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Master's Theses University of Connecticut Graduate School

5-10-2020

Sources and Fluxes of Reactive N in a Southern New England River

Veronica Rollinson [email protected]

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Recommended Citation Rollinson, Veronica, "Sources and Fluxes of Reactive N in a Southern New England River" (2020). Master's Theses. 1510. https://opencommons.uconn.edu/gs_theses/1510

This work is brought to you for free and open access by the University of Connecticut Graduate School at OpenCommons@UConn. It has been accepted for inclusion in Master's Theses by an authorized administrator of OpenCommons@UConn. For more information, please contact [email protected]. Sources and Fluxes of Reactive N in a Southern New England River

Veronica Rose Rollinson B.S., University of Connecticut, 2012 M.A.T., Sacred Heart University, 2016

A Thesis Submitted in Partial Fulfillment of the Requirements for the Degree of Master of Science At the University of Connecticut

2020

Approval Page

Master of Science Thesis Sources and Fluxes of Reactive N in a Southern New England River

Presented by: Veronica Rose Rollinson, B.S, M.A.T

Major Advisor ______Julie Granger, Ph.D.

Associate Advisor ______Craig Tobias, Ph.D.

Associate Advisor ______Claudia Koerting, Ph.D.

Associate Advisor ______Jamie Vaudrey, Ph.D.

University of Connecticut

2020

ii Acknowledgements I would like to thank Dr. Julie Granger, my advisor – who made this whole experience a possibility for me. She not only advised me on this project but helped empower me as a young woman making my place in the world and I could not be more grateful to her as a role model and friend. I also want to thank my committee members: Jamie Vaudrey who not only helped with my ability to present science but also assisted in conducting that science, Craig Tobias whose years of mentorship has gotten me to where I am today and sparked my passion for isotopes and biogeochemistry (I will forever model my life off a box model), and Claudia Koerting whose mentorship over the past decade has propelled me down many of the paths I have been through in my life. This project was made possible through funding support from Connecticut SeaGrant with encouragement from Clean Up Sounds and Harbors (CUSH). I am grateful to the University and the Department of Marine Science for their ongoing support as well as funding provided by UConn’s Feng Student Travel Fund for assistance in sending me to ALSO and LIS Conference to present this research. I could not have done this project without the support of my colleagues including the immense help from Reide Jacksin and Claire Schlink who assisted with many of the sampling trips as well as support from the entirety of the Granger, Tobias, and Vaudrey labs. Special thanks are extended to Matt Lacerra who provided assistance in a kayak trip as well as to Sydney Clark and Meredith Hastings at Brown University for assistance in running ∆17O isotopic samples. Lastly, I would like to thank my friends, my dog, and my family for the support that they have provided for me throughout this process and my life to get me to this point. Your unconditional love has helped me though so many challenges and I can’t thank you enough.

iii Table of Contents Approval Page ...... ii Acknowledgements ...... iii Table of Contents ...... iv List of Tables and Figures ...... v Abstract ...... vii 1. Introduction ...... 1 1.1 Site Description ...... 3 1.1.1 Little Narraganset Bay ...... 3 1.1.2 Pawcatuck River and Upper Watershed ...... 3 2. Methods ...... 4 2.1 Field Samplings ...... 4 2.2 Nutrient Analyses ...... 5 2.3 Isotopic analysis ...... 6 3. Results ...... 7 3.1 Weekly samplings ...... 7 3.2 Downriver samplings ...... 10 4. Discussion ...... 12 4.1 Source Attribution ...... 12 4.2 River Cycling ...... 16 4.3 Estuary Point sources and mitigation ...... 19 References ...... 22 5. Tables and Figures ...... 28

iv List of Tables and Figures Figure 1. Map of (A) watershed land use and (B) sampling locations. Figure 2. Mean daily river discharge at three discharge gauges along the Pawcatuck River.

- Figure 3. Weekly measurements of solute concentrations and NO3 isotopologues ratios at the Stillman and Westerly bridges vs. sampling date and vs. the river discharge

- - recorded at the Stillman bridge: (A) [NO3 ] vs. sampling date, (B) [NO3 ] vs.

15 15 18 discharge, (C) δ NNO3 vs. sampling date , (D) δ NNO3 vs. discharge, (E) δ ONO3 vs.

18 17 sampling date, (F) δ ONO3 vs. discharge, (G) ∆ ONO3 vs. sampling date, (H)

17 + + ∆ ONO3 vs. discharge, (I) [NH4 ] vs. sampling date, (J) [NH4 ] vs. discharge, (K)

3- 3- [PO4 ] vs. sampling date, (L) [PO4 ] vs. discharge, (M) [DON] vs. sampling date, (N) [DON] vs. discharge, (O) [PN] vs. sampling date, (P) [PN] vs. discharge. Table 1. Correlation coefficients, corresponding intercepts and significance statistics for property-property plots of riverine solutes and fluxes. Figure 4. Plot of weekly [DON] measurements vs. corresponding [DIN] at the Stillman and Westerly bridges

- Figure 5. Solute concentrations and NO3 isotopologue ratios in rainwater collections at

- + 15 18 17 Avery Point: (A) [NO3 ], (B) [NH4 ], (C) δ NNO3, (D)δ ONO3, (E) ∆ ONO3.

- + Figure 6. Nutrients at Westerly WWTF. [NO3 ] vs. (A) date and (B) facility discharge. [NH4 ] vs. (C) date and (D) facility discharge. [DON] vs. (E) date and (F) facility discharge.

3- [PO4 ] vs. (G) date and (H) facility discharge. Grey line signifies WWTF mean daily average discharge rate (A, C, E, G) Figure 7. (A) DIN fluxes at the Stillman and Westerly bridges and from the Westerly WWTF vs. sampling date. (B) Riverine DIN flux at the Stillman and Westerly bridges vs. the river discharge recorded at the Stillman bridge. (C) DON flux at the Stillman and Westerly bridges and from the Westerly WWTF vs. sampling date. (D) Riverine DON flux at the Stillman and Westerly bridges vs. river discharge recorded at the Stillman bridge. (E) Riverine TN flux at the Stillman and Westerly bridges vs. sampling date, and (F) vs. river discharge recorded at the Stillman

bridge.

v Figure 8. (A) River discharge recorded at three flow gauges along the Pawcatuck River

- 15 during three sampling campaigns in 2018-2019. (B) [NO3 ], (C) δ NNO3, (D)

18 3- δ ONO3, and (E) [PO4 ] measured downriver from its origin at Worden pond to

the Westerly Bridge during the sampling campaigns. Figure 9. Percent contribution of river sections to the total riverine DIN flux recorded at the Westerly bridge during downriver sampling trips. Table 2. DIN flux per section of the Pawcatuck River. Table 3. N loading from Westerly WWTF from May through October

- Figure 10. Modified Keeling Plot of NO3 isotopologue ratios at the Westerly and Stillman

15 - 18 bridges: (A) [δ NNO3] vs. the inverse of the NO3 flux, (B) δ ONO3 vs. the inverse

- 18 of the NO3 flux, an (C) δ ONO3 corrected for the contribution of atmospheric

- - NO3 vs. the inverse of the NO3 flux.

- Figure S1. Log conductivities at bridges vs. (A) date and (B) river discharge. (C) [NO3 ] vs log conductivities at bridges.

- Figure S2. TukeyHSD analysis of downriver seasonal transects for (1, 2, 3) [NO3 ], (4, 5, 6)

3- 15 18 [PO4 ], (7, 8, 9) δ NNO3 and (10, 11, 12) δ ONO3.

- Figure S3. (A) [NO3 ] in composite vs. grab samples from the Westerly WWTF, measured by

- - UConn. (B) Plant-reported [NO3 ] composite effluent vs. [NO3 ] measured at

+ UConn. (C) [NH4 ] in composite vs. grab samples from the WWTF, measured by

+ + UConn. (D) Plant-reported [NH4 ] composite effluent vs. [NH4 ] measured at UConn. Figure S4. Mixing curve of [DIN] with river discharge.

vi Abstract Little Narraganset Bay has seen an increase in filamentous algae, colloquially known as cladophora, over the past twenty years. The Pawcatuck River is the dominant input of freshwater to Little Narraganset bay, delineating the Connecticut and Rhode Island border. There are three point sources of nitrogen on the river system: two waste water treatment facilities (WWTF) in the estuary and one fabric processing plant within the upper-watershed in Kenyon, RI. Non-point sources include a myriad of farms specializing in turf and animal husbandry, as well as septic systems for the majority of the watershed that are not serviced by the WWTFs. Prior research has produced yearly flux estimates of nitrogen (N) yet point to different anthropogenic sources including sewage for estuarine WWTF and fertilizer from upriver as predominant nutrient sources of N algal growth. Samplings, conducted weekly for an annual cycle, at the Westerly and Stillman Bridges at the mouth of the Pawcatuck River – upstream of seawater intrusion, and seasonal down-river transects from Wakefield, RI to Westerly, RI were collected in order to track seasonal changes and determine sources of N along the river. Nutrients of N (nitrate, nitrite, and ammonium), phosphate, and nitrate isotopes were measured at each sampling point. Nitrate isotopes provided additional power for analysis to help differentiate various input sources into the river as well as to determine cycling of N on the river. Results indicate that there were significant seasonal variations in nutrient input of nitrate and ammonia linked to seasonal discharge rates where high discharge occurs in the winter and low flow occurs in the summer. With seasonal discharge rates, the WWTFs contribute negligible amounts of N to the river in winter and up to 15% of the loading in the summer. Results also signify that the overall annual flux of the Pawcatuck river into Little Narraganset Bay are consistent with prior estimates, where DIN, DON, and TN export is equal to 18 x 106, 15 x 106, and 32 x 106 moles of N per year respectively indicating there have been no major changes N flux in the upper river.

vii 1. Introduction In the past twenty years, Little Narraganset Bay – an estuary located on the southern border of Connecticut and Rhode Island, has hosted an increasing biomass of a filamentous green algae in the family Cladophoraceae, with a resultant decrease in the vascular plant, eelgrass (Zostera marina Linnaeus). The complex of filamentous green algae includes primarily Chaetomorpha linum (O.F.Müller) Kützing and a Cladophora sp. similar in appearance to Cladophora vagabunda (Linnaeus) Hoek, but not positively identified as this species. Collectively, the complex of green algae is referred to by the common name, “cladophora”, to ease communication with the general public and in reference to the family name of the two dominant species. Nuisance algae, like cladophora, is often associated with eutrophic waters where an increased prevalence is seen with respect to population density (Dodds and Gudder, 1992; D’Avanzo and Kremer, 1994; Valiela et al., 1992). Cladophora in Little Narraganset Bay are a free- floating group of algae which can aggregate in mats exceeding a foot in thickness and create problems both for the ecology of the ecosystem as well as for the human inhabitants who work on, live near, or visit the waterway (Dostie and Vaudrey, 2014). Floating cladophora mats can easily entangle boat propellers, disrupt swimmers, and foul structures; washed up mats impede beach and shellfish bed access and produce a pungent sulfidic odor upon decomposition (Dodds and Gudder, 1992; Valiela et al., 1992). These large mats exert a high biological demand of oxygen on the system, resulting in large oscillations of dissolved oxygen which are only confounded by light attenuation, temperature, and wind stress which all effect the ability of the cladophora to photosynthesize (Dodds and Gudder, 1992; D’Avanzo and Kremer, 1994). Indeed, large oscillations in dissolved oxygen, from hypoxic in the morning due to lack of photosynthesis with high respiration at night to hyper-oxic in the late afternoon when photosynthesis is at its max, have been documented during the summer in coves within Little Narraganset Bay (Westbrook and Magnano, 2016; Rubino and Karamavros, 2016). These oscillations can create a niche that limits species richness and abundance of benthic creatures who do not have the mechanisms to survive in the extreme oscillating conditions that cladophora produces (Valiela et al., 1992). Little Narragansett Bay is a relatively small and shallow estuary at the head of a large watershed (~100-fold larger area). The main conduit from the watershed into the Little

1 Narragansett Bay is the Pawcatuck River. Although the Pawcatuck River has a modest to low annual total discharge compared to similarly sized river systems in New England, it reportedly

-1 -1 exhibits relatively high N loading for size of estuary at 942.85 kg N haestuary yr , ranking it in the top 12% of most N loaded embayment’s on Long Island Sound (Savoie et al, 2017; Vaudrey et al, 2016). Fulweiler and Nixon (2005) observed this high ratio of N loading to watershed area in 2001 (a drought year) with an extrapolated estimate of 26.0 x 106 moles N yr-1 for a non-drought year, where they calculated about 90 % of the N imported into the watershed was retained therein. A land use model by Vaudrey et al. (2016) estimated the dominant sources of N loading to Little Narragansett Bay were fertilizer used by agriculture and residential lawns (35 % of the total N load; 55 % commercial and 36 % residential) and sewer and septic discharge (33 % of the total load) through groundwater penetration. An industrial point source and atmospheric deposition each account for an additional 15%, which is consistent with reports from the Rhode Island Department of Environmental Management (RI-DEM) which state that WWTF N is negligible compared to the upper watershed (Dillingham et al., 1993). However, the Vaudrey et al. (2016) study suggested that wastewater treatments facilities (WWTF) in the Pawcatuck river estuary, in particular, account for 14 % of the total N loading to Little Narragansett Bay, and given their close proximity to the Cladophora mats, implicated WWTF N discharge as a predominant fuel for algal growth in the estuary. Prior research has concluded that the Pawcatuck River exhibits excess N loading; however, questions remain regarding the sources of this N. This study is intended (a) to identify the major sources of N to the upper river, (b) to revisit N loading estimates produced by Fulweiler and Nixon (2005) and Vaudrey et al.; (2016), and (c) to determine what sources may be mitigated to diminish the loading to Little Narraganset Bay. The study is comprised of three parts (1) weekly sampling at the mouth of the Pawcatuck River, (2) seasonal downriver transects, and (3) weekly monitoring of effluent from the Westerly WWTF. For all sampling trips, we measured nutrients

- of collected water as well as N and oxygen (O) isotopes of nitrate (NO3 ) to identify sources and cycling within the system. Stable isotope analysis provides us with an added resource to track and distinguish sources and cycling within a system, complementing the nutrient analysis making this an insightful tool to differentiate sources and cycling tendencies (Kendall et al., 2007; Kreitler,

2 1975, 1979; Broadbent et al., 1980; Casciotti, 2016). We do note that although isotopic analysis can be a powerful tool, it does come with some limitations including the difficulty of differentiating source mixing effects from the isotopic fractionation of N cycling steps particularly in a fluvial system with multiple point sources and benthic niches (Lin et al. 2019). Our observations suggest that the WWTF in the winter supplies negligible amounts of N to the river while in the summer, WWTFs can account for upwards of 10 % of the N load, coinciding with low discharge. We also observe that there is a disproportionate loading of N up-river near Kenyon, RI but that the dominant N loading in summertime appears to originate from the groundwater sources derived along and released over the whole of the river.

1.1 Site Description

1.1.1 Little Narraganset Bay Little Narraganset Bay is a well-mixed saltwater bay, which divides Pawcatuck, Connecticut and Westerly, Rhode Island and has an average depth of 2 meters and an area of 3.2 km2 (Dillingham et al., 1992). Wequetequock Cove flows into the northwestern edge of the Bay, prominently tidal flushed, while the east side of the embayment is dominated by the Pawcatuck River, a 47 km long high humic concentration river (Doering et al., 1994). Between the two freshwater inputs is Connecticut’s largest wildlife management area, Barn Island, a 1,000 acre ecologically significant coastal wetland (CTDEEP, 2016). At the southern border of Little Narraganset Bay is Sandy Point Island, dividing the bay from Fishers Island Sound, a part of Long Island Sound. The main inlet, through the navigational channel, is to the west of Sandy Point - although there is a non-navigable inlet to the east which also provides exchange with open water.

1.1.2 Pawcatuck River and Upper Watershed The Pawcatuck River has a watershed area of 486 km2, 80 % of this drainage area is within Rhode Island municipal jurisdiction where private and state land trust holdings protect ~22 % of the watershed in RI from development (Dillingham et al. 1992; Heffner, 2007). The Pawcatuck River originates at Worden Pond in Wakefield, RI and extends southwest to Westerly, RI. There are two larger tributaries (Meadow Brook and Wood River) along with a multitude of smaller inlets onto the larger Pawcatuck River. At the Westerly Bridge in downtown Westerly, the river experiences sea water intrusions via a salt wedge estuary originating from Little Narraganset Bay,

3 thus we mark this as the mouth of the freshwater river system. The upper watershed is dominated by forested land (58 % of watershed land use) and smaller villages (Heffner, 2007). The area was noted as the most densely populated turf farm county in the nation in 2005; many of these farms border tributaries of the Pawcatuck River and agriculture accounts for 8.5 % of land use (EPA, 2005; Heffner, 2007). While the upper river is dominated by farmland and trees, the lower river from Bradford, RI to Westerly, RI is more urbanized (6.6 % of watershed land use; Heffner, 2007). There are three discharge permits allotted on the river, which include the Westerly WWTF, the Pawcatuck WWTF and Kenyon Industries (a fabric processing plant) (Figure 1). The two WWTFs discharge within a kilometer of each other about 2 km south of the Westerly bridge, directly before a naturally formed bottleneck on the estuarine portion of the river while the dye processing plant is about 7 km downstream from the head of the river at Worden Pond. Currently Pawcatuck WWTF does not have a limit on N loading, whereas Westerly WWTF is limited on TN (total organic N + ammonia + nitrate + nitrite) during May to October at 4.9 x 106 moles N per year and 1.8 x 106 moles N per year for ammonia (EPA 2014, 2013). Westerly WWTF is also limited on ammonia discharge from November to April to 1.0 x 107 moles N per year (EPA, 2013). Upriver at Kenyon Industries, the nitrate discharged ranges from 2.0 x 106 to 7.0 x 106 moles of N per year from May to October and from November to April respectively (EPA, 2010).

2. Methods 2.1 Field Samplings We followed four types of sampling regimens: (1) We collected weekly river samples from January 10, 2018 through to January 12, 2019 at two sites: the Stillman bridge (Station 12) near the mouth of the freshwater river and further down at the Westerly bridge (Station 13), at the head of seawater intrusion (Figure 1). (2) We collected weekly samples of waste water treatment effluent at the Westerly Waste Water Treatment Facility (W-WWTF) from June 6, 2018 to May 22, 2019. (3) We conducted three seasonal down-river surveys on May 21, 2018; November 9, 2018; and March 12, 2019 at discrete sampling stations between Worden Pond (at Rhode Island Department of Environmental Management designated boat launch) and the Westerly bridge. (4) We collected rainwater samples following rain events from a roof-top collector at the Avery

4 Point campus in Groton, CT, from September 6, 2018 to December 2, 2018, in order to define the nitrate isotopic endmembers. Weekly samplings at the bridges occurred around sunrise, before the onset of photosynthetic activity. River water was collected during all trips using a Van Dorn bottle. A portion of the collected water was carefully transferred to a 0.5 L container, avoiding turbulent mixing, and dissolved oxygen was measured using an Orion Start optical DO probe calibrated via 100 % air saturation. The remainder of the bridge sampled water was then transferred into a five- liter carboy for transport, on ice, back to the lab for processing. Samples for nutrient and nitrate isotope analyses were filtered through pre-combusted 0.45 µm GF-F filters and collected in acid washed polypropylene bottles, then stored at -20˚C pending analysis. The weekly effluent samples at the Westerly WWTF were collected by facility personnel into 0.5 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 prior to its release to the river, while composite samples are effluent collected over a 24-hour period, thus providing a concentration-weighted daily average. Samples were stored at -20˚C at WWTF and collected monthly where they were returned to the University, filtered through a 45 µm GF-F filter and frozen at -20 ˚C pending analysis. Rainwater was collected into trace metal-clean one-liter Teflon bottles that were outfitted with a glass funnel to create a vapor lock preventing evaporation. These samples were stored at - 20˚C until analysis and were not filtered. 2.2 Nutrient Analyses

- The nitrate concentration, [NO3 ], in river samples was measured by conversion to nitric oxide in a hot Vanadium III solution followed by detection on a chemiluminescent NOx analyzer

TM - (Teledyne ; Braman and Hendrix, 1989). Incident nitrite (NO2 ) in the samples was first reacted with Greiss reagents (Green et al., 1982) before injection into Vanadium (III) in order to detect

- - [NO3 ] only. The concentration of NO2 in river samples was measured by conversion to nitric oxide in hot iodine solution, followed by detection on the chemiluminescent NOx analyzer

- - (Garside, 1982). [NO3 ] and [NO2 ] in the rainwater samples were measured on the SmartChem

5 discrete nutrient autoanalyzer (Unity ScientificTM) using EPA and standard methods adapted for SmartChem (Strickland and Parsons, 1972; Standard Methods, 2018; EPA, 1993). Concentrations

+ 3- of ammonium [NH4 ], and phosphate [PO4 ], in river water were measured on a SmartChem autoanalyzer using EPA and standard methods adapted for SmartChem (Murphy and Riley, 1962; Strickland and Parsons, 1972; EPA, 1993b, 1978; Standard Methods 2017, 2017b). The concentration of total dissolved nitrogen, [TDN], in river samples was measured by

- persulfate oxidation (0.1 – 0.2 : 1 sample to persulfate oxidizing reagent) to NO3 , then measured via chemiluminescent NOx analyzer as described above (Knapp et al, 2005; Solorzano and Sharp, 1980). The persulfate reagent was first recrystallized following protocol by Gasshoff et al. (1999).

- - Dissolved organic nitrogen [DON] was calculated as the difference between [NO3 + NO2 ] and [TDN]. Flux estimates were calculated using the product of nutrient concentration and river discharge. Downriver sectional flux estimates were calculated to the nearest river flow gauge (3 in total), where each sectional flux was subtracted from the prior section to obtain net flux per section of river. This calculation does not take length of section into consideration. 2.3 Isotopic analysis

- 15 14 18 16 The nitrogen and oxygen isotope ratios of NO3 , N/ N and O/ O, were analyzed using

- the denitrifier method for samples where [NO3 ] ≥ 1.5µM (Sigman et al. 2001; Casciotti et al, 2002).

- Briefly, NO3 was converted quantitatively to a nitrous oxide (N2O) analyte by denitrifying bacteria that lack a terminal reductase (Pseudomonas chlororaphis f. sp. aureofaciens; 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

- 17 16 autosampler (Casciotti et al., 2002). The NO3 O/ O in rainwater 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). Herein, we express isotopes ratios (e.g. 15N/14N, 17O/16O, and 18O/16O) in delta notation:

� (‰) = − 1 × 1000 Equation 1

- 15 Isotope ratios were calibrated from parallel analyses of NO3 reference materials USGS-34 (δ N: -1.8 ‰ vs. air, δ18O: -27.9 ‰ vs. VSMOW) and IAEA-N3 (δ15N: 4.7 ‰ vs. air, δ18O: 25.6 ‰ vs.

6 VSMOW). Rain samples were calibrated via USGS-34 (δ15N: -1.8‰ vs. air, and δ18O: -27.9 ‰ vs. VSMOW, D17O: -0.1 ‰ vs. VSMOW) as well as USGS-35 (δ15N: 2.7 ‰ vs. air, δ18O 57.5 ‰ vs. VSMOW, D17O: 21.6 ‰ vs. VSMOW).

17 - 17 The mass independent fractionation of O in rainwater NO3 (∆ O) was calculated from coupled measurements of nitrate δ17O and δ18O in order to determine the fraction of atmospheric nitrate in river samples (from the Stillman Bridge samples only), as per Equation 2 (Thiemens, 1999): ∆ � = � � − 0.52 × � � Equation 2

- The fraction of atmospheric NO3 , in turn, was derived from a two-end-member mixing equation

- 17 17 of river water NO3 (∆ O = 0) to the respective mean atmospheric nitrate ∆ O observed in warm months (May – October) and cold months (November – April).

3. Results The mean daily discharge recorded by the USGS gauge at Stillman bridge between January 1st, 2018 and January 31st, 2019 was 2.0 x 106 m3 d-1, and ranged 25-fold, from 0.23 x 106 to 5.5 x 106 m3 d-1 (Figure 2). Discharge was greater in winter months and the lowest discharge occurred in the summer months. The 80-year mean daily discharge at this site is 1.4 x 106 m3 d-1, ranging from 0.06 x 106 m3 d-1 in Sept 2015 to a record 21.3 x 106 m3 d-1 in March 2010. Total precipitation in 2018 for Washington County, RI, was the third wettest year on record at 152 cm, compared to an 80-year mean of 114 cm (NOAA, 2019). 3.1 Weekly samplings

- [NO3 ] measured in samples collected weekly at the Stillman and Westerly bridges was lowest in winter and highest in the summer months, ranging from to 9.7 µM to as high as 73.5 µM (Figure

- 3A). These sites had comparable [NO3 ] at each sampling, except for instances where the Westerly

- bridge site experienced saltwater intrusions noted by elevated conductivities (Figure S1). [NO2 ]

- was negligible in all samples. [NO3 ] decreased with increasing river discharge at both bridge sites

15 (Figure 3B; Table 1). Values of δ NNO3 were lowest in winter and increased in summer, ranging

15 from 5.3 ‰ to 9.4 ‰ (Figure 3C). The δ NNO3 values thus decreased with increasing river

18 15 discharge (Figure 3D; Table 1). The δ ONO3 values followed a contrasting trend to δ NNO3, lower during the summer months and increasing in winter months, with values as low as 1.6 ‰ to

7 18 upwards of 6.8 ‰ (Figure 3E). Values of δ ONO3 at the bridges increased directly with discharge

17 (Figure 3F; Table 1). Measurements of ∆ ONO3 at Stillman Bridge ranged from -0.5 to 27.2 ‰,

- indicating the presence of modest fractions of uncycled atmospheric NO3 , from undetectable to

- 17 no more than 7.3 % of total NO3 (Figure 3G). Values of ∆ ONO3 correlated directly with river

- discharge (Figure 3H; Table 1), whereas the fraction of uncycled atmospheric NO3 derived from

17 ∆ ONO3 did not correlate significantly with river discharge.

- + In contrast to [NO3 ], [NH4 ] at the bridges was lowest in summer and higher in winter, ranging from below detection to 7.8 µM (Figure 1I), and correlated directly with discharge (Figure 3J).

3- - [PO4 ] mirrored [NO3 ], with lower concentrations generally observed in winter months, and higher concentrations in the summer (Figure 3K). Concentrations ranged from 0.1 µM to 2.7 µM

3- with one sample point as high as 5.9 µM during a single winter sampling. [PO4 ] appeared to correlate inversely with discharge, albeit only significantly so at the Westerly Bridge, not the Stillman bridge (Figure 3L; Table 1). Concentrations of DON at the bridge sites ranged from 9 to 56 µM and were not different between seasons (Figure 3M). [DON] did not appear to have a statistically significant trend with respect to river discharge (Figure 3N; Table 1). Nevertheless, [DON] was weakly inversely correlated to coincident [DIN] (significant at Stillman bridge only; Figure 4; Table 1). In turn, [PN] ranged from 1 to 13.4 µM, with higher values observed mostly, but not exclusively, in spring and fall months (Figure 3O). No correlation of [PN] with river discharge was apparent (Figure 3P; Table 1).

- DIN concentrations in rain water samples were highly variable, rain water[NO3 ] ranged

+ from 1.0 to 37.7 µM and [NH4 ] ranged from 1.3 to 18.77 µM (Figure 5A – 5B). Rain water [DIN]

- was relatively consistent from late summer to early winter. NO3 isotopes ranged from -6.1 to

15 18 17 1.7 ‰ for δ NNO3, from 57.8 to 75.7 ‰ for δ ONO3, and from 19.7 to 27.2 ‰ for D ONO3 (Figure

15 18 17 5C – 5E). δ NNO3 did not covary with season, whereas δ ONO3 and D ONO3 increased from summer to winter months. Nutrient concentrations measured in samples collected at weekly intervals at the Westerly WWTF, consisting of both grab and composite samples, ranged from 30 to 527 µM for

- + 3- [NO3 ], 1.3 to 1070 µM for [NH4 ], 11.7 to 1168 µM for DON and 2.7 to 26.5 µM for [PO4 ] (Figure

8 + - 6A, C, E, G). Concentrations of NH4 and NO3 in grab and composite samples were highly correlated (Figure S3A, C). Values measured at UConn were similar to those reported by the

- 3- Westerly WWTF (Figure S3B, D). [NO3 ] and [PO4 ] were higher in summer months when facility

+ - discharge was lower, whereas [NH4 ] and [DON] were lower in summer. [NO3 ] correlated

+ inversely with the facility-reported discharge rates, whereas [NH4 ] correlated directly with

+ discharge – that is, the facility-reported [NH4 ], for which there are more measurements to derive a correlation (Figure 6B, D; Table 1). There was overall an increase in [DON] with discharge, but while concentrations between composite and grab samples were similar during low flow, there was more variability during high flow – corresponding to winter months (Figure 6F; Table 1). Our

3- limited [PO4 ] measurements were not significantly correlated to plant-reported discharge (Figure 6H; 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 with the sum of [NO3 ] and

+ [NH4 ] recorded at the bridges, varied ~10-fold over the annual sampling period, ranging from

5 5 5 0.1 x 10 to 1.1 x 10 moles of NDIN per day – omitting a single outlier of 1.8 x 10 moles of NDIN per day (Figure 7A). The riverine DIN flux increased directly with river discharge (Figure 7B; Table

5 1), such that it was lowest in summer, averaging 0.2 ± 0.1 (x 10 ) moles of NDIN per day from May

6 through October. The annual DIN export was 18.2 x 10 moles of NDIN per year. The riverine DON

5 5 flux, in turn, ranged from <0.1 x 10 to 2.0 x 10 moles of NDON per day (Figure 7C), and also increased directly with discharge (Figure 7D; Table 1). The annual riverine export of DON was

6 15.4 x 10 moles of NDON per year. The total riverine N flux (the TN flux), the sum of respective

5 5 DIN, DON and PN fluxes, ranged from 0.2 x 10 to nearly 3.0 x 10 moles of NTN per day and correlated directly with discharge (Figure 7E, F; Table 1). The annual TN flux was on the order of

6 33.6 x 10 moles of NTN per year where ~20 % of the annual load is attributed to summer months between May and September. In contrast to the riverine flux, the DIN, DON and TDN (Total Dissolved N = DIN + DON) fluxes from the Westerly WWTF were relatively constant, and were substantially lower than

3 3 corresponding riverine fluxes, averaging 3.9 ± 2.6 (x 10 ) moles of NDIN per day, 1.7 ± 1.8 (x 10 )

3 moles of NDON per day, and 6.2 ± 2.9 (x 10 ) moles of NTDN per day, year round (Figure 7A-F). The

9 daily TDN loading at the Westerly WWTF was lower than the mandated allowable daily discharge

3 from May through November of 13.5 x 10 moles of NTN per day. Admittedly, we may be missing a non-negligible PN fraction of the WWTF effluent. Nevertheless, the discharge of TDN from the Westerly WWTF accounted for approximately 13 ± 12 % of the total summer N discharge (riverine and Westerly WWTP) from May through November, compared to 4 ± 4 % from November through April. Including the daily allowable discharge at the Pawcatuck WWTF of ca. 1.0 x 103 moles NTN per day, the TN flux from both WWTFs together accounted for approximately 15 % of the total N discharge from May through November, and 5% of daily discharge from November through April. These estimates do not include the unconstrained N loading contributed from storm drains downstream of the Westerly bridge. 3.2 Downriver samplings Mean daily water discharge at Stillman Bridge during our three collection trips on May 21, 2018, November 9, 2018 and March 12, 2019 was 1.9 x 106, 2.6 x 106, and 2.7 x 106 m3 d-1 respectively (Figure 8A). A one-way ANOVA was used to compare the effect of station location

- on nutrient and isotope concentrations by sample date. [NO3 ] measured at multiple river stations increased down-river from Worden pond to the Stillman bridge and was generally higher during March sampling and lowest during the November sampling at all river sites (Figure 8B). Concentrations in the source basin at Worden pond ranged from 0.4 to 6 µM between samplings.

- At the second sampling site 6 km downstream, at Biscuit City Road (Stn 2), [NO3 ] was similar to that at Worden Pond in May and November but was otherwise marked by a large increase to 67 µM in March. At the 9 km down-river site at Route 112 (Stn 3), past the Kenyon Industries

- discharge (7 km), [NO3 ] increased to 12 µM and 14 µM in May and November, respectively, and

- otherwise decreased to 38 µM in March. Thereon, [NO3 ] remained the same to slightly at the 14 km site at all sampling dates (Route 91; Stn 4), then decreased downstream of the Meadow Brook and Wood River inflows (Stns 5 and 6) at Burdickville Road (22 km; Stn 7) to concentrations of 17

- µM in May, 7 µM in November, and 29 µM in March. These tributaries had lower [NO3 ] than in

- the Pawcatuck (0.3 to 22 µM), evidently diluting [NO3 ] in the Pawcatuck river as they join the main flow of the river. The Wood River contributes significantly to water discharge (at least 14 ± 5 % of total – based on discharge at Hope Valley USGS gauge), whereas Meadow Brook

10 - constitutes a negligible addition to water discharge (and consequently [NO3 ]). From 22 km to the

- Stillman bridge down-river (Stn 12; 48 km), [NO3 ] increased slightly to final concentrations of 23 µM in May, 10 µM in November, and 32 µM in March at the Stillman Bridge.

15 δ NNO3 values incrementally increased down-river but differed among samplings dates at given river sites with the greatest range occurring in March 2019 (Figure 8C). Values at Worden Pond, measured in May 2018 only, were 4.3 ‰, increasing to 4.9 ‰ at the 6 km site downstream (Biscuit City Road; Stn 2). Values at this site in March 2019 were higher, on the order of 6.2 ‰,

- 15 and were associated with the large recorded [NO3 ]. δ NNO3 was not measured at Biscuit City Road in November due to constraints of the method for low concentrations. Past Kenyon Industries (Stn 3), values ranged from 3.4 ‰ to 6.0 ‰ between sampling dates, increasing down- river to values of 6.0 ‰ in May, 8.7 ‰ in November and 6.6 ‰ in March at the Stillman bridge.

15 - δ NNO3 values of NO3 delivered by the Wood River were between 5.7 ‰ and 6.6 ‰, similar to

- (in May and November) or higher than (in March) those of NO3 originating upstream in the Pawcatuck River.

15 18 In contrast to δ NNO3, δ ONO3 values decreased down-river to the larger tributaries and

18 then leveled off the remainder of the river to Stillman bridge in May (Figure 8D). δ ONO3 values

15 among sites were highest in November (as per δ NNO3) and were similar to each other in May and March. Values at Worden Pond were 10 ‰ in May and were not measured in November and

- March due to low concentrations of [NO3 ]. At Biscuit City Road, downstream values were

- identical in May and March at 4.6 ‰ – in spite of substantial differences in [NO3 ] between these sampling dates – decreasing in tandem to 4.0 ‰ at Station 3, past Kenyon Industries. Values at this site in November were higher, at 4.9 ‰. From station 3, δ18O values decreased slightly to Route 91 (Stn 5), then remained similar or increased slightly past the Wood River inflow (Stn 6) to values of 3.2 ‰ in May, 4.5 ‰ in November and 3.6 ‰ in March. Values in the Wood River were similar to those in the Pawcatuck at corresponding dates. From Station 6 to the Stillman bridge, values remained similar or decreased slightly, to 2.4 ‰ in May, 5.0 ‰ in November, and

18 3.7 ‰ in March. We note that there is no statistically significant variation in δ ONO3 downstream in November (TukeyHSD: F(9,20)=0.12, p=0.999) which is the only month that does not experience a decrease trend from head of the river to the mouth.

11 + [NH4 ] did not vary in a systematic fashion with distance downstream or among sampling

3- dates, ranging from 0.4 to 6.8 µM (data not shown). [PO4 ] were 0.2 to 0.5 µM in Worden Pond,

- and were comparable at Biscuit City Road downstream (Figure 8E). The steep increase in [NO3 ]

3- 3- at this site in March was thus not mirrored by [PO4 ]. [PO4 ] increased past Kenyon Industries in all samplings, to as high as 3 µM in March and was statistically different in concentration to the

3- remainder of the river in November and March (Figure S2). [PO4 ] remained elevated to Station

3- 4 in May and November but decreased by 0.5 µM in March. [PO4 ] in the Wood River and

3- Meadow Brook were similar to Worden Pond on all sampling trips. [PO4 ] after tributary input remained similar down river to Stillman Bridge on all trips, ranging between 0.6 and 0.9 µM. The DIN fluxes at Stations 2, 4 and 12 at each sampling date were computed from measured concentrations and the corresponding river discharge recorded by USGS gauges (Table 2). The flux at Station 2 varied 10-fold between samplings, from 1.5 x 103 moles of DIN d-1 in November to 2.6 x 104 moles of DIN d-1 in March. The flux at Station 4 varied 2 fold between samplings, lowest in May and highest in March, from 9.6 x 103 to 1.9 x 104 moles of DIN d-1, respectively. Similarly, the flux at Stillman bridge varied 2-fold from 1.5 x 104 to 6.7 x 104 moles of DIN d-1 in November and March, respectively. Overall, 15 ± 13 % of the total riverine DIN flux originated upstream of Station 2, 28 ± 11 % between Stations 2 and 4, and 72 ± 11 % between Stations 4 and 12 (Figure 9). In comparison, 23 ± 4 % of total riverine water discharge at these dates originated upstream of Station 2, 11 ± 2 % between Stations 2 and 4, and 67 ± 5 % downstream of Station 4 (see Figure 1).

4. Discussion 4.1 Source Attribution

- 3- At the Stillman and Westerly bridges, [NO3 ] and [PO4 ] scaled inversely with discharge, with higher concentrations of both occurring during summer at low base flow. This suggests that the bulk of nutrients during low base flow originated from groundwater, benthic remains in the

- shallow water column, and/or point sources along the river catchment. Indeed, [NO3 ] increased in proportion to conductivity (Figure S1), further suggesting a groundwater source for bulk riverine nutrients at low base flow. During wetter months, increased groundwater surface flow into the river in conjunction with recharging of the groundwater aquifer, potentially diluted the

12 groundwater and/or point source-origin nutrients, thus lowering riverine concentrations. Nevertheless, the daily DIN flux increased with discharge, indicating that nutrients were also imported to the river from groundwater surface flow, albeit, at a lower concentration than low base flow nutrients. From the slope of this correlation (Table 1), the daily DIN flux increased by ~0.26 x 105 moles of N per additional 106 m3 of discharge, suggesting that the DIN concentration of surface flow from the catchment averaged 26 µM. The relationship between [DIN] and

- discharge can thus be approximated by a two end-member mixing curve of base flow [NO3 ] and

- surface water [NO3 ] (Figure S3), assuming a 60 µM base-flow end-member approximated from

- median low base flow [NO3 ], and further assuming negligible in-river N consumption, a notion to which we return below. The inverse correlation of [DON] with [DIN] (Figure 4), in turn, suggests that [DON] is transported into the river by surface flow from the catchment. Surface run-off, which increases with increased precipitation, is apt to transport organic material from soils and surface plant materials. The import of DON from surface flow is consistent with the Pawcatuck being a high tannin river which gives us an indication of surface flow catchment input. In this respect, the lack of direct correlation of [DON] to discharge is surprising but may be masked by the relatively high variability of [DON] measurements, even between replicate water samples. This lack of relationship between [DON] and river discharge was observed in Pawcatuck River previously (Fulweiler and Nixon, 2005), corroborating our observations. These findings are further discussed below. Nutrient loads in the Pawcatuck River were investigated previously by Fulweiler and Nixon (2005). As in this study, they observed an inverse relationship of [DIN] to discharge from biweekly measurements at Stillman Bridge. Contrary to our interpretations, however, they argue that the seasonal decline in [DIN] was due to uptake by vegetation within the catchment. They observed the lowest [DIN] in spring – rather than summer (as per this study) – at which time discharge was the lowest of their annual study period. Here, we otherwise argue that increased water discharge dilutes the low base-flow nutrients deriving from groundwater and point sources in the river, such that concentrations are most elevated at low base-flow. The riverine DIN flux nevertheless increases with discharge, harboring nutrients from both groundwater and the catchment.

13 - Fulweiler and Nixon (2005) also observed that [NO3 ] and [DON] were inversely correlated, as per the current study, and further detected a slight positive correlation between [DON] and discharge, corroborating our earlier inference. They reasoned that the greater remineralization of bioavailable DON in spring, at low discharge, could explain this trend, given the greater in-river residence time of DON. While the mineralization of DON is likely accelerated during the warm season (in summer), we otherwise contend that the increased [DON] with discharge largely reflects greater import from the catchment by surface waters.

- We turn to NO3 isotope composition to further investigate relationships of nutrients to

- - 15 discharge. Like [NO3 ], the isotope ratios of NO3 also co-varied with discharge. Values of δ NNO3

- 15 decreased with discharge, suggesting that (a) NO3 added by surface flow had a lower δ NNO3

- 15 than low base flow NO3 , and, that (b) δ NNO3 potentially increased during warmer months due

18 to biological cycling in the water column. Concurrently, δ ONO3 values increased with discharge,

- 18 suggesting that (a) NO3 added by surface flow had a higher δ ONO3 than low base flow nitrate,

18 and/or that (b) δ ONO3 potentially decreased in summer due to biological cycling. We consider these scenarios in turn.

- 15 In order to evaluate whether the addition of NO3 from surface water with a lower δ N can

15 15 explain the δ NNO3 decrease with discharge, we plotted the δ NNO3 values vs. the inverse of the

- NO3 flux at Stillman bridge (i.e., an adapted Keeling Plot; Pataki et al. 2003; Wong et al. 2018; Figure 10). The data conform to a linear relationship expected for the addition of a reactant with a relatively invariant isotopic composition to the low base flow reservoir (Table 1). The intercept

- of the resulting linear regression suggests that the NO3 associated with increased discharge had a δ15N of ~6 ‰ (± 0.2 ‰; Figure 10B; Table 1). This value is lower than the value of ~8 ‰ for low

- base flow NO3 (see Figure 3; Table 1). It is, however, greater than that which we measured in

- atmospheric NO3 in rainwater, which averaged -2.5 ‰ (± 2.13 ‰, ranging from -6.1 to 1.8 ‰).

- Thus, we surmise that NO3 added from surface water did not originate dominantly from direct

- atmospheric deposition (i.e., as uncycled atmospheric NO3 ), but rather derived from that in

15 catchment soils and shallow groundwater. Indeed, the derived δ NNO3 end-member value of 6 ‰

- is typical of soil NO3 in forested catchments (e.g., Barnes and Raymond, 2009). While the net sources of reactive N to the catchment are rain and biological N2 fixation (and industrial

14 15 fertilizers), which all have relatively low δ NNO3 values, partial denitrification in soils increases

15 the δ NNO3 of the soil N reservoir to values ~5 ‰ (e.g., Houlton and Bai, 2009).

- The inference that uncycled atmospheric NO3 did not contribute substantially to the

- 17 increased NO3 flux at higher discharge is corroborated by the ∆ ONO3 measurements at Stillman Bridge. The low values observed evidence only a modest contribution of uncycled atmospheric

- - NO3 to total riverine NO3 , averaging 1.5 %, with values at high discharge not exceeding 7.3 %. This observation is echoed in a recent metanalysis of North American rivers, wherein the

- contribution of uncycled atmospheric NO3 to base flow is generally modest (Sebestyen et al.

- 2019). Thus, NO3 delivered to the Pawcatuck from groundwater surface flow evidently originated from a reservoir that was biologically cycled within the catchment soils and potentially in-river, thus losing its atmospheric ∆17O signature (Figure 3H).

18 - A Keeling plot of δ ONO3 vs. the inverse of the NO3 flux at the Stillman Bridge suggests that

- NO3 added from surface flow was on the order of 4.2 ± 0.3 ‰ (Figure 10B), compared to a low base flow value of 2.8 ± 0.2 ‰ (Table 1, Figure 3F). While the contribution of uncycled

- atmospheric NO3 to the riverine reservoir was modest, we nevertheless consider that the

18 - increase in δ ONO3 with discharge may derive in part from the uncycled atmospheric NO3 , given

18 - the particularly elevated δ ONO3 of ~80 ‰ observed in the rainwater NO3 . Indeed, when the

- 18 weighted contribution of atmospheric NO3 is subtracted from individual δ ONO3 values, the intercept of the Keeling plot decreases slightly to 3.4 ± 0.3 ‰, a value closer to that of low base

- 18 flow NO3 (Figure 10C; Table 1). The increase in δ ONO3 with discharge is thus partially explained

- 18 by the small component of uncycled atmospheric [NO3 ] with an elevated δ ONO3. Otherwise the

- 18 bulk NO3 added with increasing discharge had an δ ONO3 signature similar to or slightly higher

- than that at low base flow. This value is in the range generally observed for soil NO3 (Kendall et

- al., 2007). It has traditionally been ascribed to that expected for newly nitrified NO3 , based on

18 an empirical metric stipulating that the δ ONO3 produced by nitrification derives from the

18 18 fractional contribution of the reactants, namely 1/3 δ O of O2 + 2/3 δ O of H2O (Anderson and

18 Hooper, 1983; Hollocher 1984). Considering that the δ OH2O of Pawcatuck river water is -7 ‰

18 and the δ OO2 of atmospheric oxygen is 24.5 ‰ (Kroopnick and Craig, 1972), the expected

18 nitrification δ ONO3 value here is on the fortuitous order of 3.5 ‰. This empirical metric, however,

15 demonstrably overlooks important isotope effects associated with O-atom incorporation into the

- - 18 NO3 molecule during nitrification, which otherwise give way to nitrified NO3 whose δ ONO3 is close to that of ambient water (Boshers et al., 2019; Buchwald and Casciotti, 2010; Sigman et al.,

18 2005). This consideration explains frequent observations of relatively low δ ONO3 in some soils and saturated systems, which are not explained by simple fractional contribution of reactants

18 - (Veale et al., 2019; Xue et al., 2009). Thus, we posit the δ ONO3 of the NO3 imported into the river with increased discharge, which is typical of that in soils and shallow groundwater, does not

- strictly indicate that surface NO3 originated from proximate nitrification therein, as generally presumed, but also signals that it underwent partial denitrification, resulting in a coupled

15 18 - increase in its δ N and δ O relative to source values (Granger, 2008). We note here that NO3

18 assimilation by plants does not directly influence the δ ONO3 of soil NO3 since the process does not result in direct fractionation with the residual nitrate pool (Kalcsits et al., 2014). 4.2 River Cycling Our analysis thus far has not addressed any potential influence of in-river biological cycling

15 on incident nutrient concentrations and isotopologue ratios. The increase in δ NNO3 and

18 coincident decrease in δ ONO3 during summer months may not only arise from a decreased

- contribution of NO3 from the catchment, but could also implicate increased biological cycling in

- summer, modifying the isotope composition of low base flow NO3 relative to its groundwater

- and/or point source values. We note here that the isotopic composition of surface flow NO3 inferred from the intercepts of respective Keeling plots is not invalidated even if low base flow

- NO3 is modified seasonally by in-river cycling. The expectation of increased biological activity in

+ summer months is consistent with the incident decrease in [NH4 ], correspondent with lower discharge, which can be explained by increased algal assimilation. Fulwieler and Nixon (2005)

+ similarly observed lower [NH4 ] in the summer, even though lower discharge occurred in the

+ spring, supporting our contention that increased seasonal biological cycling underlies [NH4 ]

- dynamics rather than river discharge. The extent to which the coincident NO3 pool is also assimilated – and fractionated – during summer months is unclear. The fraction of the incident

- NO3 pool assimilated by algae may be relatively modest, even in summer, on the basis that the phytoplankton biomass is relatively small in this high tannin river, with median chlorophyll-a

16 concentrations ~ 1.9 µg L-1 at Stillman bridge and 4.3 µg L-1 at Westerly bridge in summer (data not shown), and corresponding PON of ~ 2.5 µM. Similarly, the spatial extent of macroalgae and macrophytes populations in the Pawcatuck River near Westerly, RI is also limited, relegated to the edges and shallow reaches. Thus, given its low biomass, algal growth may rely predominantly

+ on recycled/reduced N substrates, specifically NH4 . A prior knowledge of the concentrations and isotopologue ratios from groundwater and point

- source of NO3 at low base flow is required in order to unambiguously assess the extent to which biological cycling in summer influences these properties in-river. Nevertheless, if a sizeable

- 15 fraction of incident NO3 pool were assimilated into biomass during summer months, the δ NNO3

18 - - and δ ONO3 of base flow NO3 would increase in proportion to the fraction of NO3 assimilated

15 (Granger et al, 2004; 2010). The δ NNO3 increase down-river, observed during the seasonal

- surveys, could, by itself, post as evidence of progressive NO3 assimilation down-river, yet it was

18 + not matched by a coincident increase in δ ONO3. Moreover, [PN], chlorophyll-a and/or [NH4 ] did not increase down-river, as would otherwise be expected for the progressive conversion of the

- + 15 NO3 pool into the particulate and/or recycled NH4 pools. The down-river increase in δ NNO3 and

18 relatively invariant or decreasing δ ONO3 could otherwise result from nutrient spiraling, a

15 concept that we discuss further below. However, the downstream increase in δ NNO3 was apparent in all seasons, not only in summer. On the presumption that water-column and benthic N cycling was substantially reduced in March when river waters were glacial (average

15 temperature of 5.9 ˚C), we surmise that the increase in δ NNO3 down river may reflect spatial

15 differences in the δ NNO3 of groundwater sources and/or point sources. The watershed upriver hosts a number of turf farms as well as an industrial point source in close proximity to the river

15 which may contribute relatively low δ NNO3 (Deutsch et al, 2005). In the lower part of the river, the population density and associated septic systems are greater and agricultural lands in the watershed are dominated by animal husbandry – both associated with relatively higher

15 15 groundwater δ NNO3 (Kasper et al, 2015). Thus, the δ NNO3 increase downriver could reflect spatial differences in that groundwater and point sources.

18 - Corresponding δ ONO3 values offer limited but coherent insights with regard to NO3 sources

15 vs. in-river cycling. As alluded earlier, nutrient cycling could account for a downriver δ NNO3

17 18 increase not matched by a corresponding δ ONO3 increase. On the one hand, direct denitrification in sediments would expectedly impart coincident increases in the δ15N and δ18O

- of residual NO3 at the sediment depth of denitrification, which could be communicated to the overlying river water column via hyporheic flows (Ensign and Doyle, 2006; Harvey et al, 2013).

18 The δ ONO3 thus released could subsequently decrease due to cycling and re-cycling in the water

- 18 column, wherein nitrification produces NO3 with a δ O approaching that of ambient water. Alternatively, coupled nitrification-denitrification in sediments can also decouple N and O isotope

15 ratios, increasing the δ NNO3 in proportion to the N loss to denitrification, while subsequently

18 + decreasing the δ ONO3 due to the concurrent nitrification of NH4 (e.g., Granger et al, 2011; 2013). Both of these denitrification scenarios are consistent with the concept of riverine nutrient spiraling, which stipulates that nutrients are continually assimilated in the water column then decomposed (and denitrified) in sediments, then returned to the water column where they can undergo assimilation into new biomass then decomposition returning to the sediment (Ensign

- and Doyle, 2006; Harvey et al, 2013). While only a limited fraction of the incident riverine NO3 is

- evidently assimilated into algae at a given time, the NO3 pool is nevertheless cycled: A small fraction is likely assimilated whilst an equivalently small fraction is produced concurrently by

+ 15 nitrification. Given diminutive pools of incident NH4 and PN, we then posit that the δ NNO3 in-

15 river derives dominantly from δ NNO3 of sources and may be fractionated as a function of N loss

18 - to denitrification in sediments. In turn, δ ONO3 may be determined dominantly by that of NO3

18 sources in winter, when biological activity is dampened. In summer, the δ ONO3 may otherwise

18 be sensitive to the balance of δ ONO3 produced by nitrification vs. the O isotopic discrimination

18 from N loss to denitrification in sediments. The δ ONO3 in summer may thus not reflect that of

- NO3 sources on the basis that riverine nutrients can undergo multiple cycles of assimilation and recycling between water column and benthos (i.e., nutrient spiraling), thus progressively erasing

18 15 the δ ONO3 of original sources. In all, the increase in δ NNO3 downstream, coupled to an invariant

18 or decreasing in δ ONO3, can be explained by (a) spatial differences in groundwater and point sources and/or (b) nutrient spiraling in summer.

- Fulweiler and Nixon (2005) point out that there are no known sources of [NO3 ] on the river,

- our down-river surveys of [NO3 ] and its isotopes provide some constraints on sources within the

18 catchment. Overall, the Pawcatuck River is relatively pristine (low [DIN]) at the source in Warden Pond. We see a disproportionate contribution of [DIN] proximate to Kenyon Industries and turf farms near Kenyon, RI. This accounts for 30 % of the total riverine DIN flux though account for

3- only 10 % of the total riverine discharge. This is mirrored by a remarkable [PO4 ] concentration increase near Kenyon Industries, signaling a point source of phosphate. The [DIN] flux contribution from station 4 to station 12 accounts for 72% of the total riverine flux which is slightly lower than or commensurate with the proportional discharge for that section accounting for 89 % of the discharge (Figure 8). Thus, the Pawcatuck River is experiencing dilution by more pristine river and tributaries, notably the Wood River. Annual fluxes of TN and of DIN, measured at the Stillman Bridge were 33 x 106 and 18 x 106 moles yr-1, respectively. These values were about double to estimated made from weekly measurements at Stillman Bridge in 2001 by Fulweiler and Nixon (2005) of 16 x 106 and 7.2 x 106

-1 - moles yr , respectively. The observed [NO3 ] at low base flow in 2001 was surprisingly similar to that observed here, on the order of 50 – 60 µM, suggesting similar inputs from groundwater and point sources for the two sampling years. Regional atmospheric deposition of DIN has decreased

- by 67 % since 2000, from ~18 µM to 6 µM NO3 in 2018 (National Atmospheric Deposition Program, 2019). We would thus expect to observe a lower riverine N flux. However, 2001 was a drought year, with evidently lower river discharge (yearly average discharge = 1.5 x 106 m3 d-1 in 2001 vs

3 -1 - 0.73 m d in 2018) – thus proportionally lower [NO3 ] inputs above base flow were added from the catchment. 4.3 Estuary Point sources and mitigation The Westerly and Pawcatuck WWTF accounted for a relatively modest fraction of the total TN and DIN discharge to the estuary, from negligible in the winter to on the order of 8 % in the summer (2.5 and 0.4 x 106 moles N yr-1respectively), during the growing season. Fulweiler and Nixon (2005) otherwise estimated the WWTF TN flux during normal flow to be ~18 % of the total river flux, based on annual estimates derived from indirect sources. Corresponding estimates based on a land-use model by Vaudrey et al. (2016) suggested that WWTF effluent contributes ~12 % of TN (3.1 x 104 kg N yr-1) discharged to the estuary on an annual basis.

19 We note that we have not considered the discharge of phosphate into the estuary, though measured, given that N was the limiting nutrient in the estuary. Nevertheless, there are relatively important sources up-river, specifically, a considerable point source from Kenyon Industries which is commensurate with P release of ~ 950 moles P d-1 (EPA, 2010). As noted by Fulwiler and Nixon (2005), while [DIN] in the Pawcatuck River could be construed as elevated, [DIN] and the total N flux in the Pawcatuck River are modest relative to other New England rivers. For instance, the Pawtuxet River, RI, which has a similar discharge, has an annual TN flux on the order of 64 x 106 moles yr-1 (Nixon et al, 1995). Similarly, the Woonasquatucket River has a discharge four times lower than the Pawcatuck yet approximately half the N loading at 10 x 106 moles yr-1. We note that these estimates are over 25 years old and while population has not changed significantly (U.S. Census Bureau, 2020) WWTF have subsequently lowered their N discharge indicating that these are likely an over estimation of current N flux especially since groundwater has a longer residence time (> 50 years). However, the Pawcatuck River has a high watershed area with discharge predominantly comprised of groundwater. The river discharges into a relatively small estuarine area with poor flushing of subsurface waters, thus contributing to severe local eutrophication of Little Narragansett Bay – as reasoned by Fulweiler and Nixon (2005). A considerable fraction of the total N flux from the Pawcatuck River derives from DON, whose reactivity is unclear, with potentially upwards of 50 % bioavailable on pertinent time scales (Seitzinger et al, 2002). Nevertheless, DIN dominated the total N flux in summer (69 %), when N loading most influences macro-algal growth. Thus, specific reductions in low base flow DIN should be the target of any mitigation efforts. In addition, there is a disproportionate input (~16 % of total river DIN flux) of DIN up-river between stations 2 and 4 relative to the catchment drainage (~11 % of total river discharge flux; between Kenyon Industries and Wood River Junction), especially in direct comparison to other sections of the river where DIN flux percentages per section of river are consistently less than river discharge flux percentage. This may merit further investigation as to whether reductions are feasible within the 7 – 15 km section of the river. We suggest a closer investigation of potential point sources down-river of station 4, as there are

20 tributaries that drain agricultural land and industrial sites unseen in our resolution of down-river transects. All these mitigations would be aimed at reducing macroalgal growth in the bay, with the caveat that bay dynamics are not explicitly included in this study and therefore warrant their own monitoring beyond watershed investigations. The mitigations suggested in this study would benefit from a better understanding of Cladophora with respect to its niche in the ecosystem, it’s physiological limitations and its effect on the system with respect to nutrient cycling. This study indicates that there is a disproportionate loading in the upper river in comparison to water discharge, therefore we suggest mitigation strategies to target this region to limit direct loading to adjacent waterways in the upper river as well as decreasing the N loading to the groundwater aquafer via updates to septic systems and/or implementation of larger regional sewer system beyond the estuary section of the river.

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22 ESRI. 2018. U.S. State shapefiles. Prepared by ESRI and TomTom North America, Inc. EPA. 1978. Method 365.3: Phosphorous, All Forms (Colorimetric, Ascorbic Acid, Two Reagent). Environmental Monitoring Systems Laboratory Office of Research and Development. Cincinnati, OH. EPA. 1993. Method 353.2, Revision 2.0: Determination of Nitrate-Nitrite Nitrogen by Automated Colorimetry. Environmental Monitoring Systems Laboratory Office of Research and Development. Cincinnati, OH. EPA. 1993b. Method 350.1: Determination of Ammonia Nitrogen by semi-automated Colorimetry. Environmental Monitoring Systems Laboratory Office of Research and Development. Cincinnati, OH. Fulweiler, R.W. & Nixon, S.W. 2005. Annual export of nitrogen, phosphorus, and suspended solids from a southern New England Watershed. Biogeochemistry 76: 567- 593. DOI: 10.1007/s10533-005-0444-7 Garside, C. 1982. A chemiluminescent technique for the determi- nation of nanomolar concentrations of nitrate and nitrite in seawater. Mar. Chem. 1982, 11 (2), 159−167. Granger, J., & Wankel, S. D. 2016. Isotopic overprinting of nitrification on denitrification as a ubiquitous and unifying feature of environmental nitrogen cycling. Proceedings of the National Academy of Sciences, 113(42), E6391- E6400. https://doi.org/10.1073/pnas.1601383113 Granger, J., Prokopenko, M. G., Sigman, D. M., Mordy, C. W., Morse, Z. M., Morales, L. V., … & Plessen, B. 2011. Coupled nitrification-denitrification in sediment of the eastern Bering Sea shelf leads to 15N enrichment of fixed N in shelf waters. Journal of Geophysical Research: Oceans, 116(C11). Granger, J., Sigman, D. M., Needoba, J. A., & Harrison, P. J. 2004. Coupled nitrogen and oxygen isotope fractionation of nitrate during assimilation by cultures of marine phytoplankton. Limnology and Oceanography, 49(5), 1763-1773. Grasshoff, K., M. Ehrhardt, K. Kremling, and L. G. Anderson (Eds.) 1999. Methods of Seawater Analysis, 600 pp., John Wiley, Hoboken, N. J.

23 Green, L.m Wagner, D., Glogowski, J., Skipper, P., Wishnok, J., and Tannenbaum, S. 1982. Analysis of Nitrate, Nitrite, and [15N]Nitrate in Biological Fluids. Analytical Biochemistry 126, 131-138. Harvey, J. W., J. K. Böhlke, M. A. Voytek, D. Scott, and C. R. Tobias. 2013. Hyporheic zone denitrification: Controls on effective reaction depth and contribution to whole-stream mass balance, Water Resour. Res., 49, doi:10.1002/wrcr.20492. Heffner, L. 2007. Land Conservation as a Strategy for Water Quality Protection in Estuaries: Estimating Nutrient Fluxes in the Wood-Pawcatuck River Watershed and Recommendations for Further Conservation Actions. The Nature Conservancy’s Global Marine Initiative Hollocher, Thomas. 1984. Source of the oxygen atoms of nitrate in the oxidation of nitrite by Nitrobacter agilis and evidence against a P-O-N anhydride mechanism in oxidative phosphorylation. Archives of biochemistry and biophysics. 233. 721-7. 10.1016/0003- 9861(84)90499-5. Houlton, B. Z., & Bai, E. 2009. Imprint of denitrifying bacteria on the global terrestrial biosphere. Proceedings of the National Academy of Sciences of the United States of America, 106(51), 21713–21716. doi:10.1073/pnas.0912111106 Kalcsits, Lee & Buschhaus, Hannah & Guy, Robert. 2014. Nitrogen isotope discrimination as an integrated measure of nitrogen fluxes, assimilation and allocation in plants. Physiologia plantarum. 151. 10.1111/ppl.12167. Kasper, J., Denver, J., & York, J. 2015. Suburban Groundwater Quality as Influenced by Turfgrass and Septic Sources, Delmarva Peninsula, USA. J. Environ. Qual. 44:642–654 doi:10.2134/jeq2014.06.0280 Kendall, Carol & Elliott, Emily & Wankel, Scott. 2008. Tracing Anthropogenic Inputs of Nitrogen to Ecosystems. 10.1002/9780470691854.ch12. Knapp, A. N., D. M. Sigman, and F. Lipschultz. 2005. N isotopic composition of dissolved organic nitrogen and nitrate at the Bermuda Atlantic Time-series Study site, Global Biogeochem. Cycles, 19, GB1018, doi:10.1029/2004GB002320.

24 Kracht, O. 2011. Simultneous N, C and S isotope ratio determination on a Delta V isotope ratio MS using a flash elemental analyzer. ThermoFisher Scientific Application Note: 30194 Kroopnick, K.; Craig, H. 1972. Atmospheric oxygen: isotopic composition and solubility fractionation. Science (Washington, DC, U. S.), 175 (4017), 54−55. Lin, J., Bohlke, J., Huang, S., Gonzales-Meler, M., Sturchio, N. 2018. Seasonality of nitrate sources and isotopic composition in the Upper Illinois River. Journal of Hydrology 568 (2019) 849-861. Mulholland, Margaret & Lomas, Michael. 2008. Nitrogen Uptake and Assimilation. 10.1016/B978-0-12-372522-6.00007-4. Murphy, J.; Riley, J. P. 1962. A modified single solution method for the determination of phosphate in natural waters. Anal. Chim. Acta 1962, 27, 31−36. National Atmospheric Deposition Program (NRSP-3). 2019. NADP Program Office, Wisconsin State Laboratory of Hygiene, 465 Henry Mall, Madison, WI 53706. NOAA National Centers for Environmental information. 2019. Climate at a Glance: County Time Series, published November 2019, retrieved on November 15, 2019 from https://www.ncdc.noaa.gov/cag/ Prokopenko, M. G., Hirst, M. B., De Brabandere, L., Lawrence, D. J. P., Berelson, W. M., Granger, J., … & Thamdrup, B. 2013. Nitrogen losses in anoxic marine sediments driven by Thioploca-anammox bacterial consortia. Nature, 500(7461), 194-198. Rubino, R. & Karamavros, E. 2016. Daily Oxygen Fluctuations in Wequetequock Cove. Savoie, J.G., Mullaney, J.R., and Bent, G.C. 2017. Analysis of trends of water quality and streamflow in the Blackstone, Branch, Pawtuxet, and Pawcatuck Rivers, Massachusetts and Rhode Island, 1979 to 2015: U.S. Geological Survey Scientific Investigations Report 2016–5178, 43 p, https://doi.org/10.3133/sir20165178. Sigman, Daniel & Karsh, Kristen. 2009. Ocean process tracers: Nitrogen isotopes in the ocean. Encyclopedia of Ocean Science. Sigman, Daniel & Granger, Julie & Lehmann, Moritz & DiFiore, P. & van Geen, Alexander. 2005. Coupled Nitrogen and Oxygen Isotope Measurements of Nitrate Along the Eastern North Pacific Margin. Global Biogeochemical Cycles. 19. 01-. 10.1029/2005GB002458.

25 Solorzano, L., and J. H. Sharp. 1980. Determination of total dissolved nitrogen in natural waters, Limnol. Oceanogr., 25, 751–754. Standard Methods for the Examination of Water and Wastewater. 2018. 4500-NO3-Nitrogen (Nitrate). DOI: 10.2105/SMWW.2882.089 Standard Methods for the Examination of Water and Wastewater. 2017. 4500-P Phosphorus. DOI: 10.2105/SMWW.2882.093

Standard Methods for the Examination of Water and Wastewater. 2017. 4500-NH3 (Ammonia). DOI: 10.2105/SMWW.2882.087 Stephen D. Sebestyen, Donald S. Ross, James B. Shanley, Emily M. Elliott, Carol Kendall, John L. Campbell, D. Bryan Dail, Ivan J. Fernandez, Christine L. Goodale, Gregory B. Lawrence, Gary M. Lovett, Patrick J. McHale, Myron J. Mitchell, Sarah J. Nelson, Michelle D. Shattuck, Trent R. Wickman, Rebecca T. Barnes, Joel T. Bostic, Anthony R. Buda, Douglas A. Burns, Keith N. Eshleman, Jacques C. Finlay, David M. Nelson, Nobuhito Ohte, Linda H. Pardo, Lucy A. Rose, Robert D. Sabo, Sherry L. Schiff, John Spoelstra, and Karl W. J. Williard. 2019. Unprocessed Atmospheric Nitrate in Waters of the Northern Forest Region in the U.S. and Canada. Environmental Science & Technology 2019 53 (7), 3620-3633 DOI: 10.1021/acs.est.9b01276

Strickland, J. D.; Parsons, T. R. 1972. A Practical Handbook of Seawater Analysis, 2nd ed.; Fisheries Research Board of Canada: Ottawa, Ontario. Thiemes, M. 1999. Mass-independent isotope effects in planetary atmospheres and the early Solar system. Science, 283, 341-345 United States Environmental Protection Agency. 2010. Authorization to Discharge under the Rhode Island Pollutant Discharge Elimination System. Kenyon Industires. Permit No. RI0000191. URIEDC_RIGIS. 2019. Watershed Boundary Dataset HUC 12. Credit: USDA-NRCS, USEPA, RI- DEM. U.S. Bureau of the Census. 2017 TIGER/Line Shapefiles (Area_Water). Washington, D.C.: Bureau of the Census, 2017.

26 U.S. Census Bureau, Population Division. 2020. Annual Estimates of the Resident Population for Counties in Rhode Island: April 1, 2010 to Jul 1, 2019 (CO-EST2019-ANNRES-44) U.S. Geological Survey. 2011. 20140331, NLCD 2011 Land Cover (2011 Edition). U.S. Geological Survey, Sioux Falls, SD. U.S. Geological Survey. 2005. National Atlas of the United States for: Steams and Waterbodies of the United States. Sioux Falls, SD. http://nationalatlas.gov Valiela, I., Foreman, K., LaMontagne, M., Jersh, D., Costa, J., Peckol, P., DeMeo-Anderson, B., D’Avanzo, C., Babion, M., Sham, C., Lajtha, K., 1992. Couplings of Watersheds and Coastal Waters : Sources and Consequences of Nutrient Enrichment in Waquoit Bay , Massachusetts. Estuaries 14, 443–457. Vaudrey, J.M.P., C. Yarish, J.K. Kim, C.H. Pickerell, L. Brousseau, J. Eddings, M. Sautkulis. 2016. Long Island Sound Nitrogen Loading Model. University of Connecticut, Groton, CT. [email protected] Veale, Nate & Visser, Ate & Esser, Bradley & Singleton, Michael & Moran, Jean. 2019. Nitrogen Cycle Dynamics Revealed Through δ18O-NO3− Analysis in California Groundwater. Geosciences. 9. 95. 10.3390/geosciences9020095. Westbrook, H. & Magnano, M. 2016. Dominance of Cladophora Oxygen Demand at Night- Time in Wequetequock Cove. Available at: http://marinesciences.uconn.edu/marn-4001- posters/. Xue, Dongmei & Botte, Jorin & De Baets, Bernard & Accoe, Frederik & Oertel née Nestler, Angelika & Taylor, Philip & Cleemput, Oswald & Berglund, Michael & Boeckx, Pascal. 2009. Present Limitations and Future Prospects of Stable Isotope Methods for Nitrate Source Identification in Surface and Groundwater. Water research. 43. 1159-70. 10.1016/j.watres.2008.12.048.

27 5. Tables and Figures A

B

Figure 1. A) Land cover within the Pawcatuck – Wood River Watershed. Boundary of watershed outlined in black and main channel of Pawcatuck river outlined in dark blue (insert: location of watershed in relation to Southern New England). B) Location of water flow gauges maintained by the USGS (In Kenyon (X), Wood River Junction (Y), and Westerly, RI (Z/Station 12)), sampling stations, and key landmarks along the Pawcatuck River. Where Landmarks include (A) start of the Pawcatuck River at Worden Pond, (B) Usquepaug River and Checkasheen Brook Inlet, (C) Paquisent Brook Inlet, (D) Kenyon Industries, (E) Beaver River Inlet, (F) Shannock Dam, (G) Tanery Brook Inlet, (H) White Brook Inlet, (I) Meadow Brook Inlet, (J) Sandbarrens, (K) Wood River Inlet/ Wood River State Park, (L) Phantom Bog, (M) Poquiant Brook Inlet, (N) Kedinker Island, (O) Newton Swamp, (P) Chapman Pond Inlet, (Q) Mile Brook Inlet, (R) Ashaway River Inlet, and (S) Shunock River Inlet.

28

Figure 2. Mean daily river discharge over the sampling period at three respective flow gauges on the river. Round markers indicate dates of weekly samplings at the Stillman and Westerly bridges and square markers denote river transects.

29

- 15 Figure 3. [NO3 ] at Stillman (S) and Westerly (W) bridges vs. (A) sampling date and (B) river discharge. δ NNO3 at the bridges vs. 18 17 (C) sampling date and (D) river discharge. δ ONO3 vs. (E) date and (F) river discharge. Δ ONO3 vs. (G) date and (H) river + discharge. [DON] vs. (I) date and (J) river discharge. [PN] vs. (K) date and (L) river discharge. [NH4 ] vs. (M) date and (N) river 3- discharge. [PO4 ] vs. (O) date and (P) river discharge. Mean daily river discharge recorded at the Stillman Bridge is plotted in black (A, C, E, G, I, K, M, O).

30 Table 1. Linear Regressions with statistical analysis, used to interpret graphs where all slopes with a p-value over 0.05 have a dashed line to indicate the significance.

Figure Location X Y Slope Intercept r2 p-value P≤0.05 - 3B Stillman Discharge [NO3 ] -9.6 51.7 0.53 < 0.01 . - 3B Westerly Discharge [NO3 ] -7.9 47.8 0.44 < 0.01 . 15 3D Stillman Discharge d NNO3 -0.3 7.9 0.16 < 0.01 . 15 3D Westerly Discharge d NNO3 -0.4 8.1 0.27 < 0.01 . 18 3F Stillman Discharge d ONO3 0.5 2.8 0.35 < 0.01 . 18 3F Westerly Discharge d ONO3 0.6 2.6 0.29 < 0.01 . 17 3H Stillman Discharge ∆ ONO3 0.1 -0.1 0.15 0.01 .

3H Westerly Discharge % atm. 0.1 1.3 0.00 0.87 + 3J Stillman Discharge [NH4 ] 0.6 0.5 0.27 < 0.01 . + 3J Westerly Discharge [NH4 ] 0.7 0.7 0.29 < 0.01 . 3- 3L Stillman Discharge [PO4 ] 0.0 0.8 0.00 0.78 3- 3L Westerly Discharge [PO4 ] -0.1 1.0 0.09 0.05

3N Stillman Discharge [DON] 1.3 26.4 0.02 0.39

3N Westerly Discharge [DON] -1.2 27.8 0.01 0.64

3P Stillman Discharge [PN] 0.3 2.4 0.06 0.14

3P Westerly Discharge [PN] -0.1 3.5 0.01 0.59

4 Stillman [DIN] [DON] -0.2 34.0 0.17 0.01 .

4 Westerly [DIN] [DON] -0.3 33.7 0.08 0.10 - Stillman [NO3 ] [DON] -0.2 34.4 0.05 0.18 - -2 Westerly [NO3 ] [DON] 3.0 x 10 26.0 0.00 0.80 - -2 5B WWTF comp WWTF discharge [NO3 ] -2.7 x 10 514 0.43 < 0.01 . - -2 5B WWTF grab WWTF discharge [NO3 ] -3.0 x 10 545 0.34 < 0.01 . - -2 5B WWTF comp reported WWTF discharge [NO3 ] -2.4 x 10 439 0.31 < 0.01 . + -2 5D WWTF comp WWTF discharge [NH4 ] 2.8 x 10 -114 0.12 0.07 + -2 5D WWTF grab WWTF discharge [NH4 ] 2.3 x 10 -65.5 0.06 0.13 + -2 5D WWTF comp reported WWTF discharge [NH4 ] -4.5 x 10 -257 0.27 < 0.01 . -2 5F WWTF comp WWTF discharge [DON] 2.6 x 10 -8.3 0.10 0.06 -2 5F WWTF grab WWTF discharge [DON] 1.3 x 10 10.3 0.08 0.31 3- -2 5H WWTF comp WWTF discharge [PO4 ] 5.0 x 10 8.0 0.05 0.30 3- -2 5H WWTF grab WWTF discharge [PO4 ] 4.0 x 10 3.7 0.08 0.12 4 6B Stillman Discharge DIN flux 2.6 x 10 Forced zero 0.85 < 0.01 . 4 6B Westerly Discharge DIN flux 2.7 x 10 Forced zero 0.86 < 0.01 . 4 6D Stillman Discharge DON flux 3.0 x 10 Forced zero 0.91 < 0.01 . 4 6D Westerly Discharge DON flux 2.7 x 10 Forced zero 0.94 < 0.01 .

6F Stillman Discharge TN flux 602 Forced zero 0.93 < 0.01 .

6F Westerly Discharge TN flux 598 Forced zero 0.95 < 0.01 . - 15 4 9A Stillman 1/(NO3 flux) d NNO3 2.6 x 10 6.6 0.43 < 0.01 . - 15 4 9A Westerly 1/(NO3 flux) d NNO3 5.5 x 10 6.2 0.57 < 0.01 . - 18 4 9B Stillman 1/(NO3 flux) d ONO3 -1.5 x 10 4.2 0.08 0.07 - 18 4 9B Westerly 1/(NO3 flux) d ONO3 -4.5 x 10 4.7 0.16 0.04 . - 18 4 9C Stillman 1/(NO3 flux) d ONO3-corr 1.7 x 10 3.7 0.58 <0.01

31

Figure 4. [DON] vs. [DIN] at Stillman (S) and Westerly (W) bridges.

32

- + Figure 5: Nutrients (A) [NO3 ] vs. date, and (B) [NH4 ] vs. date of rainwater samples collected at Avery Point, Groton, CT. Isotopic 15 18 17 analysis of (C) δ NNO3, (D) δ ONO3, and D ONO3 vs. date at Avery Point, Groton, CT. Mean daily river discharge recorded at the Stillman Bridge is plotted in black on all charts.

33

- Figure 6. Composite (Co), Grab (Gr) and lab reported (Re) Nutrients at Westerly WWTF. [NO3 ] vs. (A) date and (B) facility + 3- discharge. [NH4 ] vs. (C) date and (D) facility discharge. [DON] vs. (E) date and (F) facility discharge. [PO4 ] vs. (G) date and (H) facility discharge. Grey line signifies WWTF mean daily average discharge rate (A, C, E, G).

34

Figure 7. (A) DIN fluxes vs. date at the Stillman (S) and Westerly (W) brides and from the Westerly wastewater treatment facility (WWTF). (B) Riverine DIN flux vs. river discharge at the Stillman and Westerly bridges. (C) DON fluxes vs. date at the Stillman and Westerly brides and from the WWTF. (D) Riverine DON flux vs. river discharge at the Stillman and Westerly bridges. (E) Riverine TN flux vs. date and (F) river discharge at the Stillman and Westerly bridges.

35

- 15 18 3- Figure 8. (A) River discharge, (B) [NO3 ], (C) δ NNO3, (D) δ ONO3, and (E) [PO4 ] vs. downstream from origin at Worden pond on three sampling campaigns in 2018 - 2019. Sampling station identifiers Open symbols denote samples collected from tributaries discharging into the Pawcatuck river, dashed lines denote flow gauges at Kenyon, RI (X), Wood River Junction, RI (Y), and Westerly, RI (Z) respectively, letters denote key landmarks, pink outlines denote inlet onto river. Where Landmarks include (A) start of the Pawcatuck River at Worden Pond, (B) Usquepaug River and Checkasheen Brook Inlet, (C) Paquisent Brook Inlet, (D) Kenyon Industries, (E) Beaver River Inlet, (F) Shannock Dam, (G) Tanery Brook Inlet, (H) White Brook Inlet, (I) Meadow Brook Inlet, (J) Sandbarrens, (K) Wood River Inlet/ Wood River State Park, (L) Phanotm Bog, (M) Poquiant Brook Inlet, (N) Kedinker Island, (O) Newton Swamp, (P) Chapman Pond Inlet, (Q) Mile Brook Inlet, (R) Ashaway River Inlet, and (S) Shunock River Inlet.

36

100%

80%

60%

40%

20%

0% May November March

Section1 Section2 Section3

Figure 9. Percentage contribution of river sections to the total riverine DIN flux recorded at the Westerly bridge during downriver sampling trips.

37 Table 2. DIN Flux per section of the Pawcatuck river.

DIN flux (moles N d-1) Station 2 Station 4 Station 12 Date (7 km) (15 km) (48 km) May 21, 2018 4.0 x 103 1.0 x 104 2.9 x 104 November 9, 2018 1.5 x 103 1.1 x 104 1.4 x 104 March 12, 2019 2.6 x 104 1.9 x 104 4.1 x 104

38 - Table 3. NO3 , DIN, and DON loading from Westerly WWTF during warm/ lower river discharge (May – September).

Min Max Mean Median Plant discharge (106 m3 d-1) 0.6 x 10-2 1.4 x 10-2 0.9 x 10-2 0.7 x 10-2 - NO3 Flux (moles N d-1) 1.9 x 103 3.5 x 103 2.5 x 103 2.5 x 103 DIN Flux (moles N d-1) 1.9 x 103 3.8 x 103 2.8 x 103 2.7 x 103 DON Flux (moles N d-1) 9.2 x 101 2.0 x 103 8.6 x 102 8.1 x 102

39

15 18 18 Figure 10. Keeling Plots of NO3 flux for (A) δ NNO3, (B) δ ONO3 and (C) Rain corrected δ ONO3 for Westerly and Stillman Bridges.

40

- Figure S1. Log conductivities at bridges vs. (A) date and (B) river discharge. (C) [NO3 ] vs log conductivities at bridges.

41 1 2 3

4 5 6

7 8 9

10 11 12

- 3- 15 18 Figure S2. Results of TukeyHSD analysis on (1, 2, 3) [NO3 ], (4, 5, 6) [PO4 ], (7, 8, 9) δ NNO3, and(10, 11, 12) δ ONO3 vs. distance downstream for each sampling trip in May 2018 (1, 4, 7, 10), November 2018 (2, 5, 8, 11), and March 2019 (3, 6, 9, 12).

42

- - Figure S3. (A) [NO3 ] in composite vs. grab samples from the Westerly WWTF, measured by UConn. (B) Plant-reported [NO3 ] - + composite effluent vs. [NO3 ] measured at UConn. (C) [NH4 ] in composite vs. grab samples from the WWTF, measured by + + UConn. (D) Plant-reported [NH4 ] composite effluent vs. [NH4 ] measured at UConn.

43

Figure S4: Mixing curve of [DIN] with river discharge at Stillman (S) and Westerly (W) Bridges.

44