potter Hill Dam (Station 11), we cannot deduce the fractional contribution from this portion of the watershed confidently.
Nevertheless, assuming a d 15 N input from the urbanized catchment of 10 ‰ and a mean δ 15 NNO3 of 6‰ at Potter Hill Dam, compared to 8 ‰ at the Westerly Bridge, the DIN added to the river within 675 this reach would amount to ~50 % of the total riverine N load. Otherwise assuming a d 15 N input of 15 ‰, the DIN contributed from the urbanized reach would otherwise amount to ~20 % of total.

Riverine contributions 680
Estimates of the annual N loading from the Pawcatuck River into Little Narragansett Bay for 2018, compiled from our weekly measurements at the Stillman bridge, were 20.2 x 10 6 moles yr -1 for DIN and 40.3 x 10 6 moles yr -1 for TN, albeit, with uncertainty associated with the TN loading estimate given the variability of our DON measurements (Table 2) (Table   2). This estimate is 25% lower than that for 2018, yet more than double estimate for 2002, highlighting the impact of inter-annual variability in river discharge on the annual N load.
Extrapolating DIN discharge for other years with various river flows allows for a comparison to independent estimates of nitrogen loads from the watershed. Vaudrey et al. (2017)  TN loading estimate is thus 12.1 x 10 6 moles TN yr -1 higher than the land-use model estimate, leaving 740 45 % of TN apparently unaccounted for. However, our TN measurements include both labile and nonlabile N, while the land-use model represents reactive TN and does not account for non-labile species.
On the basis that refractory humified allochthonous organic material dominates the DON pool in the Pawcatuck River, a value of 45 % of TN being non-labile could be consistent with this system, if only ~10 % of riverine DON were labile on pertinent time scales. Seitzinger et al. (1997) otherwise 745 estimated that 40 -70 % of DON from the Delaware River was reactive on pertinent time scales. The data at hand do not permit us to resolve this quandary, although characterizing the reactivity of DON from the Pawcatuck River is evidently crucial to mitigating eutrophication in the bay.
Rhode Island met an ambitious goal of a 50 % reduction in N loading to Narragansett Bay in 2012 relative to 1995-1996 loads, but the Pawcatuck River was not included in these reduction priorities. 750 This oversight is evident in the loads we currently see to the Pawcatuck River relative to loads in rivers draining to Narragansett Bay, located just east of the Pawcatuck River watershed. In the early 1980s through the early 2000s, the TN load normalized to watershed area for the Woonasquatucket, Moshassuck, Blackstone, Taunton, and Pawtuxet rivers ranged from 9.3 to 14.9 kg ha -1 yr -1 , with an average of 11.8 kg ha -1 y -1 (as reviewed in Narragansett Bay Estuary Program, 2017; Nixon et al., 1995;755 Nixon et al., 2008). Compared to this time period, the Pawcatuck River's current load of 7.4 kg N ha -1 yr -1 is relatively low. However, by the late 2000s, these riverine loads were substantially reduced to an average of 6.6 kg ha -1 yr -1 , and have continued to decline, achieving an average of 4.8 kg ha -1 yr -1 in recent N budgets developed for the 2013-2015 time period (as reviewed in Narragansett Bay Estuary Program, 2017; Krumholz, 2012). One exception to the success achieved in riverine loads to 760 Narragansett Bay is the Ten Mile River, which drains an urbanized watershed in East Providence, RI, with a 2013-2015 load estimate of 10.2 kg ha -1 yr -1 . However, the Ten Mile River watershed is relatively small (~20 % of the Pawcatuck River watershed), accounting for only 7% of the riverine load to Narragansett Bay (Narragansett Bay Estuary Program, 2017). Export from pristine temperate zones prior to human disturbances is estimated to have been on the order 1.3 kg N ha -1 yr -1 (Howarth et al. 765 1995;1996). Most Narragansett Bay rivers are moving toward this pristine condition, whereas the Pawcatuck River has shown an increase in N load over time.

Point source loading from the WWTFs and Kenyon Industries
The Westerly and Pawcatuck WWTFs downstream of the Stillman and Westerly bridges 770 accounted for a relatively modest fraction of the total annual nitrogen loading into the estuary, approximately 7 % (Table 2). This estimate does not consider loading from the catchment downstream of the Stillman and Westerly bridges, which would modestly lower the relative contributions of the WWTFs. Fulweiler and Nixon (2005) otherwise estimated that the WWTFs accounted for 18 % of annual N loading into the estuary, albeit, relying on a WWTF loading estimate 775 of 6 x 10 6 moles TN yr -1 , a flux notably higher than that of 1.45 x 10 6 moles TN yr -1 reported by Rhode

Island Department of Environmental Management and the Connecticut Department of Energy and
Environmental Protection for 2002 (Vaudrey et al., 2017). Replacing the higher WWTF load in the Fulweiler and Nixon's (2005) estimate with the lower load reported by the States yields a match to the current study, indicating that the WWTFs accounted for about 5 % of the total annual load. 780 Independent estimates by Vaudrey et al. (2017) derived from a land-use model suggest that WWTF effluents contribute ~13 % of the riverine-plus-WWTFs TN discharged to the estuary on an annual basis, but this load included only reactive nitrogen and did not estimate the non-labile fraction measured in this and Fulweiler and Nixon's (2005) studies; including an estimate of the non-labile fraction brings the annual contribution from WWTFs down to 8 % of the TN. 785 The annual N loading into the Pawcatuck River from Kenyon Industries, as monitored by the Rhode Island Department of Environmental Monitoring from 2011-2013, was 2.7 x 10 6 moles TN per year, thus accounting for 6 % of the annual riverine-plus-WWTFs loading to the estuary, an input comparable to that of the WWTFs. Loading by Kenyon Industries is notable in that it is approximately equivalent to the amount of fertilizer applied to agricultural, hay, and pasture lands throughout the 790 whole watershed (Vaudrey et al., 2017).
The overgrowth of nuisance macroalgae in Little Narragansett Bay is presumably fueled predominantly by nutrients delivered during warmer months, at which time riverine N loading is at a relative minimum (Table 2). While the fraction of TN loading to the estuary by the WWTFs was negligible during colder months (< 5 %), this proportion increased to 21 % during the warmer months 795 in 2018, from May 1st to into October 31 st . The estimated contribution from Kenyon Industries similarly increased to 16 % of total N loading during the warmer months. The influence of these point sources on algal growth during the warm season is likely to be even greater, considering that an important fraction of the total N flux from the Pawcatuck River derived from DON (38 % from May to https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. November), of which only a fraction may be bioavailable on pertinent time scales. Assuming a median 800 reactivity of riverine DON of 50 % (Seitzinger et al., 1997), the WWTFs and Kenyon industries could account for as much as 25 and 19 % of labile N loading to the estuary during the warm season, respectively, given a riverine DIN loading of 2.6 x 10 6 mol NDIN from May through October. Thus, we estimate that the WWTFs contributed between 21 -25 % of N loading to the estuary during the warm season, and Kenyon Industries contributed 16 -19 %. 805

implications for the mitigation of eutrophication
Our analysis suggests that the Pawcatuck River is strongly impacted by anthropogenic N input.
Compounding the problem, the drainage basin of the river is large relative to the receiving estuary, explaining the severe eutrophication therein. The DIN concentrations and NO3isotope ratios indicate substantive inputs of reactive N to the river from agricultural and/or point sources along the upper 810 river catchment, and from urbanized sources along the lower reach of the river. The reactive N loaded annually into Little Narragansett Bay from the Pawcatuck River is highly influenced by the amplitude of river discharge, increasing with discharge due to the additional import of reactive N by shallow flow. Loading during the warmer months in 2018 was thus substantially lower than in colder months due to lower summertime precipitation, rendering point source discharges from Kenyon Industries 815 and WWTFs more important to the total N loading to the estuary during the major growing season.
Reductions in summertime discharge by Kenyon Industries and the Westerly WWTF offer the most expeditious targets to decrease N loading into the estuary, albeit, at considerable cost. The disproportionate loading from the catchment of the upper river also begs more tempered applications of agricultural fertilizers at adjacent turf farms and expansion of riparian buffers, in order 820 to effect reductions in shallow and deeper groundwater N concentrations. In the more populated portion of the watershed, N reductions could be achieved by augmenting linkage of households to the sewer line, transitioning traditional septic systems to advanced, N-removing septic systems, and encouraging the dismantling of outdated, legacy cesspools (Amador et al. 2017;Narragansett Bay Estuary Program, 2017). Within the watershed draining directly to the estuarine portion of the 825 Pawcatuck River south of the Westerly Bridge, 90 % of the households are connected to sewer (Vaudrey et al., 2017). In the remainder of the watershed, where groundwater drains to a freshwater body (wetland, pond, river) prior to entering the estuary, only 21 % of people are on sewer. This distribution reflects the urban nature of the watershed near the coast and the more rural character https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. of the watershed further inland. Finally, restricting the use of lawn fertilizers and lessening the extent 830 of impervious surfaces in and around Westerly would further aid in reducing loading from storm water.
While reductions in N loading are necessary to mitigate eutrophication in Little Narragansett Bay, target N loads have yet to be adopted by Rhode Island or Connecticut. A TN load of 50 kg ha -1 estuary yr -1 (3.6 x 10 3 mol ha -1 estuary yr -1 ), which is generally supportive of eelgrass, has been proposed by the 835 scientific community (Hauxwell et al., 2003;Latimer and Rego, 2010). In 2018, the DIN and TN loads to Little Narragansett Bay were 37 x 10 3 and 74 x 10 3 moles N ha -1 yr -1 , respectively (given a 583 ha area of estuary downstream of the Westerly Bridge), suggesting that an astounding 10 to 20-fold reduction in N loading may be required to recover eelgrass beds. We consider that a fraction of this N load may escape the estuary directly and not be retained therein, reducing the effective annual 840 estuarine N load. Moreover, seasonality of nitrogen delivery coupled with the warm summer growing season may point the way towards targeted summer reductions that could have greater impact on the eutrophic status of the system. Regardless, immediate mitigation efforts are necessary at this junction, not purely to realize reductions in N loading, but, more soberly, to prevent further increases in N loading to the Pawcatuck River and continuing degradation of the river and estuary. 845

Conclusions
Our findings illustrate the utility NO3isotopologue ratios in differentiating among N sources, with implications for the management of N loading from of the watershed. In particular, the seasonal and flow-dependent nature of N loading and cycling uncovered herein presents important considerations for mitigation efforts. 850 Our interpretations of NO3isotopologues dynamics also move beyond the traditional source attribution framework in an effort to reconcile with current theory of riverine N biogeochemistry.
Nutrient spiraling theory offers a powerful conceptual basis to differentiate the influences of N sources vs. cycling on NO3isotopologue distributions. Continued inquiry in the context of this framework is bound to yield novel and unexpected insights on N isotopologue cycling and, more 855 fundamentally, on river biogeochemistry. https://doi.org/10.5194/bg-2020-390 Preprint. Discussion started: 4 November 2020 c Author(s) 2020. CC BY 4.0 License. will also be archived online in the PANGAEA data repository.
Author contributions. VRR and JG conceive the research question, designed the study approach, led the field survey, ensured data curation and conducted formal analysis. SC, MLB, CPK, LAT, and HCW assisted with data collection and analysis. CMM assisted with statistical analyses. CRT and HH provided use of specialized facilities. JG and JMPV secured funding for the investigation. VRR 870 and JG wrote the first draft of the paper, and all co-authors contributed to writing review and editing.
Competing interests. The authors declare that they have no conflicts of interest.