The fate of wastewater-derived nitrate in the subsurface of the Florida Keys: Key Colony Beach, Florida

Estuarine, Coastal and Shelf Science 58 (2003) 517–539

The fate of wastewater-derived nitrate in the subsurface of the Florida Keys: Key Colony Beach, Florida Erin M. Griggsa,1, Lee R. Kumpa,*, J.K. Bo¨hlkeb a

Department of Geosciences, The Pennsylvania State University, 535 Deike Building, University Park, PA 16802, USA b United States Geological Survey, 431 National Center, 12201 Sunrise Valley Drive, Reston, VA 20192, USA Received 13 May 2002; accepted 14 April 2003

Abstract Shallow injection is the predominant mode of wastewater disposal for most tourist-oriented facilities and some residential communities in the US Florida Keys National Marine Sanctuary. Concern has been expressed that wastewater nutrients may be escaping from the saline groundwater system into canals and surrounding coastal waters and perhaps to the reef tract 10 km offshore, promoting unwanted algal growth and degradation of water quality. We performed a field study of the fate of wastewaterderived nitrate in the subsurface of a Florida Keys residential community (Key Colony Beach, FL) that uses this disposal method, analyzing samples from 21 monitoring wells and two canal sites. The results indicate that wastewater injection at 18–27 m depth into saline groundwater creates a large buoyant plume that flows quickly (within days) upward to a confining layer 6 m below the surface, and then in a fast flow path toward a canal 200 m to the east within a period of weeks to months. Low-salinity groundwaters along the fast flow path have nitrate concentrations that are not significantly reduced from that of the injected wastewaters (ranging from 400 to 600 lmol kg
1). Portions of the low-salinity plume off the main axis of flow have relatively long residence times (>2 months) and have had their nitrate concentrations strongly reduced by a combination of mixing and denitrification. These waters have dissolved N2 concentrations up to 1.6 times air-saturation values with d15 N½N2  ¼ 0:5
5&, d15 N½NO
3  ¼ 16
26&, and calculated isotope fractionation factors of about
124&, consistent with denitrification as the predominant nitrate reduction reaction. Estimated rates of denitrification of wastewater in the aquifer are of the order of 4 lmol kg
1 N day
1 or 0.008 day
1. The data indicate that denitrification reduces the nitrate load of the injected wastewater substantially, but not completely, before it discharges to nearby canals. Published by Elsevier Ltd. Keywords: nitrogen isotopes; wastewater; nutrients; phosphate; nitrate; groundwater; Florida Keys; denitrification

1. Introduction There is growing concern about the health of the world’s reef tracts, and a number of possible causes for reef decline have been proposed (e.g. Smith & Buddemeier, 1992). One of the suggestions is that reefs near inhabited areas are being degraded in response to an increase in anthropogenic nutrification through wastewater disposal. This concern prompted us to investigate the fate of wastewater nutrients in the Florida Keys, US.

1

Now at ERM Corporation, Exton, PA 19341, USA. * Corresponding author. E-mail address: [email protected] (L.R. Kump). 0272-7714/03/$ - see front matter Published by Elsevier Ltd. doi:10.1016/S0272-7714(03)00131-8

The Florida Keys is a chain of islands off the southern tip of Florida, composed of permeable Pleistocene reefal and oolitic limestone. Thus far, field studies have not provided unequivocal evidence that the wastewaterderived nutrients reach the reefs of the Florida Keys (approximately 10 km offshore) through groundwaters or surface waters (Bo¨hlke et al., 1999; Shinn, Reuse, & Reich, 1994; Szmant & Forrester, 1996). However, wastewater nutrients do appear in canals and nearshore waters closer to the point of injection, and concentrations of nutrients decrease (rapidly) along offshore transects toward the reefs (Lapointe & Clarke, 1992; Szmant & Forrester, 1996). While the focus of the current research is on the fate of nitrogen, previous work has addressed the uptake of

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wastewater phosphate after injection into the subsurface (Corbett, 1999; Corbett, Kump, Dillon, & Chanton, 2000; Dillon, Corbett, Chanton, Burnett, & Kump, 2000; Elliott, 1999; Machusak & Kump, 1997; Monaghan, 1996). Phosphate is substantially removed from the water by adsorbing onto calcite and eventually precipitating as an insoluble phase, likely carbonate fluorapatite (Corbett et al., 2000; Elliott, 1999). As demonstrated below, phosphate concentrations rapidly diminish to ambient levels as the wastewater interacts with the limestone substrate except on the most rapid flow path, where elevated levels persist. Our goals in this study included characterizing the hydraulic properties of the subsurface, ascertaining the amount of nitrate removal, identifying the mechanisms for removal, determining factors that influence removal mechanisms, and constructing a nitrate mass balance for the system. Toward these ends, groundwater samples were collected from the area surrounding a wastewater treatment and injection facility in Key Colony Beach (KCB), FL. Nutrient and nitrogen isotopic analyses were performed to document the pathways and extents of nitrate transformation in the subsurface. Here, we report that although a considerable amount of denitrification occurs in the subsurface, the buoyancy of the wastewater plume and the proximity of surface-

water discharge sites (canals) create favorable conditions for the release of wastewater nitrogen, and to a lesser extent, phosphate, to surface waters. 1.1. Hydrology of the Florida Keys The Florida Keys extend into a 240 km arc from just south of Miami, FL, to the southwest, terminating in Key West, with the remote Marquesas Keys and Dry Tortugas to the west (Fig. 1). The present-day coral reef ecosystem is located approximately 10 km off the seaward coast of the Keys. The bedrock of the Florida Keys in the vicinity of our study site is the Key Largo limestone (KLL) (Hoffmeister & Multer, 1968), a Pleistocene reef tract now infilled with carbonate sands and mud, cemented, and subjected to subaerial diagenetic recrystallization (Halley, Vacher, & Shinn, 1997). The typical KLL of the Upper Keys has a high degree of primary and secondary porosities (45% or greater), permitting rapid flow of water through the subsurface (Shinn et al., 1994). Using a Dupuit–Ghyben–Herzberg analysis, Vacher, Wightman, and Stewart (1992) estimated the hydraulic conductivity of the KLL to be 1400 m day
1. Groundwater beneath the Keys is subject to tidal forcing by both Florida Bay and the Atlantic Ocean. Near the Keys, groundwater tidal fluctuations in

Fig. 1. Map of the Florida Keys, showing the position of the KCB field site.

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Florida Bay are greater than surface-water fluctuations (Machusak & Kump, 1997) because of the propagation of the larger amplitude Atlantic side tide through the subsurface. This potentially creates a tidally reversing direction of groundwater flow (tidal pumping; Halley, Vacher, Shinn, & Haines, 1994). Florida Bay sea level is on an average 10–20 cm higher than the adjacent Atlantic, though, which creates a mean hydraulic gradient driving net transport from Florida Bay to the Atlantic side of the Keys (Halley et al., 1994). The existence of tidal pumping is supported by dye-tracer experiments, which indicate a net horizontal flow from Florida Bay to the Atlantic (Reich, Shinn, Hickey, & Tihansky, 1998). Moreover, the groundwater beneath the Keys tends to be slightly hypersaline, an observation consistent with recharge from Florida Bay, a more restricted body of water that experiences substantial seasonal salinity fluctuations (e.g. Bo¨hlke et al., 1999; Machusak & Kump, 1997). 1.2. Field site: Key Colony Beach The city of KCB is an island on the ocean side of the Keys, located in the middle Keys near Marathon, FL (Fig. 2). KCB has a largely residential community with an official permanent population of 1000, which can

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increase to 3600 residents in the winter. The population fluctuations lead to variations in wastewater production rate (Fig. 3). Once an undeveloped, low-lying small island (Shelter Key), KCB was created through dredging in the early 1950s that expanded the size of the island by a factor of 4. Fill brought the island to 2 m above mean sea level. Thus, there is a 3–6 m carbonate mud cap on the KLL bedrock that is original mud overlain by fill. The island is dissected by a series of canals dredged to variable depths. The city operates a 28-year-old centralized secondary wastewater treatment plant. It is one of a very few municipal sewage collection and treatment facilities in the Keys. Originally, the plant discharged treated sewage directly into Shelter Bay (Fig. 2). In 1994, the plant began injecting the treated effluent into the groundwater system at a rate of 500–1200 m3 day
1 through six gravity-driven injection wells enclosed in a concrete trough. The inlet heights of the six wells are different, and in our observations, virtually the entire wastewater flows down the central well (which has the lowest inlet elevation). Each injection well is cased to 18 m and open to 27 m below the land surface (approximately 26 m below sea level). In July 1999 (midway through our study), the plant was upgraded to provide advanced wastewater treatment (AWT), which for the

Fig. 2. KCB, FL, and an expanded view of the vicinity of the wastewater treatment plant, showing the position of the injection and monitoring wells and the seismic refraction transects (see Fig. 4).

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Fig. 3. Variation of the monthly wastewater injection rate at KCB, FL in the year 1999. Sampling intervals for this study are shown as ovals.

state of Florida is defined as producing water with 3 mg l
1 (200 lmol kg
1) N and 1 mg l
1 (30 lmol kg
1) phosphorus. As of October 1999, the plant was not at full AWT, with nitrate at approximately 400 lmol kg
1 N and phosphate at 30 lmol kg
1 P. 1.3. Characterization of the subsurface The subsurface of KCB consists of consolidated KLL overlain by a layer of Holocene mud. Monitoring-well boreholes drilled at KCB in 1998 penetrated the mud– limestone interface at approximately 6 m depth in all seven boreholes. Five additional boreholes were drilled in 1999. The mud–limestone contact in all 12 drill cores was at approximately 6 m, except for wells J and K, where the contact was at approximately 3 m, and well L with a contact at 4 m (Table 1; Griggs, 2000). Coral dominates the bedrock in these shallow regions, indicating that the topographic highs are former patch reefs. The coral rock is relatively free of the large dissolution

2. Materials and methods 2.1. Monitoring wells

Table 1 Mud–limestone interface depths Well

Depth (m)

Uncertainty (m)a

A B C D E F G H I J K L

6.09 6.09 6.09 6.09 6.09 5.79 6.09 6.55 6.64 3.35 3.35 4.42


0.25 1.5 +1.0,
0.25 +0.3,
0.5 +1.5,
0.2
0.06 +0.9
0.3 +0.25 0.3
0.3
0.3

Depth data from Elliott (1999). Positive uncertainties result when the core recovery, in the drill core section where the mud–limestone interface was encountered, is less than 100%. Negative uncertainties result when the mud–limestone interface may have occurred at a depth shallower than the first recoverable drill core section. a

voids common elsewhere, but does exhibit unconnected borings of up to 7 cm in length created by burrowing clams. Core recovery from the interval just below the mud– limestone interface was high for all wells. Deeper in the cores, recovery varied but decreased significantly from the upper portion. Low core recovery may be a result of large voids in the KLL, sand layers, or friable bedrock ground-up by the drill bit and flushed out of the hole by tap-water drilling fluid. A seismic refraction survey (Griggs, 2000) along three east–west transects (northern, middle, and southern) revealed that the bedrock–mud interface dips to the south, from 3 to 6 m, as it extends over 150 m from the northern to the southern transect (Fig. 4). Moreover, subsurface undulations of up to 3 m exist in each transect. This agrees with the almost 3 m difference in mud–limestone interface depths determined from the drill cores. The northern and middle transects exhibit the largest and most frequent undulations. The southern transect mud–rock interface is at a greater depth and has undulations of smaller amplitude. Most mud–limestone interface depths recorded by the drill cores are similar to those estimated from the seismic survey, but some differ by as much as 1.5 m (Fig. 4). Uncertainties in the seismic first-arrival times are as high as 6 ms and may account for the 1.5 m discrepancies. We argue below that the dominant flow path for wastewater occurs deeper than the paleo-topographically high patch reefs, and that the groundwater in these lower-porosity knobs has a longer subsurface residence time and thus greater extents of biogeochemical transformation (especially denitrification).

The 12 monitoring wells form approximately north– south or east–west transects (Fig. 2). The depths of the boreholes are variable, with the deepest at 20 m. In general, the monitoring wells were installed by drilling until a sandy layer was reached through which the drill could not penetrate. Each hole contains two to three 1.3-cm (0.5-in.)-diameter polyvinyl chloride (PVC) piezometers of different lengths (sampling depths). The bottom 1.5 m of each piezometer is slotted for sampling. (see Table 2 for specific sampling depth information on individual wells). The piezometers are referred to in the text by the well letter (A, B, C, . . .) and the approximate sampling depth in meters; for example, A-9 is well A, 9 m depth. The sampling interval of each borehole was grouted with coarse quartz sand. Non-sampling intervals were grouted with Portland cement (wells A–G) or fine-grained sand (wells H–L). The sand was used

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Fig. 4. Seismic refraction survey depth transects for the mud–limestone interface. Position of transects shown in Fig. 2 (two seismic lines were run for each transect, and are shown in Fig. 2). Letters correspond to monitoring wells and are placed at the depth of the mud–limestone interface, as recorded in the drilling log. Most wells continue to 18 m depth. Slotted (sampling) intervals are shown. Well K falls outside of the seismic survey.

because of earlier problems we had with artificially high pH conditions created by unset cement. The persistence of unstable salinity gradients (e.g. well L) created by wastewater injection indicates that the fine sand formed an effective seal. Portland cement was poured into the borehole near the surface and extended to the land surface to create a cement mound. Finally, each well was covered with a rubber tension well cap (Fig. 5). The positions of the original seven wells (A–G) were surveyed in June 1998 (Elliott, 1999). The positions of

the five wells installed in May 1999 were triangulated using measurements to existing wells. The piezometric water-surface elevation of the wells ranged from about 80 to 95 cm below the (relatively level) land surface, with the more saline wells exhibiting the deeper piezometric surface as expected (Griggs, 2000). The water-surface elevation of all wells fluctuated with a semidiurnal tidal amplitude of approximately 30 cm. The tidal cycle in the wells lagged behind that of the nearby canal by about an hour, and was damped

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Table 2 Field and nutrient analyses results for the June 1999 water samples

Sample A-9 A-18 B-9 B-18 C-9 C-30 D-9 D-14 E-9 E-18 F-8 F-14 F-18 G-9 G-18 H-9 H-13 H-18 I-9 I-14 I-18 J-6 J-13 J-18 K-5 K-14 L-6 L-9 Wastewater

Sample depth interval (m) 7.6–9.1 16.8–18.3 7.6–9.1 16.8–18.3 7.6–9.1 18.3–19.8 7.6–9.1 12.2–13.7 7.6–9.1 16.8–18.3 6.1–7.6 12.2–13.7 16.8–18.3 7.6–9.1 16.8–18.3 7.6–9.2 11.3–12.8 16.8–18.3 7.6–9.2 12.2–13.7 16.8–18.3 4.3–5.8 11.9–13.4 16.8–18.3 4.6–5.2 7.0–7.6 4.3–5.8 7.6–9.2

pH

Salinity

Conductivity (mS cm
1)

7.69 7.46 7.52 7.50 7.25 7.49 7.48 7.52 7.73 7.42 7.33 7.98 7.62 7.34 7.49 7.25 7.43 7.35 7.01 7.35 7.24 7.27 7.56 7.54 7.23 7.55 7.75 7.58 6.71

8.3 18.8 5.5 7.9 6.1 37.4 5.9 37.2 8.3 32.7 10.9 35.7 36.9 7.9 36.2 4.8 4.7 4.9 4.6 4.8 4.8 3.4 37.0 37.7 4.8 6.8 17.0 11.2 4.4

14.24 30.40 16.07 13.77 10.79 56.00 10.47 55.80 14.24 50.00 18.44 54.10 55.10 13.60 36.20 8.55 8.35 8.84 8.31 8.68 8.61 6.23 55.30 56.40 8.52 12.06 27.40 18.98 7.97

Alkalinity (meq kg
1) 3.40 3.10 3.80 3.10 3.40 2.40 3.40 3.00 3.10 3.80 5.00 2.40 2.60 4.40 2.70 3.30 4.00 3.10 3.60 3.60 3.60 3.30 2.80 2.60 5.10 3.70 8.00 3.50 2.40

P Sulfide (lmol kg
1) 8 0 249 0 0 22 16 47 28 156 20 31 203 62 BDL BDL BDL BDL BDL BDL 3 8 28 86 BDL >300 3 0

NH+ 4 (lmol kg
1)

NO
3 (lmol kg
1)

PO3
4 (lmol kg
1)

20 21 149 47 8 12 2 23 31 203 240 12 8 153 27 2 7 15 3 1 2 7 13 12 80 1 852 41 16

531 208 1 657 800 BDL 630 BDL 47 BDL 4 BDL BDL BDL BDL 728 458 663 835 482 785 465 BDL BDL 336 624 BDL 495 677

2.5 41.6 1.2 43.9 39.4 BDL 3.2 BDL 1.3 BDL 0.8 BDL BDL 2.1 1.3 45.8 27.1 48.1 79.7 30.9 61.3 6.4 BDL BDL BDL 1.7 0.4 BDL 80.4


1
1
1
BDL, below method detection limits (NH+ 4 =0.3 lmol kg , NO3 =0.4 lmol kg , phosphate=0.3 lmol kg ).

(the well tidal amplitude was about 60% of that in the canal; Elliott, 1999). An SF6 tracer study was conducted in June 1998 to determine groundwater travel times from the wastewater injection wells to the monitoring-well network (Dillon et al., submitted for publication). Results of this experiment indicate that the predominant direction of flow is to the southeast and that the velocities are locally variable. The SF6 peak travel times to wells within the network range from several days to at least 2 months (e.g. 7 days to C-9, 10 days to B-18, 55 days to D-9, 60 days to A-9, 60 days to the canal west of the injection well, and 70 days to the canal east of the injection well). 2.2. Sample collection For this study, two sampling rounds were conducted at KCB. The June 1999 round was conducted from June 3 to 4. The October 1999 round extended from October 9 to 10. The water sampling procedure followed our EPA-approved Quality Assurance Project Plan (Monaghan, 1996). Prior to entering the field, plastic 250 ml HDPE screw-cap bottles were washed with phosphate-free soap

and tap water. Subsequently, they were rinsed with 1 N HCl, allowed to stand for 10 min, and then rinsed three times with distilled deionized water. After drying, each bottle was capped and labeled with the well letter and depth of the sample to be placed in the bottle. In the field, three well volumes of water were drawn from each well using Tygon tubing and a high-capacity diaphragm pump. Then, a peristaltic pump was used to collect water samples into the prepared sample bottles. Each sample bottle was rinsed three times with sample water before filling to 75% capacity (to allow for expansion upon freezing). Because chlorine, added during wastewater treatment, presents an interference with some nutrient analyses, each sample was dechlorinated following standard method 4500-NH3 (APHA, 1995). Duplicate samples were taken for all the wells and 10% were taken in triplicate. All the samples were immediately placed in a cooler and then frozen for transportation to Pennsylvania State University. We were unable to follow these protocols with the shallow wells with sampling intervals in the mud layer. These wells were quickly pumped dry, and recovered very slowly. Moreover, samples taken from the mud layer tended to be more saline than the waters below

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Fig. 5. Schematic representation of a monitoring-well borehole and piezometer nest (not drawn to scale).

(Elliott, 1999). These observations indicate that the mud layer is of low permeability and likely prevents vertical movement of water, including percolation of rainwater through the mud layer into the groundwater system below. In this study, only piezometers with sampling intervals located in the limestone were sampled. Wells H and I were not sampled during October 1999. At that time, a 32P tracer study was being done and wells H and I, close to the point of injection, had to be bypassed. During October 1999, additional water samples were collected at selected wells for analyses of major dissolved gases (N2, Ar, CH4, O2) and N isotopes of nitrate and dissolved N2. The nitrate isotope samples were immediately transferred to coolers and then frozen later in the same day. The dissolved-gas samples were collected in duplicate in 150 ml pre-weighed serum bottles with blue butyl rubber stoppers. Each serum bottle was filled completely, a pellet of dry KOH (approximately 100– 150 mg) was dropped into the full bottle as a preservative and to ionize free CO2, and then the stopper was

inserted with a 23-gauge syringe needle to release excess water. The samples were stored, with the bottles upsidedown, at room temperature and shipped the following day to the USGS in Reston, VA, for analysis. In the field, temperature, pH, salinity, conductivity, and sulfide measurements were made on samples from each well. Temperature and pH were measured using a Hanna HI model 9023 portable pH meter standardized with NBS buffers. The pH measurements in the field were reproducible to within 0.1 pH units. Conductivity was measured using an Orion model 115 conductivity meter and salinity was determined from conductivity and temperature relationships using algorithms for the computation of fundamental properties of seawater (Fofonoff & Millard, 1983), with a precision of 0.1. A hydrogen sulfide CHEMets kit, which employs the methylene blue colorimetric comparison method (EPA methods 376.2; Kopp & McKee, 1983), was used to determine the sulfide concentration. The precision of this method, using visual comparison, is 10%.

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In the laboratory at Pennsylvania State University, the samples were thawed and filtered through pre-ashed Whatman GF/F glass microfiber filters. Subsequently, one sample from each well depth was analyzed for total alkalinity following EPA method 310.1 (Kopp & McKee, 1983), where the sample is acidified to a pH 4.5, with a precision, on replicate analysis, of 0.1 meq kg
1. After alkalinity analysis, each sample was refrigerated at 4  C until nutrient analysis could be performed, always within 1 week and usually within 1 day of thawing. 2.3. Nutrient analyses All the samples were analyzed in replicate for the inorganic compounds nitrate, phosphate, and ammonium using a Technicon Autoanalyzer II. Nitrate analysis followed EPA method 353.2 (Kopp & McKee, 1983). The results are a measure of the combined nitrate and nitrite in the sample. Samples containing sulfide (as determined in the field) were acidified with HCl and then aerated prior to analysis. Method detection limit (MDL) was 0.4 lmol kg
1. Phosphate spectrophotometric analysis was done by EPA method 365.1 (Kopp & McKee, 1983). This method determines the amount of soluble reactive phosphate, which includes all species of the inorganic phosphoric acid system and, perhaps, some organically bound phosphate. It is referred to simply as ÔphosphateÕ in this document. The MDL for phosphate was 0.3 lmol kg
1. Analysis for ammonium followed EPA method 350.1 (Kopp & McKee, 1983), with an MDL of 0.3 lmol kg
1. Total nitrogen and total phosphorus were determined by simultaneous oxidation of nitrogen and phosphorus compounds with persulfate, following the method of Grasshoff, Kremling, and Ehrhardt (1999). The organic N and P concentrations were obtained by subtracting the inorganic concentration from the total concentration. A Shimadzu TOC-5000A was used to measure the non-purgeable dissolved organic carbon (DOC). Three to four replicate samples were analyzed and the average taken as the value for that sample. Precision of this method is 2%. 2.4. Dissolved gas analyses and estimation of excess nitrogen Dissolved gas analyses were performed in the USGS Dissolved-Gas Laboratory in Reston, VA (see http:// water.usgs.gov/lab/dissolved-gas). The analyses were done by gas chromatography (HP5890) on aliquots of low-pressure headspace, using He as the carrier gas. Headspace was created in the serum bottles by removing 10–12 ml of water, and then the sample and headspace were allowed to re-equilibrate to room temperature.

Subsequently, a sample of the headspace gas was extracted for analysis. Argon, N2, and O2 were quantified with a thermal conductivity detector. Methane was quantified by flame ionization detection. Gas chromatography data were calibrated by analyses of mixed tank gases, and checked by analyses of air-equilibrated water samples prepared in the laboratory and treated as samples. Laboratory-equilibrated water samples yielded N2 and Ar concentrations with uncertainties of 1% or better. The concentrations of N2 and Ar were used to estimate the amount of N2 produced by denitrification (microbial reduction of nitrate to N2 gas). The total dissolved N2 in each groundwater sample (N2,total) was considered to include three major potential components (Bo¨hlke, Wanty, Tuttle, Delin, & Landon, 2002; Heaton & Vogel, 1981; Vogel, Talma, & Heaton, 1981; Wilson, Andrews, & Bath, 1990) N2;total ¼ ðN2;equil Þ þ ðN2;excess-air Þ þ ðN2;denit Þ

ð1Þ

where (N2,equil) is N2 that entered the groundwater system through equilibration with the atmosphere at a given temperature, pressure, and salinity during recharge, (N2,excess-air) the N2 added by entrapment of air bubbles during recharge, and (N2,denit) the excess N2 that is attributed to denitrification. The total dissolved Ar in each sample (Artotal) was considered to include only the atmospheric components Artotal ¼ ðArequil Þ þ ðArexcess-air Þ

ð2Þ

The solubilities of N2 and Ar in recharge water of varying temperatures and salinities were calculated using the solubility equations of Weiss (1970) at 1 atm total pressure and 100% relative humidity ðN2;equil Þ ¼ fðT; SÞ and ðArequil Þ ¼ fðT; SÞ

ð3Þ

and the excess air component was described by ðN2;excess-air Þ ¼ 83:6ðArexcess-air Þ

ð4Þ

where T is the temperature, S the salinity, and 83.6 is the N2/Ar ratio of unfractionated air. Because there are three unknown quantities (equilibration temperature, concentration of excess air, and concentration of N2 from denitrification) and only two measurements with which to evaluate those quantities (N2,total and Artotal), three alternative sets of calculations were made by making different assumptions: (1) no excess air; (2) equilibration T ¼ 25  C (approximate mean annual air temperature); and (3) equilibration T ¼ 30  C (approximate temperature of wastewater in June and October 1999). Eqs. (1–4) were solved simultaneously to obtain the unknown quantities in each case. The concentration
0 of NO
3 prior to denitrification ½NO3  was estimated by

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combining the concentrations of NO
3 and N2,denit in each sample. The extent of reaction (n) was calculated as n ¼ 2½N2;denit =½NO
3

0

ð5Þ

2.5. Isotope analyses and evaluation of fractionation effects Nitrogen isotope analyses were done on NO
3 and N2 in selected samples in the USGS Reston Stable Isotope Laboratory (see http://isotopes.usgs.gov; Bo¨hlke & Denver, 1995; Bo¨hlke et al., 2002). For isotope analysis of NO
3 , water samples were titrated to pH >10 and freeze-dried. The dried salts were sealed into evacuated quartz glass tubes with Cu2O, Cu, and CaO, and the tubes were baked at 850  C and cooled slowly to produce pure N2. For isotope analysis of dissolved N2, lowpressure headspace gas was expanded from the serum bottles after GC analysis into evacuated glass tubes, which were sealed and baked, as were the salts. Isotope measurements were done in dual-inlet mode on a Finnigan MAT 251 instrument and calibrated by analyses of air-equilibrated water samples and nitrate solutions prepared in the laboratory and treated as samples. Values of d15N were normalized to 0.0& for atmospheric N2,+4.7& for IAEA-N3, and +180& for USGS-32 (Bo¨hlke & Coplen, 1995), with average uncertainties of about 0.1& for dissolved N2 and 0.3& for NO
3 . Analyses of water samples equilibrated with air in the laboratory and treated as groundwater samples yielded d15 N ¼ þ0:78  0:03&, approximately consistent with published results (Hu¨bner, 1986; Klots & Benson, 1963). Isotope fractionation resulting from denitrification was evaluated from the concentrations and d15N values of the NO
3 and denitrification-derived N2 (N2,denit) in each sample. The d15N value of N2,denit was calculated by assuming that the total N2 in the sample was a mixture of equilibrated atmospheric N2 with d15 N ¼ þ0:78& (as measured in samples that were not denitrified and in the standards), an excess air component with d15 N ¼ 0:0&, and the N2,denit component. The apparent isotope fractionation factor (e) was calculated by applying the Rayleigh distillation equation (Clark & Fritz, 1997) 0

15
e ¼1000 lnfðd15 N½NO
3  þ 1000Þ=ðd N½NO3  þ 1000Þg= 0


lnð½NO
3 =½NO3  Þ

ð6Þ

Theoretical reaction curves were generated for a range of values of e to give calculated d15N values for the residual reactant NO3
1 (remaining after a given amount of denitrification) and the accumulating product N2,denit.

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3. Results 3.1. Wastewater distribution The water-quality results from the June and October 1999 sampling rounds are presented in Tables 2–4. The wastewater had the lowest salinity of all samples, S ¼ 4:4 in June and S ¼ 3:8 in October, and contained no measurable sulfide. In October, the main fixed-nitrogen component of the wastewater was nitrate (85%), followed by organic nitrogen (15%) and ammonium (<1%) (Table 4). A decrease in the wastewater nitrate concentration occurred between the June and October sampling rounds, presumably because of the upgrades to the treatment plant in July (Table 5). Otherwise, the spatial distributions of nutrient concentrations were similar in June and October; so, only the October results are displayed (Figs. 6–9). Subsurface salinities clearly reflected the impact of wastewater injection at depth. Salinities below 10 were found in the deepest monitoring wells near the point of injection, and extended to the mud–bedrock interface at approximately 6 m subsurface depth (Fig. 6; in this and other cross-sections, the upper surface is the land surface, and the saturated–unsaturated interface is approximately 80 cm below the land surface). Low-salinity waters extended in all directions below the mud–bedrock interface. The data indicate that lower-salinity waters exist at shallow depths further from the point of injection to the south than to the north (Fig. 6a). The east–west asymmetry of our monitoring-well network makes it impossible to argue, on the basis of salinity alone, for preferred flow to the southeast (Fig. 6b), although tracer studies indicate that the most rapid flow is to the southeast (Dillon et al., submitted for publication). The highest nitrate concentrations occurred close to the point of injection; concentrations generally decrease to the north and south (Fig. 7a). Similarly, the highest nitrate concentrations were found near the point of injection on the east–west transect (Fig. 7b), but nearwastewater concentrations were detected at our furthest monitoring well to the east, a few meters below the mud– bedrock interface. There appears to be a zone of elevated nitrate concentration centered at about 9 m and extending to the east, with lower concentrations both above and below this zone. Concentrations apparently drop off more rapidly to the west than to the east, although there is only one monitoring well (well E) to the west of injection site. The phosphate profiles (Fig. 8) resemble qualitatively those of nitrate. Phosphate was not detected in most of the hypersaline wells (except G-18), but had detectable concentrations in all brackish water samples. The highest concentrations were closest to the point of injection and concentrations decreased with distance. Wells B-18 and C-9 along the east–west transect had the highest concentrations of phosphate.

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Table 3 Field and inorganic nutrient analyses results for the October 1999 water samples Sample

pH

Salinity

Conductivity (mS cm
1)

A-9 A-18 B-9 B-18 C-9 C-20 D-9 D-14 E-9 E-18 F-8 F-14 F-18 G-9 G-18 J-6 J-13 J-18 K-5 K-8 L-9 East canal West canal Wastewater 8/25 East canal 8/25 West canal 8/25 Wastewater

7.64 7.46 7.83 7.24 7.24 7.45 7.73 7.44 7.74 7.23 7.22 7.81 7.62 7.39 7.22 7.20 7.54 7.54 7.06 7.37 7.42 7.93 8.05 6.98

6.7 9.5 2.5 4.1 2.7 26.4 5.8 27.9 8.2 19.7 11.5 31.5 39.7 9.0 39.9 2.9 27.9 35.5 3.4 7.1 12.1 33.5 34.7 3.8 N/A N/A N/A

11.76 16.34 4.67 7.37 5.02 41.30 10.13 42.60 14.22 32.10 19.32 48.50 58.70 15.36 59.40 5.37 42.80 53.80 6.18 12.45 20.20 51.00 52.10 6.96 N/A N/A N/A

Alkalinity (meq kg
1)

P Sulfide (lmol kg
1)

NHþ 4 (lmol kg
1)

NO
3 (lmol kg
1)

PO3þ 4 (lmol kg
1)

3.60 4.20 2.00 4.400 4.00 3.000 4.20 3.00 3.20 3.60 4.90 2.80 2.80 4.00 2.90 4.80 3.00 2.40 5.60 4.40 7.20 2.30 4.70 4.40 3.00 2.80 3.60

3.1 BDL 39.0 BDL BDL BDL BDL 9.4 15.6 12.5 >300 9.4 3.1 46.8 46.8 59.3 BDL 10.9 9.4 BDL 218.3 BDL BDL BDL BDL BDL BDL

16.3 3.3 71.4 3.1 2.9 BDL 6.2 13.7 20.3 19.1 218.5 16.3 18.3 37.2 30.8 61.6 1.4 10.3 26.2 2.4 121.4 1.4 3.2 1.3 BDL BDL 2.1

262 244 BDL 370 333 53 301 BDL 91 BDL BDL BDL BDL 66 BDL 180 23 BDL 149 471 279 5 2 447 20 10 359

5.6 35.5 5.6 30.4 22.0 BDL 5.1 BDL 1.0 1.1 BDL 2.5 BDL 0.5 1.5 16.8 BDL N/A 1.3 1.7 BDL N/A N/A 28.6 N/A N/A N/A


1

1
1 BDL, below method detection limits (NHþ 4 =0.3 lmol kg , NO3 =0.4 lmol kg , phosphate=0.3 lmol kg ); N/A, not available.

Table 4 Dissolved total and organic nutrient analyses results for the October water samples Sample

Total N (lmol kg
1)

Organic N (lmol kg
1)

Total P (lmol kg
1)

Organic P (lmol kg
1)

Organic C (lmol kg
1)

A-9 A-18 B-9 B-18 C-9 C-20 D-9 D-14 E-9 E-18 F-8 F-14 F-18 G-9 G-18 J-6 J-13 J-18 K-5 K-8 L-9 Wastewater

402 369 104 404 392 81 291 34 163 38 184 26 28 150 32 225 30 11 256 491 494 527

123.6 121.4 31.9 30.6 56.1 28.1 BDL 20.6 51.2 19.2 BDL 9.5 10.1 47.0 1.3 BDL 5.8 1.0 80.7 16.7 92.9 78.6

5.5 35.3 4.8 31.3 25.4 0.2 21.4 BDL 5.4 BDL 1.4 2.1 0.3 1.1 1.1 BDL BDL 1.9 3.3 7.9 0.3 34.4

BDL BDL 0.0 0.8 3.4 0.2 16.3 BDL 4.4 BDL BDL BDL N/A 0.7 BDL BDL BDL N/A 2.0 6.2 BDL 5.7

283 310 297 271 288 174 289 133 250 195 243 208 123 255 132 304 139 108 300 254 267 406


1

1
1 BDL, below detection limits; MDL: NHþ 4 =0.3 lmol kg , NO3 =0.4 lmol kg , phosphate=0.3 lmol kg ; N/A, not available.

E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539 Table 5 A comparison between the June and October 1999 wastewater compositions Analyses

June wastewater

October wastewater

pH Salinity Alkalinity (meq kg
1) Sulfide (lmol kg
1) Ammonium (lmol kg
1) Nitrate (lmol kg
1) Phosphate (lmol kg
1)

6.71 4.4 2.40 BDL 16.4 677 80.4

6.98 3.8 4.40 BDL 1.3 447 28.7

527

distribution was the opposite to that of nitrate, with highest concentrations to the west of the point of injection and in regions of higher salinity (low wastewater influence) near the bedrock–mud interface. Methane concentrations were relatively high in the wastewater and in wells near the bedrock–mud interface (Table 6). 3.2. Argon and N2 concentrations, recharge temperatures, excess air, and excess N2

BDL, below method detection limit.

Ammonium was detected in all wells except C-18, closest to the point of wastewater injection (Fig. 9). Its

Concentrations of Ar and N2 were approximately in equilibrium with air in samples of wastewater and hypersaline deep groundwater (F-18), whereas most of

Fig. 6. October 1999 salinity measurements along transects. Dotted lines refer to areas of unconstrained contours. (a) North–south transect, (b) east– west transect. Arrows refer to the location of wastewater injection.

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E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

Fig. 7. October 1999 nitrate concentrations (lmol kg
1) along transects. Dotted lines refer to areas of unconstrained contours. (a) North–south transect, (b) east–west transect. Arrows refer to points of wastewater injection. BDL, below detection limit.

the fresh and brackish groundwaters had significantly lower Ar/N2 ratios (Table 6, Fig. 10). The relatively low Ar/N2 ratios of the fresh and brackish groundwaters could be the result of either (1) dissolution of excess N2 and Ar from air bubbles entrained during wastewater injection or (2) production of excess N2 by denitrification in the groundwater after injection, or both. These possibilities cannot be resolved completely from the concentrations of N2 and Ar alone, but the data can be used to estimate limits on the effects of denitrification in individual samples. The concentrations of excess N2 from denitrification (N2,denit) were bracketed by calculations based on several different assumptions about

recharge temperatures and excess air concentrations: (1) the apparent equilibration temperature and concentration of N2,denit in each sample were calculated exactly by assuming that no excess air was dissolved during injection (this appears to be approximately true for some of the brackish-water samples that were near air– water equilibrium) and (2) the amounts of excess air and N2,denit in each sample were calculated exactly by assuming an equilibration temperature of 25  C (approximately equal to the mean annual air temperature in the Florida Keys and to the average gas-equilibration temperature of saline groundwaters in the region; J.K. Bo¨hlke et al., unpublished data). The results of these

E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

529

Fig. 8. October 1999 phosphate concentrations (lmol kg
1) along transects. Dotted lines refer to areas of unconstrained data. (a) North–south transect, (b) east–west transect. Arrows refer to points of wastewater injection. BDL, below detection limit.

calculations are summarized in Table 7 and displayed in Fig. 11. The assumption that wastewater injection did not introduce significant amounts of excess air yields somewhat variable estimates of equilibration temperatures (18–26  C, average 22.72.3  C) and widely varying concentrations of N2,denit. Those temperatures could be reasonable if the samples represent injection at various times of the year, with a slight bias toward the colder months when wastewater injection rates generally were higher (Fig. 3). The average monthly temperatures in the Keys range from around 20 to 29  C, with an annual mean of 25  C and monthly lows during the winter of approximately 18  C (National Climatic Data

Center website: http://lwf.ncdc.noaa.gov/oa/ncdc.html). The assumption of no excess air is supported by several analyses of brackish water that has N2 and Ar concentrations near air–water equilibrium (e.g. B-18 and A-18; Fig. 10), but it does not account for the positions of some oxygenated samples like C-9 that appear to be enriched in N2Ar. Calculations based on the average annual air temperature in the region (25  C) yield estimates of excess air and concentrations of N2,denit that were not greatly different from those obtained by assuming no excess air for most samples (Table 7). Similar calculations based on the assumption that the equilibration temperature was 30  C (the wastewater temperature in October 1999) yielded

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E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

Fig. 9. October 1999 ammonium concentrations (lmol kg
1) along transects. Dotted lines refer to areas of unconstrained contours. (a) North–south transect, (b) east–west transect. Arrows refer to points of wastewater injection. BDL, below detection limit.

unrealistic results, including negative values for excess air and N2,denit (data not shown). The data indicate that there was some variation in both the temperatures of equilibration and the amounts of excess air, but that most samples probably were recharged at temperatures somewhat lower than the wastewater temperature at the time of sampling (30  C in October 1999). Samples most likely to have had small amounts of excess air include C-9, D-9, J-6, and K-5. Despite the uncertainties in the assumptions used to calculate the concentrations of N2,denit, the data indicate that denitrification can account for all or most of the apparent nitrate deficit in most of the fresh and brackish groundwaters (Fig. 11). Moreover, the calculated con-

centrations of N2,denit in some groundwaters (e.g. E-9, G-9, L-9) are too high to have resulted from even complete denitrification of the October 1999 wastewater sample (indeed, some samples have significant nitrate remaining), implying that those groundwaters contained components of wastewater that were injected earlier in the year when nitrate concentrations were higher. Allowing for known variations in the initial concentrations of nitrate in wastewater during the previous months, the gas data indicate that denitrification could account for the entire nitrate loss in the subsurface in most cases. The most conspicuous exception to this is B9, which is almost fresh, but had relatively little nitrate or N2,denit. The reason for this anomaly is not known;

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Table 6 Analyses of dissolved gases and the isotopic composition of nitrate and nitrogen gas in samples taken at KCB in October 1999, with nitrate isotopes for three earlier wastewater samples Well

NO3 (d15N&)

O2

CH4

Ar (lmol kg
1)

N2

Ar/N2

N2 (d15N&)

A-9 A-18 B-9 B-18 C-9 C-20 D-9 D-14 E-9 E-18 F-8 F-14 F-18 G-9 G-18 J-6 J-13 J-18 K-5 K-8 L-9 Wastewater WW-6/99 WW-8/99 WW-11/99

23.0 16.4 N/A 15.9 17.8 24.8 23.5 N/A N/A N/A N/A N/A N/A N/A N/A 19.3 N/A N/A 25.6 23.5 21.1 16.5 20.2 20.0 7.4

1 2 1 59 101 N/A 1 1 1 1 2 N/A 1 1 N/A 1 1 N/A 1 N/A 2 204

0.21 0.00 1.73 0.01 0.05 N/A 0.46 0.91 0.25 0.17 2.37 N/A 0.19 0.39 N/A 1.07 0.27 N/A 0.62 N/A 1.05 4.12

12.74 12.12 12.68 12.60 13.76 N/A 13.85 10.78 11.90 11.04 11.80 N/A 10.02 12.85 N/A 13.96 10.70 N/A 13.54 N/A 12.34 11.39

593 481 557 506 593 N/A 727 484 766 562 712 N/A 384 783 N/A 685 470 N/A 678 N/A 643 450

0.0215 0.0252 0.0228 0.0249 0.0232 N/A 0.0190 0.0222 0.0155 0.0196 0.0166 N/A 0.0262 0.0164 N/A 0.0204 0.0228 N/A 0.0200 N/A 0.0192 0.0253

1.23 0.80 1.96 0.62 0.55 N/A 1.92 1.52 5.03 4.84 4.53 N/A 0.80 3.96 N/A 1.96 1.84 N/A 0.90 N/A 1.41 0.75

possibly this sample represents relatively uncontaminated recharge, or a situation in which intermediate nitrogen compounds (e.g. N2O) were relatively abundant. 3.3. Nitrogen isotopes The d15N values of nitrate in four wastewater samples ranged from about 7 to 20& (averaging 166&;

Fig. 10. Relationship between Ar and N2 concentrations in wastewater and groundwater samples collected in October 1999 (data in Table 6). Curves indicate Ar and N2 concentrations in equilibrium with moist air at 1 atm, with varying salinity, temperature, and excess air.

Table 7). This is within the range found previously for nitrate in sewage-derived wastewaters in the Florida Keys (J.K. Bo¨hlke, unpublished data) and elsewhere (Fogg, Rolston, Decker, Louie, & Grismer, 1998; Schroeder, Martin, & Bo¨hlke, 1996; Smith, Howes, & Duff, 1991; Wolterink, 1979). The d15N values of nitrate in the fresh and brackish groundwaters ranged from about 16 to 26&, generally higher than the wastewater values and qualitatively consistent with 15N enrichment caused by partial denitrification. The d15N values of total dissolved N2 in wastewater and hypersaline groundwaters were +0.7–0.8&, consistent with air equilibration as indicated by the Ar and N2 concentrations in the same samples. The d15N values of total dissolved N2 in fresh and brackish groundwaters ranged from about +0.5 to +5.0&, indicating that nonatmospheric components of N2 were present. Assuming recharge with no excess air, groundwaters containing N2,denit had estimated values of d15N[N2,denit] from about
1 to +17& (Table 7). If denitrification and dilution were the only processes affecting the concentrations of nitrate in the groundwaters, then the data indicate that the denitrified groundwater samples 0 had initial d15 N½NO
values of around 6–17& 3 (averaging 123&), similar to the range of values measured in the wastewater samples. In this case, the calculated isotope fractionation factor (e) for denitrification in the groundwater was about
133& (Fig. 12a), within the range commonly observed in other

532

Excess air ¼ 0 cm3 STP kg
1

Well A-9 A-18 B-9 B-18 C-9 C-20 D-9 D-14 E-9 E-18 F-8 F-14 F-18 G-9 G-18 J-6 J-13 J-18 K-5 K-8 L-9

Recharge temperature ( C)

N2,denit (lmol kg
1)

22.4 24.0 24.1 23.8 19.7

104 0 69 0 68

18.4 23.9 25.5 25.6 24.8

Recharge T=25  C [NO3]0 (lmol kg
1)

[NO3]0 (d15N&)

n

e (&)

14.2 16.4 10.2 15.9 12.3

0.45 0.00 1.00 0.00 0.29


14.6


1.3

474 245 139 370 468

201 75 309 143 260

4.9 5.6 11.3 17.0 11.1

703 150 710 281 520

12.8 5.5

0.57 1.00 0.87 1.00 1.00

23.7 21.1

0 293

9.3

0 652

18.9 24.3

153 64

6.1 8.6

486 150

11.0

0.63 0.85

20.3

161

1.3

470

9.0

22.2

173

3.1

625

11.1

N2,denit (d15N&) 3.3 10.2

17.0 11.0

Air excess (cm3 STP kg
1)

N2,denit (lmol kg
1)

N2,denit (d15N)

[NO3]0 (d15N)

414 245 118 370 333

16.3 16.4

n

e

0.37 0.00 1.00 0.00 0.00


14.3

0.45 1.00 0.87 1.00 1.00


11.0

1.5 0.5 0.5 0.6 3.2

76 0 59 0 0

12.2

4.0 0.5
0.3
0.3 0.1

121 65 315 146 258

8.6 6.5 11.1 16.4 11.1

542 131 721 292 516

0.5 2.2

0 249

11.1

0 564


8.3

3.7 0.3

78 58

12.4 9.6

336 138

16.1

0.47 0.84


5.1

0.68


14.2

2.8

105

2.3

358

11.9

0.59


15.2

0.55


12.1

1.5

142

3.9

563

12.4

0.50


12.1


15.9
12.3

0.00 0.90

4.8

[NO3]0 (lmol kg
1)

15.9 17.8 16.8 6.5 16.4 11.1

0.88

E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

Table 7 Calculated concentrations and isotopic compositions of N2,denit and initial nitrate [NO3]0, denitrification progress (n), and apparent isotope fractionation factor (e), assuming (1) no excess air or (2) equilibration at 25  C

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533

Fig. 11. Relationship between salinities and nitrate concentrations in groundwater and wastewater (data in Tables 3 and 7). Nitrate concentrations are given as measured (a) or as reconstructed using N2,denit values assuming (b) no excess air, and (c) recharge T=25  C.

field and laboratory studies (Hu¨bner, 1986; Mariotti, Landreau, & Simon, 1988). Results of calculations based on the assumption that all samples recharged at 25  C 0 yield d15 N½NO
3  ¼ þ14  3& and apparent fractionation factors of about
124&, similar to the results for no excess air (Fig. 12b). The systematic variation of d15N values of NO3
, N2, and [NO3
]0, and the value of the apparent fractionation factor derived from these calculations provide strong evidence in support of denitrification as the major cause of the apparent NO3
deficits in the fresh and brackish mixed groundwaters.

4. Discussion The wastewater of KCB exhibits low salinity, elevated nitrate and phosphate, and low ammonium concentrations compared with the saline groundwaters into which it is injected. Thus, the effluent can be readily traced through the subsurface using the distribution of

salinity and nutrient concentrations. However, although salinity behaves conservatively, nutrient concentrations behave non-conservatively. Thus, the pattern and rate of movement of the wastewater as it travels away from the point of injection influences how the chemical composition of the plume changes along its path. 4.1. Subsurface salinity The salinity distributions demonstrate how easily the freshwater plume can be traced in the hypersaline groundwater (Fig. 6). Salinities in wells I and H, located to the north and south of the main injection well (Fig. 6a), indicate that the low-salinity wastewater injected at a depth of approximately 18–27 m is transported upward (presumably because of its lower density) to the mud–limestone contact within a few meters of the injection well. The wastewater continues traveling laterally underneath the mud–limestone contact in both directions.

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the October round, the population and rate of wastewater input were higher, but had not yet peaked. Many wells at KCB demonstrated a reduction in salinity between sampling rounds (Tables 2–3). The wells showing the strongest reduction (highest ratio of June/October salinities) are those nearest the point of injection (wells A, B, C, and D) or on the presumed fast east–west flow path (wells E, J, K, and L). Wells at a distance north and south of the point of injection (wells F and G) actually show higher October salinities, perhaps reflecting some entrainment of hypersaline groundwaters from greater depth or distance from the point of injection. 4.2. Non-conservative behavior of nutrients If the concentrations of nitrate and phosphate were behaving conservatively and decreasing after injection by simply mixing with the nutrient-poor saline groundwater, then the samples would plot along mixing lines between wastewater and ambient groundwater endmembers. Most of the data for both June and October 1999 plot below these mixing lines (which have different wastewater endmembers), indicating non-conservative behavior, i.e. transformation or removal of phosphate (Fig. 13a) and nitrate (Fig. 13b).

Fig. 12. Relationship between denitrification progress and the d15N values of nitrate and N2,denit in ground-water samples collected in October 1999 (data in Table 7). To facilitate the evaluation of isotope fractionations, all plotted data were normalized to make the initial isotopic composition of the nitrate in each sample equal to 0&. Calculations were done by assuming (a) no excess air, and (b) recharge T=25  C.

The majority of freshwater input to the KCB groundwater system is clearly through wastewater injection. There does not appear to be any percolating meteoric input, given the low permeability and brackish salinity of the Holocene mud overlying the permeable KLL. The groundwater salinity at KCB differed in June and October 1999, presumably because of a change in the rate of wastewater injection. During the June sampling round, the local population was smaller, therefore the rate of effluent injection was lower (Fig. 3). During

Fig. 13. The June and October 1999 concentrations of (a) phosphate and (b) nitrate plotted against salinity. The mixing lines connect the concentrations of the wastewater (for each month) and ambient groundwater endmembers; conservative mixing would yield intermediate waters along the lines.

E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

4.3. Transport and reaction of phosphate From column experiments, Elliott (1999) determined that phosphate undergoes rapid adsorption onto calcite surfaces or precipitation of a metastable phosphate phase. Column experiments indicate that the uptake of phosphate caused a reduction of phosphate concentrations to ÔequilibriumÕ (steady-state) [??] values of 25 lmol kg
1. Our results (Figs. 8 and 13a) and those of Elliott suggest that the near-field bedrock (wells A-18, B-18, and C-9) has equilibrated with the wastewater input, but that the equilibration front has not extended beyond well J about 30 m to the east along the fast flow path. Nevertheless, low but detectable phosphate concentrations were found at our most distant well to the east at shallow depths, indicating that some wastewater-derived phosphate probably is reaching the canal to the east.

4.4. Transport and reaction of nitrate Nitrate depletion is evident in most of the wells (Fig. 13b). Possible microbially mediated nitrate removal mechanisms from groundwater include denitrification and dissimilatory nitrate reduction to ammonium (Smith & Duff, 1988). Both of these mechanisms can be related to oxidation of organic matter in suboxic to anoxic systems. Denitrification has been shown to occur in a variety of aquifers containing reactive substrates such as organic carbon (Hiscock, Lloyd, & Lerner, 1991; Korom, 1992) or sulfide phases (Ko¨lle, Strebel, & Bo¨ttcher, 1985). Denitrification is commonly considered the main nitrate removal mechanism in groundwater. However, some studies have shown that dissimilatory nitrate reduction to ammonium can account for up to 30–50% of nitrate removal in some settings (Bengtsson & Annadotter, 1989; Kasper, 1983; Koike & Hattori, 1978; Koike & Sorensen, 1988; Steenkamp & Peck, 1981). The apparent deficits in the measured nitrate concentrations in fresh and brackish groundwaters are similar in magnitude to the concentrations of excess N2 attributable to denitrification (N2,denit), assuming either no excess air or constant 25  C recharge (Table 7, Fig. 11). The N isotopic compositions of coexisting NO3
and N2 are consistent with isotope fractionation effects of denitrification in closed groundwater systems (Fig. 12). Thus, we conclude that denitrification probably is the major process accounting for subsurface NO3
loss from the injected wastewater. Because of known variations in the concentrations of NO3
in the wastewater over time, the initial NO3
concentrations of individual samples before denitrification are not known precisely except by reconstructing the NO3
and N2 data. Therefore, it is not known

535

whether small amounts of NO3
also might have been reduced to NH4+. Assuming all the measured NH4+ was produced this way yields initial NO3
concentrations that are not substantially different from those calculated by assuming denitrification was the only NO3
-consuming reaction. However, it is also possible that the high NH4+ concentrations in some of the shallow brackish waters were not related to wastewater NO3
, having formed by anaerobic degradation of organic N in the overlying fine-grained sediments. Saline groundwaters throughout the region of the Florida Keys (e.g. well F-18) typically have NH4+ concentrations of 10– 100 lmol kg
1, presumably as a result of natural organic matter mineralization (J.K. Bo¨hlke et al., unpublished data; Shinn et al., 1994). The cumulative electron consumption for denitrification in individual fresh and brackish groundwater samples near the KCB injection site ranged from 0 to about 1600 lmol kg
1 as e
. Potential electron donors for denitrification of wastewater in the subsurface include DOC or H2S in the saline groundwater, DOC in the wastewater, and solid organic C or sulfide S in the aquifer. Saline-reduced groundwaters in the Florida Keys region have DOC and H2S concentrations as high as several hundred lmol kg
1 (J.K. Bo¨hlke et al., unpublished data; Shinn et al., 1994). Such concentrations could be high enough to contribute substantially to denitrification. However, the known concentrations in saline endmembers at the KCB site are less than 140 lmol kg
1 DOC (Fig. 14a) and 100 lmol kg
1 H2S (Tables 2 and 3), which would be minor sources of electrons for denitrification, especially when diluted by mixing with the wastewater. Furthermore, of all the samples collected at the site, the saline groundwaters had the lowest DOC concentrations (Table 4), so consumption of saline groundwater DOC cannot be related directly to denitrification in the brackish groundwaters. In general, the relationship between salinity and the concentration of N2,denit indicates that denitrification probably was not dependent on dissolved constituents in the saline groundwater (Fig. 14b), as large amounts of denitrification occurred in groundwaters that were almost fresh. Although DOC concentrations generally decreased with distance from the injection wells, there was no relationship between the concentrations of DOC and N2,denit. Instead, there was an inverse relationship between DOC and salinity that was independent of the concentration of N2,denit (Fig. 14). These relationships indicate that DOC in the wastewater also was not a major electron donor for denitrification along the groundwater flow path. The difference between the DOC concentrations of the October 1999 wastewater sample and the freshest groundwaters is about 100 lmol kg
1, and this difference is maintained during mixing (Fig. 14a). This may indicate that DOC was an electron donor only for O2

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E.M. Griggs et al. / Estuarine, Coastal and Shelf Science 58 (2003) 517–539

could account for up to a 50% reduction in the DOC content of the wastewater plume. DOC depletions in the wells are generally somewhat less than 50% (Fig. 14a), suggesting that other reactions are consuming dissolved oxygen or that the wastewater sample was atypical. The extent of denitrification ðn ¼ 1
½NO3 =½NO3 0 Þ ranges from 0 to 1 (Table 7), with a spatial pattern that essentially mirrors that of nitrate. The lowest values of n are found near the point of injection (in wells A-18 and B-18 n ¼ 0:0), and modest extents of denitrification are found along the fast flow path (n ¼ 0:3
0:6 in wells A-9, C-9, and D-9; gas analyses were not performed in well K8). Denitrification is more complete near the bedrock– mud interface (n ¼ 0:6
0:7 in wells J-6 and K-5) and off the main flow path (n ¼ 0.8–1.0 in wells E-9, E-18, J-13, D-13, F-8, and G-9) where residence times are longer. Apparently, time is an important factor limiting the extent of denitrification, implying that substrate availability is not the limiting factor. On the other hand, we have not identified the substrate that is supporting denitrification in the KCB subsurface. There is no evidence for enhanced denitrification above vs. below the fast flow path, ruling out a source of substrate from the overlying sediment. We are left with inconspicuous solid organic carbonsulfide or iron phases as potential substrates for denitrification. This material could have been deposited with, or formed diagenetically within the KLL, or it could include organic compounds sorbed previously from the wastewater plume.

4.5. Denitrification rate estimates

Fig. 14. Relationship between salinity and (a) DOC concentrations and (b) concentrations of N2,denit for the October 1999 data only.

reduction, unless the wastewater sample from October 1999 had a higher than average DOC concentration. The wastewater contained 200 lmol kg
1 of O2, approaching atmospheric equilibrium of 260 lmol kg
1 at 25  C. Only two of the wells, B-18 and C-9, had substantial amounts of dissolved oxygen remaining. According to tracer study data (Dillon et al., submitted for publication), B-18 and C-9 are the first wells to receive the injected effluent, with tracer peak arrivals after 10 and 7 days, respectively. The rest of the wells were essentially anoxic (Table 6). Assuming a 1 : 1 stoichiometry for O2 reduction by organic C (i.e. assuming the canonical ÔCH2OÕ organic composition and oxidation to CO2), aerobic decomposition

The rate of denitrification can be estimated by different methods involving different estimates of groundwater ages, for example (1) from the mass of N2,denit and the overall average residence time of fresh water in the subsurface and (2) from the N2,denit concentrations and ages of individual groundwater parcels derived from tracer velocities. The average residence time of fresh water in the subsurface can be estimated from the average wastewater injection rate (800 m3day
1, Fig. 3) and the overall volume of fresh water in the saturated zone. The limited coverage of the study area by monitoring wells makes an estimate of the volume of subsurface wastewater highly uncertain. Salinity distributions represented in Fig. 6 support a wastewater plume approximately 10 m thick that extends primarily to the southeast. The southeast quadrant of a 10 m thick disk of radius 150 m and 45% porosity is 80,000 m3. The average residence time is thus 100 days. This average residence time is long enough to account for the observation that the range of initial NO3
concentrations of the October 1999 groundwaters includes values that imply injection before and after conversion of the plant to AWT in July 1999, and it could be consistent with

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variation in the apparent water–air equilibration temperatures derived from dissolved-gas data. An upper limit on the average rate of denitrification can be obtained by assuming complete denitrification of the wastewater within the residence time of wastewater in the subsurface. Accordingly, using the average concentrations of injected NO3
(400 lmol kg
1) and a residence time of 100 days gives an average denitrification rate of 4 lmol kg
1 N day
1. The average rate likely is less than this because denitrification clearly is not complete: wastewaters containing high concentrations of nitrate are present far from the point of injection and near to a canal where discharge is likely occurring (Fig. 7). Estimates for denitrification rates can also be made by comparing the concentrations of N2,denit and groundwater travel times estimated from tracer data (Dillon et al., submitted for publication). Of the five wells with measurable peak arrival times in the SF6 tracer study, four were analyzed for dissolved nitrogen. Assuming recharge with no excess air, or recharge at 25  C (Table 7), it is estimated that well B-18 (tracer travel time ¼ 10 days) had no detectable N2,denit, whereas well C-9 (7 days) had 0– 68 lmol kg
1, well A-9 (60 days) had 76–104 lmol kg
1, and well D-9 (55 days) had 121–201 lmol kg
1 of N2,denit. Corresponding zero-order denitrification rates for these wells (in lmol kg
1 N day
1) would be 0 for B-18, 1010 for C-9, 3.00.5 for A-9, and 5.91.5 for D-9. Linear regression of these travel times and estimated N2,denit concentrations give a denitrification rate of approximately 4.3 lmol kg
1 N day
1. The corresponding first-order rate constant for denitrification in these samples, based on the reaction progress variable (n in Table 7), is approximately 0.008 day
1. These values are similar to the mass-balance estimate above and fall within the wide range of observed values in other groundwater systems (e.g. Bates & Spalding, 1998; Bo¨hlke et al., 2002; Bragan, Starr, & Parkin, 1997; DeSimone & Howes, 1998; Korom, 1992; Schnabel, Shaffer, Stout, & Cornish, 1997; Tesoriero, Liebscher, & Cox, 2000). There is orders-ofmagnitude variability in published groundwater denitrification rates, probably resulting from a combination of variability in substrate availability, other environmental factors, and the technique being used to make the estimate. Our results indicate that denitrification requires groundwater residence times of several weeks to substantially affect groundwater nitrate concentrations in the wastewater-contaminated aquifer at KCB.

537

denitrification, presumably because of the relatively short residence time of waters in the subsurface of the study area. As previously mentioned, wells J-6 and K-5 were not included in the tracer study, but do appear to be along a fast flow path; the decrease in nitrate concentration from June to October, which reflects the improved quality of the treated wastewater, indicates that the turnover time of wastewater is relatively rapid in these locations. In comparison with other wells known to be along the fast flow path, however, the amount of nitrate removal is greater than expected for wells J-6 and K-5. We suggest that wells J-6 and K-5 may lie above the fast flow path, in areas of lower porosity bedrock highs (fossil patch reefs). The residence times in the patch reefs at the shallowest sampling intervals of wells J and K are thought to be relatively long. However, the changes in the salinities and nutrient concentrations from June to October 1999 indicate that substantial turnover of the water at these wells occurs within 90 days. 4.7. Groundwater quality improvement The upgrade of the KCB wastewater treatment plant was completed in July 1999, approximately 90 days prior to the October 1999 sampling round. According to the SF6 tracer study, only five of the wells and both canals included in the study should have received the improved wastewater within 90 days. Four of these five wells along the main flow path did record higher nitrate values in June than October, which indicates that those wells were indeed receiving improved wastewater (Tables 2 and 3). Only well A-18 did not show a decrease in nitrate. Salinity in well A-18 decreased over this interval, as it did in most other wells near the point of injection, indicating perhaps that the wastewater plume had expanded in response to higher rates of wastewater discharge and was having greater influence on the 18 m depth in well A from June to October 1999. Canal waters were not sampled in June 1999, but samples collected in August and analyzed as part of this study have higher nitrate concentrations than those from October (Table 3). This is consistent with the expectation based on improvements in wastewater quality, but the data are too few to make more conclusive statements about canal or groundwater response to the improvements in wastewater treatment.

4.6. Influences of groundwater flow paths on geochemistry 5. Conclusions According to the SF6 tracer experiment, the main direction of effluent flow after injection is upwards and to the east or southeast (Dillon et al., submitted for publication). Groundwaters determined to be along this fast flow path clearly have been relatively less affected by

Wastewater-derived nutrients injected into the KLL at KCB, FL, react non-conservatively in the subsurface before the buoyant wastewater plume discharges into surrounding canals. Most of the phosphate burden of

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the wastewater plume is removed before discharge to surface waters, but substantial quantities of wastewater nitrate may be escaping to canal waters. Phosphate is largely scavenged onto the limestone substrate, although low concentrations of phosphate are detected at our more distant wells. Gas and nitrogen isotopic analyses confirm that microbially mediated denitrification is primarily responsible for the observed decreases in nitrate concentration beyond the effects of dilution. A main determinant in the extent of nitrate removal is the flow path of the wastewater. The faster moving groundwaters tend to have less nitrate depletion than those waters with longer residence times. Nutrient concentrations in the wells farthest from the point of injection give some indication of nutrient concentrations that may be entering the surrounding surface waters. Although residence times of wastewater in the system are generally long enough for substantial denitrification to take place, waters flowing along the main flow path have insufficient time for complete nitrate removal. Groundwater at most distant well in our study along the main flow path (K-8) is virtually unchanged in its nitrate content compared with wastewater, and phosphate is also detected (although most of the other distant wells have undetectable phosphate concentrations). Well K is located only 50 m from the east canal. Therefore, it is considered likely that some of the groundwater entering the canal has elevated nitrate (and phosphate) concentrations similar to those sampled in well K. Thus, despite enhanced wastewater treatment and biogeochemical removal in the subsurface, wastewater disposal at KCB does appear to be introducing nutrients into surface waters of the Florida Keys.

Acknowledgements E.M.G. and L.R.K. acknowledge support from the US Environmental Protection Agency, grant X98429297-0-PA. J.K.B. acknowledges support from the National Research Program, Water Resources Discipline, US Geological Survey (USGS). We benefited from the assistance and wisdom of our colleagues, including K. Dillon, W. Burnett, and J. Chanton (Florida state), R. Corbett (East Carolina), and E. Shinn, C. Reich, and D. Hickey (USGS, St. Petersburg). P. Widman, M. Doughten, E. Busenberg, J. Hannon, and S. Mroczkowski (USGS, Reston, VA) assisted with analyses of dissolved gases and isotopes.

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