Coastal waters globally are increasingly impacted due to the anthropogenic
loading of nitrogen (N) from the watershed. To assess dominant sources
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
(NO3-) isotope ratios (15N /14N, 18O /16O, and
17O /16O) at the mouth of the river and from the larger of two
wastewater 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 more important at lower river flows during
warmer months, comprising ∼ 23 % and ∼ 18 % of N loading,
respectively. Riverine nutrients derived predominantly from deeper
groundwater and the industrial point source upriver in summer and from
shallower groundwater and surface flow during colder months – wherein
NO3- associated with deeper groundwater had higher
15N /14N ratios than shallower groundwater. Corresponding
NO3-18O /16O ratios were lower during the warm season,
due to increased biological cycling in-river. Uncycled atmospheric
NO3-, detected from its unique mass-independent NO3-17O /16O vs. 18O /16O fractionation, accounted for
< 3 % of riverine NO3-, even at elevated discharge.
Along-river, NO3-15N /14N ratios showed a correspondence
to regional land use, increasing from agricultural and forested catchments
to the more urbanized watershed downriver. The evolution of
18O /16O isotope ratios along-river conformed to the notion of
nutrient spiraling, reflecting the input of NO3- from the
catchment and from in-river nitrification and its coincident removal by
biological consumption. These findings stress the importance of considering
seasonality of riverine N sources and loading to mitigate eutrophication in
receiving estuaries. Our study further advances a conceptual framework that
reconciles with the current theory of riverine nutrient cycling, from which
to robustly interpret NO3- isotope ratios to constrain cycling and
source partitioning in river systems.
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 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 wastewater treatment facilities (WWTFs), septic systems,
industrial discharge, fertilizer applied to turf and agricultural lands, and
atmospheric sources from industry and fossil fuel use (Valiela et al., 1997;
McClelland et al., 2003; Latimer and Charpentier, 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 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 (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 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 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. NO3- is otherwise removed from
anoxic groundwater by denitrification, reduced to inert N2. Reactive N
is further cycled and attenuated in-river. The hyporheic zone, where
groundwater interchanges with stream and river water, creates a complex
environment that can stimulate nitrification and denitrification, as oxic
and anoxic pockets exist in close proximity (Sebilo et al., 2003; Harvey et
al., 2013). Reactive nitrogen can be further attenuated by benthic
denitrification within the river channel (Sebilo et al., 2003; Kennedy et
al., 2008; Mulholland et al., 2008).
Identifying sources of N to rivers can be difficult due to the expanse and
heterogeneity of the watershed, the long integration time of deeper
groundwater, and the degree of biological N cycling in groundwater and
in-river. While measurements of N concentrations along the river channel in
relation to regional land use can offer insights in this regard, N sources
can be further resolved using complementary measurements of the naturally
occurring N and oxygen (O) isotope ratios of riverine NO3-
(15N /14N and 18O /16O, respectively). Henceforth, we
express the isotope ratios in delta notation:
δ‰=isotope
ratio of sampleisotope ratio of reference-1×1000.
The reference for δ15N is N2 in air and for δ18O is Vienna Standard Mean Ocean Water (VSMOW). The N and O isotope
ratios of NO3- provide constraints on N sources and cycling in
part because respective N sources cover discrete ranges of δ15N
and δ18O values (Kendall et al., 2007). Reactive N species from
atmospheric deposition, biological N2 fixation, and industrial N2
fixation share overlapping ranges of δ15N values (≤0 ‰), which differ appreciably from those of livestock
and human waste (8 ‰–25 ‰; Kendall, 1998; Böhlke,
2003; Xue et al., 2009). In contrast, the δ18O signatures of
atmospheric NO3- (60 ‰–80 ‰) are distinct
from those of industrial NO3- (∼ 25 ‰)
and from NO3- produced by nitrification, which aligns closely with
that of ambient water (Boshers et al., 2019, and references therein). Atmospheric NO3- is
further distinguished by a mass-independent δ17O vs. δ18O fractionation that is not manifest in industrial and biological
NO3- (Savarino and Thiemens, 1999).
The isotope ratios of NO3- also provide constraints on N cycling
because N and O isotopologues are differentially sensitive to respective
biological N transformations (reviewed by Casciotti, 2016), implicating different mass
balance considerations within the N cycle that permit differentiation of N
sources from cycling. Briefly, in riverine systems where NO3- is
the dominant N pool, δ15NNO3 integrates across values of
reactive N delivered from the watershed, minus NO3- removed by
benthic denitrification (if associated with N isotopic fractionation; Sebilo
et al., 2003). Values of δ15NNO3 are additionally
sensitive to isotopic fractionation due to internal cycling in-river –
assimilation and remineralization to NO3- via nitrification – in
systems where riverine N is otherwise partitioned comparably between
oxidized and reduced pools (i.e., NO3- vs. ammonium and particulate N;
Sebilo et al., 2006). Riverine δ18ONO3, in turn,
integrates across values of exogenous NO3- delivered to the river
from the watershed and from atmospheric deposition, those of NO3-
produced in-river by nitrification, minus the NO3- lost
concurrently to denitrification and assimilation (see Sigman and Fripiat,
2019). Interpreted in tandem, NO3- N and O isotopologue ratios
thus offer complementary constraints to identify important source terms and
characterize cycling.
Here we present a study of annual N loading from the Pawcatuck River to the
Little Narragansett Bay in southern New England (USA), wherein we exploit
measurements of the N and O isotope ratios of riverine NO3- to
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: vast seagrass beds of Zostera marina (eelgrass) historically
established in Little Narragansett Bay were overtaken in the early 1990s by
extensive mats of filamentous macroalgae dominated by the Cladophoraceae clade, whose
substantial biomass has been linked to frequent events of nighttime 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 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
NO3- isotopologue 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 measurements of samples collected from seasonal along-river
surveys. Utilizing NO3- isotopologue 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 temperate
estuaries.
MethodsSite description
The Pawcatuck River watershed (∼ 760 km2) is located predominantly
in the state of Rhode Island (RI) with a small portion in eastern
Connecticut (Fig. 1). The river originates at Worden Pond in 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 flat, hosting terrain with forests
and wetlands (73 %) and relatively low human population (∼ 56 400;
based on a dasymetric analysis of the 2010 US Census Bureau population
data in the watershed; Vaudrey et al., 2017) – owing in part to state land
trust holdings that protect ∼ 22 % of the watershed in RI from
development (Dillingham et al., 1992; US Geological Survey, 2011).
Agricultural areas comprise 8 % of land use (US Geological Survey,
2011) and are mostly located in the upper watershed, which hosts a number of
turf farms. In 2005, Washington county – where the Pawcatuck River
originates – was noted as having the highest density of turf farms in the
United States (US Environmental Protection Agency, 2005). Urbanized and
developed land usage comprises 13 % of the total watershed with the
majority of the urban areas concentrated on the lower 19 km portion of the
river, between Bradford and Westerly (US Geological Survey, 2011).
(a) Map of the Pawcatuck River watershed and associated land use
(URIEDC_RIGIS, 2019; US Geological Survey, 2011). (b) Map
of sampling locations, landmarks including the Westerly and Pawcatuck wastewater treatment facilities (WWTFs), and river discharge gauges along the
Pawcatuck River (US Geological Survey, 2005; US Census Bureau, 2017).
Three nutrient discharge permits are allotted along the river by the RI
Department of Environmental Management (RI DEM; Fig. 1): Kenyon
Industries, a fabric processing plant, is located approximately 7 km
downstream from Worden Pond. Two WWTFs discharge into the estuary and are
located 1 km downstream of the Westerly Bridge, approximately 47 km
downstream of Worden Pond.
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 samplings. We collected weekly river
samples (a) from 10 January 2018 through to 12 January 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 (Figs. 1, S1). (b) We obtained samples of wastewater
treatment effluent collected weekly at the Westerly Wastewater Treatment
Facility (W-WWTF) from 6 June 2018 to 22 May 2019. (c) We conducted three
seasonal along-river surveys on 21 May, 9 November 2018, and 12 March 2019 at 15 discrete sampling stations between Worden Pond and the
Westerly Bridge. Additionally, we performed a highly resolved sampling
(approximately every 0.75 km) of the lower river from Potter Hill Dam
(Station 11) to Westerly (Station 15) aboard kayaks in May 2017 (Fig. 1).
No samples were collected directly from the retention ponds or their outflow
at Kenyon Industries (Fig. 1). (d) We collected rainwater samples
following rain events from a rooftop collector at the Avery Point campus in
Groton, CT (approximately 18 km west of the Pawcatuck River), from 6 September to 2 December 2018, in order to define regional NO3-
isotopic end-members.
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 midday. During each
sample collection, river temperature and dissolved oxygen concentrations
were 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 NO3- isotope
ratios were filtered through pre-combusted 25 mm GF/F glass 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 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 h 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
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 NO3- isotope ratio analyses.
Nutrient analyses
The NO3- concentration, [NO3-], in river and WWTF
samples was measured by conversion to nitric oxide in a hot Vanadium(III)
solution followed by detection on a chemiluminescent NOx analyzer
(Teledyne™; Braman and Hendrix, 1989). Incident nitrite in the samples
was first reacted with Griess reagents (Strickland and Parsons, 1972) before
injection into the hot Vanadium(III) solution in order to detect
NO3- only. The concentration of nitrite, [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). For the rainwater samples, [NO3-] and
[NO2-] were measured on a SmartChem discrete nutrient autoanalyzer
(Unity Scientific™) using standard protocols adapted for the SmartChem instrument
(Strickland and Parsons, 1972; US Environmental Protection Agency, 1993b;
4500-NO2-, 2018; 4500-NO3-, 2018). Concentrations of
ammonium, [NH4+], and phosphate, [PO43-], in river and
WWTF samples were measured on a SmartChem autoanalyzer using standard
protocols (Murphy and Riley, 1962; Strickland and Parsons, 1972; US
Environmental Protection Agency, 1978, 1993a; 4500-NH3, 2018; 4500-P, 2018).
The concentration of total dissolved nitrogen, [TDN], in filtered river and
WWTF samples was measured by persulfate oxidation to NO3- and then
measured via a 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 ≤ 0.3 % of the TDN signal. The concentration of
dissolved organic nitrogen, [DON] was calculated as the difference between
[TDN] and dissolved inorganic nitrogen, [DIN], where [DIN] = [NO3-] + [NO2-] + [NH4+].
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 fluorescence detection on a Turner Designs Trilogy laboratory fluorometer (Arar
and Collins, 1997).
NO3- isotope ratio analyses
The nitrogen and oxygen isotope ratios of NO3-, 15N /14N,
18O /16O, and 17O /16O were analyzed using the denitrifier
method in samples where [NO3-] ≥ 1.5 µM (Sigman et al.,
2001; Casciotti et al., 2002; Kaiser et al., 2007). 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
gas chromatograph–isotope ration mass spectrometer (GC-IRMS) prefaced with a custom-modified Gas Bench II device with two cold
traps and a PAL autosampler (Casciotti et al., 2002). The NO3-17O /16O in rainwater (as well as 18O /16O) 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).
Coupled δ15NNO3 and δ18ONO3 analyses at
UConn and Brown University were calibrated from parallel analyses of
NO3- reference materials USGS-34 (δ15N: -1.8 ‰ vs. air; δ18O: -27.9 ‰ vs. VSMOW) and IAEA-N3 (δ15N:
+4.7 ‰ vs. air; δ18O: +25.6 ‰ vs. VSMOW). Samples were analyzed in triplicate among
two or more batch analyses. Reproducibility averaged
0.2 ‰ for δ15NNO3 and
0.3 ‰ for δ18ONO3. Coupled analyses of
δ18ONO3 and δ17ONO3 of rainwater
NO3- and some of the river samples were calibrated with USGS-34
(Δ17O: -0.1 ‰ vs. VSMOW) and USGS-35
(δ18O +57.5 ‰ vs. VSMOW; Δ17O: +21.6 ‰ vs. VSMOW). The mass-independent
fractionation of NO3-17O vs. 18O (Δ17O vs.
VSMOW) is calculated from Thiemens (1999):
Δ17O=δ17O-0.52×δ18O.
The analytical reproducibility for Δ17ONO3 averaged
0.3 ‰ based upon the pooled standard deviation of
repeated measures of reference materials. The fraction (%) of atmospheric
NO3- in river water was derived from a two-end-member mixing
equation of river water NO3- (Δ17O = 0) with the
corresponding atmospheric NO3-Δ17O value (19.7 ‰ to
27.2 ‰; Sect. S1), with an associated uncertainty of
∼ 1 % based on the pooled standard deviations of Monte Carlo error
propagations.
Particulate nitrogen analyses
Particulate nitrogen (PN) in river samples was collected on pre-combusted 25 mm GF/F glass fiber filters that were freeze-dried and 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. Samples
were calibrated with aliquots of recognized reference materials USGS 40 and
41 (δ15N =-4.52 ‰ and +47.57 ‰ vs. air,
respectively), achieving an analytical precision of ∼ 0.3 ‰.
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 concentration data, annual fluxes (L) of respective
constituents were calculated using Beale's ratio estimator (Beale, 1962;
Quiblé et al., 2006), which includes a bias correction factor (the term
in parentheses) to account for the covariance between load and river flow
values (Eq. 3):
3aL=CQ‾μqQ‾n1+1ndSCQCQ‾⋅Q‾1+1ndSQ2Q‾2,3bSCQ=1nd-1∑i=1ndCiQi-ndQ‾CQ‾,3cSQ2=1nd-1∑i=1ndQi-ndQ‾2.
The term μq is the mean of all river discharge measurements,
Ci 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 the number of observations of Ci. Overbars denote sample
arithmetic means. We further conducted bootstrap analyses to provide
estimates of the uncertainty of the respective fluxes.
ResultsWeekly river samplings
The concentration of NO3- measured 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 73.5 µM, with a median
value of 30.4 µM (Fig. 2a). Comparable concentrations were detected
at the Westerly Bridge at each sampling, except for instances where the site
experienced saltwater intrusions, evidenced by elevated conductivities (data
not shown) – at which times [NO3-] at the Westerly Bridge was
lower due to lower concentrations in the seawater end-member. The
concentration of NO2- was negligible in all samples. At both
bridge sites, [NO3-] decreased with increasing river discharge
(Fig. 2b; Table 1). The [NO3-] at the Stillman Bridge, upstream
of potential seawater intrusion, also correlated directly with conductivity
(Fig. 3a). Values of δ15NNO3 were lowest in winter and
increased in summer, ranging from 5.3 ‰ to 9.4 ‰ – thus decreasing with increasing river discharge
(Fig. 2c–d; Table 1). Values of δ18ONO3 followed a
contrasting trend, being lower during the summer months and increasing in
winter months, with values ranging from 1.6 ‰ to 6.8 ‰, notwithstanding a single outlying value of 8.1 ‰ (Fig. 2e). Values of δ18ONO3 at
the bridges increased directly with discharge (Fig. 2f; Table 1).
Measurements of Δ17ONO3 at the Stillman Bridge ranged from
-0.5 ‰ to 1.9 ‰. The fraction of atmospheric
NO3- estimated based on the Δ17ONO3 values of
precipitation recorded at Avery Point (Sect. S1; Fig. S2) indicated that
uncycled atmospheric NO3- was 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 NO3- was otherwise < 3%,
notwithstanding a single sample in which atmospheric NO3-
accounted for ∼ 7 % of total riverine NO3- (Fig. 2g).
Values of Δ17ONO3 nevertheless correlated with river
discharge (Fig. 2h; Table 1).
Weekly measurements of solute concentrations and NO3-
isotopologue 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 in the left-hand
panels is the river discharge (×106 m3 d-1).
[NO3-] vs. (a) sampling date and (b) discharge; δ15NNO3 vs. (c) sampling date and (d) discharge; δ18ONO3 vs. (e) sampling date and (f) discharge; Δ17ONO3 vs. (g) sampling date and (h) discharge; [NH4+]
vs. (i) sampling date; (j) [NH4+] vs. discharge;
[PO43-] vs. (k) sampling date and (l) discharge; [DON] vs. (m) sampling date and (n) [DON] vs. discharge. Statistical fits of least-square
linear regressions are reported in Table 1.
Correlation coefficients, corresponding intercepts, coefficients of
determination (r2), and statistical probability of least-squared
regression analyses from property–property plots of riverine solutes and
fluxes. Statistically significant relationships are signaled by an asterisk
(p value ≤0.05*; ≤0.01**).
The concentration of NH4+ recorded weekly at the bridges was
consistently lower than corresponding [NO3-]. In contrast to
[NO3-], [NH4+] at the bridges was lowest in summer and
higher in winter, ranging from below detection to 7.8 µM, and
correlated directly with discharge (Fig. 2i–j). The [PO43-]
ranged from 0.1 to 2.7 µM with one sample as high as 5.9 µM during a single sampling event, exhibiting higher concentrations
occurring in summer months, thus mirroring [NO3-] (Fig. 2k).
Concentrations of PO43- appeared to correlate inversely with
discharge, yet only at the Westerly Bridge and not the Stillman Bridge
(Fig. 2l; Table 1).
The concentration of DON at the bridge sites ranged from 9 to 56 µM,
appeared similar among seasons, and did not show a statistically significant
relationship to river discharge (Fig. 2m–n; Table 1). Nevertheless, [DON]
and coincident [DIN] were inversely correlated, albeit weakly so, and
significantly so only at the Stillman Bridge (Fig. 3b; Table 1). In turn,
[PN] exhibited median values of 2.6 µM from May through October and 2.9 µM during the colder season, showing no values greater than 7 µM; no correlation of [PN] with river discharge was evident (Fig. S3a–b;
Table 1). Concentrations of chlorophyll a, which we measured only from June
through December, ranged from 0.5 to 12.1 µg L-1, with higher values occurring in late summer to early fall.
Chlorophyll a showed no correlation with discharge (Fig. S3e–f; Table 1).
(a) [NO3-] in weekly samples at the Stillman Bridge vs.
conductivity (log scale). (b) Weekly [DON] at the Stillman and Westerly
bridges vs. [DIN].
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+] recorded at
the bridges, varied ∼ 10-fold over the annual sampling period, ranging
from 0.1 to 1.1 (×105) mol of NDIN per day – omitting a single
outlier of 1.8×105 mol of NDIN per day (Fig. 4a). The
riverine DIN flux increased directly with river discharge, such that it was
lowest in summer, averaging 0.2 ± 0.1 (×105) mol of NDIN
per day from May through October (Fig. 4b; Table 1). The riverine DON
flux, in turn, ranged from < 0.1 to 2.0 (×105) mol of
NDON per day and also increased directly with discharge (Fig. 4c–d;
Table 1). The total riverine N flux (TN flux), which is the sum of
respective DIN, DON, and PN fluxes, ranged from 0.2 to 3.0 (×105) mol
of NTN per day and correlated directly with discharge (Fig. 4e–f;
Table 1).
Weekly estimates of N fluxes at the Stillman and Westerly bridges
and from the Westerly WWTF vs. sampling date (superposed onto river
discharge) and vs. the mean daily river discharge recorded at the Stillman
Bridge. The secondary axis in the left-hand panels is the river discharge (×106 m3 d-1). DIN flux vs. (a) sampling date and (b) mean
daily river discharge; DON flux vs. (c) sampling date and (d) mean daily
river discharge; TN flux (the sum of DIN, DON, and PN) vs. (e) sampling date
and (f) mean daily river discharge. Statistical fits of least-square linear
regressions are reported in Table 1.
WWTF samples
Nutrient concentrations measured in samples collected weekly at the Westerly
WWTF, consisting of both grab and composite samples, ranged from 30 to 527
µM for [NO3-], 1.3 to 1070 µM for [NH4+],
11.7 to 1168 µM for [DON], and 2.7 to 26.5 µM for
[PO43-] (Fig. 5a, c, e, g). Concentrations of NO3- and
NH4+were similar in grab vs. composite samples (Fig. S4a–b).
The [NO3-] and [NH4+] measured at UConn were similar to
those reported by the Westerly WWTF (Fig. S4c–d). The [DON] measured at
UConn showed poor correspondence to the facility-reported [TON] (total
organic nitrogen) for the few corresponding sampling dates, although these
sample types may arguably not be comparable as the UConn analyses did not
include [PN] (Fig. S4e).
Nutrient discharged from the Westerly WWTF: [NO3-] vs.
(a) date and (b) facility discharge; [NH4+] vs. (c) date and (d) facility discharge; [DON] and facility-reported [TON] vs. (e) date and (f) facility discharge; [PO43-] vs. (g) date and (h) facility
discharge. The grey line corresponds to the WWTF mean daily water discharge with
reference to the secondary axis in the left-hand panels (×104 m3 d-1). Statistical fits of least-square linear regressions are reported
in Table 1.
Both [NO3-] and [PO43-] were higher in summer months
when facility discharge was lower, at which time [NH4+] was lower.
The concentration of NO3- correlated inversely with the
facility-reported discharge, whereas [NH4+] correlated directly
with discharge (Table 1). There was an apparent increase in [DON] with
discharge, albeit with high variability during high flow in winter months,
whereas facility-reported [TON] did not correlate with discharge (Fig. 5f;
Table 1). Our limited [PO43-] measurements were not significantly
correlated with facility-reported discharge (Fig. 5h, j; Table 1).
In contrast to the riverine N fluxes, which increased with river discharge,
the DIN and TON fluxes from the Westerly WWTF were remarkably constant and
were substantially lower than corresponding riverine fluxes, averaging 3.2×103 mol of NDIN per day, 1.0×103 mol of NTON per
day, and 4.1×103 mol of NTN per day in 2018 (Fig. 4a–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 13.5×104 mol of NTN per day.
Along-river samplings
Samples collected at stations along the length of the river showed both
spatial and seasonal patterns in nutrients and NO3- isotope ratios
(Figs. 6, S5). On average, [NO3-] differed among sampling dates
(F2,12=122.4, p< 0.0001) and was lower during the November 2018 sampling than during the May 2018 and March 2019 samplings at all river
sites (Tukey honestly significant difference (HSD), both p< 0.05; Fig. 6b; Table S2).
[NO3-] tended to increase along river sections (F3,9=32.1, p< 0.0001), but the specific patterns varied among sampling
dates (F6,12=107.2, p< 0.0001). In the source basin at
Worden Pond, [NO3-] ranged from 0.4 to 6 µM among sampling
events and increased to date-specific maxima of 10 to 65 µM between
Stations 2 and 4 (Biscuit City Road to Wood River Junction). Concentrations
decreased downstream of the Wood River inflow (between river sections 2 and
3) to values as low as 7 µM in November and as high as 29 µM
in March (Tukey HSD, both p< 0.01), although these two sections of
the river were similar in May 2018 (Tukey HSD, both p>0.9).
[NO3-] increased between Potter Hill Dam (Station 11) and the
Stillman and Westerly bridges (between sections 3 and 4) during all sampling
campaigns (Tukey HSD, all p< 0.05), with final concentrations of 10 µM in November and 32 µM in March (Fig. 6b).
Solute concentrations and nitrate isotopologue ratios observed
along the Pawcatuck River from its origin at Worden Pond to the Westerly
Bridge during along-river sampling campaigns. (a) Mean daily river discharge
recorded at three flow gauges along the Pawcatuck River during the sampling
campaigns: (b) [NO3-], (c)δ15NNO3, (d)δ18ONO3, and (e) [PO43-] measured at stations
along-river. Open circles designate the lower Wood River that inflows to the
Pawcatuck.
Values of δ15NNO3 differed among sampling dates
(F2,15=16.6, p< 0.001) and along the river (F3,9=26.2, p< 0.001; Fig. 6c). On average, values were lower in March 2019, at which time [NO3-] was relatively elevated, than in May
and November 2018 (Tukey HSD: both p< 0.001), although a sample in
the uppermost river section (Station 2) had higher δ15NNO3
values in March than in May (see Discussion). Values during the March 2019
sampling ranged from 3.4 ‰ at Station 3 to 6.7 ‰ at the bridges (river Sect. 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. NO3- delivered by the Wood River (Station 6) had δ15NNO3 values similar to or greater than those of NO3-
originating upstream in the Pawcatuck River.
In contrast to δ15NNO3, δ18ONO3 values
tended to decrease downriver (F3,9=8.6, p< 0.01; Fig. 6d), despite relatively large variability. Relative maxima between 3.2 ‰ and
5.0 ‰ were apparent at Stations 3 and 4 (river section 2), decreasing to values oscillating between 2.7 ‰ and 4.8 ‰ toward the bridges (F3,9=8.6, p< 0.01; Fig. 6d). The δ18ONO3 values upriver were
generally higher in November (in contrast to δ15NNO3) 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.
The concentration of NH4+ did not vary systematically across river
sections (F3,9=3.2, p= 0.08), ranging from 0.4 to 6.8 µM
(Fig. S5). [NH4+] was greater during the May 2018 sampling than
during March or November, and this effect did not vary significantly across
the river (F6,12=2.1, p= 0.12).
The concentration of PO43- varied over both space (F3,9=45.2, p< 0.0001) and time (F2,12=72.0, p< 0.0001), and these effects were non-additive (F6,12=32.2, p< 0.0001). [PO43-] was relatively homogeneous across the
river sections in May 2018 (Tukey HSD, all pairwise p>0.05) but
increased in river section 2 in both March and November compared to
neighboring sections up- and downstream (Tukey HSD, all p< 0.01;
Fig. 6e). Across all sampling dates, [PO43-] ranged from 0.2 to
0.5 µM in and near Worden Pond (river section 1) and peaked at values
between 0.7 and 2.9 µM, in river section 2. Further downstream,
[PO43-] ranged from 0.5 to 1.2 µM. [PO43-] in
the Wood River (Station 6) was relatively low and similar to that at Worden.
DiscussionNutrient source attribution
At the Stillman and Westerly bridges, concentrations of NO3- –
the principal component of DIN – scaled inversely with discharge, wherein
higher concentrations occurred during summer at low base flow. This
relationship suggests the bulk of riverine DIN during low base flow
originated from groundwater and point sources along the river catchment.
Given that there is only one documented point source upstream of Stillman
and Westerly bridges, we surmise that DIN at low base flow originated
predominantly from groundwater and partially from discharge at Kenyon
Industries. That [NO3-] at the Stillman Bridge increased in
proportion to conductivity also suggests a groundwater source for bulk
riverine nutrients at low base flow, although an analogous trend could
admittedly arise from loading by a point source.
During wetter months in winter, increased input of shallow groundwater and
surface runoff (henceforth collectively referred to as “shallow flow”)
diluted the low-base-flow [NO3-], thus lowering riverine
concentrations, a dynamic documented in other temperate rivers (Mulholland
and Hill, 1997; Dubrovski et al., 2010). Nevertheless, the daily DIN flux
increased with discharge, indicating that DIN is also imported to the river
by shallow flow, albeit at a lower concentration than low-base-flow DIN.
From the slope of the DIN flux-to-discharge relationship, the daily DIN flux
increased by 2.6 ± 0.3 (×104) mol of N per additional 106 m3 of discharge, suggesting the DIN concentration of shallow flow from
the catchment averaged 26 ± 3 µM (Table 1). The relationship
between [DIN] and river discharge, which we initially presumed linear, is
then better described by a two-end-member mixing curve comprised of low-base-flow [DIN] mixing with shallow-flow [DIN]:
[DIN]i=[Qi-2.2×108⋅26×10-6+2.2×108⋅64×10-6]/Qi.
The term Qi is mean river discharge on day i in units of L d-1, from
which the subtracted value of 2.2×108 L d-1 is the asymptote of
low base flow (i.e., the lowest river discharge observed in 2018), and
[DIN]i is the corresponding concentration in units of mol L-1.
The low-base-flow [DIN] end-member of 64 ± 9 µM derives from
the best fit of the equation to the data. 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. The annual DIN load returned
by this function based on the continuous (15 min interval) discharge records
at the Stillman Bridge for 2018 is 21.4×106 mol N yr-1, slightly
higher than our direct estimate of 20.2 (±2.0) × 106 mol N yr-1, yet reassuringly within the uncertainty (Table 2).
Estimates of annual N loading into Little Narragansett Bay from the
Pawcatuck River at Stillman Bridge (inclusive of Kenyon Industries) and from
the Westerly and Pawcatuck wastewater treatment facilities downstream.
RiverDIN fluxTON fluxTN fluxTN fluxTN% of% ofdischarge106 mol N yr-1106 mol N yr-1106 mol N yr-1106 mol N yr-loadingTN fluxTN flux106 m3 yr-1(May–Nov)kg N ha-1 yr-1annual(May–Nov)Stillman Bridge (2018)70220.1 ± 2.020.2 ± 2.840.3 ± 6.84.8 ± 0.67.4a––Kenyon Industries–––2.7b1.0c–616Westerly WWTFd–1.4 ± 0.11.1 ± 0.92.5 ± 0.11.2 ± 0.3–620Pawcatuck WWTFe–––0.3< 0.1–< 12Stillman Bridge (2002)f3037.26.2g16.0–3.1a––Land use modelh–––17.8–3.6j––Mixing model (2013–2015)i40813.7–27.3k––––
a Based on a watershed area of 760 km2. b DEM-monitored loading in 2012. c Permitted seasonal
loading. d Measured. e Reported. f Fulweiler and Nixon (2005). g As DON only. h Vaudrey et al. (2016). i Eq. (4). j Based on a watershed
area of 660 km2 established from Arc Hydro. k Assuming that DON
loading is equivalent to coincident DIN loading.
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, 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. 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 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.
Fulweiler and Nixon (2005) also observed that [DON] and [DIN] were inversely
correlated, as in the current study, and further detected a positive
correlation between [DON] and discharge, corroborating our earlier inference
that such a relationship should be manifest. They reasoned that the greater
remineralization of bioavailable DON in summer, at low discharge, could
explain this trend, given the greater in-river residence time of DON at low
base flow. While the mineralization of DON may be significant during the
warm season (e.g., Brookshire et al., 2005), we otherwise contend that the
increased [DON] with discharge may reflect import from the catchment via
shallow flow.
The mean [DIN] imported by shallow flow inferred herein is relatively low
(∼ 26 µM), in the range of 15 to 70 µM generally observed
in surface and shallow groundwater of undeveloped catchments across the US,
and substantially lower than the range of 100 to 700 µM observed in
shallow streams draining agriculture catchments (Dubrovsky et al., 2010).
However, it is greater than the [DIN] of ≤ 5 µM recorded in
shallow streams draining pristine forested catchments in the northeast
USA, which are otherwise dominated by DON (Dickerman et al., 1989; Hedin
et al., 1998). The [DIN] of ∼ 64 µM recorded here at low base
flow, which likely reflects that of deeper groundwater (barring a
substantial point source input) is also within the range reported for
groundwater NO3- in undeveloped catchments, albeit at the higher
end of this range (of 7 to 75 µM; Mueller et al., 1995) and falling
within the range reported for groundwater NO3- in southern RI (0–91 µM; Moran et al., 2014). The mean low-base-flow [DIN] observed
here is substantially lower than concentrations typical of groundwater in
agricultural catchments but higher than the [DIN] that was observed in the
groundwater reservoir of the upper Wood River in the 1980s (median ≤ 11 µM; Dickerman et al., 1989; Dickerman and Bell, 1993) –
suggesting that anthropogenic input to the deeper groundwater N reservoir of
the Pawcatuck watershed has increased over time.
Corroborating insights from NO3- isotope ratios
We turn to the N and O isotope composition of NO3- to further
investigate relationships of nutrients with river discharge and to
characterize N sources and cycling in the river. Like [NO3-], the
isotope ratios of NO3- co-varied with discharge. Values of δ15NNO3 decreased with discharge, suggesting that (a) NO3- added by shallow flow had lower δ15NNO3
values than low-base-flow NO3-, and/or (b) δ15NNO3 values at low base flow increased during warmer months
compared to their groundwater end-member due to biological cycling in-river.
Concurrently, δ18ONO3 values increased with discharge,
suggesting that (c) NO3- added by shallow flow had higher δ18ONO3 values than low-base-flow NO3-, and/or (d) δ18ONO3 values decreased in summer due to biological
cycling. We consider these hypotheses in turn.
Mixing curve of low-base-flow [NO3-] with shallow-flow
[NO3-] superposed onto weekly measurements of [NO3-] vs.
the corresponding mean daily discharge at the Stillman Bridge (Eq. 4).
Sources of DIN in shallow flow evidenced from δ15NNO3 values
In order to evaluate whether the lower δ15N DIN values observed
at higher discharge can be explained by the addition of relatively low
δ15N DIN by shallow flow, we plotted the δ15NNO3 values recorded at the Stillman Bridge vs. the inverse of
the corresponding NO3- flux (i.e., an adapted Keeling plot;
Keeling, 1958, 1961; Fig. 8a). Because we lack measurements of the δ15N values of the 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
NO3- captures 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 δ15NNO3 data conform to a linear relationship
expected for the addition of DIN with a lower mean δ15N 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 mean δ15N value of 6.7 ± 0.2 ‰ (Table 1), compared to a low-base-flow value of ∼ 8 ‰ observed at the bridges. The average δ15NNO3 value of atmospheric NO3- in rainwater was -2.5±2.1 ‰ (Sect. S1; Fig. S2), indicating that
NO3- added by shallow flow did not predominantly originate from
direct atmospheric deposition as uncycled atmospheric NO3-. While
the δ15NNO3 of atmospheric NO3- could
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 δ15NNO3 end-member value of 6.7 ‰ is in the
upper range observed for soil NO3- in temperate forested
catchments (Mayer et al., 2002; Barnes and Raymond, 2009). While the net
sources of reactive N to forested soils are atmospheric deposition and
biological N2 fixation – which have relatively low δ15N
values (≤ 0 ‰) – partial denitrification in soils
and shallow groundwater increases the δ15N of the soil N
reservoir to values of ∼ 5 ‰ (Amundson et al.,
2003; Houlton et al., 2006; McMahon and Böhlke, 2006; Houlton and Bai,
2009). The NO3- imported by shallow flow draining urbanized
systems has comparatively higher δ15NNO3 values (≥ 10 ‰; e.g., Divers et al., 2014), while NO3- in
soils and shallow groundwater in agricultural systems generally falls within
a lower range of values between 2 ‰–4 ‰ (Green et al.,
2008; Böhlke et al., 2009; Lin et al., 2019). The watershed of the
Pawcatuck River is largely forested, yet hosts agricultural and urbanized
sections that ostensibly contributed to the mean δ15NNO3
end-member imported by shallow flow. Thus, while DIN added to the river by
shallow flow at high discharge had a mean δ15N value consistent
with expectations for a largely forested catchment, inputs from agricultural
and urbanized catchments may be rendered undiscernible due to their opposing
contributions to the mean δ15NNO3 value in shallow flow.
Modified Keeling plot of NO3- isotopologue ratios at the
Westerly and Stillman bridges vs. the inverse of the daily NO3-
flux: (a)δ15NNO3 vs. the inverse of the NO3-
flux; (b)δ18ONO3 and δ18ONO3 corrected
for atmospheric NO3- vs. the inverse of the NO3- flux.
Negligible fraction of uncycled atmospheric NO3- confirmed by O isotope ratios
The inference that uncycled atmospheric NO3- did not contribute
substantially to the increased NO3- flux at higher discharge is
corroborated by the Δ17ONO3 measurements at the Stillman
Bridge. The low values of NO3- in rainwater observed evidenced
only a slight contribution of < 3 % uncycled atmospheric
NO3- to total riverine NO3- in a few samples, suggesting
efficient processing of atmospheric NO3- in soil shallow
groundwater (Mengis et al., 2001; Barnes et al., 2008). This observation is
further echoed in a recent meta-analysis of North American rivers, wherein
the contribution of uncycled atmospheric NO3- to base flow was
inferred to be generally modest (Sebestyen et al., 2019). The NO3-
delivered 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 Δ17O
signature.
A Keeling plot of δ18ONO3 values vs. the inverse of the
NO3- flux at the bridges suggests that NO3- added by
surface flow had a mean δ18ONO3 value of 4.5 ± 0.4 ‰ (Fig. 8b; Table 1), compared to a mean low-base-flow
value of 2.8 ± 0.2 ‰. Although the contribution
of uncycled atmospheric NO3- to the riverine reservoir was modest,
we nevertheless consider that the increase in δ18ONO3
values with discharge may derive in part from uncycled atmospheric
NO3-, given the direct relationship of Δ17ONO3
to discharge, and considering the characteristically elevated δ18ONO3 values of 60 ‰–80 ‰ observed in the
local rainwater NO3-. Indeed, when the weighted contribution of
atmospheric NO3- is subtracted from individual δ18ONO3 values (attributed from corresponding Δ17O
measurements, accounting for precipitation-dependent differences in the mean
Δ17O and δ18ONO3 values of rainwater), the
intercept of the Keeling plot decreases slightly to 3.8 ± 0.2 ‰, nevertheless remaining greater than the δ18ONO3 of low-base-flow NO3- (Table 1). The increase in
δ18ONO3 with increasing discharge is thus partially
explained by the small component of uncycled atmospheric [NO3-]
with elevated δ18ONO3 values.
The δ18ONO3 signature of 3.8 ‰ for
NO3- added with increasing discharge (minus the uncycled
atmospheric NO3-) is in the range generically observed for soil
NO3- (Kendall et al., 2007; Michener and Lajtha, 2007). It has
traditionally been ascribed to that expected for newly nitrified
NO3-, based on an empirical metric stipulating that the δ18ONO3 values produced by nitrification derive from the
fractional contribution of the reactants, namely 1/3δ18O of
O2+2/3δ18O of H2O (Andersson and Hooper, 1983;
Hollocher 1984; Kendall et al., 2007). Considering that the δ18OH2O of Pawcatuck River water is -7 ‰ and
the δ18OO2 of atmospheric oxygen is ∼ 23.5 ‰ (Kroopnick and Craig, 1972), the nitrification δ18ONO3 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
into the NO3- molecule during nitrification and isotopic exchange
of the nitrite intermediate with water, which otherwise give way to
nitrified NO3- whose δ18ONO3 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 δ18ONO3 in some soils and saturated systems, which are not
explained by simple fractional contribution of 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 NO3- imported into the
river with increased discharge, which is typical of that in soils and
shallow groundwater, does not strictly indicate that shallow-flow
NO3- originated from proximate nitrification therein, as generally
presumed. It also signals that NO3- underwent partial
denitrification in soils and shallow groundwater, resulting in a coupled
increase in its δ15N and δ18O 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 δ18ONO3 end-member, specifically reducing it
from its soil value due to the nitrification of incident NH4+.
Thus, δ18ONO3 values imported by shallow flow, once
adjusted for modest contributions of uncycled atmospheric NO3-,
fall within the range typically observed in soils, potentially modified by
nitrification in-river.
Values of δ15NNO3 at low base flow reflect groundwater DIN
The higher δ15NNO3 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 δ15NNO3 values in deeper
groundwater are generally higher than in shallower groundwater above, more
fractionated by denitrification in the saturated zone (e.g., Böhlke et
al., 2006). Alternatively, the higher NO3- isotope ratios at low
base flow may result from increased biological cycling in summer –
modifying the isotope composition of low-base-flow NO3- relative
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. Fulweiler and
Nixon (2005) similarly observed lower [NH4+] in the summer but
saw no correlation to 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 NO3- pool is also assimilated
during summer months – and isotopically fractionated – is unclear. The
fraction of the NO3- pool assimilated by algae may be 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 limited 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 of the
river, which may assimilate NO3- as well as reduced N substrates.
Nevertheless, the inference that the riverine NO3- pool is
minimally assimilated, even in summer, appears consistent with along-river
distribution of NO3- isotope ratios. If a sizable fraction of the
incident NO3- pool were assimilated into biomass during summer
months, both the δ15NNO3 and δ18ONO3
values of low-base-flow NO3- would expectedly increase in
proportion to the fraction of NO3- assimilated (Granger et al.,
2004; Johannsen et al., 2008). However, the δ15NNO3
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 δ18ONO3
values. Similarly, [PN] and chlorophyll a did not increase along-river, as
would otherwise be expected for the progressive and sizable conversion of
the NO3- pool into the particulate pools (Fig. S6c–d). Thus, we
rule out a dominant influence of algal assimilation in fractionating the
riverine NO3- isotope ratios.
A more nuanced framework from which to interpret the NO3- isotope
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 Ensign and Doyle, 2006; Harvey et al.,
2013). A small fraction of the NO3- pool is likely assimilated
during the growing season, resulting in the production of PN with a lower
δ15N than coincident NO3- due to N isotope
fractionation during assimilation (Needoba et al., 2003; Fig. 9).
Considering the small summertime pools of PN and NH4+relative
to the NO3- pool, δ15NNO3 values will be
minimally fractionated by assimilation. Moreover, the concomitant recycling
of PN and its subsequent nitrification will ostensibly regenerate
NO3- with a δ15NNO3 value roughly equivalent to
that assimilated into organic material then ammonified – given an
approximate steady state between NO3- assimilation and
nitrification – such that δ15NNO3 values will not incur a
progressive increase from continual assimilation along-river. These dynamics
will result in little net change in riverine δ15NNO3
values relative to the mean catchment end-member.
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 and then
mineralized and released back into the water column or denitrified. (a) The
δ15N of the riverine NO3- reservoir integrates the
NO3- and NH4+ delivered continually from groundwater
(δ15NNO3-GW and δ15NNH4-GW), minus the
NO3- removed concurrently by sedimentary denitrification – the
δ15N of which depends on the sedimentary isotope fractionation
communicated to the water column reservoir, 15εD.
Given the small size of the respective PN and NH4+ pools relative
to NO3-, ammonification and subsequent nitrification produce
NO3- with a δ15NNO3-N value approximating that
lost concurrently to assimilation (δ15NNO3–15εA), notwithstanding the NH4+ input from
groundwater. The input of groundwater NH4+ (δ15NNH4-GW) implicitly subsumes the input of reactive
allochthonous PN and DON. (b) The riverine δ18ONO3
integrates the NO3- input from groundwater and precipitation
(δ18ONO3-GW) and from in-river nitrification (δ18ONO3-N), minus NO3- lost to algal assimilation and
sedimentary denitrification – whose respective values depend on the net
isotope effects associated with assimilation and denitrification,
18εA and 18εD.
The NO3- isotope ratios could, however, be influenced by
denitrification in-river (Kellman and Hillaire-Marcel, 1998; Fig. 9). While
direct benthic denitrification does not communicate an isotope enrichment to
NO3- in the overlying water column due to a reservoir effect
(Brandes and Devol, 1997; Sebilo et al., 2003; Lehmann et al., 2005),
δ15N- and δ18O-enriched NO3- from the
sediment depth of denitrification can be entrained into the water column by
hyporheic flows in the riparian zone (Sebilo 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 NO3- lost concurrently to denitrification,
thus contributing to an increase in δ15N of the water column
reactive N reservoir (Brandes and Devol, 1997; Granger et al., 2011). The
along-river increase in δ15NNO3 values could then result
from isotopic fractionation by sedimentary denitrification. Yet a downstream
increase in δ15NNO3 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 δ15NNO3 values along-river arises
principally from differences in the δ15N of DIN input from
respective reaches of the catchment – although some influence of benthic
denitrification on riverine δ15NNO3 values cannot be ruled
out. We thus interpret the riverine δ15NNO3 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.
Influence of in-river biological cycling on δ18ONO3 values at low base flow
The δ18ONO3 values along-river can also be interpreted
within the framework of nutrient spiraling. As with δ15NNO3, the riverine δ18ONO3 values integrate
the contribution of NO3- imported from the catchment (including
uncycled atmospheric NO3-), the NO3- produced by
nitrification in-river, and the NO3- consumed by assimilation and
by denitrification (Fig. 9). Without continual exogenous input from the
catchment, δ18ONO3 values of an initial NO3-
reservoir would theoretically converge downriver onto a steady-state value
dictated by the δ18ONO3 of newly nitrified NO3-
and the effective isotope effect for NO3- consumption, by
assimilation and denitrification: for instance, assuming a δ18ONO3 value of -6 ‰ for newly nitrified
NO3- (δ18OH2O+1 ‰;
Casciotti et al., 2008; Sigman et al., 2009; Buchwald and Casciotti, 2010;
Granger et al., 2013; Boshers et al., 2019), a canonical NO3-
assimilation isotope effect of 5 ‰ (Needoba et al.,
2003), and no influence of sedimentary denitrification on water column
δ18ONO3, values downriver would asymptote to -1 ‰. The δ18ONO3 values of 2.8 ‰ observed at the bridges during low base flow thus
suggest that the NO3- introduced continuously along the catchment
had δ18ONO3 values greater than -1 ‰, assuming roughly equivalent in-river assimilation and nitrification
fluxes. These greater δ18ONO3 values may also signal some
influence of sedimentary denitrification in fractionating the water column
δ18ONO3. Observations of decreasing along-river values are
then consistent with the notion of higher catchment δ18ONO3 end-member values converging onto lower values determined
by the ratio of nitrification to consumption in-river – and associated
isotopic fractionation. Within this framework, δ18ONO3
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) similarly
observed higher δ18ONO3 values during the cold season in
streams draining forested watersheds in the northeastern USA. The riverine
δ18ONO3 values thus afford insights into N sources and
cycling that are consistent with expectations for nutrient spiraling.
Regional N sources to the Pawcatuck River
Observations from the along-river surveys provide insights into the
contribution of different 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 can be identified
from concentration decreases. Reaches of the river that exhibit
disproportionate loading present potential targets for mitigation. As
detailed above, we interpret changes in δ15NNO3 values
along-river to primarily reflect differences in the δ15N of DIN
inputs from respective reaches of the catchment, thus serving to identify
dominant regional N sources.
Surface water in Worden Pond had relatively low nutrient concentrations,
which remained similarly low at Biscuit City Road (Station 2) in two of
three samplings. The otherwise extremely elevated [NO3-] at
Station 2 in March 2018 decreased downstream at Station 3 by > 40 %, more than can be explained by either dilution from additional inflow
or denitrification. This elevated concentration may then reflect the
inadvertent sampling of a groundwater plume or a localized reach of
slow-flowing water, rather than the mean river composition. Monitoring at
Biscuit City Road from 2007–2016 by the Wood-Pawcatuck Watershed Association
similarly reveals relatively low median [NO3-] values of ∼ 1 µM during fall samplings, punctuated by stochastic instances of
elevated concentrations, as high as 41 µM (Fig. S6; Wood-Pawcatuck
Watershed Association, 2020). The δ15NNO3 value recorded
at this station during the March 2019 sampling was 6.2 ‰
and the δ18ONO3 was 4.8 ‰, values
consistent with either a groundwater plume or a slow-flowing reach of the
river.
Both [NO3-] and [PO4-] increased thereafter at Kenyon and
Wood River Junction (Stations 3 and 4, respectively) in all sampling
campaigns. Associated δ15NNO3 values were relatively low
during the March 2019 sampling (≤ 4 ‰) –
coincident with more elevated [NO3-] – potentially signaling the
input of DIN by shallow flow from proximate turf farms (Kreitler et al.,
1978; Katz et al., 1999; Townsend et al., 2003; Deutsch et al., 2005). Input
of uncycled atmospheric NO3- by surface flow due to regional snowmelt, which could also explain lower δ15NNO3 values, is
not supported by the corresponding δ18ONO3 values, which
would otherwise be disproportionately high. Moreover, there was little to no
accumulated snow in March 2019. The increased nutrient concentrations
observed at Stations 3 and 4 in all sampling campaigns also likely derived
in part from the retention ponds at Kenyon Industries, in light of a
permitted discharge of 7500 mol N and 950 mol P per day (US
Environmental Protection Agency, 2010). Corresponding δ15NNO3 values at Stations 3 and 4 during the May and November 2018 samplings were ∼ 6 ‰, which could indicate
input from deeper agricultural groundwater or could reflect discharge by
Kenyon Industries, for which we do not have end-member values.
Inflow from the less impacted Wood River evidently diluted nutrient
concentrations in the Pawcatuck River (Station 7). The Wood River
contributes significantly to the total discharge of the Pawcatuck River
(≥ 14 ± 5 % of total – based on discharge at Hope Valley USGS
gauge), draining a more forested watershed that harbors fewer agricultural
areas than the lower Pawcatuck River. The [NO3-] in all sampling
campaigns remained relatively invariant downstream of the Wood River inflow
through the largely forested catchment to Potter Hill Dam (Station 11),
while δ15NNO3 values increased marginally. The increases
in [NO3-] and δ15NNO3 thereafter to the Stillman
and Westerly bridges indicate DIN input from groundwater in the more
populated portion of the watershed. The population density and associated
septic systems increase considerably in the vicinity of the town of Westerly
(Wood-Pawcatuck Watershed Association, 2016). Septic leachate and urban
runoff are typically associated with relatively higher δ15N
values, on the order of 8 ‰ to 15 ‰ (Kendall et al., 2007;
Böhlke et al., 2009; Kasper et al., 2015). Thus, changes in land use
along the catchment best explain the δ15NNO3 increase in
the lower portion of the river.
In all, the substantial difference in [DIN] between Stations 2 and 5 signals
disproportionate input 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 three 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 substantial even assuming some biological
attenuation. This portion of the river also contributed disproportionately
to the riverine PO43- burden, although we do not explicitly
consider this contribution in relation to the total discharge into the
estuary, given the complex geochemistry of PO43- that involves
adsorption and release from authigenic particles in sediments (Froelich,
1988).
The increase in [DIN] and δ15NNO3 values in the lower
portion of the river, in light of the large 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 δ15N
input from the urbanized catchment of 10 ‰ and a mean
δ15NNO3 of 6 ‰ at Potter Hill Dam,
compared to 8 ‰ at the Westerly Bridge, the DIN added to
the river within this reach would amount to ∼ 50 % of the total
riverine N load. Otherwise, assuming a δ15N input of 15 ‰, the DIN contributed from the urbanized reach would
otherwise amount to ∼ 20 % of the total.
N loading into Little Narragansett BayRiverine contributions
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 20.2×106 mol yr-1 for DIN and 40.3×106 mol yr-1 for TN, albeit with uncertainty associated with
the TN loading estimate given the variability of our DON measurements (Table 2). These values are considerably larger than those estimated from biweekly
measurements at the Stillman Bridge for 2002 by Fulweiler and Nixon (2005),
which were 7.2×106 mol yr-1 for DIN and 16.0×106 mol yr-1 for TN. The greater N loading in 2018 could arise from (a) an increase in groundwater concentrations and/or point source discharge,
evident at low base flow, and/or (b) increased N loading by shallow flow.
The latter could result from increased atmospheric N deposition, greater
annual precipitation, and/or higher surface N concentrations imported by
shallow flow. We examine these hypotheses in turn.
In 2002, the [DIN] at low base flow, which reflects that associated with
deeper groundwater and point source discharge, was ∼ 50 µM
(Fulweiler and Nixon, 2005) – thus lower than the value of ∼ 64 µM observed by us – suggesting lower DIN inputs from deeper groundwater
and/or point sources in 2002. This inference is supported by monitoring data
of the Wood-Pawcatuck Watershed Association from 1989 to 2017, which
documented a discernible increase in riverine [NO3-] of
approximately 1 µM per year at Bradford (Station 8) in the month of
October and a slighter rate of increase 0.3 µM per year in May
(Fig. S7a). These trends are not explained by a secular change in monthly
precipitation (Fig. S7b). Given that mean river discharge is generally
higher in May than in October, the greater rate of increase in October
suggests an increase in [DIN] of deeper groundwater entering the river –
and/or an increase in point source discharge up-river. Assuming a 10 µM difference in [DIN] during low base flow at the Stillman Bridge in 2018
compared to 2002, and a year-round discharge of deeper groundwater of 0.25×106 m3 d-1 (based on the mean low-base-flow discharge), 0.9×106 mol of additional DIN was potentially delivered in 2018 from the
increased groundwater or point source [DIN]. This greater DIN input from
deeper groundwater and/or point sources only explains a small fraction of
the additional loading of 12.9×106 mol DIN estimated for 2018
compared to 2002.
Regional atmospheric deposition of DIN has decreased ∼ 67 % since
2000, from ∼ 95 to 36 µM DIN in 2018 (NOAA National
Atmospheric Deposition Program, 2019), which should have resulted in a lower
riverine N flux given similar annual precipitation. However, 2002 was a
drought year, whereas 2018 was the third wettest year on record in
Washington County, RI, with total precipitation at 152 cm compared to an
80-year mean of 114 cm (NOAA National Centers for Environmental information,
2019). River discharge was thus substantially lower in 2002, at 303×106 m3 yr-1, compared to 702×106 m3 yr-1 in
2018 (Table 2). The larger riverine N loading in 2018 is thus explained by
greater precipitation and consequent discharge above low base flow,
importing additional DIN (and DON) into the river via shallow flow. Assuming
a comparable [DIN] delivered by shallow flow between then and now (26 ± 3 µmol L-1), the greater discharge in 2018 entails an
additional DIN influx of 10.1 (±1.2) × 106 mol yr-1,
accounting for most of the estimated difference of 12.9 × 106 mol DIN yr-1 between 2018 and 2002. The greater discharge in 2018 ostensibly
resulted in increased DON influx into the river, although the variability
of our DON measurements precludes a robust estimate of this additional flux.
In all, the greater DIN loading in 2018 compared to 2002 is explained in
small part by an apparent increase in [DIN] at low base flow – deriving
from a parallel increase in groundwater [DIN] or a potential increase in
point source discharge by Kenyon Industries – and in greater part by a
substantial difference in annual river discharge and associated import of
DIN (and DON). Parenthetically, Eq. (4) returns a load of 10.1 × 106 mol DIN yr-1 for 2002 when adjusting low base flow to 50 µM,
compared to the observed flux of 7.2 × 106 mol DIN yr-1 (Fulweiler
and Nixon, 2005), suggesting that the import of DIN from shallow groundwater
may also have increased in the last 2 decades.
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) utilized a land use model to estimate the
TN load from the Pawcatuck River at 14.6 × 106 mol TN yr-1,
based on precipitation from 2013–2015. In comparison, using the mixing
curve algorithm (Eq. 4) and river flow for the 2013–2015 period, and
further assuming that DIN accounted for roughly half of TN as in our
measurements, DIN loading is otherwise estimated as 13.7 × 106 mol DIN yr-1 and TN loading as 27.3 × 106 mol TN yr-1 for this
period (Table 2). The TN load thus estimated is 12.7 × 106 mol TN yr-1 higher than the land use model estimate, leaving nearly 45 % of
TN apparently unaccounted for. However, our TN measurements include both
labile and non-labile 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 – an N pool that is largely unavailable to micro-organisms
(Stevenson 1994) – a value of 45 % of TN being non-labile could be
consistent with this system. Seitzinger et al. (1997) otherwise estimated
that 40 %–70 % of DON from the Delaware River was reactive on pertinent
timescales. 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. 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 these rivers ranged from 9.3 to
14.9 kg ha-1 yr-1 (as reviewed in Narragansett Bay Estuary Program, 2017; Nixon
et al., 1995, 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, these riverine loads were substantially reduced, achieving an
average of 4.8 kg ha-1 yr-1 in recent N budgets developed for the
2013–2015 time period, with one exception (i.e., the Ten Mile River;
Narragansett Bay Estuary Program, 2017; Krumholz, 2012). Export from
pristine temperate zones prior to human disturbances is estimated to have
been on the order of 1.3 kg N ha-1 yr-1 (Howarth et al., 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 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 of 6 × 106 mol TN yr-1, a flux notably higher than that of 1.45 × 106 mol TN yr-1 reported by the 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 (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. Independent estimates by Vaudrey et
al. (2017) derived from a land use model suggest that WWTF effluents
contribute ∼ 13 % of the riverine-plus-WWTF 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 the Fulweiler
and Nixon (2005) study; including an estimate of the non-labile fraction
brings the annual contribution from WWTFs down to 8 % of the TN.
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 × 106 mol TN yr-1, thus accounting for 6 %
of the annual riverine-plus-WWTF loading to the estuary, an input
comparable to that of the WWTFs (Table 2). 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 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 in 2018, from 1 May to 31 October.
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 November), of which
only a fraction may be bioavailable on pertinent timescales. Assuming a
median 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 × 106 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 %.
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 NO3- isotope
ratios indicate substantive inputs of reactive N to the river from
agricultural and/or point sources along the upper 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 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 to affect 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 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 connected to sewer. This distribution
reflects the urban nature of the watershed near the coast and the more rural
character of the watershed further inland. Finally, restricting the use of
lawn fertilizers and lessening the extent 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 haestuary-1 yr-1
(3.6 × 103 mol haestuary-1 yr-1), which is generally
supportive of eelgrass, has been proposed by the scientific community
(Hauxwell et al., 2003; Latimer and Rego, 2010). In 2018, the DIN and TN
loads to Little Narragansett Bay were 37 × 103 and 74 × 103 mol 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 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 a 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.
Conclusions
Our findings illustrate the utility of NO3- isotopologue ratios in
differentiating among N sources, with implications for the management of N
loading of the watershed. In particular, the seasonal and
flow-dependent nature of N loading and cycling uncovered herein presents
important considerations for mitigation efforts.
Our interpretations of NO3- isotopologue dynamics also move
beyond the traditional source attribution framework in an effort to
reconcile with the current theory of riverine N biogeochemistry. Nutrient
spiraling theory offers a powerful conceptual basis to differentiate the
influences of N sources vs. cycling on NO3- isotopologue
distributions. Continued inquiry in the context of this framework is bound
to yield novel and unexpected insights into N isotopologue cycling and, more
fundamentally, into river biogeochemistry.
Data availability
All data were submitted to Rhode Island Department of Environmental
Management and the Connecticut Department of Energy and Environmental
Protection and are submitted to the PANGAEA data repository 10.1594/PANGAEA.931561 (Rollinson et al., 2021).
The supplement related to this article is available online at: https://doi.org/10.5194/bg-18-3421-2021-supplement.
Author contributions
VRR and JG conceive the research question, designed the study approach, led
the field survey, ensured data curation, and conducted formal analysis. SCC,
MLB, CPK, LAT, and HCW assisted with data collection and analysis. CMM
assisted with statistical analyses. CRT and MGH provided use of specialized
facilities. JG and JMPV secured funding for the investigation. VRR 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 conflict of interest.
Acknowledgements
We thank Clare Schlink, Reide Jacksin, Lindsey Potts, Danielle Boshers-Snow,
Anna Alvarado, Peter Ruffino, and Matt Lacerra for assistance with fieldwork
and/or laboratory analyses. We are also grateful to Nicholas De Gemmis and the
Jacobs Group at Westerly Wastewater Treatment facility for providing us
with weekly samples. Water quality monitoring by the Wood-Pawcatuck
Watershed Association provided important historical data that facilitated
our interpretations. Monitoring of water quality in Little Narragansett Bay
by Clean Up Sound and Harbors (CUSH) inspired our effort to determine
sources of nutrients to the estuary. Comments by the two anonymous reviewers
helped improve the manuscript.
Financial support
This research has been supported by the NOAA Connecticut Sea Grant, University of Connecticut (grant no. NA18OAR4170081 project number R/ER-30) and the NSF Career Award (grant no. OCE-1554474).
Review statement
This paper was edited by Jack Middelburg and reviewed by two anonymous referees.
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