Denitrification in a Sand and Gravel Aquifer - PubMed Central Canada

Vol. 54, No. 5

APPLIED AND ENVIRONMENTAL MICROBIOLOGY, May 1988, p. 1071-1078 0099-2240/88/051071-08$02.00/0 Copyright ©3 1988, American Society for Microbiology

Denitrification in RICHARD L.

a Sand and Gravel SMITH'* AND JOHN H. DUFF2

Aquifer

Water Resources Division, U.S. Geological Survey, MS 408, Lakewvood, Colorado 80225,1 and Water Resources Division, U.S. Geological Survey, MS 496, Menlo Park, California 940252 Received 16 November 1987/Accepted 12 February 1988

Denitrification was assayed by the acetylene blockage technique in slurried core material obtained from a freshwater sand and gravel aquifer. The aquifer, which has been contaminated with treated sewage for more than 50 years, had a contaminant plume greater than 3.5-km long. Near the contaminant source, groundwater nitrate concentrations were greater than 1 mM, whereas 0.25 km downgradient the central portion of the contaminant plume was anoxic and contained no detectable nitrate. Samples were obtained along the longitudinal axis of the plume (0 to 0.25 km) at several depths from four sites. Denitrification was evident at in situ nitrate concentrations at all sites tested; rates ranged from 2.3 to 260 pmol of N20 produced (g of wet sediment)-' h-'. Rates were highest nearest the contaminant source and decreased with increasing distance downgradient. Denitrification was the predominant nitrate-reducing activity; no evidence was found for nitrate reduction to ammonium at any site. Denitrifying activity was carbon limited and not nitrate limited, except when the ambient nitrate level was less than the detection limit, in which case, even when amended with high concentrations of glucose and nitrate, the capacity to denitrify on a short-term basis was lacking. These results demonstrate that denitrification can occur in groundwater systems and, thereby, serve as a mechanism for nitrate remoyal from groundwater.

Among the substances entering groundwater systems from anthropogenic origins, nitrate is very prominent. It is highly soluble, mobile, and represents a health hazard to human infants at relatively low concentrations (6). Point source contamination of groundwater supplies by nitrate is an increasingly common phenomenon in both rural and urban areas as a result of localized fertilizer application and disposal of human and animal wastes (9, 11, 19, 27, 28, 30, 35). In addition, in agricultural regions, nitrate contamination can occur in large areas due to combined fertilization and irrigation practices (30, 35). Once nitrate enters an aquifer, it probably remains in the groundwater unless it is removed or transformed by biological processes. As dissimilatory activities, denitrification and nitrate reduction to ammonium are the microbial processes most likely to affect nitrate concentrations in groundwater. Indirect evidence, such as N03-, N2, and 02 concentrations (10, 28, 39), 15N/14N ratios (11), nitrate injection and withdrawal tracer tests (38), and isolation of denitrifying bacteria (14, 41), has suggested that denitrification occurs in groundwater. However, only recently has the activity been measured directly in groundwater samples (33, 40). Even less is known about nitrate reduction to ammonium in groundwater. In the rumen and in digested sludge, reduction to ammonium is the predominant nitrate-reducing process (18, 37), and even in sediments in which denitrification predominates, ammonium can account for as much as 30% of the nitrate reduced (15). Therefore, this mechanism needs to be considered as a potential nitrate sink in groundwater as well. In this study, we directly examined sediments from a sand and gravel aquifer that had been contaminated with nitrate for denitrification and dissimilatory nitrate reduction to ammonium. Denitrification was the predominant nitratereducing mechanism within the aquifer, essentially depleting the nitrate in the core of the contaminant plume 250 m downgradient from the contaminant source. *

Corresponding author.

MATERIALS AND METHODS

Study site. The study area consisted of a sand and gravel aquifer located on Cape Cod, Mass., and is the site of a multidisciplinary research effort organized by the U.S. Geological Survey Toxic Waste-Ground Water Contamination Program. The aquifer has been continuously contaminated with secondarily treated sewage from a sewage treatment plant since 1936. Sewage effluent from the plant is discharged onto rapid-infiltration sand beds, where it percolates into the ground to the water table, which is located 6 m beneath the surface of the sand beds. Groundwater flow in the region is horizontal in a southerly direction at an average rate of 0.2 to 0.5 m per day (23). As a result, the sewagecontaminated groundwater has formed a plume (based on specific conductivity) that is more than 3.5-km long, 0.9-km wide, and 23-m thick (Fig. 1). The nature of the contaminant plume and the hydrogeology of the aquifer have been previously described in more detail (23, 24). Sampling. Groundwater samples were obtained from observation wells that were constructed with 5-cm-diameter polyvinyl chloride pipe and 0.6- to 0.9-m-long polyvinyl chloride slotted screen. A group of four to five wells was located at each sampling site, and each well was screened at a different depth. Well clusters sampled were S314, S316, S344, and F347. These are designated A, B, C, and D, respectively (Fig. 1). The wells were sampled by using a stainless steel submersible pump (model SP81; Keck Geophysical Instruments, Inc., Okemos, Mich.) equipped with Teflon tubing (E. I. du Pont de Nemours Co., Inc., Wilmington, Del.). Samples were taken after 3 to 5 well volumes were pumped and the specific conductance had stabilized. For activity measurements, 1-liter glass bottles were flushed with 3 to 5 liters of sample, completely filled, and capped to avoid an air headspace. The bottles were transported to a laboratory in an ice bath, and incubations were initiated on the day of collection. For nitrate and ammonium determinations, samples were filtered through 1071

1072

APPL. ENVIRON. MICROBIOL.

SMITH AND DUFF

0

500

1000

1500

meters

FIG. 1. GroundWater study site on Cape Cod, Mass. The shaded area, other than those indicating ponds, delineates the extent of the contaminant plume caused by infiltration of treated sewage into the aquifer. Solid lines represent the altitude of the water table (meters above sea level in November 1979); arrows show the direction of groundwater flow. The sampling locations for this study are indicated as points A to D.

0.45-jm-pore-size filters (Metricel; Gelman Sciences, Inc., Ann Arbor, Mich.). Aquifer sediments were collected by using an auger drilling rig equipped with a splitspoon core barrel. Cores (0.6 m in length) were obtained by drilling a hole to the desired depth and then driving the core barrel into undistulrbed sediment beneath the bottom of the hollow-stem auger. During the augering, a disposable knockout plate was attached to the cutting bit of the hollow-stem auger to prevent sediment from entering the hollow stem. Cores were taken immediately adjacent to each well screen at sites A, B, C, and D (Fig. 1). Core material was transferred to widemouth bottles, which were completely filled; and the sediment was resaturated by using freshly collected water from the adjacent well to remove air pockets that resulted from the transfer process. The samples were kept on ice until they were dispensed later the same day. Incubations. Denitrification was assayed by the acetylene blockage technique (3, 22, 42). Triplicate 125-ml, widemouth Erlenmeyer flasks were tared and flushed for 10 min with a continuous stream of 02-free N2. Then, approximately 30 ml of sediment -(50 g [wet weight]) was added to each flask by using a plastic syringe that had the hub end cut off. The sediment was slurried with 50 ml of water obtained from the well adjacent to the sediment core. The bottle containing the water sample was flushed with 02-free N2 during the dispensing process; the subsample was transferred to the Erlenmeyer flasks with a syringe and filtered through a 0.45,Im-pore-size filter (Metricel; Gelman) into each flask. The flasks were flushed for an additional 15 min with 02-free N2 and then sealed with recessed butyl rubber stoppers and weighed. The sediment slurries were preincubated at room temperature (22 to 25°C) overnight (6 to 8 h) on a reciprocating shaker, and then 9 ml of acetylene was added by syrihge to the headspace of each flask. The flasks were placed back on the shaker for 1 h, after which 100 to 500 [lI

of headspace gas was removed and analyzed for N20. This sample was designated as the zero time point. N20 was analyzed at several time points during the incubations, which usually lasted about 48 h. There was no sulfide, which would interfere with the added acetylene, at any of the sites. The effect of various amendments and manipulations on denitrifying activity in the aquifer was also examined. To test whether the activity was carbon- or nitrate-limited, additional sets of sediment slurries from each of the sampling sites described above were amended with 0.1 ml of an 02-free solution of glucose, nitrate, or glucose-nitrate at the time of the acetylene additions to the flasks. For each substrate the final concentration was 1.0 mM. Killed controls were prepared by boiling sediment slurries for 10 min; the effect of air in the headspace was tested by omitting the N2 sparge. In another sample set, 9 ml of 02-free N2 was added in place of the acetylene. Denitrification was also tested in well water alone by transferring 100 ml of an unfiltered water sample to 125-ml Erlenmeyer flasks and incubating the samples as described above. Dissimilatory nitrate reduction to ammonium was tested by two different methods. The first was a 15-day incubation with sediment slurries to determine concomitant rates of nitrate consumption and ammonium production. In this case the flasks were 250-ml Erlenmeyer flasks, the sediment volume was 60 ml, and the volume of the water sample (unfiltered) added was 100 ml. After an overnight preincubation, 25 ml of water was removed from each flask with a syringe and replaced with 25 ml of 02-free N2. These subsamples were filtered through 0.45-,um-pore-size filters (Metricel; Gelman), acidified with 20 ,ul of concentrated HCI, and frozen for nitrate and ammonium analyses. The flasks were incubated on a shaker at room temperature for 15 days, and subsequently 25-ml portions were removed and filtered every 3 to 4 days. Sediment slurry incubations with [15N]nitrate was the second technique used to assay dissimilatory nitrate reduction. For each core sample, 30 ml of sediment was added to each of six 125-ml Erlenmeyer flasks that had been flushed with 02-free He. The flasks were then completely filled with water obtained from the adjacent well and stoppered. After an overnight preincubation, 9 ml of sample was displaced with O2-free He and 0.1 ml of a solution of [15N]KNO3 (6.8 g/liter; 76.7 atom% 15N; Monsanto Research Corp.) was added to each flask. Activity in three flasks from each set was stopped immediately by adding 1 ml of concentrated H2SO4. The other flasks were incubated at room temperature for 60 h on a shaker, and then the activity was also stopped by adding 1 ml of concentrated H2SO4. The sediment slurries were transferred to acid-washed Erlenmeyer flasks. Plastic conical analyzer cups (Curtin Matheson Scientific, Inc.) that contained 1 ml of 0.5 N H2SO4 were inserted into test tubes. The collar on the plastic cups held them in place at the open end of the test tube. The test tubes were then placed in the Erlenmeyer flasks that contained the sediment slurries, 7.0 ml of 10 N NaOH was added to each flask, and the flasks were immediately stoppered with butyl rubber stoppers that had been wrapped in Saran Wrap. The stoppers were secured in place with wire, and the flasks were placed in a 55°C oven. After 7 days, the plastic cups were removed from the flasks, the contents of the cups were dried at 550C, and the (NH4)2SO4 residue was analyzed for 15N. Recovery of ammonium chloride standards (150 jig of N per flask) by this technique was 87 + 7% (standard deviation). Analytical techniques. Oxygen was analyzed by the iodometric method with the azide modification (1). Nitrate and

DENITRIFICATION IN GROUNDWATER

VOL. 54, 1988

1073

TABLE 1. Chemistry of groundwater at the sampling sites Site"

Depth (m) Specific Dissolved Dissolved Dissolved beneath conductance oxygen nitrate ammonium (>M) (>iM) (,uS) water table (p1M)

390

NDb

930

61

2.1 7.0 14.4 21.3

260 260 280 145

166 30 22 18

450 520 540 560

0 0 404 0

1.5 6.1 11.0 16.4 26.8

350 350 250 300 250

120 10 0 0 4

1,760 1,320 1,380 20 330

2 2 2 61 98

2.1 9.1 14.9 21.0

193 305 335 297

42 0 0 0

320 0 0 0

0 609 577 439

2.4 10.4 16.8 27.1

103 373 306 262

548 4 4 4

140 0 0 0

9 464 340 108

Sewage effluent A

B

C

D

a Samples from sites A, B, and C were collected in May 1985; samples from site D and sewage effluent were collected in November 1984. b ND, Not determined.

ammonium were determined with an Autoanalyzer (Technicon Instruments Corp., Tarrytown, N.Y.) by using the cadmium reduction method (1) and the salicylate hypochlorite method (32), respectively. Nitrous oxide was determined by using a gas chromatograph (model 417; Packard Instrument Co., Inc., Rockville, Md.) equipped with a 63Ni electron capture detector. The gases were separated with a Porapak N column (outer diameter, 0.32 cm) at 90°C. The column was divided into two lengths (0.66 and 1.33 m) separ-ated by a six-port backflush valve to vent the acetylene. Carrier gas was a mixture of methane (5%) and argon (95%) at a flow rate of 30 ml/min. The 15N content of ammonium was analyzed by Isotope Services Inc., (Los Alamos, N.M.) by using an isotope ratio mass spectrometer (Nuclide RMS 3-60). Ammonium N was oxidized to N, with alkaline hypobromite with an automated injection system

(5). RESULTS Chemical profiles. The monitoring wells sampled for this study lie along an approximately 250-m-long longitudinal transect of the contaminant plume at the Cape Cod study site. The specific conductance of groundwater samples obtained from these wells was 250 ,uS or greater, except for samples obtained from the deepest well at site A and the shallowest well at site D (Table 1). By comparison, the specific conductance of the sewage effluent was 390 ,uS (Table 1) and that of the uncontaminated groundwater in the area was 50 ,uS (data not shown). The predominant nitrogencontaining compound in the sewage effluent was nitrate, the concentrations of which fluctuated. During this study, effluent nitrate concentrations ranged between 900 and 1,000 ,uM (Table 1). The effluent also contained ammonium, al-

/ /

E

0

1S _

1.5m

z

Groundwater

Q5

0

15

30

45

HOURS

FIG. 2. Time course of nitrous oxide production by core samples and well water samples collected from two depths at site B when incubated in the presence of acetylene. For comparative purposes, the units of volume are liters of sediment and liters of groundwater. Samples were collected in November 1984; depths are meters beneath the water table. Data points represent the means of triplicate flasks.

though the concentrations were considerably lower than those of nitrate. In the contaminated groundwater, the relative ratio of nitrate to ammonium decreased with increasing distance away from the infiltration sandbeds, and therefore, ammonium predominated at sites C and D (Table 1). On a vertical axis, the center of the contaminant plume at sites C and D was anoxic and contained no detectable nitrate. Only in the shallowest well at each of these two sites were oxygen and nitrate present. Denitrification. Well water and core material obtained from site B both produced nitrous oxide when incubated in the presence of acetylene at ambient nitrate concentrations (Fig. 2). The time course of N20 production was linear for approximately 30 h for all of the samples tested, with no apparent time lag. The composite denitrification rate for the two slurried core samples (0 to 25 h) was 42.1 nmol of N20 produced (liter of sediment)-' h- 1, or approximately 40-fold higher than the denitrification rate for groundwater alone. Nitrous oxide production by sediment slurries for the deeper of the two samples (6.1 m beneath the water table) was inhibited completely in the killed control and substantially TABLE 2. Effect of various manipulations on N2O production in sediment slurries collected from site B" N20 produced (nmol/g of wet Incubation conditions sediment +

SD)'

1.12 ± 0.06 0 ± 0 Killed control ......... Air headspace ......... 0.10 ± 0.02 No acetylene ........ 0.24 ± 0.12 "1 The sample was collected from site B at a depth of 6.1 m beneath the

Endogenous ........

water table in November 1984. b N20 produced after a 48-h incubation by the

technique.

acetylene blockage

1074


SMITH AND DUFF

0 C

0CE

4

I

I e

3

I

-0

3

I

2

HOURS

40 HOURS

FIG. 3. Nitrous oxide production by slurried core material obtained from site B at a depth of 1.5 m beneath the water table in November 1984. Samples were incubated at ambient nitrate concentration (0) and amended with 1 mmol of nitrate per liter (0), 1 mmol of glucose per liter (O), or 1 mmol of nitrate per liter and 1 mmol of glucose per liter (l). Error bars represent ±1 standard error.

inhibited by the presence of air in the headspace of the flask (Table 2). When the headspace was 02-free N2, some N20

also produced in the absence of acetylene (Table 2). To delineate some of the factors controlling denitrification, the effect of selected amendments on nitrous oxide production was examined. The rate of denitrification in sediment slurries of core material obtained from 1.5 m beneath the water table at site B was unaffected by the addition of 1.0 mM nitrate (Fig. 3). However, during the first 25 h of incubation, the addition of 1.0 mM glucose or 1.0 mM glucose-1.0 mM nitrate increased the rate of denitrification approximately fivefold over that of unamended samples. After 35 h, the rate of denitrification increased exponentially in all samples amended with glucose (data not shown). Similar patterns of response to glucose and nitrate additions occurred in samples that contained high ambient nitrate concentrations (e.g., samples from 1.5, 6.1, and 11.0 m beneath the water table at site B). Samples that originated from zones that contained low nitrate concentrations had a substantially different response to the glucose and nitrate additions. For example, the endogenous rate of nitrous oxide production was very low for core material collected from site D, at a depth of 10.4 m below the water table, and was unaffected by the addition of 1.0 mM glucose (Fig. 4). The addition of 1.0 mM nitrate or 1.0 mM nitrate-1.0 mM glucose greatly increased denitrification, but only after a 25-h time lag; the activity increased exponentially after that point in time. In general, denitrification in samples in which the nitrate concentration was low was stimulated by nitrate; denitrification in samples in which the nitrate concentration was high was stimulated by glucose addition. Denitrification in unamended core samples was assayed a second time along the same transect; depth profiles of the

FIG. 4. Nitrous oxide production by slurried core material obtained from site D at a depth of 10.4 m beneath the water table in November 1984. Samples were incubated at ambient nitrate concentration (0) and amended with 1 mmol of nitrate per liter (0), 1 mmol of glucose per liter (O), or 1 mmol of nitrate per liter and 1 mmol of glucose per liter (U).

rates of activity are shown in Fig. 5. The highest rates [260 pmol of N20 produced (g of wet sediment)-' h-1] were for samples obtained from site A, at a depth of 7.0 m below the water table; the lowest rates [2.3 pmol of N20 produced (g of wet sediment)-' h-'] were for samples obtained from site D, at a depth of 16.8 m below the water table. The depth profiles demonstrated that the maximum rate of nitrous oxide pro-

was

WaterTable

O 5t

SteV

_A

aera

eSite B 15

20

0

60 120 180 240 N20 PRODUCTION I pmolig sedimentr1houi'l

FIG. 5. Depth profile of rates of denitrification potential in slurried core material at ambient nitrate concentrations. Samples from sites A (0), B (U), and C (0) were assayed in May 1985; the sample from site D (LI) was assayed in November 1984. Each point represents a best-fit linear regression of 5 to 6 time points from triplicate flasks during a 48-h incubation.

VOL. 54, 1988


TABLE 3. Ammonium production relative to nitrate consumption in sediment slurry incubations from site B" Depth (m) beneath water table

1.5 6.1 11.0 16.4 26.8

Rate [nmol (g of wet sediment) ' day l

Nitrate consumption

17.0 33.1 9.5 0 8.3

+ + + +

2.5 4.8 3.9 0 2.7

SDI of":

Ammonium production

0.1 +0.1 ± 0.6 ± 0.1 ± 0.9 ± 1.4

-1.8 -0.4 -3.5 1.0

Samples were collected in November 1984. " Incubation time was 15 days; rates were calculated from the concentration at the final time minus the concentration at time zero.

duction decreased sequentially from site A through site D. This sequence also represents an increasing length of travel time for nitrate and other sewage effluent constituents that entered the aquifer at the infiltration sand beds. Dissimilatory nitrate reduction. Core samples from all four sites were also tested by two different methods to determine whether dissimilatory nitrate reduction to ammonium was occurring in the aquifer. The first technique involved monitoring changes in in situ nitrate and ammonium concentrations during incubations with sediment slurries. Results were similar for all samples tested, as exemplified by samples obtained from site B (Table 3). Nitrate consumption rates were higher than the corresponding rates of nitrous oxide production in the presence of acetylene, but there was not an accompanying increase in ammonium concentration. The only exception was a slight increase for samples collected from the deepest location at site B (Table 3); however, the ammonium produced was significantly less than the nitrate consumed. Indeed, for most samples, a net loss of ammonium occurred during the incubation. The nitrate consumption rate (0 nmol [g of sediment]-' day-1) for samples collected from site B, at a depth of 16.4 m below the water table, reflects the fact that the initial nitrate concentration for this sample was at the limit of analytical detection (Table 1). During these incubations, ammonium may have been consumed or incorporated into microbial biomass as fast as it was produced from nitrate; therefore, a second set of assays with '5N-enriched nitrate was conducted. For most samples from sites A through D, a significant 15N enrichment of the ammonium pool did not occur after a 60-h incubation (Table 4). The ammonium concentration in about one-half of the samples was very low (Table 1), so a slight enrichment in '5N atom% of the ammonium pool (for example, site B, at a depth of 11.0 m below the water table) represents only a small quantity of nitrate reduced to ammonium. A substantial quantity of '5N-enriched ammonium was produced by samples obtained from 16.4 m below the water table at site B (Table 4). Most of the added [15N]nitrate in this case was reduced to ammonium. However, as noted previously, the ambient nitrate concentration at site B was very low, and therefore, this result does not have a great deal of significance to in situ nitrate reduction at the sites sampled. DISCUSSION The sewage effluent that was discharged onto the rapidinfiltration sand beds at the study site was well oxygenated and contained high levels of nitrate and relatively low concentrations of dissolved organic carbon (12 mg of C/liter) (Table 1) (36). Since the surface soils in the area are highly

1075

permeable, the effluent percolates into the ground very quickly and minimal ponding occurs. Water samples obtained from porous cup lysimeters placed in the unsaturated zone at different depths in the sand beds had the same nitrate and ammonium concentrations as the effluent. In addition, groundwater samples obtained from wells screened immediately beneath the sand beds contained greater than 20 ,uM dissolved oxygen down to a depth of 2.4 m beneath the water table (R. L. Smith, unpublished data). Thus, the effluent is oxygenated as it travels down through the unsaturated zone, as is the contaminant plume in the immediate vicinity of the infiltration sand beds (also see Table 1, site A). Biological processes (in particular, nitrate reduction) did not seem to have any significant effect on effluent nitrate concentrations during transport through the unsaturated zone. When the effluent enters the aquifer, it is not diluted to any significant extent, as the specific conductance of the effluent was only slightly greater than that of the contaminated groundwater. As the contaminated groundwater subsequently moves downgradient with regional groundwater flow, it sinks and becomes overlain by uncontaminated water because of rainfall recharge to the aquifer (23). This is the reason for the lower specific conductance found in the shallowest wells at sites C and D (Table 1). Between site A, which is located immediately adjacent to a sand bed, and site D, which is downgradient, the nature of the plume changed considerably. At site A the groundwater was oxygenated and contained high concentrations of nitrate; whereas at site D, the core of the contaminant plume was anoxic, or nearly so, did not contain any nitrate, but did contain high concentrations of ammonium. This scenario suggests that a nitrate reduction zone (denitrification or dissimilatory reduction of nitrate to ammonium, or both) has been established within the aquifer in this region of the contaminant plume. For perspective, this distance from the sand beds to site D represents 1.5 to 3.0 years of groundwater travel time, as opposed to 50 years of travel time for the entire plume. When assayed by the acetylene blockage technique, aquifer core samples from this region of the contaminant plume indeed displayed a significant potential for denitrification. The activity was linear during 30- to 40-h incubations, did not display an initial time lag, and was inhibited by the presence of air in the incubation flask. There was no evidence to indicate that dissimilatory nitrate reduction to ammonium was occurring within this zone of the aquifer. The source of the ammonium at sites C and D must be from either organic nitrogen within the contaminant plume or TABLE 4. Reduction of lt5N]nitrate to ammonium in sediment slurry incubations" Depth (m) beneath

Site

water

A

B

Ammonium (atom'-,

taible 0.66 0.77 0.43 1.02

1.5 6.1 11.( 16.4 26.8

0.46 0.49 0.55 0.98

Samples were collected in May 1985.

"ND.

Not determined.

60 h

0h

2.1 7.0 14.4 21.3

'5N + SD) at:

0.08 0.13 0.05 0.04

1.21 t).54 1.25 + 0.44 0.48 + 0.04

± 0.04 ± 0.04 + 0.13 ± 0.07 N D"

0.59 + 0.22 0.60 ± 0.15 1.35 0.19 6.59 ± 1.78 ND

+ + + +

1.60 ± 0.69

1076

SMITH AND DUFF

higher effluent ammonium concentrations at some point in the past. It should be noted that the mobilities of nitrate and ammonium at this site were not the same. Ammonium is sorbed onto aquifer solids and therefore is transported nonconservatively at a slower rate than nitrate (7). Denitrification, then, is the predominant nitrate-reducing mechanism within this aquifer. As the bulk of the denitrifying activity was associated with core solids (Fig. 2), the population of denitrifying organisms must be attached to particulate surfaces rather than existing as free-living organisms in the interstitial groundwater. The same is true for the entire microbial community, for greater than 96% of the organisms within the contaminant plume are particle bound (12). While this situation is probably no different from that for the sediments in most surface water systems, the result is an interesting contrast between groundwater and surface water sedimentary environments. In surface water environments, denitrification is largely limited by the downward diffusion of nitrate into the sediments and is usually restricted to a fairly narrow zone (2, 17, 26). However, in an aquifer where groundwater is flowing (most often in a relatively horizontal direction), nitrate supply is controlled by the rate of water movement past the particulate surfaces. Thus, the hydrologic conditions that affect the establishment and the extent of the denitrifying zone are quite different. The rates of denitrification measured in this aquifer in unamended samples ranged from 2.3 to 260 pmol of N2O produced (g of wet sediment)-' h-1. Assuming a porosity of 0.38 liters of groundwater (liter aquifer)-' and using a conversion factor of 1.2 kg of wet sediment (liter of aquifer)-1, the higher rate results in the complete reduction of 1 mM nitrate in 25 days. This extrapolation does not actually represent the real situation, because as a liter of groundwater moves through the aquifer, the nitrate concentration and the degradable organic carbon pool diminish, which lowers the rate of denitrification. The rates measured in this study were not necessarily in situ rates. The reason was because at least some of the dissolved oxygen in the groundwater was lost while the flasks were being sparged during the setup of the incubations. For samples in which dissolved oxygen controlled the in situ rate of denitrification, this oxygen loss could have resulted in a rate estimate that was higher than the in situ rate. This may have been the situation for samples from site A, which contained 30 ,uM 02 at the depth that had the highest rate of denitrification (7 m beneath the water table) (Table 1 and Fig. 5). Ronner and Sorensson (29) found that denitrification does not occur above a threshold of 9 ,uM 02 in the water column of the Baltic Sea, while Nakajima et al. (25) reported that the denitrification rate at 20 to 30 ,uM 02 was about one-half the activity under anoxic conditions in the sediments of a eutrophic lake. Sediment slurry incubations were employed in this study because of the limitations imposed by sampling with a splitspoon core barrel. Techniques to obtain a core sample with the interstitial groundwater intact are being developed. These techniques will allow incubations with whole cores. There have been only a few previous studies in which denitrification rates in groundwater systems were assessed directly. Using the acetylene blockage technique, Slater and Capone (33) measured a rate of 350 pmol of N20 produced (g of dry sediment)-' h-1 for a core obtained from the surface aquifer on Long Island, N.Y., while Ward (40) reported rates of 78,000 to 96,000 pmol of N20 produced (g of dry sediment)-1 h-1 in nitrate-amended core samples obtained adjacent to a septic tile field. Using a completely different


approach, Trudell et al. (38) conducted an in situ injectionwithdrawal test, measuring rates of decrease in injected nitrate concentrations relative to that of a conservative tracer (bromide) in a sandy aquifer in Ontario, Canada. The resulting rates were 0.3 to 4.6 ,umol of N20 produced (liter of groundwater)-' h-V, which can be converted to 95 to 1,457 pmol of N20 produced (g of wet sediment)-' h-1 by assuming that the conversion factors for the Cape Cod aquifer are appropriate for the aquifer in Canada. For the most part, these rates from groundwater systems were difficult to compare with those from surface water sediments or soils, as the latter were usually reported on an areal basis. The rates measured in this study were lower than the rates in a loamy soil (8) or in a marine sediment (13), 25,000 to 50,000 and 417 to 2,290 pmol of N2 produced g-1 h-', respectively, but were in the range reported for individual soil aggregates, 1 to 460 pmol of N20 produced g-' h-' (31). Although the rates in the Cape Cod aquifer may be at the low end of the activity scale, they were sufficient to deplete most of the nitrate in the contaminant plume by the time the groundwater reached site D. These results demonstrate that denitrification can occur in groundwater systems and thereby serve as a mechanism for nitrate removal from groundwater. In this study, denitrification was carbon-limited but not nitrate-limited at those sites that contained measurable nitrate concentrations. In fact, amending a sample from such a site with additional nitrate usually had no effect on the endogenous rate (Fig. 3). This carbon limitation may be the reason that denitrification is the predominant nitrate-reducing mechanism within the aquifer. Tiedje et al. (37) hypothesized that the competition for nitrate between denitrification and dissimilatory nitrate reduction to ammonium is controlled by the ratio of available carbon to electron acceptor concentration. They propose that as the ratio becomes smaller (i.e., nitrate becomes abundant relative to carbon) denitrification is favored. This seems to be the case in estuarine and salt marsh sediments, where the relative proportion of N2 production increases as nitrate increases (20, 21). At site D, where the nitrate concentration was very low, initial rates of denitrification were also very low, even in sediment slurries to which nitrate and glucose were added (Fig. 4). These latter conditions have been used to assay denitrification capacity or denitrifying enzyme quantity (37). In surface sediments the denitrification capacity frequently is fairly substantial, even in situations in which nitrate is never present (16, 26). Even sandy soils which have 20% 02 in the air-filled spaces exhibit the presence of denitrifying enzymes (37). However, at this groundwater site in Cape Cod, which is essentially anoxic (Table 1), the denitrifying capacity was very low. Only after a 24-h adaptation period was denitrification evident in samples obtained from site D, at which time the activity increased exponentially indicative of growth by denitrifying organisms. In summary, denitrification occurred in a sand and gravel aquifer in response to external loading of nitrate and organic carbon. Rates of nitrate-reducing potential decreased with decreasing nitrate concentration, which also correlated with increasing distance from the nitrate source. Denitrification was the predominant nitrate-reducing mechanism in the aquifer and served as a means of removing combined nitrogen from the contaminant plume. The production of N2 as opposed to that of NH4' is the preferred pathway when the concern centers on nitrate contamination in drinking water. There is also well-documented evidence that denitrifying conditions can support the degradation of a number of aromatic compounds (4, 34, 43). Therefore, the denitrifying


VOL. 54, 1988

zone within the Cape Cod aquifer may be removing or transforming some of the xenobiotic compounds that are entering the aquifer. In general, because the additiori of nitrate to a contamination zone is relatively easy, denitrification needs to be considered as a potential mechanism for in situ treatment of groundwater contaminated with certain types of organic compounds.

18. 19.

ACKNOWLEDGMENTS We thank D. R. LeBlanc and his colleagues at the U.S. Geological Survey office in Boston, Mass., for help and technical assistance. B. L. Howes for the use of his laboratory and gas chromatograph. and M. L. Ceazan for nitrate and ammonia analyses. We also thank R. W. Harvey for reviewing the manuscript.

20.

LITERATURE CITED

22. 23.

1. American Public Health Association, American Water Works Association, and Water Pollution Control Federation. 1985. Standard methods for the examination of water and wastewater, 16th ed. American Public Health Association, Washington, D.C. 2. Andersen, T. K., M. H. Jensen, and J. Sorensen. 1984. Diurnal variation of nitrogen cycling in coastal, marine sediments. 1. Denitrification. Mar. Biol. 83:171-176. 3. Balderston, W. L., B. Sherr, and W. J. Payne. 1976. Blockage by acetylene of nitrous oxide reduction in Pseudoinonas peifectomnarinaiis. Appl. Environ. Microbiol. 31:504-508. 4. Bossert, I. D., M. D. Rivera, and L. Y. Young. 1986. p-Cresol biodegradation under denitrifying conditions: isolation of a bacterial coculture. FEMS Microbiol. Ecol. 38:313-319. 5. Bremner, J. M., and R. D. Hauck. 1982. Advances in methodology for research on nitrogen transformations in soils. Agronomy 22:467-502. 6. Burden, R. J. 1982. Nitrate contamination of New Zealand aquifers: a review. N.Z. J. Sci. 25:205-220. 7. Ceazan, M. L., E. M. Thurman, and R. L. Smith. 1987. The role of cation exchange in the transport of ammonium and nitrate in a sewage-contaminated aquifer, p. B53-B58. In B. J. Franks (ed.), U.S. Geological Survey program on toxic waste-groundwater contamination: Proceedings of the Third Technical Meeting, Pensacola, Florida. U.S. Geological Survey Open-File Report 87-109. U. S. Geological Survey, Tallahassee, Fla. 8. Firestone, M. K., and J. M. Tiedje. 1979. Temporal change in nitrous oxide and dinitrogen from denitrification following onset of anaerobiosis. Appl. Environ. Microbiol. 38:673-679. 9. Foster, S. S. D., A. C. Cripps, and A. Smith-Carington. 1982. Nitrate leaching to groundwater. Philos. Trans. R. Soc. London 296:477-489. 10. Gillham, R. W., and J. A Cherry. 1978. Field evidence of denitrification in shallow groundwater flow systems. Water Pollut. Res. Can. 13:53-71. 11. Gormly, J. R., and R. F. Spalding. 1979. Sources and concentrations of nitrate-nitrogen in groundwater of the Central Platte Region, Nebraska. Ground Water 17:291-301. 12. Harvey, R. W., R. L. Smith, and L. George. 1984. Effect of organic contamination upon microbial distributions and heterotrophic uptake in a Cape Cod, Mass., aquifer. Appl. Environ. Microbiol. 48:1197-1202. 13. Horrigan, S. G., and D. G. Capone. 1985. Rates of nitrification and nitrate reduction in nearshore marine sediments at near ambient substrate concentrations. Mar. Chem. 16:317-327. 14. Howard, K. W. F. 1985. Denitrification in a major limestone aquifer. J. Hydrol. 76:265-280. 15. Kaspar, H. F. 1983. Denitrification, nitrate reduction to ammonium, and inorganic nitrogen pools in intertidal sediments. Mar. Biol. 74:133-139. 16. Kaspar, H. F. 1985. The denitrification capacity of sediment from a hypereutrophic lake. Freshwater Biol. 15:449-453. 17. Kaspar, H. F., R. A. Asher, and I. C. Boyer. 1985. Microbial nitrogen transformations in sediments and inorganic nitrogen fluxes across the sediment/water interface in the South Island

21.

24.

25.

26. 27.

28. 29. 30. 31.

32.

33. 34. 35. 36.

1077

West Coast, New Zealand. Estuarine Coastal Shelf Sci. 21: 245-255. Kaspar, H. F., J. M. Tiedje, and R. B. Firestone. 1981. Denitrification and dissimilatory nitrate reduction to ammonium in digested sludge. Can. J. Microbiol. 27:878-885. Katz, B. G., J. B Lindner, and S. E. Ragone. 1980. A comparison of nitrogen in shallow groundwater from sewered and unsewered areas, Nassau County, New York, from 1952 through 1976. Ground Water 18:607-616. King, D., and D. B. Nedwell. 1985. The influence of nitrate concentration upon the end-products of nitrate dissimilation by bacteria in anaerobic salt marsh sediment. FEMS Microbiol. Ecol. 31:23-28. King, D., and D. B. Nedwell. 1987. The adaptation of nitratereducing bacterial communities in estuarine sediments in response to overlying nitrate load. FEMS Microbiol. Ecol. 45:15-20. Knowles, R. 1982. Denitrification. Microbiol. Rev. 46:43-70. LeBlanc, D. R. 1984. Sewage plume in a sand and gravel aquifer, Cape Cod. Massachusetts. U.S. Geological Survey Water Supply Paper 2218. U.S. Government Printing Office, Washington, D.C. LeBlanc, D. R. (ed.). 1984. Movement and fate of solutes in a plume of sewage-contaminated groundwater, Cape Cod, Massachusetts: U.S. Geological Survey Toxic Waste Ground-Water Contamination Program. U.S. Geological Survey Open-File Report 84-475. U.S. Geological Survey, Boston. Nakajima, M., T. Hayamizu, and H. Nishimura. 1984. Effect of oxygen concentration on the rates of denitratification and denitritification in the sediments of an eutrophic lake. Water Res. 18:335-338. Oremland, R. S., C. Umberger, C. W. Culbertson, and R. L. Smith. 1984. Denitrification in San Francisco Bay intertidal sediments. Appl. Environ. Microbiol. 47:1106-1112. Porter, K. S. 1980. An evaluation of sources of nitrogen as causes of ground-water contamination in Nassau County, Long Island. Ground Water 18:617-625. Robertson, F. N. 1979. Evaluation of nitrate in the groundwater in the Delaware Coastal Plain. Ground Water 17:328-337. Ronner, U., and F. Sorensson. 1985. Denitrification rates in the low-oxygen waters of the stratified Baltic Proper. Appl. Environ. Microbiol. 50:801-806. Saffigna, P. G., and D. R. Keeney. 1977. Nitrate and chloride in groundwater under irrigated agriculture in central Wisconsin. Ground Water 15:170-177. Sexstone, A. J., N. P. Revsbech, T. B. Parkin, and J. M. Tiedje. 1985. Direct measurement of oxygen profiles and denitrification rates in soil aggregates. Soil Sci. Soc. Am. J. 49:645-651. Skougstad, M. W., M. J. Fishman, L. C. Friedman, D. E. Erdmann, and S. S. Duncan (ed.). 1979. Methods for determination of inorganic substances in water and fluvial sediments, p. 419-424. In U.S. Geological Survey techniques of water-resources investigations, book 5. U.S. Government Printing Office, Washington, D.C. Slater, J. M., and D. G. Capone. 1987. Denitrification in aquifer soil and nearshore marine sediments influenced by groundwater nitrate. Appl. Environ. Microbiol. 53:1292-1297. Sleat, R., and J. P. Robinson. 1984. The bacteriology of anaerobic degradation of aromatic compounds. J. AppI. Bacteriol. 57:381-394. Spalding, R. F., J. R. Gormly, B. H. Curtiss, and M. E. Exner. 1978. Nonpoint nitrate contamination of groundwater in Merrick County, Nebraska. Ground Water 16:86-95. Thurman, E. M., L. B. Barber, and D. R. LeBlanc. 1986. Movement and fate of detergents in groundwater: a field study. J. Contam. Hydrol. 1:143-161.

37. Tiedje, J. M., A. J. Sexstone, D. D. Myrold, and J. A. Robinson. 1982. Denitrification: ecological niches, competition, and survival. Antonie van Leeuwenhoek J. Microbiol. Serol. 48:569583. 38. Trudell, M. R., R. W. Gillham, and J. A. Cherry. 1986. An in-situ study of the occurrence and rate of denitrification in a shallow unconfined sand aquifer. J. Hydrol. 83:251-268.

1078

SMITH AND DUFF

39. Vogel, J. C., A. S. Talma, and T. H. E. Heaton. 1981. Gaseous nitrogen as evidence for denitrification in groundwater. J. Hydrol. 50:191-200. 40. Ward, T. 1985. Characterizing the aerobic and anaerobic microbial activities in surface and subsurface soils. Environ. Toxicol. Chem. 4:727-737. 41. Whitelaw, K., and J. F. Rees. 1980. Nitrate-reducing and ammonium-oxidizing bacteria in the vadose zone of the Chalk


Aquifer of England. Geomicrobiol. J. 2:179-187. 42. Yoshinari, T., R. Hynes, and R. Knowles. 1977. Acetylene inhibition of nitrous oxide reduction and measurement of denitrification and nitrogen fixation in soil. Soil Biol. Biochem. 9:177-183. 43. Young, L. Y. 1984. Anaerobic degradation of aromatic compounds, p. 487-523. In D. T. Gibson (ed.), Microbial degradation of organic compounds. Marcel Dekker, Inc., New York.