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a Daniel B. Warnell School of Forest Resources, Athens, GA 30602
b Dep. of Crop and Soil Sciences, The Univ. of Georgia, Athens, GA 30602
c USDA Forest Service, Northeast Research Station, P.O. Box 267, Irvine, PA 16329
d Nutter and Associates, 1073 S. Milledge Ave., Athens, GA 30605
* Corresponding author (lmorris{at}arches.uga.edu)
Received for publication March 13, 2000.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: PVC, polyvinyl chloride
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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One acceptable avenue of N removal from the forest ecosystem involves the microbially mediated denitrification of NO3 into gaseous nitrogen (N2) and nitrous oxide (N2O), which are harmlessly carried into the atmosphere. Rates of denitrification are usually low in undisturbed forests (Groffman and Tiedje, 1989; Groffman et al., 1991; Struwe and Kjoller, 1994) but are known to increase in sites irrigated with high N concentration wastewater. Federal and state guidelines for forest land treatment system design allow for up to 25% of applied N to be removed through denitrification (USEPA, 1981) and forested land treatment systems are often designed assuming denitrification of 15 to 20% of applied N.
The actual extent to which N is lost from forested land treatment systems by denitrification is poorly documented. Results from available studies are site and effluent specific. In some cases denitrification has been shown to be an insignificant source of N removal in forests irrigated with municipal effluent. For instance, Barton et al. (1999) found that in situ denitrification rates in a New Zealand Monterey pine (Pinus radiata D. Don) forest irrigated with municipal wastewater averaged only 2.4 kg ha-1 yr-1 N. Smith et al. (1994) reported similarly low rates for Monterey pine forests irrigated with municipal effluent in Australia. In contrast, Burton and Hook (1979) estimated denitrification losses in a maple (Acer saccharum Marsh.)–beech (Fagus grandifolia Ehrh.) forest in Michigan irrigated with 10 cm wk-1 of municipal effluent were about 225 kg ha-1 N or about 55% of input. Denitrification losses were negligible when this same site was irrigated with 5 cm wk-1 of wastewater because this lower rate of irrigation failed to saturate the soil. Kim and Burger (1997) reported denitrification rates as high as 20 kg ha-1 yr-1 N in mature Appalachian hardwood forests of the eastern USA irrigated with municipal wastewater. Russell et al. (1993) reported that N evolution rates varied from 105 kg ha-1 yr-1 N to >2000 kg ha-1 yr-1 N in forests irrigated with meat processing effluent.
A large degree of spatial and temporal variability in denitrification rates can make it difficult to accurately estimate denitrification from forested landscapes (Groffman and Hanson, 1997; Gold et al., 1998; Jacinthe et al., 1998; Myrold, 1998). Estimates of denitrification based on sampling and direct measurement in upland areas may underestimate landscape-scale denitrification because of the high rates of denitrification that can occur in relatively small areas of riparian zones. Saturated soil conditions characteristic of these areas coupled with high concentrations of NO3–N draining from upslope areas and an abundance of soluble C will promote denitrification (Schipper et al., 1993). Studies have shown that large losses of NO3–N can occur from water flowing through forested riparian zones. For instance, Jacobs and Gilliam (1985) found NO3–N decreased from >7 mg L-1 near the edge of a riparian forest to <0.1 mg L-1 near the stream edge in the Coastal Plain of North Carolina, a loss that they attributed largely to denitrification. Verchot et al. (1997) measured NO3–N concentration change within a forested riparian zone adjacent to an agricultural field in the Piedmont of North Carolina and found that subsurface flow NO3–N concentrations decreased from >8 mg L-1 to near 0 mg L-1 as water moved through the riparian zone. Several processes contribute to these decreases, and although they are often attributed largely to denitrification, processes such as vegetative uptake and soil fixation are also important (Lowrance et al., 1984; Pinay et al., 1993). Direct measurement of denitrification in riparian zones receiving high NO3–N inputs suggest rates can exceed 4000 kg ha-1 yr-1 N under ideal conditions (Schipper et al., 1993).
Studies of denitrification from riparian zones suggest that NO3–N concentrations are the most important factor limiting denitrification rate with soluble C concentrations of secondary importance. For instance, Schipper et al. (1993) investigated factors regulating denitrification in a riparian zone beneath an irrigated slope in New Zealand. These investigators found that up to 77% of the variation in denitrification rate could be explained by differences in NO3–N concentration and denitrifying enzyme activity. Some evidence for a temperature effect was noted, but a relationship was not developed. Jordon et al. (1998) studied the effect of NO3–N and carbohydrate addition on denitrification rates in a riparian zone forest. In their study, the forest received NO3–N in drainage from a corn (Zea mays L.) field. Using large in situ chambers, these investigators found that denitrification rates ranged from 4.8 to 408 µg N2O–N d-1. Addition of NO3–N alone, or in combination with soluble C as sucrose, stimulated denitrification rate more than 10-fold at all locations. Addition of sucrose alone stimulated denitrification in some locations but not in others, suggesting that NO3–N limited denitrification throughout the riparian forest and that, in places, soluble C also limited denitrification. In a 2-yr study of denitrification rates in the Catskill Mountains of New York, Ashby et al. (1998) found that denitrification rates were positively correlated with soil organic matter concentration, total P, and water-filled pore space, which were highest in toe-slope and stream-edge positions. Rates of denitrification in these natural riparian zones were low, averaging from 0.75 to 1.4 kg ha-1 yr-1 N. Denitrification rates following amendment of these sites with glucose and NO3–N suggested that both the supply of readily available C and NO3–N limited denitrification from these sites. In the Coastal Plain of Georgia, Ambus and Lowrance (1991) showed that factors limiting denitrification from forested riparian zones differed by soil. More poorly drained soils with higher C contents were able to denitrify more added NO3–N than better drained soils, suggesting that C was less limiting in poorly drained soils, but in both cases denitrification was increased by the addition of NO3–N.
Most studies of denitrification from forested land treatment systems were completed within a relatively short time (<5 yr) after irrigation was begun. These systems are likely to still be responding to the changes in water and N loading associated with the initiation of land treatment. They are unlikely to represent evolution rates in older systems that are nearer steady-state and more likely to be approaching maximum rates of wastewater and N loading allowed for in system design. In this study, we examined evolution of N2 + N2O–N from a forested land treatment system that has been operationally irrigated with municipal wastewater for >13 yr. Our specific objectives were to (i) quantify and compare N evolution rates in irrigated and nonirrigated areas of forest, (ii) evaluate the influence of landscape position on N evolution rates, and (iii) investigate relationships between soil properties and denitrification rates that would suggest what factors limit denitrification from riparian zones following long-term effluent application. We expected to find generally high rates of denitrification from this site, particularly in riparian zone and toe-slope landscape positions. Soils in this region have well developed argillic B horizons that would result in saturated zones at the base of irrigated slopes. Forest productivity is high and decomposition rapid. It seemed likely that C availability would be sufficient to sustain high rates of denitrification.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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The site is contained within the headwaters of Pates Creek in the Piedmont physiographic region. The underlying geology is dominated by granitic gneiss with some fracturing and jointing. Dominant soils in the uplands are well drained Typic Kanhapludults with A horizon textures ranging from fine-sandy loam to sandy-clay loam. The A horizon is shallow due to past erosion history and rarely exceeds 0.15-m depth. The B horizon is sandy clay to clay-textured. Subsoil (B) horizon saturated hydraulic conductivity averages 8.4 x 10-4 cm s-1. Poorly and very poorly drained soils are restricted to narrow zones along streams (Nutter, 1986).
Climatic conditions in the Georgia Piedmont favor year-round application of wastewater. Average precipitation is 1035 mm yr-1, with March generally the wettest month (154 mm) and October the driest (64 mm). Average annual temperature is 16.7°C, with no daily average temperature below freezing. Snowfall averages <10 cm yr-1 and soil freezing is rare, particularly in forests. Potential evapotranspiration is estimated to be 890 mm yr-1 (Nutter, 1986).
Wastewater loading is limited by N and water-assimilative capacities of the site. Based on site design investigations and the nutrient and hydraulic budgets, wastewater irrigation is limited to 594 kg ha-1 yr-1 N and 6.4 cm wk-1 water (Nutter et al., 1996). Actual N loading has not reached this design rate and was 407 kg ha-1 yr-1 N during 1997.
Experiment 1—Upland Areas
Establishment of Sample Plots
In the summer of 1995, paired gas sampling plots were established at three different landscape positions at three locations within the land treatment system. Hilltop, midslope, and toe-slope positions were established at each location. Hilltop positions were located at the top of the hill or on convex-shaped shoulders and had slopes from 4 to 6%. Midslope positions were located midway between hilltops and drainages and had slopes averaging 6 to 12%. Toe-slope positions were convex-shaped areas located just above riparian zones and had slopes ranging from 2 to 5%. For each landscape position and location, one plot of the plot pair was located within the irrigated area and the second plot was located within nonirrigated buffer zones that were left adjacent to property boundaries and roads.
Gas Sampling Procedures
Gas was sampled using a static chamber technique. Stainless steel chamber bases (0.5 x 1.0 m) were permanently installed on each plot to minimize disturbance during sampling. Rectangular Plexiglas chamber tops (100 cm length, 50 cm width, and 20 cm height) were attached to each base with a water seal to form a contained atmospheric head space of approximately 100 L above the soil surface during sampling events. Construction of chamber tops and bases and sampling procedures are described by Dulohery et al. (1996).
Plexiglas chamber tops were transported to the field before sampling events. At the start of an incubation, chamber tops were placed on the base and a reflective insulator was placed over the chamber. Initially, the chamber sampling port was left open and air allowed to escape from the chamber until chamber atmosphere had equilibrated with atmospheric pressure. The port was then sealed and the chamber fan was run for 30 s to mix air within the chamber. Chamber air was then sampled from the septum port using a hypodermic needle and stored in glass serum vials. For chamber gas sampling, 30-mL gas samples were collected using a 50-mL gas-tight syringe with a butyl rubber septum and stored in 10-mL crimp-top (West Company, Phoenixville, PA) evacuated vials. The atmosphere within each chamber was sampled at time zero after sealing the chambers and again after 1 and 2 h. Initial work conducted at this site found evolution rates to be linear at least through this 2-h period. Chamber tops were removed from bases between sampling events.
Gas flux was measured approximately bimonthly for 1 yr beginning in July 1995 and continuing to July 1996. To determine short-term changes in gas flux following irrigation, gas was sampled 1, 3, and 5 d after irrigation during the initial monitoring period in September 1995.
Incubation Core Collection for N2/N2O Ratio Determination
To provide in-field estimates of N2/N2O ratio, soil cores were collected within a 0.3-m distance of each chamber for laboratory incubation by the acetylene inhibition method (Tiedje, 1982). Cores were collected using 5-cm lengths of aluminum tube, which could be easily pressed into the ground to 5-cm depth to remove a section of the soil intact. Cores were sealed in plastic bags, transferred to an ice-pack-filled cooler, and returned to the laboratory.
Laboratory Analyses of Gas Concentrations
Gas samples were analyzed for N2O concentration on a Shimadzu GC-14A gas chromatograph (Shimadzu Scientific Instruments, Columbia, MD), using a 10-port valve with a sample loop of fixed 2-mL volume. Column arrangement consisted of 91 cm of Poropak N with 80/100 mesh in series with 305 cm Poropak Q with 100/120 mesh. An electron capture detector (63Ni) was used with a 95:5 argon–methane carrier gas mixture flowing at 40 mL min-1, a column temperature of 70°C, and a detector temperature of 340°C.
To index overall microbial activity, monitoring of the upland sites during 1995–1996 included measurement of CO2 evolution. The same Shimadzu GC-14A gas chromatograph using a thermal conductivity detector, a He carrier at a flow rate of 35 mL min-1, a 50°C column, and 110°C detector were used to measure CO2 in gas samples. The mass of a particular gas in the chamber was calculated using the Ideal Gas Law and Dalton's Law of partial pressures.
Core Incubation for Determination of N2/N2O Ratios
In the laboratory, cores were removed from the plastic bags and placed in small PVC chambers fitted with septum ports. These cores were incubated at room temperature (
21°C) for 5 h. Samples were collected from the headspace at the beginning of the incubation, 2.5 h following the beginning of incubation, and at the end of the 5-h incubation period. Following this initial 5-h incubation, the cores were opened to the atmosphere, allowed to equilibrate, and then resealed. Then, 10% of the headspace from each core was removed and replaced with acetylene (C2H2) gas. Acetylene gas inhibits the production of N2 gas causing only the production of N2O gas during further denitrification (Tiedje, 1982). The 5-h incubation was repeated and the headspace again sampled at the beginning, middle, and end of the 5-h incubation period. The amount of N2 produced was estimated as the amount of N2O produced during the second incubation minus the amount produced during the first. The resulting N2/N2O ratio was applied to the amount of N2O measured in the field to estimate N2 evolution.
Statistical Analyses
Differences in measured variables between irrigated and nonirrigated buffer zone plots were evaluated using mean difference (Md) and standard error of the mean difference (SE) as:
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Experiment 2—Riparian Zones
Establishment of Sample Plots
Riparian zone plot pairs were established at four locations within the land treatment site in irrigated areas and in nonirrigated buffer zones in 1997. Each plot-pair was located within the intermittently saturated zone adjacent to a stream or wetland and were matched based on surrounding hydrology (Table 1) and initial soil gleying. Except as noted below, the four plot pairs were characterized by an oxidized surface horizon underlain by a reduced zone (gleyed soil matrix) less than 0.5-m depth below the soil surface.
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Gas Sampling Procedures
Riparian zones were sampled monthly for 1 yr beginning in September 1997 and continuing to September 1998 using the same chambers and procedures similar to those previously described for sampling upland areas in Experiment 1. Sampling differed in two ways. First, a smaller 3-mL gas sample was collected and stored in 2-mL vials. Second, samples were collected only at the start and end of the 2-h field incubation period.
Temperatures were measured with Taylor model 5367 max–min thermometers placed outside of each chamber within a 0.3-m distance. Air temperature was measured 20 cm above the soil surface and soil temperature was measured at 5-cm soil depth.
Soil Sampling Procedures
Soil sampling was conducted during the riparian zone monitoring to evaluate the relationship between soil properties and rates of N2 and N2O evolution. For this sampling, approximately 100–500 g of loose soil was collected from the top 10 cm of soil adjacent to each chamber location during each sampling event and returned to the laboratory for analysis of inorganic N (NH4–N and NO3–N), soluble C, pH, combustible organic matter content, and gravimetric water content. In addition, total N was determined twice (September 1997 and September 1998) and soil textural class and clay content were determined once for samples collected in May 1998. These measurements, along with bulk density, water filled porosity calculations and the previously described temperature readings, were then used to evaluate plot pairing (Table 1) as well as for evaluating the extent to which N2O and N2 gas evolution rates could be predicted from these measures.
Laboratory Analyses of Gas Concentrations
Gas samples were analyzed for N2O–N concentrations using the same Shimadzu GC-14A gas chromatograph (Shimadzu Scientific Instruments, Columbia, MD), column arrangement and detector previously described for Experiment 1. Carbon dioxide concentrations were not measured for riparian zone gas samples.
Core Incubation for Determination of N2/N2O Ratios
Procedures used for determination of N2/N2O ratios were similar to those used in Experiment 1. Differences were as follows. For riparian zone sampling, 15-cm lengths of thin-walled polyvinyl chloride (PVC) pipes rather than aluminum tubing was used to collect samples. Rather than placing tubes in small chambers fitted with septum ports as done in Experiment 1, tubes were fitted with stoppers with septum ports. The headspace from each tube was equilibrated with room atmospheric pressure by inserting a stainless steel hypodermic needle through each septum. After several minutes of air pressure equilibration the needles were removed and cores were placed within a Precision TS-31213-AN-10 Model 816 Low Temperature Incubator (Precision Scientific Group, GCA Corp., Chicago, IL). The cores were then incubated at the field-recorded temperature for 5 h and gas samples were collected from the headspace of each core at the beginning, during, and at the end of the 5-h incubation. Cores were then reopened, vented, resealed, and headspace air pressure equilibrated to room air pressure, as before. Then, 10% of the headspace from each core was removed and replaced with acetylene (C2H2) gas. The 5-h incubation was then repeated and again headspace gas was sampled at the beginning, middle, and end of the incubation period.
After completion of incubations, headspace volume, bulk density, gravimetric water content, volumetric water content, and water filled porosity were measured in the intact cores. Head-space volume was calculated with the use of a Tensimeter (Soil Measurement Systems, Tucson, AZ), which measured the change in pressure inside the sealed intact core tube that resulted after the injection of a known volume of air. For the measurement of bulk density, water contents, and water-filled porosity, the PVC cap and rubber stopper were removed from the ends of the tubes and each tube plus intact core was weighed, dried at 65°C, and reweighed. The soil cores were then removed and the empty tubes were washed, dried, and weighed. Bulk density (
b), gravimetric water content (
g), volumetric water content (v), and water-filled porosity (s) were then calculated according to Jury et al. (1991) as:
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b was the bulk density of intact core (g cm-3), Ms was the mass of dried intact core (g), Vt was the volume of intact core (cm3), g was the gravimetric water content of intact core (g H2O g-1 soil), Mw was the mass of H2O of intact core (g), v was the volumetric water content of intact core (cm3 H2O cm-3 soil), w was the density of H2O assumed to be constant (1.00 g cm-3), and m was the density of minerals, assumed to be constant (2.65 g cm-3).
Laboratory Analyses of Soil Samples
Analysis of total N in soil sampled from riparian zones was conducted using a block total Kjeldahl digestion followed by measurement on a Technicon Autoanalyzer using continuous flow colorimetry (Technicon Industrial Systems, 1978). Inorganic N (NH4–N and NO3–N) were measured using colorimetric methods on a Technicon Autoanalyzer system, following a 2 mol L-1 KCl extraction procedure (Technicon Industrial Systems, 1973). A second extraction was conducted with 0.5 mol L-1 K2SO4 (Allison, 1965; Nelson and Sommers, 1996) for measurement of soluble C (defined as C passing through a 0.2-µm filter) on a Shimadzu TOC 5000 carbon analyzer (Shimadzu Scientific Instruments, Columbia, MD). Gravimetric water content was estimated after drying to constant weight at 105°C (Jury et al., 1991). Combustible organic matter was estimated by the weight lost from the oven-dry sample after furnace muffling for 24 h at 500°C. Soil pH was measured on 20 g of field moist sample mixed with 20 mL of distilled water using a glass calomel pH electrode, and again with 20 mL of 0.01 mol L-1 CaCl2. The Bouyoucus hydrometer method (Gee and Bauder, 1986) was used to determine particle-size distribution and clay content of the riparian plot soils.
Statistical Analyses
Differences in measured variables between irrigated and nonirrigated buffer zone plots were evaluated using the mean difference (Md) and standard error of the mean difference (SE) as described for Experiment 1 using a significance level of P = 0.05. Multiple linear regression was used to evaluate relationships between soil variables and denitrification rates with a confidence level of 0.05 along with calculation of the correlation coefficient (r) using the Corel Quattro Pro 8 software package.
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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With the exception of the July 1996 sampling date, no significant difference in denitrification rate occurred among landscape positions in either irrigated or nonirrigated buffer zones (Fig. 1). On this sampling date, denitrification rate was lower in the hilltop position of the wastewater-irrigated site. One unusually high value was observed for the midslope position in July 1995 but this was entirely due to unusually high denitrification at one location.
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There was some evidence that N evolution rates were highest immediately following irrigation (Fig. 2) but a clear trend of declining rates between irrigation events did not exist.
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Relationship between Soil Characteristics and Denitrification Rate
None of the soil parameters measured monthly (extractable NH4–N and NO3–N, extractable soluble C, pH, combustible organic matter, temperature, water content, and water filled porosity) or more stable measures (soil texture and bulk density) were closely correlated with soil N evolution rates. The only parameter that showed any predictive ability was soil temperature. However, a linear regression of N evolution vs. soil temperature had an r2 value of only 0.18. Low correlations were partially due to infrequent and unusually high N2O gas concentrations measured at some irrigated plots that were 10 to 100 times greater than the normally observed range. Using monthly means to reduce the effect of these observed outliers resulted in an r2 of 0.54 for N evolution vs. soil temperature.
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Denitrification at the E.L. Huie Land Treatment Facility was increased by irrigation and did vary between upland and riparian zones. The overall range of measured values from 5 to 38 kg ha-1 yr-1 N is lower than often reported for land treatment sites and lower than we expected. Ryden et al. (1981) found between 47 and 69 kg ha-1 yr-1 N loss from denitrifying activity by direct measurement after pasture plots were irrigated with municipal wastewater. Russell et al. (1993) found N evolution rates to be between 12 and 240 g ha-1 h-1 N following irrigation of forest with a meat-processing effluent. This corresponds to annual evolution rates of 100 to >2000 kg ha-1 yr-1 N. Our rates were much closer to the rates reported by Barton et al. (1999) of 2.4 kg ha-1 yr-1 N for soils that did not reach the critical moisture threshold for denitrification.
Under natural precipitation, both upland forests and forested riparian zones have been found to have seasonal peaks in denitrification rate in spring and autumn with reduced rates generally occurring in summer when soil moisture is lowest (Zak and Grigal, 1991; Pinay et al., 1993; Groffman et al., 1993; Hanson et al., 1994) and the potential for vegetation uptake of N that can reduce NO3–N concentrations of soil water is greatest (Lowrance, 1992). On our site, effluent irrigation maintained soil moisture throughout the year. This, coupled with relatively stable soil NO3–N concentrations, resulted in peak denitrification rates in summer when soil temperatures were warmest.
Low rates of N2 + N2O–N soil evolution at our site may be the result of continuous wastewater irrigation over a long period and reduction of the necessary soluble C energy source. Insufficient soluble C has been previously shown to limit denitrification in forested land treatment sites. For instance, Lance et al. (1976) reported that soil denitrification could be increased during sewage effluent applications by controlling the available soil NO-3 and C, both of which were limiting denitrification rates. Some evidence suggests this as a possible reason for low denitrification on our study site. Irrigated areas of the E.L. Huie Land Treatment Facility have noticeably less forest floor than nonirrigated areas. Since growth of the forest in irrigated areas is greater than in nonirrigated areas (unpublished data), this indicates more rapid litter turnover and, potentially, greater soluble C to support denitrification. However, we found relatively small differences in CO2 evolution rates between irrigated and nonirrigated upland areas except during summer (July sampling periods). During the July sampling periods, CO2 evolution from irrigated uplands area was about double that of nonirrigated areas (Fig. 2). This increased decomposition and more rapid turnover did not translate to greater pools of soluble C in riparian zones. Levels of soluble C and organic matter were lower in riparian zone plots in irrigated areas than in nonirrigated areas during this period as well as throughout the rest of the year. Instead, it appears that long-term irrigation has depleted riparian zone soil C necessary to support denitrification activity.
Many of the physical and chemical parameters often associated with denitrifying potential were sampled during the riparian zone monitoring portion of this study. The best relationship was with temperature, which had an overall r2 of 0.18 for all sample data and an r2 of 0.54 when monthly means were used to reduce the effect of outliers. Our results are similar to those of other investigators have also found correlations between temperature and denitrification rate to be weak (Schipper et al., 1993; Parsons et al., 1991; Pinay et al., 1993). Several explanations for unexpectedly weak correlations between denitrification rate and temperature have been proposed, but the two most common are that seasonal differences in NO3–N concentrations obscure relationships with temperature and the relatively dry soil conditions during summer limit denitrification when temperatures are warmest. Neither of these explains our results. Little seasonal difference in NO3–N concentrations occurred on our study sites and riparian zones within irrigated areas did not dry out during summer.
No significant relationship occurred between any other measured soil characteristic and N2 + N2O–N riparian zone soil evolution rates. It is possible that difference in surface soil texture or other characteristics (Table 1) were large enough to mask differences associated with these measured variables; however, it seems more likely that the inability to establish a strong relationship is simply a function of the generally low rates of denitrification we observed. Again, our results are similar to those of Barton et al. (1999), who also had difficulty relating denitrification rates to soil factors and could explain only 29% of the observed variation in denitrification rates using soil properties. We were much less successful in explaining riparian zone denitrification than Schipper et al. (1993), who explained up to 77% of the in situ variation in riparian zone denitrification using NO3–N and denitrifying enzyme activity, or Pinay et al. (1993), who were also able to explain up to 77% of the variation in denitrification from a riparian soil receiving high nitrate drainage using a linear combination of soil moisture, extractable glucose equivalent, and NO3–N concentrations.
One problem in most field denitrification research is the difficulty in obtaining reliable field measurements. Groffman and Tiedje (1989) reported a high level of spatial variability for denitrifying activity within forested research plots. This level of variability can be associated with difficulty in sampling small areas when the majority of denitrifying activity is located within infrequently sampled anaerobic microsites of high C availability (Jacinthe et al., 1998; Myrold, 1998). With the exception of a few sample dates, results from this study were relatively consistent among replicated plots, suggesting that the use of large static chambers decreased sample variability related to the inconsistent sampling of micro sites with small sampling cores.
Finally, N2 + N2O–N gas evolution from this site ranged from 5 to >38 kg ha-1 yr-1 N. Using average values for denitrification of 5.0, 9.5, 4.1, and 38.5 kg ha-1 yr-1 N for hilltop, midslope, toe-slope, and riparian zones, respectively, along with estimates of the percentage of the treatment site in each landscape position (estimated from topographic maps), it is possible to compute an area-weighted denitrification rate. On this area-weighted basis, N loss through denitrification accounted for only 2.4% of the total N applied during wastewater irrigation. This is much less than the 10 to 25% rate suggested for use in design of forested land treatment systems in the USEPA design manual (USEPA, 1981). It is unlikely that this high rate of denitrification will be reached or sustained for the long-term without active intervention to improve rate limiting conditions and promote denitrification activity.
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CONCLUSIONS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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ACKNOWLEDGMENTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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