Introduction

Human-caused eutrophication in coastal ecosystems and the need to understand the fate of excess nitrogen is a complex global problem that is influenced by ongoing climate change and changes in atmospheric deposition of nitrogen (Paerl and others, 2014; Cloern and others, 2016; Valiela and others, 2016). There is substantial interest and urgency in understanding the fate and transport of nitrogen that is released from septic systems in high-density residential neighborhoods on Cape Cod that subsequently infiltrates the nearshore groundwater and subterranean estuaries in multiple embayments. This nitrogen is a pollutant that at sufficient concentrations and under certain conditions can cause eutrophication, macroalgal blooms, and associated environmental problems. Nitrogen loading to coastal embayments on Cape Cod, Massachusetts, is one of the more pressing environmental challenges faced by coastal communities in the northeastern United States (Valiela and others, 1992; Nixon, 1995; Short and Burdick, 1996; Valiela and others, 2000; Howarth and others, 2002; Valiela and others, 2016).

External nitrogen loading increases the growth of phytoplankton and macroalgae in coastal marine systems in the northeastern United States (Howarth, 1988; Valiela, 2006). In recent decades, the relative contribution to nitrogen loading from wastewater inputs has increased while the contribution of atmospherically deposited nitrogen has decreased (Valiela, Collins, and others, 1997; Valiela, McClelland, and others, 1997). Increased supply of nitrogen causes decreased water clarity from increased algal growth and encourages the growth of epiphytes on submerged aquatic vegetation that reduces light penetration to the detriment of more desirable plant species, for example, eel grass, and damages ecosystems favoring finfish and shellfish (Thomsen and others, 2012; Rasmussen and others, 2013; Nelson, 2017). Nitrogen inputs to coastal ecosystems also are partially responsible for ocean acidification through biogeochemical processes associated with eutrophication (Cai and others, 2011; Gledhill and others, 2015; Rheuban and others, 2019) that can harm vulnerable calcifying species that are important fisheries in the northeastern United States (Ekstrom and others, 2015; Hare and others, 2016). Nitrogen additions also can stimulate microbial decomposition of organic matter in coastal marsh sediments, potentially reducing their capacity for carbon sequestration (Bulseco and others, 2019). Because tourism, retirement, and fishing are the mainstays of the Cape Cod economy (Cape Cod Chamber of Commerce, 2013), degradation of the coastal aquatic ecosystems is an important issue on Cape Cod.

In 2000, recognizing the importance of nitrogen in eutrophication of Cape Cod waters, the Massachusetts Department of Environmental Protection established the Massachusetts Estuaries Project to evaluate nitrogen loading and capacity in 89 Cape Cod embayments (Massachusetts Department of Environmental Protection, 2018). In approximately one-half of the embayments to date, these evaluations have established regulatory numeric nitrogen loading limits, or total maximum daily loads (TMDLs) (Massachusetts Department of Environmental Protection, 2018). Currently, Massachusetts towns covered by the Massachusetts Estuaries Project are implementing nitrogen remediation actions that would meet the TMDL requirements.

On Cape Cod, principal sources of nitrogen are from onsite disposal of domestic sewage (through cesspools and septic systems) and lawn fertilizer (Valiela and others, 2000). Nitrogen transport from these sources to coastal water is largely through groundwater because streams are uncommon on the sand deposits of the narrow peninsulas. In much of the Cape Cod aquifer, where subsurface conditions are aerobic and the predominant form of nitrogen is nitrate, transport is likely to be conservative. Aerobic onsite wastewater-treatment systems (septic system leaching fields) have been shown to be inefficient at removing inorganic nitrogen (Costa and others, 2002; Van Cuyk and others, 2001). Along the flow path from onshore groundwater to submarine groundwater discharge sites, some portion of inorganic nitrogen commonly is attenuated by conversion to gaseous forms but a substantial fraction can be transported through the subsurface to the estuary (Valiela and others, 1992; Paerl, 1997; Charette and Sholkovitz, 2002). For example, nitrate reduction can occur where groundwater encounters onshore or offshore anaerobic sediments containing organic carbon or other electron donors (originating from infiltrating seawater or decomposition of particulate phases). Nitrogen attenuation reactions are not considered in the conservative assumptions used in determining TMDLs (Massachusetts Department of Environmental Protection, 2018) and are less well documented in the subterranean estuary than in many other groundwater settings affected by agricultural and wastewater nitrogen inputs.

Where subsurface conditions are anaerobic, nitrate, nitrite, and ammonium can be attenuated, that is, converted to gaseous forms (nitric oxide [NO], nitrous oxide [N₂O], and nitrogen gas) through the bacterially mediated processes of denitrification and anaerobic ammonium oxidation (anammox) (Thamdrup and Dalsgaard, 2002; Strous and Jetten, 2004). Nitrate also can be microbially converted to ammonium through the process of dissimilatory nitrate reduction to ammonium (DNRA) (Giblin and others, 2013). Recently, it has been shown that bacterial denitrification occurs in sandy coastal sediments where tidal cycles resulted in fluctuating oxic to anoxic conditions (Marchant and others, 2017). Denitrification generally refers to microbial respiration, whereby nitrate nitrogen is reduced to gaseous forms of nitrogen (N₂O and N₂) in response to the oxidation of an electron donor such as organic matter, sulfide, methane, or reduced iron oxides (Knowles, 1982). Nitrogen gas can be released to the atmosphere or reduced to ammonia through nitrogen fixation in coastal sediments (Newell and others, 2016). One of the intermediate gaseous products of bacterial denitrification, nitric oxide that is produced from nitrite, can be chemically denitrified in the presence of Fe(II) or denitrified by fungi; in both cases, it is converted to N₂O that can subsequently be released to the atmosphere (Wankel and others, 2017). Denitrification is common in aquatic anaerobic or suboxic environments in freshwater, wetlands, and estuarine and marine waters, as well as groundwaters where geochemical conditions are suitable (Smith and others, 1991; Korom, 1992; Seitzinger and others, 2006). Denitrification was reported to be an important nitrate reducing mechanism in a nitrate-contaminated aquifer on Cape Cod (Smith and Duff, 1988; Smith and others, 2004).

In DNRA, microbes oxidize organic matter and use nitrate as an electron acceptor for respiration, reducing it first to nitrite and then to ammonium. Denitrification and DNRA compete for nitrate. Denitrification is usually the dominant nitrate reduction process; however, certain factors can favor DNRA (Giblin and others, 2013). DNRA may be favored when concentrations of organic matter are high and concentrations of nitrate are low (Tiedje and others, 1983; Dong and others, 2011) or when salinity increases (Giblin and others, 2010). DNRA retains inorganic nitrogen in the sediment as ammonium, which could later be converted back to nitrate in an aerobic environment.

In anammox, ammonium and nitrite are converted into nitrogen gas, and nitrite is used as a terminal electron acceptor (Thamdrup and Dalsgaard, 2002; Strous and Jetten, 2004). The anammox reaction does not require organic carbon, but it does require nitrite that, in anaerobic environments, would most likely be available if nitrate was present and anammox was coupled with ongoing denitrification or DNRA (where nitrite was produced). Similarly, fungal-mediated denitrification or abiotic (Fe(II)-mediated chemodenitrification) reactions (Wankel and others, 2017) converting nitrite to N₂O require nitrate conversion to nitrite. Gaseous forms of nitrogen are released to the atmosphere and therefore are unavailable to algae, potentially mitigating eutrophic conditions. Conditions conducive to attenuation and losses of inorganic nitrogen have been documented in landfill plumes and in large plumes formed by land disposal of municipal sewage (Smith and others, 1991; Christensen, and others, 1994; DeSimone and Howes, 1995; Repert and others, 2006) and in many other aquifer settings. Denitrification and anammox processes have been documented in a wastewater-contaminated plume in the Cape Cod aquifer system (Smith and Duff, 1988; Smith and others, 2004, 2015).

The current study was conducted to address questions that arose in a previous study at the Eel River, a coastal embayment in East Falmouth, Massachusetts, on western Cape Cod (Colman and others, 2018). The goal of the previous study was to identify pathways of nitrogen transport and areas of potential nitrogen attenuation in the subterranean estuary through bacterially mediated denitrification or anammox. Colman and others (2018) found evidence for nitrate reduction in some portions of the subterranean estuary but concluded that further study was needed to determine the distribution of nitrate reduction in shallow groundwater below the intertidal and nearshore subtidal zones. The goal of this follow-up project was to investigate nitrogen attenuation in the subterranean estuary below the intertidal and nearshore subtidal zones at the same study site to refine the conceptual model of nitrogen attenuation at the Eel River subterranean estuary. This study includes geochemical and hydraulic information in the subterranean environment of the intertidal and adjacent subtidal zones with a focus on the sediment/water interface and shallow offshore zones where fresh and saline water mix. This report includes details of field investigations, analytical methods, results, and discussion of geochemical indicators of nitrate reduction. The results presented in this report may help inform the design of future nitrogen remediation investigations in peninsular, coastal areas with high-housing density adjacent to saline surface water that may be targeted for sewer installation to prevent coastal eutrophication.

Geographic, Geologic, and Hydrologic Setting

The Seacoast Shores neighborhood occupies a peninsula that extends southward from the mainland area of Cape Cod towards Vineyard Sound (fig. 1). The peninsula is about 2.6 kilometers (km) long and as much as 615 meters (m) wide and is bordered by saline tidal surface-water bodies that include the Eel River on the west and Eel Pond and the Childs and Seapit Rivers on the south and east. Land-surface altitudes generally decrease from about 10 m on the north to less than 1 m on the south. The topography generally is flat, with steep bluffs along the sides of the peninsula and a few low-lying areas in the interior.

The peninsula and its adjacent narrow saltwater bays are similar to other peninsulas and bays along the southwestern shore of Cape Cod near East Falmouth. The bays occupy valleys that were cut into the glacial outwash plain at the end of the Pleistocene Epoch. The bays were subsequently drowned when sea level rose, leaving the intervening areas of the outwash plain as long, narrow peninsulas interspersed with coastal embayments. The aquifer materials are glacial outwash sand and gravel that generally grades downward into fine-grained glaciolacustrine sand, silt, and clay (Oldale, 1992; Masterson and others, 1997; Hull and others, 2019). The unconsolidated sediments are about 90 m thick and lie on crystalline, granitic bedrock (Fairchild and others, 2013).

The Seacoast Shores peninsula is a part of the Sagamore flow lens, a part of the Cape Cod aquifer system consisting of several hydrogeologic units (outwash plains, terminal and ground moraines, and ice contact deposits) primarily composed of sand and gravel, with some silt and clay. The units were deposited during the late glaciation of New England. The regional aquifer is a shallow, unconfined hydrologic system in which groundwater flows radially outward from the top of a water-table mound that underlies western Cape Cod (Masterson and others, 1997; Walter and Whealan, 2005; Walter and others, 2019). The horizontal hydraulic conductivity of the sand and gravel is estimated to be about 60–110 meters per day and that of the interbedded fine sands and silts in the aquifer less than 15 meters per day (Walter and Whealan, 2005; Hull and others, 2019). The study area is located along the southern coastline, where the regional groundwater system discharges to Vineyard Sound. The peninsular freshwater aquifer is bounded at the top by the water table, at the bottom by the freshwater/saltwater interface, and at the lateral boundaries by the tidal saltwater bodies. The freshwater aquifers beneath the peninsulas are seaward extensions of the much larger regional groundwater-flow system that also discharges to the tidal saltwater bodies.

The Seacoast Shores peninsula receives an estimated 114 centimeters per year (cm/yr) of precipitation. Slightly more than one-half of the precipitation recharges the aquifer at the water table (LeBlanc and others, 1986); the remainder is lost to evapotranspiration. Surface runoff is negligible owing to the sandy soils and low topographic relief of the area. Water supply is public water obtained from wells and a lake inland from the peninsula. Data from the Cape Cod Commission (2015) indicated that about 1,000 dwellings on the peninsula account for about 10.5 cm/yr of wastewater disposal averaged over the total 1.2 square kilometer (km²) area of the peninsula (Colman and others, 2018). The concentration of nitrate in domestic wastewater reaching groundwater on Cape Cod is estimated to be about 3,000 micromoles per liter (µmol/L) (Eichner and Cambareri, 1992). Colman and others (2015) estimated that the volume-averaged concentration of dissolved nitrogen in groundwater in the Seacoast Shores peninsula based on wastewater inputs and dilution by recharge from precipitation should be about 500 µmol/L. These estimates indicate an annual loading of 3,675 kg N/km² to the aquifer below the Seacoast Shores peninsula. Colman and others (2018) summarized the historical residential development on the Seacoast Shores and the legacy of onsite domestic wastewater disposal that led to contamination of the subterranean aquifer with nitrogen.

Previous Investigations and Conceptual Flow Model

Previous studies on Cape Cod have examined the Seacoast Shores groundwater system, the Cape Cod subterranean estuary, tidal mixing at Waquoit Bay, and the fate of nitrogen in the subterranean estuary at the Seacoast Shores peninsula.

Well Installation

Pushpoint samplers (MHE Products, East Tawas, Michigan) for collection of water samples were installed vertically into the bottom sediments beneath the Eel River estuary to the desired depth to sample groundwater along transects A–A′ and D–D′ (fig. 4). Each sampler consisted of a stainless steel 3.175-cm-diameter, 15.24-cm-long, 0.010-slot well screen connected to 152.4-cm lengths of 3.175-cm-diameter stainless steel pipe. Multiple pushpoints at 12 approximately evenly spaced locations from 1 m onshore to 13.5 m offshore were installed along these transects. The positions of offshore locations are reported relative to the shoreline. The depths and horizontal distances from the shoreline of all well points in transects A–A′, B–B′, C–C′, and D–D′ are shown in figure 5 and reported in Huntington and others (2024). The latitude and longitude of the location of the shoreline for each transect are listed for the points described as 0.0 feet along each east-west transect (Huntington and others, 2024). The distances and directions from the shoreline referred to in this report correspond to the values on the x-axes for all relevant figures. Four pushpoints were installed in each of the 12 clusters. The samplers were installed so that the bottoms of the well screens were positioned at 30, 61, 91, or 122 cm below the sediment/water interface. Three additional permanent Solinst well point clusters (transect C–C′, fig. 4) were installed parallel to and about 3 m northeast of transect D–D′. These clusters were located 4.3, 9.8, and 14.3 m from the shore. The three Solinst well points in these clusters were installed so that the bottom of the well screens were positioned at depths of 1.83, 3.35, and 4.88 m below the sediment/water interface (transect C–C′, fig. 4) by using a slide-hammer or an air-driven post driver.

Another transect of well clusters that was installed in the earlier study also was sampled during this study. That transect (B–B′, fig. 4) consisted of 12 well clusters of stainless steel pushpoint samplers (MHE Products Inc.) and permanent well points (AMS Inc., American Falls, Idaho), from the onshore site (F741) spaced evenly and extending 13.7 m offshore (F745) parallel to the other transects and midway between transects A–A′ and D–D′. The deepest samplers in each cluster were AMS permanent well points, and the remainder were MHE pushpoints. These well clusters each included four wells installed with the bottom of the well screens positioned 30, 61, 91, or 122 cm below the sediment/water interface. The installation of those wells is described in Colman and others (2018).

Measurement of Hydraulic Head and Interpretation of Flow Direction

Water-level altitudes were measured during October 24–26, 2016, in the nine drive-point wells at the three well clusters (transect C–C′; sites F773, F774, and F776, fig. 4) at low, high, and mean tides to determine hydraulic-head gradients. The measurements were made with HOBO dataloggers (Onset Computer Corp., Bourne, Massachusetts). The values were plotted in vertical cross sections relative to the bottom of the Eel River by using graphing software. Contour lines were hand-drawn at varying intervals, and arrows were added perpendicular to the contours to indicate direction of flow.

A pressure-transducer tide gage was installed to monitor the water level in the Eel River near the study area, and hydraulic heads were measured during October 24–26, 2016. The gage consisted of an open-bottom polyvinyl chloride (PVC) standpipe that was secured to a permanent dock piling at a nearby private residence. The tide gage installation is described in Colman and others (2018, app. 1). Measurements of tidal elevation were not recorded at the Eel River during water-quality sampling in 2015 and 2016. Tide data for the sampling periods in this study were retrieved for the U.S. Geological Survey (USGS) tide gage for Popponesset Bay, Mashpee Neck Road, near Mashpee, Massachusetts (USGS site 413601070275800), from the USGS National Water Information System (NWIS) database (U.S. Geological Survey, 2018).

Water-Quality Sampling and Laboratory Analyses

Well clusters in transect B–B′ were sampled once during November 9–10, 2015. Well clusters in transect D–D′ were sampled once during June 14–15, 2016. Well clusters in transect A–A′ were sampled once during June 15–16, 2016. Well clusters in transect C–C′ were sampled once on October 24, 2016.

Samples for water-quality analyses were collected by peristaltic pump and polyethylene tubing. Field water-quality parameters were measured by using parameter-specific meters for water temperature (in degrees Celsius [°C]), temperature-corrected specific conductance (SpC) (µS/cm), and pH. Methods for water temperature and pH are described in the USGS national field manual (U.S. Geological Survey, variously dated). Methods for SpC are described in a U.S. Geological Survey Techniques and Methods book chapter (U.S. Geological Survey, 2005). Dissolved oxygen concentration (DO as O₂) in milligrams per liter (mg/L) was analyzed colorimetrically with the indigo Rhodazine-D method (less than 1 mg/L) (Chemetrics, Inc., 2015) or by using a dissolved oxygen probe and meter (greater than 1 mg/L) (U.S. Geological Survey, variously dated). Alkalinity, in microequivalents per liter (µeq/L), was analyzed by titration (Fishman and Friedman, 1989).

Samples for analysis of nitrate plus nitrite, ammonium, total Kjeldahl nitrogen, and total phosphorus concentrations were filtered (0.45 micrometer), preserved with sulfuric acid to pH <2, refrigerated at <4 °C, and shipped to analytical laboratories. Nitrate plus nitrite was analyzed by hydrazine reduction and colorimetric analysis (EPA Method 353.1) (Wendt, 1995). Ammonium was analyzed by the alkaline phenol method and colorimetric analysis (EPA Method 353.1) (U.S. Environmental Protection Agency, 1983b). Total Kjeldahl nitrogen was analyzed by Kjeldahl digestion with a mercury catalyst followed by colorimetric analysis (EPA Method 353.2) (O’Dell, 1993). Dissolved organic nitrogen concentrations were determined from the difference between total Kjeldahl nitrogen and ammonium nitrogen concentrations. Total phosphorus was analyzed by Kjeldahl digestion with a mercury catalyst followed by molybdate colorimetric analysis (EPA Method 365.4) (U.S. Environmental Protection Agency, 1983a). DOC was analyzed with a Shimadzu TOC–VCPH analyzer using high temperature catalytic oxidation followed by nondispersive infrared detection of carbon dioxide (EPA Method 415.3) (Potter and Wimsatt, 2005). The median relative percent difference for duplicate samples analyzed for nutrients was 4.2 percent (range 0.6 to 32 percent). In some cases, analyses of constituent concentrations were reported as beneath the level of quantitation. In the associated data release (Huntington and others, 2024), these results were reported with a “<” symbol (for example, <0.019). Values that were beneath the level of quantitation were plotted as 0 in the figures in this report.

Samples for analysis of dissolved nitrogen gas and argon gas concentrations were collected in 12.4-milliliter (mL) glass vials with a threaded cap and butyl rubber septum. Vials were filled from the bottom with a continuous stream of pumped groundwater, gently overflowed with greater than 5 vial volumes, capped quickly without bubbles or headspace, and stored under water at less than 4 °C. In the laboratory at the Woods Hole Oceanographic Institution, samples were analyzed by membrane inlet mass spectrometry (Kana and others, 1994; Young and others, 2013; Szymczycha and others, 2017). Sample water was pumped through a section of capillary-diameter, gas-permeable, silicon tubing with the external space under vacuum. Extracted gas passed through a liquid nitrogen trap to remove water vapor and carbon dioxide, through a copper reduction tube at 600 °C to remove oxygen, and to a quadrupole mass spectrometer. The quadrupole mass spectrometer signals representing dissolved gas concentrations were recorded for the three masses of nitrogen gas (28, 29, 30) and the primary mass for argon gas (40). A deionized water bath, held at constant temperature (10 °C) and circulated continuously with a head space at 100-percent relative humidity served as a calibration solution for the gases and was sampled repeatedly before and after each set of 15 to 20 field samples. Readings for field samples were corrected for minor instrument drift. The relative standard deviation of three replicated analyses of the calibration bath was 0.02 percent for ²⁸N₂ and 0.01 percent for ⁴⁰Ar. Duplicate samples collected from 12 locations using the same collection method, with duplicates collected sequentially from a stream of pumped groundwater, resulted in an average relative standard deviation of 2.0 percent, with a range of 0.04 to 9.0 percent. Data from samples that degassed substantially during sample handling were excluded; minor degassing may have contributed to overall uncertainties of the evaluated dissolved gas data, as described in the section “Evaluation of Nitrate Reduction to Nitrogen Gas.”

Additional information on methods is provided in the accompanying data release (Huntington and others, 2024). Reported concentration units for various parameters in the data release were converted for discussion in this report. Quality assurance procedures were previously described (Colman and others, 2018).

Determination of Nitrogen Attenuation

The primary method for quantifying permanent removal of fixed nitrogen from the subterranean estuary in this study was to evaluate the distribution of nitrogen gas produced by microbial nitrate reduction ± anaerobic ammonium oxidation (denitrification ± anammox). Other microbial nitrogen transformation processes such as dissimilatory nitrate reduction to ammonium (DNRA) or assimilation are more difficult to quantify using porewater data where potentially large reservoirs of solid-phase ammonium and organic nitrogen also exist, and they are more likely to represent temporary fixed nitrogen sinks that can be oxidized, mobilized, and discharged subsequently.

Measured concentrations of dissolved nitrogen gas (N₂) and argon gas (Ar) were used to estimate concentrations of excess nitrogen gas produced by nitrate reduction that could indicate nitrogen attenuation. Interpretation of denitrification or other microbial metabolic pathways for nitrogen gas production in groundwater must be conducted in the context of conditions in which the groundwater was recharged that control “background” gas concentrations in the absence of microbial nitrogen gas production (for example, Böhlke and others, 2002; Griggs and others, 2003; Szymczycha and others, 2017). The following processes can contribute to dissolved nitrogen gas in groundwater: (1) equilibration with the atmosphere and soil gas during recharge to the aquifer, with solubility determined by temperature, salinity, and atmospheric pressure (elevation), (2) dissolution of air bubbles entrained during recharge, resulting in “excess air” (EA), and (3) nitrogen gas production by microbial processes (denitrification or anammox). Argon gas occurs in groundwater due to processes 1 and 2 and, depending on various assumptions, can be used as a tracer to estimate nitrogen gas concentration resulting from those processes. Dissolved nitrogen gas and argon gas concentrations were evaluated as follows (Böhlke and others, 2002):

where N_2ASW and Ar_ASW refer to “air saturated water,” or the molar concentrations of nitrogen gas and argon gas from equilibration with the atmosphere at recharge. N_2MIC is the concentration of nitrogen gas produced by microbial processes. N_2EA and Ar_EA refer to concentrations of nitrogen gas and argon gas added to the equilibrium values by dissolution of air bubbles (excess air). Observations commonly indicate that excess air results from complete or nearly complete dissolution of bubbles, resulting in an excess air composition that is similar to atmospheric air, thus allowing estimation of the excess air component of total nitrogen gas as follows (Böhlke and others, 2002): where N_2AMF and Ar_AMF are atmospheric mole fractions of N₂ and Ar, respectively.

In this study, the concentration of microbial nitrogen gas (N_2MIC) in each water sample was used to estimate the concentration of nitrate that was lost by nitrate reduction in the water parcel represented by the sample. The N_2MIC was assumed to have accumulated after the water in the sample entered the saturated zone during recharge, and it was assumed initially that 1 mole of N_2MIC (2 moles of N₂-N_MIC) were produced from 2 moles of nitrate, as in denitrification. The initial nitrate concentration in the water at the point of recharge was estimated as the sum of the measured nitrate concentration and the estimated concentration of nitrate lost (converted to nitrogen gas) in the water sample. Estimated concentrations of nitrate lost and initial nitrate could be smaller if nitrate reduction was coupled with ammonium oxidation (anammox), which could yield 1 mole of N_2MIC (2 moles of N₂-N_MIC) from 1 mole of nitrate. Additional details of these calculations and assumptions are presented in the context of the dissolved gas data in the section “Evaluation of Nitrate Reduction to Nitrogen Gas.”

Hydrogeologic and Geochemical Observations

Observations reported in this section include water levels measured at wells, hydraulic head gradients, field water-quality characteristics measured during sampling, and chemical analytical results from laboratory analyses. All well-location and water-level data collected in this investigation are stored in the National Water Information System database (U.S. Geological Survey, 2018), available at https://doi.org/10.5066/F7P55KJN. All water-quality data are listed in the data release associated with this publication (Huntington and others, 2024).

Evaluation of Nitrate Reduction to Nitrogen Gas

A common method for evaluation of nitrogen gas and argon gas concentrations in groundwater, and estimation of microbial nitrogen gas production, includes plotting the data as scatterplots of nitrogen gas in relation to argon gas (fig. 11). In this plot, the solid lines indicate theoretical dissolved gas concentrations in equilibrium with air (ASW, air-saturated water; Weiss, 1970) for water having salinities and recharge temperatures roughly bracketing those of the samples (0 and 30 practical salinity units [psu]; 10, 15, and 20 °C), and dotted lines indicate the effect of adding 2 mL (standard temperature and pressure, STP) of excess air per liter due to dissolution of bubbles during recharge or sample processing.

The nitrogen and argon gas concentrations in the freshest (salinity <0.1 psu) endmember groundwater samples collected from transects A–A′, B–B′, and D–D′ are clustered in an array that resembles a hypothetical trajectory for air-saturated water (ASW) plus variable amounts of excess air (EA ≈ 0–3 ccSTP/L) at relatively constant recharge temperature (T ≈ 11–12 °C) (fig. 11). Average groundwater recharge temperatures commonly are similar to average air temperatures in humid temperate regions (Stute and others, 1995). Apparent recharge temperature can deviate significantly from average air temperature where the vadose zone is thin, and, because of warming through time of air and soil temperature, estimated recharge temperature can vary with date of recharge (Szymczycha and others, 2017). Thus, the observed scatter of points around the estimated recharge temperature and EA line is likely due to variations in recharge temperature as well as procedural and analytical errors. As the average air temperature at the nearby Barnstable airport from 1991 to 2020 was 11.1 °C (Northeast Regional Climate Center, http://www.nrcc.cornell.edu, accessed on April 11, 2022), the clustering of gas data indicating a similar recharge temperature with variable excess air (fig. 11) suggests that gas concentrations in the fresh groundwater endmember samples were largely controlled by physical processes, and that microbial nitrogen gas was relatively minor.

In high-salinity groundwater samples (salinity ≥ 25 psu), gas concentrations plot in an array that is roughly parallel to the ASW curve for saline water (30 psu), with a relatively wide range of apparent recharge temperatures (fig. 11). Data suggest relatively little EA content, and little or no clearly resolvable N_2MIC in the high salinity samples, consistent with general lack of nitrate in saline estuarine recharge. The pattern observed for samples with intermediate salinity (0.1 < salinity ≤ 25 psu) (fig. 11) contrasts strongly with the pattern for the fresh and saline endmember samples. Gas concentrations in many of these intermediate-salinity samples show evidence of substantial microbial production of nitrogen gas, resulting in a large shift to the right along the nitrogen gas axis (fig. 11). Previous studies have provided evidence for microbial removal of nitrate in subterranean estuaries on Cape Cod, Massachusetts, where nitrate conveyed by fresh groundwater approaches discharge areas near the land/sea margin where reducing, saline groundwater also is present (Kroeger and Charette, 2008; Colman and others, 2018).

Quantification of nitrate losses, as freshwater parcels transit toward the estuary and mix prior to discharge with saline groundwater that has recharged through Eel River sediment, requires a different approach to that commonly used for fresh groundwater samples (Böhlke and others, 2002; Griggs and others, 2003; Szymczycha and others, 2017). In estuarine research, it is common practice to examine behavior of dissolved constituents across the salinity gradient, to determine whether concentrations deviate from conservative mixing between representative freshwater “endmember” concentrations and saline endmember concentrations. Concentrations above such mixing lines are evidence of production of the constituent of interest during transit through the subterranean estuary, while concentrations below the conservative mixing line provide evidence for loss or consumption of the constituent of interest. In groundwater settings where dispersion is limited, substantial variability in constituent concentrations can occur at relatively small spatial scales, and there can be uncertainty regarding which freshwater and saltwater samples best represent the endmembers for any given mixed salinity sample. Those features present a challenge for analysis of conservative mixing. Nonetheless, the ranges of concentrations and concentration changes in groundwater environments commonly are large enough that chemical transformation can be readily detected, particularly where a large number of samples have been collected across each salinity zone (Kroeger and Charette, 2008).

In the current study, we applied conservative mixing analysis for dissolved gas samples in the subterranean estuary beneath the Eel River (fig. 12A–12E). Samples were divided into groups representing a “fresh” endmember (<0.1 psu) that displayed evidence of varying excess air but no clear evidence of N_2MIC, a “saline” endmember (>25.0 psu) that presumably was recharged from the estuary with little or no nitrate and also displayed little or no evidence of N_2MIC, and “mixed” samples (≥0.1 to 25 psu) with evidence of varying N_2MIC concentrations. Whereas the individual gas concentrations vary substantially because of varying physical factors such as recharge temperature and excess air, the N₂/Ar ratios are less affected by those factors and relatively more sensitive to addition of microbial nitrogen gas. Data from the “fresh” and “saline” endmembers were used to determine a “baseline” relation between salinity and the N₂/Ar ratio for samples most likely to have no N_2MIC (fig. 12C). The N₂/Ar baseline equation was constrained by the mean values of psu and N₂/Ar in the fresh and saline endmembers and by the assumption that the mean values of calculated N_2MIC concentrations in those groups should be 0:

Samples that plot substantially above this N₂/Ar mixing baseline indicate excess (non-atmospheric) nitrogen gas that may be a result of nitrate reduction (that is, N_2MIC). The apparent concentration of N_2MIC in each sample was estimated from:

Calculated N_2MIC concentrations are summarized in figure 12D, where small positive and negative apparent values (for example, −25 µmol/L < N_2MIC < 25 µmol/L) are related in part to uncertainties in the data and assumptions, as described below. The relation between calculated N_2MIC and salinity indicates that nitrate loss occurred over a wide range of salinity values from nearly fresh to greater than 24 psu (fig. 12D).

For the “fresh” samples (n=37), equations 3 and 4 yielded a mean value of 0±13 (stdev) µmol/L for N_2MIC and was consistent with a mean recharge temperature of 11.5 °C and excess air concentration of 1.1 ccSTP/L. For the “saline” samples (n=10), these equations yielded a mean value of 0±8 (stdev) µmol/L for N_2MIC, which was consistent with a mean recharge temperature of 13.8 °C and excess air concentration of 0.6 ccSTP/L. For the “mixed” samples (n=82), equations 3 and 4 yielded variable concentrations of N_2MIC ranging up to 217 µmol/L. The mean estimated fresh groundwater recharge temperature was similar to the mean annual air temperature in the area (11.1 °C), as discussed above; whereas the slightly higher mean estimated saline groundwater recharge temperature may have been related to the fact that many of the shallow saline samples were collected when infiltrating surface water from the estuary would have been warmer than the mean annual air temperature (for example, June 2016). The apparent difference in endmember excess air concentrations was consistent with previous coastal groundwater studies indicating subaqueous (saline) recharge is less likely to incorporate excess air than terrestrial (fresh) infiltration recharge (Böhlke and Krantz, 2003; Kroeger and Charette, 2008; Colman and others, 201818). The current calculations differ slightly from the approach used by Colman and others (2018), which was based on the approximation that all samples had 1±1 ccSTP/L of excess air.

Relatively high and variable excess air concentrations in fresh samples were indicated by an array of sample data in the upper part of the Ar-N₂ diagram (fig. 11). For comparison with results of the N₂/Ar baseline equation (eq. 3), N_2MIC concentrations for individual fresh samples also were calculated by assuming they were all recharged at a constant temperature and with variable excess air. For a constant recharge temperature of 11.53 °C, the mean N_2MIC concentration in fresh samples was 0±17 µmol/L, indicating more variability with this approach than with equation 3, for the same assumed mean value of 0 for N_2MIC.

For the saline samples, an alternative approach to equation 3 was based on the assumption that no excess air was incorporated during recharge beneath the estuary. In this case, N_2MIC concentrations were mostly positive, with mean value of 12±10 µmol/L. Thus, some N_2MIC may have been present in saline samples, but the calculated concentrations were not significantly different from zero and it is possible that trace amounts of excess air were introduced as a result of imperfect sample handling.

For each sample, the concentration of N_2MIC derived from equations 3 and 4 was used to quantify the concentration of nitrate that may have been reduced to nitrogen gas by denitrification, where 2 moles of nitrate are consumed to produce 1 mole of N₂:

If anammox was an important process of nitrogen attenuation, then each mole of N_2MIC could represent as little as 1 mole of nitrate loss, with concurrent consumption of ammonium. For plotting and interpretation of the data, results were considered to be insignificant if less than 25 µmol/L as N_2MIC (figs. 12D and 12E) and 50 µmol/L as N₂-N_MIC, based on the standard deviations of the “fresh” and “saline” endmember values and additional uncertainties that could be related to sample handling (for example, air contamination, leakage, gas loss) or the calculations (for example, incompletely constrained assumptions). Calculation uncertainties were related to the fact that multiple variables affect the total concentration of nitrogen gas in a groundwater sample, including recharge temperature, salinity, amount of excess air incorporated, fractionation of excess air, production of excess nitrogen gas from denitrification (or anammox), amount of degassing, and fractionation during degassing; thus, measurements of Ar, N₂, and salinity do not provide unique estimates of the nitrogen gas yield from nitrate reduction. Additional measurements, such as other noble gases, can help reduce those uncertainties in some situations (for example, Böhlke and others, 2007).

For each sample with dissolved gas data, measured nitrate plus nitrite concentrations were combined with the calculated concentration of nitrate loss (assuming all N₂-N_MIC was from denitrification) to obtain an estimate of the nitrate concentration the sample would have had in the absence of nitrate loss (apparent initial nitrate concentration). For these calculations, zero values were assigned to censored data (less than 7 µmol/L for measured nitrate plus nitrite and less than 50 µmol/L for nitrate loss). When plotted against salinity, these results provide useful constraints on major factors affecting the distribution of nitrate in the shallow subterranean estuary (fig. 13).

In figure 13A, measured nitrate plus nitrite (“NO₃⁻”) concentrations in samples analyzed for argon gas and nitrogen gas plot within the triangular area defined by (1) saline water with no initial nitrate, (2) freshwater with no initial nitrate, and (3) freshwater with high initial nitrate (maximum near 700 µM). With few exceptions, calculated concentrations of N₂-N_MIC (nitrate loss) and initial nitrate (assuming denitrification) also plot within the same triangular area (figs. 13B and 13C). These patterns are consistent with the interpretation that freshwater was recharged on land with variable concentrations of nitrate (more or less contaminated by wastewater) and then the nitrate was variably reduced microbially and variably diluted by mixing with saline water in the subterranean estuary. These data do not indicate clearly where the nitrate reduction occurred (only that it had occurred between recharge and sampling), nor do they indicate a common factor that controlled the distribution of nitrate reduction (which could have been promoted by electron donors in the freshwater, saline water, or sediment). Some of the largest concentrations of N₂-N_MIC were in samples that were nearly fresh, indicating that aqueous electron donors in saline water likely were not the predominant controlling factor in those cases.

Some of the nearly saline samples with no measurable nitrate apparently had significant concentrations of N₂-N and initial nitrate that plot outside the triangular area described above. Assuming no problems with sample handling, storage, and analyses, those data potentially could be consistent with either (1) dilution and denitrification of a freshwater component with initial nitrate concentration substantially higher than implied by the other samples (for example, around 1,200 µmol/L) or (2) anammox rather than denitrification as the predominant nitrate reduction process in a freshwater component more similar to the other samples (maximum around 700 µmol/L). Option 1 could be consistent with nitrate concentrations up to 1,700 µmol/L reported previously for deeper fresh groundwater beneath the study site (Colman and others, 2018), although flow paths consistent with this hypothesis have not been identified. Option 2 could be consistent with apparent ammonium deficits in some samples with moderate to high salinities, in comparison to conservative groundwater mixing (dilution) curves. If anammox were indicated as a predominant process in any of the samples, then it could have contributed partially to nitrate loss in other samples as well. Additional studies involving isotopic tracers, microbial data, and field experiments could help resolve these possibilities (for example, Smith and others, 2015).

Results summarized in figure 13 were aggregated to yield mean values and standard deviations for the concentrations of measured nitrate plus nitrite (158±202), nitrate loss (60±110 µmol/L), and initial nitrate (219±191 µmol/L) for the complete sample set with reported nitrogen gas and argon gas concentrations. These averages include saline samples that were recharged with little or no nitrate as well as fresh and mixed samples that were recharged with variable nitrate concentrations. The ratio of nitrate loss to initial nitrate in this evenly weighted sample set would be equivalent to a fractional nitrate loss of 27 percent. It is not possible to assess the accuracy and uncertainty of these aggregated data as representations of overall nitrogen removal in the subterranean estuary. Nonetheless, there is a high degree of certainty that a detectable quantity of land-derived groundwater nitrate was reduced to microbial nitrogen gas within the subterranean estuary before being discharged, as indicated by nitrogen gas concentrations that were commonly greater in the mixed salinity sample set than in the endmember sample sets (figs. 11 and 12C; table 1).

Patterns and Controls of Nitrogen Transport and Attenuation

Some of the major processes affecting the distribution of nitrate can be resolved by comparing the distributions of salinity, measured nitrate, and nitrate lost along transects A–A′, B–B′, and D–D′ (fig. 14). The extent, or completeness, of denitrification at a given sampled location can be inferred from the proportion of estimated nitrate loss to initial nitrate (sum of nitrate measured and lost). In most locations in the region from A03 to A08 (from 2.5 to 8.5 m from the reference point on the shoreline), nitrate removal from fresh and brackish groundwater was less than 50 µmol/L and maximum nitrate concentrations (500 and 700 µmol/L) were similar to those associated with the underlying groundwater moving upwards in the interval from 2 to 10 m below the sediment/water interface measured by Colman and others (2018) (fig. 3). In contrast, from A09 to A12 (from 10 to 14 m from shore), fresh and brackish groundwater commonly had little or no measured nitrate because of partial or complete nitrate reduction.

The dissolved gas data suggest the following broad patterns regarding nitrate reduction in groundwater sampled in the present study, as detailed in the following paragraphs and as shown in figure 14. Relatively little nitrate reduction was evident in the nearshore fresh groundwater, as indicated by gas concentrations that are largely explained by solubility and by bubble dissolution during recharge (fig. 11). Colman and others (2018) reported substantial N_2MIC in shallow fresh groundwater in the upland more than 30 m from the shore (fig. 3). In the highest salinity zones, there may be weak evidence for minor nitrate reduction (24±20 stdev µmol/L, assuming no excess air); this could indicate a small nitrate source in estuarine surface water or estuarine sediment porewaters, possibly associated with coupled nitrification-denitrification near the sediment/water interface. The major occurrences of nitrate reduction were in or near areas where nitrate-bearing terrestrial fresh groundwater encountered more reducing saline groundwater recharged beneath the Eel River estuary, where electron donors may have been supplied by infiltrating surface water and shallow subestuarine sediments (for example, Kroeger and Charette, 2008).

Geochemical indicators of nitrate reduction were observed principally during the June 2016 sampling of transect A–A′, November 2015 sampling of transect B–B′, and October 2016 sampling of transect C–C′. Samples with the highest concentrations of excess nitrogen were observed 9 m or more from shore. Estimated nitrate loss was greatest in the zone from 10 to 14 m offshore and depths of 61 to 122 cm in the subterranean estuary of transect A–A′ and ranged from 230 to 434 µmol/L of nitrate (fig. 14A). Estimated nitrate loss in the zone from 10 to 14 m offshore and depths of 30 to 122 cm in the subterranean estuary of transect B–B′ ranged from 54 to 275 µmol/L of nitrate (fig. 14B). Four of the 12 samples collected from these locations in transect B–B′ had detectable nitrate at concentrations ranging from 58 to 430 µmol/L (fig. 8E). We cannot rule out the possibility that some fraction of the excess nitrogen gas was derived from anammox rather than denitrification. Almost all samples collected from transect D–D′ in June 2016 had estimated nitrate loss of less than 50 µmol/L (fig. 14C). Where estimated nitrate loss was less than 50 µmol/L (generally within about 8 m of the shoreline position at mean high water), there may have been some nitrate reduction but the concentrations of N₂-N_MIC were within the estimated combined uncertainties of the data and assumptions.

A limited number of analyses in this study in transect C–C′ showed evidence of partial to complete nitrate removal at 178 cm below the sediment/water interface in the subtidal zone from 2.5 to 14 m offshore. The nitrate concentrations in these samples were 66, 138, and 0 µmol/L at 2.75, 8.25, and 13.75 m from shore, respectively. Colman and others (2018) measured partial nitrate removal of 71 µmol/L and nitrate concentration of 346 µmol/L in a similar location to the 178-cm depth of transect C–C′ at 152-cm depth and 9 m offshore (fig. 3).

Dissolved oxygen concentrations generally were low (less than 25 µmol/L as O₂) in the regions of the highest nitrate loss (figs. 7B and 8B), which is consistent with an environment suitable for denitrification activity. Alkalinity was elevated relative to conservative mixing of fresh and saline endmembers in some areas of high nitrate loss (figs. 7D and 8D), which could be consistent with potential denitrification activity because alkalinity is a biproduct of microbial oxidation of organic matter; however, the data are incomplete and appear to be somewhat inconsistent with simple reaction stoichiometries. In relatively saline groundwater from the shallowest sampled depths in the subterranean estuary (30 cm) and from 9 to 13.5 m offshore, ammonium concentrations (figs. 7F, 8F, and 9F) were elevated in comparison to those of either fresh groundwater or overlying saline surface water, based on ammonium measurements in the Eel River reported by Colman and others (2018). Elevated ammonium concentrations in saline groundwater likely were generated by anaerobic biodegradation of sedimentary organic matter. However, some mixed-salinity samples with detectable N_2MIC appeared to be relatively depleted in ammonium when compared with conservative mixtures of fresh and saline groundwater endmembers, possibly indicating ammonium loss (for example, anammox) or sediment-water exchange processes. DOC concentrations were relatively low in fresh groundwaters that had high initial nitrate concentrations, including some that were denitrified and some that were not denitrified. This could be viewed as evidence for the importance of electron donors from solid phases, rather than aqueous species, as controls of localized nitrate reduction (as in the subtidal area of transect A–A′ (figs. 7H and 14A).

Hydraulic gradients and vertical profiles of SpC indicate that fresh groundwater originating from recharge under the peninsula migrates towards the estuary as depicted in the conceptual dashed yellow lines in figure 3. Figure 3 is modified from the earlier study (Colman and others, 2018) to illustrate our current understanding of groundwater flow in this subterranean estuary and how that compares to the earlier conceptual model shown in figure 2. In two of the three transects in the current study (A–A′ and B–B′), nitrate concentrations in the near-shore intertidal zone generally were relatively low initially and may have been partially diluted by mixing with saltwater; this nitrate was minimally or not denitrified and some of it may have been discharging to the estuary. In the third transect (D–D′), nitrate concentrations were substantially lower. The nitrate in the intertidal zones of these transects was minimally or not denitrified and some of it may have been discharging to the estuary. Slightly farther offshore in a narrow region from 5 to 6 m from shore, generally there was little evidence of mixing with infiltrating saltwater and low nitrate concentrations and minimal evidence of denitrification; small amounts of nitrate may have been discharging to the estuary in this zone. Farther from shore in the subtidal area from about 7 to 12.5 m offshore, salty porewater overlays the freshwater plume, indicating no freshwater discharge directly to the estuary and possible horizontal offshore movement of groundwater. Groundwater in this freshwater zone was depleted in nitrate and this is consistent with evidence for nearly complete nitrate reduction. Farther offshore, Colman and others (2018) reported fresh groundwater with relatively high nitrate concentrations from about 1.5 to 11 m below the sediment/water interface, but data were not available at shallower depths to indicate whether mixing, reaction, or discharge might have occurred there (fig. 3). Thus, there appears to be a large body of fresh groundwater containing substantial nitrate underlying the estuary farther offshore than our measurements, but there are insufficient data to determine where or how it discharges.

Summary

This study was conducted to follow up on a previous study (Colman and others, 2018, https://doi.org/10.3133/sir20185095) that investigated the fate and transformations of nitrogen in the groundwater below the Seacoast Shores peninsula. The previous study found evidence of nitrate reduction near the water table in fresh groundwater onshore and near the freshwater/saltwater transition zone beneath a 7–8 meter (m)-thick freshwater zone in the subterranean aquifer. That study documented variable concentrations of nitrate in fresh and brackish shallow groundwater beneath intertidal and subtidal portions of the estuary but did not assess the distribution of nitrate reduction in those areas. The current study focused on evidence for the distribution of nitrate reduction in shallow groundwater from 0 to 1.5 m beneath the intertidal and subtidal portions of the estuary.

The distributions of fresh and saline groundwater in the subtidal area of the subterranean estuary under the Eel River are consistent with shallow intertidal and subtidal saltwater cells where infiltrating saline and underlying freshwater mix. However, these distributions do not conform to the simplified conceptual model of groundwater flow in the subterranean estuary at Waquoit Bay, East Falmouth, Massachusetts, where a deep saltwater wedge discharges to the estuary. At the location of this study in the Eel River subterranean estuary, to a distance offshore of at least 18 m, lateral subsurface fresh groundwater flow apparently prevented a deep saltwater wedge from discharging to the surface near the shoreline (Colman and others, 2018). Studies elsewhere have documented similar offshore flow of fresh groundwater lenses beneath estuaries, in some cases extending hundreds of meters offshore (for example, Bratton and others, 2004, 2009; Russoniello and others, 2013). The distribution of salinity determined in this study, and in Colman and others (2018), although variable over time, indicates a nearshore intertidal cell and a subtidal saltwater cell where rising fresh groundwater mixes with infiltrating saline water, and an intermediate zone in the intertidal area closer to mean low water where freshwater discharges without mixing with saline water. Specific conductance data were collected to about 18 m offshore in the Colman and others (2018) study and to about 13.5 m in this study, and both studies seem to indicate fresh groundwater discharges into the estuary farther offshore beyond the shallow subtidal saltwater cell.

Hydraulic-head gradients measured 2 to 5 m below the sediment/water interface in October 2016 indicated that fresh groundwater flow generally was upwards (towards the estuary) under all tide conditions, providing a mechanism for transport of nitrate from deeper nitrate-rich groundwater (that was documented in Colman and others, 2018) into the shallower subestuarine sediments where fresh and saline porewaters mixed.

Nitrate-bearing fresh and brackish groundwater in some of the shallowest samples (less than 61 centimeters below the sediment/water interface) might indicate direct discharge of groundwater nitrate to the estuary in portions of the intertidal and subtidal zones. In other areas, fresh and brackish groundwater was overlain by nitrate-free saline water, indicating little or no direct local nitrate discharge. Variability of nitrate concentrations in shallow intertidal and subtidal groundwater was related to a combination of (1) variable original nitrate concentrations in fresh groundwater recharge, (2) dilution by mixing with nitrate-free saline water recharged beneath the estuary, and (3) nitrate reduction along groundwater flow paths.

The primary geochemical indicator of nitrate reduction (excess nitrogen gas) indicated only minimal attenuation of nitrate by microbial denitrification or anammox in the intertidal zone. Other nitrate reduction processes such as dissimilatory nitrate reduction to ammonium and assimilation were not evaluated quantitatively, in part because they involve interactions with solid phases that were not sampled. There was no aqueous evidence for substantial ammonium production in mixed-salinity samples, and the distributions of initial nitrate, nitrate remaining, and concentration of microbially produced nitrogen gas, as N₂ (N_2MIC) could reasonably be explained as a result of mixing and local denitrification (± anaerobic ammonium oxidation [anammox]) of freshwater sources of nitrate. In the subtidal zone, data were consistent with complete removal of nitrate by denitrification in fresh groundwater in one of the transects measured in June 2016. Anammox was not measured directly and it cannot be ruled out as a potential microbial source for some of the excess nitrogen gas. Excess nitrogen gas, determined from analysis of dissolved gases (nitrogen gas and argon gas), indicated that microbial conversion of nitrate to nitrogen gas was likely occurring in the intertidal zone and especially in the subtidal zones. The estimated amounts of nitrate loss were less than 50 to 139 micromoles per liter (µmol/L) in the intertidal zone and less than 50 to 434 µmol/L in the subtidal zone. Estimated amounts of nitrate reduction were highest in subtidal area greater than 10 m from the shore in one of two of the transects sampled in June 2016 (A–A′), in one transect sampled in November of 2015 (B–B′), and in another sampled in October 2016 (C–C′). In the highest salinity zones, there may be weak evidence for minor nitrate reduction (24 ± 20 standard deviation µmol/L, assuming no excess air). Although these values are considered to be within the uncertainties of the calculations, they could potentially indicate a minor nitrate source in estuarine surface water or estuarine sediment porewaters, possibly associated with coupled nitrification-denitrification near the sediment/water interface.

The primary zone where evidence for nitrate reduction was detected was in subtidal fresh groundwater that was overlain by a thin layer of saline and mixed groundwater. The fact that geochemical evidence for nitrate reduction was less pronounced in other areas of the subterranean estuary that were suboxic but rich in nitrate may indicate more reactive sources of electron donors in shallow subtidal sediments than in shallow intertidal sediments. Fresh groundwater with high initial nitrate concentrations had relatively low dissolved organic carbon (DOC) concentrations in areas with and without N_2MIC, indicating that wastewater DOC may have been depleted by microbial reactions before reaching the shallow subterranean estuary. The concentration and chemical nature of DOC also are important variables that can limit nitrate reduction if the concentration is too low or if the carbon is resistant to decomposition (Devito and others, 2000; Rivett and others, 2008).

The transport of nutrient-rich groundwater through a hydrogeologically complex subterranean estuary can be influenced by variable transmissivity due to the occurrence of heterogeneous confining layers that can restrict flow (Mulligan and others, 2007; Bokuniewicz and others, 2008; Michael and others, 2011; Russoniello and others, 2013), tidal and current dynamics (De Simoni and others, 2005; Rezaei and others, 2005) that affect mixing between fresh groundwater discharge and saltwater infiltration, and the biogeochemical processes that can attenuate nutrient transport. Previous geologic borings on the Seacoast Shores peninsula encountered fine sands and silts directly upgradient of the study area (Colman and others, 2018). Natural heterogeneities of the subsurface lithology combined with tidal and hydrologic effects, therefore, make it difficult to rely on a single conceptual model that describes groundwater flow in these complex subterranean estuaries in Cape Cod embayments.

Further study could better quantify the seasonal, spatial, and tidal variation in excess nitrogen gas in the subterranean estuary adjacent to the Seacoast Shores peninsula, because large variations were observed in this and the previous study. Densely spaced vertical and longitudinal water-level measurements and sampling of groundwater along potential or suspected flow paths from septic sources to the nearshore subterranean estuary could be used to follow nitrate along these logical flow paths to better discern where transformations may occur. This information could be supplemented with isotopic and conservative tracers to confirm suspected flow paths and timing of transport. Further studies using molecular techniques (environmental DNA) or isotopic enrichment experiments could also be employed to determine the relative importance of denitrification, anammox, DNRA, or other nitrate attenuation processes. The ultimate fate of deep fresh groundwater and associated nitrate prior to discharge might be determined by extending such studies farther offshore and possibly all the way across the Eel River.