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Dep. of Forest Ecology and Management, Univ. of Wisconsin-Madison, 1630 Linden Dr., Madison, WI 53706-1598
Corresponding author (krbrye{at}students.wisc.edu)
Received for publication March 6, 2000.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: DOC, dissolved organic carbon • ET, evapotranspiration • ETL, equilibrium-tension lysimeter • IC, inorganic carbon • OC, organic carbon • TC, total carbon
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Exponentially decreasing microbial biomass with depth has been documented in agricultural soils (Fraser et al., 1988; Roder et al., 1988; Paul and Clark, 1989; Tessier et al., 1998). However, that is not to say microbial activity does not exist at 1 m or more if given sufficient C substrate to support microbial growth. Dissolved OC has been identified as one potential limiting factor for denitrification in soils (Drury et al., 1991). Additionally, denitrification has been shown to be proportional to dissolved OC in unamended soils (Katz et al., 1985).
Though denitrification has been considered negligible in many grassland ecosystems (Woodmansee, 1978; Seastedt and Hayes, 1988), more recent data suggest that denitrification can be responsible for a much larger loss of N from deep, fertile soils, including tallgrass prairies (Groffman et al., 1993; Groffman and Turner, 1995). Denitrification is limited to soil zones where O2 concentrations are low and available nitrate N and C are sufficiently high (Groffman et al., 1993). Denitrification under anaerobic soil conditions is also regulated by the supply of mineralizable organic matter and OC (Burford and Bremner, 1975; Lind and Eiland, 1989; Yeomans et al., 1992). Water-soluble C contained in soils or found in the deep soil solution could provide an index for the denitrification potential of nitrate N (Burford and Bremner, 1975). Subsoil denitrification potential has been suggested as an important indication of whether an ecosystem can denitrify excess nitrate N in the vadose zone before reaching ground water aquifers (Sotomayor and Rice, 1996). Sotomayor and Rice (1996) also recognized that few studies have assessed subsoil denitrification potential below the crop rooting zone.
The potential magnitudes of N and C leaching losses have been inherently difficult components of their respective nutrient cycles to measure in situ. Martin et al. (1994) recognized the need for nitrate leaching data over complete annual cycles, not just for growing season periods. Reducing the time period for examining the effects of various management practices to several months associated with a growing season seriously limits the inferences that can be drawn about the effects that various management practices may have on solute leaching.
In many regions of the USA, especially in the Midwest, crop establishment occurs after one of the wettest periods of the year. Spring thaws of winter snow accumulation have great potential to move solutes deeper into the soil profile, out of reach of any root system that would be produced by the next crop. Depending on residual soil storage of N and postharvest tillage and fertilization practices, a significant quantity of nitrate N could leach from the system over the winter and during the spring seasons while no crop exists to capture the highly mobile nutrients (Watts and Martin, 1981; Martin et al., 1994).
Technological advancements in past decades and in recent years have allowed some of the obstacles to measuring water and solute movement through soil, even frozen soil, to be overcome with lysimetry (Prunty and Montgomery, 1991; Cameron et al., 1992; Brye et al., 1999). Bergstrom (1990) acknowledged that lysimetry offers a reasonable method to carry out investigations under field conditions that are subject to actual environmental influences. The use of both tension and tensionless lysimeters has provided a means to quantify solute leaching losses, particularly nitrate N losses, from mainly fertilized agroecosystems (Tyler and Thomas, 1977; Bergstrom, 1987; Prunty and Montgomery, 1991; Shipitalo and Edwards, 1993; Baker and Timmons, 1994; Martin et al., 1994). However, a need still exists to establish year-round inorganic N and soluble C leaching losses from common agroecosystems, concentrating on the time between harvest of one crop and planting of the next.
The objectives of this study were threefold: (i) to quantify in situ inorganic N and soluble C leaching year-round from undisturbed soil profiles using equilibrium-tension lysimeters (ETLs), (ii) to determine the influence of management practices on inorganic N and soluble C leaching losses during the entire year, and (iii) to determine the likelihood of deep soil denitrification based on available nitrate N and organic C supplies. We hypothesized that land use significantly influences N and C leaching and that deep soil denitrification potential can be determined by examining year-round patterns of inorganic N and soluble OC concentrations and leaching at depth in the soil profile.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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A randomized complete-block design was established for maize tillage treatments of conventional chisel-plowed and no-tillage in fall 1994 (Brye et al., 2000). A 105-d relative maturity hybrid maize variety was planted in both tillage treatments. Pelletized ammonium nitrate (NH4NO3) was broadcast by hand immediately following planting at a rate of 190 kg N ha-1 yr-1. Fall tillage occurred following harvest in the chisel-plowed treatment.
Four 7- x 7-m plots were established at the prairie site in spring 1995 (Brye et al., 2000). The prairie was restored from agricultural influence in June 1976; the current vegetation is classified as a mesic tallgrass prairie. The prairie was last burned on 18 Apr. 1998. A more detailed description of the sites can be found in Brye (1997) and Wagai et al. (1998).
Soil and Climatic Data
Ambient air temperatures, soil temperatures at 10, 30, 70, and 120 cm, and precipitation were monitored at both study sites (Brye, 1997). A micrometeorological weather station located <150 m from the agricultural site provided air temperatures, solar radiation, wind speed, humidity, and additional precipitation measurements. The equivalent depth of water in the snow pack at the agricultural and prairie sites was measured approximately weekly by manual coring and measuring the equivalent depth of water once the snow core had melted in the laboratory (Brye et al., 2000).
Equilibrium-Tension Lysimeters
Six stainless steel ETLs (0.25 x 0.76 m) were established in field plots during summer and fall of 1995. Two ETLs were installed in the fertilized no-tillage, fertilized chisel-plowed, and tallgrass prairie field plots at 1.4 m below the soil surface. A portable, regulated vacuum system provided continuous suction to the 0.2-µm stainless steel porous plate of the ETLs (Brye et al., 1999). Heat dissipation sensors were placed immediately above the porous plate of each ETL and in the surrounding bulk soil to continuously monitor the matric potential at the two locations (Reece, 1996; Brye et al., 1999). The regulated vacuum system was adjusted manually several times a week to provide suction that was slightly more negative (i.e., a few kPa) than the matric potential recorded in the surrounding bulk soil with the heat dissipation sensors.
The lysimeters were sampled under vacuum every 2 wk between March and December and once every 4 wk during the rest of the year (Brye et al., 1999). Leachate was collected from the lysimeter's collection reservoir, which can contain 
23 L or 110 mm of water, through a sampling tube that extends from a drain port on the lysimeter to the soil surface. The initial leachate up to 1 L was collected into a 1-L high-density polyethylene bottle and transported back to the laboratory where the leachate volume was measured. Any remaining leachate from the lysimeters, >1 L, was collected, volumes were recorded, and the leachate was discarded. Subsample aliquots of the initial 1 L of leachate were filtered through glass fiber filter paper (Whatman G6) and stored at 4°C until chemical analysis could be performed.
Porous Cup Samplers
Porous cup samplers were installed at 1.2 m in each of the four plots of the prairie, fertilized no-tillage, and chisel-plowed maize ecosystems in spring 1995 for nitrate N concentration measurements. Porous cup samplers were left in the field plots over winter. Some samplers ceased functioning the following spring and collected no leachate for the season. Leachate was collected from the porous cup samplers every 2 wk. Subsample aliquots of porous cup sampler leachate solutions were filtered through glass fiber filter paper (Whatman G6) and stored at 4°C until chemical analysis could be performed.
Soil Water Profiles
Soil water profiles were measured every 7 d from March through October using a neutron hydroprobe (Campbell Pacific Nuclear, Martinez, CA; Model 503) (Brye et al., 2000). Soil water profile measurements were replicated four times in each ecosystem and used to construct water budgets for all three ecosystems.
Chemical Analyses
The ETL leachate was analyzed for inorganic N (NO-3–N and NH+4–N) and soluble C. Porous cup sampler leachate was analyzed for nitrate N only. Inorganic N analysis was performed colorimetrically using a Lachat (Milwaukee, WI) continuous-flow ion analyzer (Lachat, 1993a,b). Total carbon (TC) and inorganic carbon (IC) were determined by high-temperature catalytic combustion using a Rosemount–Dohrmann (Santa Clara, CA) DC-190 total dissolved C analyzer. The DOC was determined by difference between TC and IC. Leachate samples were exposed to atmospheric carbon dioxide (CO2) concentrations for less than 0.3 h. Equilibration with atmospheric CO2 was assumed negligible.
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RESULTS AND DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Drainage
Drainage through 1.4 m of undisturbed soil was measured for nearly 4 yr between June 1995 and April 1999 in the restored tallgrass prairie, fertilized no-tillage maize, and fertilized chisel-plowed maize ecosystems (Fig. 1)
. Water movement occurred in two phases, a drainage phase and a nondrainage phase, where the drainage phase lasted from 3 to 8 mo, but typically occurred between January and July (Brye et al., 2000). Mean drainage (± standard error [SE]) recorded over 4 yr for the prairie, fertilized no-tillage, and fertilized chisel-plowed ecosystems totaled 461 (SE ± 14), 1116 (SE ± 96), and 1575 (SE ± 261) mm, respectively (Table 1). This drainage represented 16, 33, and 47% of the measured precipitation plus melted drifted snow that the prairie, no-tillage, and chisel-plowed ecosystems experienced during this time. However, various soil physical properties could affect the flow paths of water through the soil and ultimately affect the collection efficiency of the lysimeters.
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More than 95% of the time, drainage fluxes less than 2.2 mm d-1 for the prairie and less than 5.0 mm d-1 for the agroecosystems were recorded during the drainage phases. A single extreme rainfall event in June 1996 produced the largest drainage flux from all three ecosystems (Brye et al., 2000). In response to receiving more than 100 mm of rainfall in less than 3 d, drainage fluxes of 8, 39, and 40 mm d-1 were recorded by the ETLs for the prairie, fertilized no-tillage, and fertilized chisel-plowed ecosystems, respectively. The low drainage flux observed in the prairie ecosystem suggests that substantial macropore flow might result in an underestimation of the prairie drainage flux by the ETLs.
A detailed water balance for this study, describing overland runoff and ET, is described in a companion paper (Brye et al., 2000). An evaluation of the water balance for the prairie and agroecosystems demonstrated that runoff from each ecosystem was negligible during the summer, but significant during winter and spring months (Brye et al., 2000). During 1996 and 1997 there were three runoff events. Runoff from the no-tillage plots was 15 mm greater than runoff from the chisel-plowed plots in both years and greater than the prairie in 1996 (Brye et al., 2000). Greater runoff, though not substantially more, and presumably less infiltration from the no-tillage plots compared with the chisel-plowed plots, could partially explain smaller drainage in the no-tillage agroecosystem.
As suggested by Martin et al. (1994), solute leaching losses could be significant between cropping periods, therefore drainage must be significant to transport soluble constituents through the soil and out of the root zones of crops. Table 3 summarizes winter drainage, precipitation, and frost penetration recorded for the prairie, fertilized no-tillage, and fertilized chisel-plowed ecosystems. Between the months of December and March, ET was small to negligible, causing drainage to account for an extensive fraction of precipitation as snow that subsequently melted and infiltrated the partially frozen soil profile (Table 2 of Brye et al., 2000). Because of the relatively small plot sizes and the proximity of the agricultural field site to a property boundary fence, snow drifts accumulated on the field plots, affecting both tillage treatments evenly (Brye et al., 2000). The drifting snow eventually added extra water, in excess of natural precipitation that fell, to the soil profile of both tillage treatments. Of the total precipitation plus melted drifted snow, the drainage fraction was 0.16, 0.33, and 0.47 for the prairie, no-tillage, and chisel-plowed ecosystems, respectively. Even during the winter months, the volumetric soil water content was higher under the no-tillage treatment than under the chisel-plowed treatment (Brye et al., 2000). The larger winter drainages recorded for the chisel-plowed than for the no-tillage agroecosystem are attributable to extra water added to the system from the accumulation of drifting snow. Snow drift thicknesses typically ranged from 0 to 25 cm at the prairie and from 0 to 46 cm at the agricultural site. The equivalent water depth of the snow pack during the winter ranged from 0 to 62 mm at the prairie and from 0 to 132 mm at the agricultural site (Brye et al., 2000). The water added from melted snow passed through the chisel-plowed soil profile quicker, while extra snow melt remained in the no-tillage soil profile longer (Brye et al., 2000). Differing ET rates (Table 2 of Brye et al., 2000) alone are insufficient to account for the differences in drainage between the prairie and agroecosystems.
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Inorganic Nitrogen Concentrations
Nitrate N concentrations measured in the ETL-collected leachate increased seasonally in response to N fertilization in the agroecosystems, while the prairie consistently maintained less than 0.7 mg L-1 of NO-3–N in the soil solution (Fig. 2)
. During 1996, 1997, and 1998, growing season (i.e., April through October) nitrate N concentrations averaged 21, 14, and 23 mg L-1 for the fertilized no-tillage and 16, 11, and 10 mg L-1 for the fertilized chisel-plowed maize agroecosystem (Fig. 2). Similar to nitrate N concentrations in ETLs, nitrate N concentrations measured in solutions collected from porous cup samplers in the prairie averaged less than 0.3 mg L-1 (Fig. 3)
. Porous cup sampler nitrate N concentrations were generally higher than ETL nitrate N concentrations in 1995, similar to ETL concentrations in 1996, and lower than ETL concentrations in 1997 (Fig. 3).
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Dissolved Carbon Concentrations
There were no clear seasonal trends with dissolved C concentrations as there were with nitrate N concentrations (Fig. 4)
. The DOC concentrations remained relatively uniform, between 5 to 20 mg DOC L-1 more than 60% of the time, between 1996 and 1999 regardless of season or ecosystem. However, on two sample dates for the prairie and fertilized chisel-plowed agroecosystem, total dissolved C concentrations were very large and were nearly 100% DOC following minor precipitation events of less than 25 mm (Fig. 4). On 16 June 1996, total dissolved C concentrations were 100 mg C L-1 for one prairie lysimeter replicate and 82 and 45 mg C L-1 for each fertilized chisel-plowed agroecosystem lysimeter replicate. On 26 Oct. 1998, total dissolved C concentrations were 453 and 252 mg C L-1 for the two prairie lysimeter replicates. Duplicate TC and IC measurements performed on each leachate sample from these sample dates were consistent, giving no indication that these high concentrations were from an erroneous concentration measurement. The organic fraction of total dissolved C was generally less variable than IC concentrations (Fig. 4). At times, the chisel-plowed agroecosystem maintained consistently higher dissolved OC concentrations than the other ecosystems (Fig. 4). This may reflect higher rates of organic matter decomposition from tillage.
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Many denitrifying microorganisms are facultative anaerobes that can use nitrate (NO-3–N) as a terminal electron acceptor to usurp energy from the substrate decomposition pathway (Burford and Bremner, 1975). However, Cook and Allan (1992a) also recognized that not all DOC is sufficiently labile to support microbial activity. In addition to sufficient substrate concentration and quality and terminal electron acceptor availability (i.e., NO-3), denitrifiers require an extremely low concentration of dissolved oxygen (<0.4 mg L-1) commonly associated with saturated soils (Wilson and Bouwer, 1997). Otherwise, the amount of energy per unit substrate gained with oxygen as the terminal electron acceptor exceeds the energy per unit substrate gained with nitrate as the terminal electron acceptor, and facultative anaerobic denitrifiers will reduce oxygen in O2 instead of N in NO-3.
In terms of dissolved C, subsurface denitrification potential exists for all three ecosystems. Dissolved C concentrations are similar among the prairie and fertilized agroecosystems and would be sufficient to support facultative anaerobic denitrifying activity (W. Hickey, personal communication, 1998). In terms of available N, the very low concentrations of nitrate N in the prairie would limit the prairie's overall denitrifying capacity. However, both dissolved C and available N exist in sufficiently high concentrations that, during periods of low oxygen concentration, both fertilized agroecosystems would possess significant subsurface denitrification potential. Higher soil water contents due to less surface evaporation and increased denitrifying bacteria populations in no-tillage soils, compared with conventionally tilled soils, might suggest that denitrification losses would be larger in no-tillage agroecosystems compared with conventionally tilled agroecosystems (Doran, 1980; Rice and Smith, 1982).
Denitrification potential can be estimated from OC and NO-3–N concentrations in lysimeter leachate solutions. In the dissimilatory half-reaction of an OC substrate, the C has an oxidation state assumed to be zero with four associated electrons (e). The OC substrate would be oxidized to CO2, with a resulting C oxidation state of +IV with zero associated electrons, releasing four e in the process (Harris and Arnold, 1995). In the nitrate N reduction half-reaction, the N in NO-3 has an oxidation state of +V and has zero e. The reduction reaction of NO-3 to 1/2N2O consumes four e, leaving the N in N2O with an oxidation state of +I with eight associated e (Harris and Arnold, 1995). Coupling these two half-reactions and keeping track of the electrons transferred provides a means to calculate how much nitrate N could potentially be reduced to N2O if all of the OC collected in lysimeter leachate solutions is assumed available for denitrifying microorganisms under the proper (i.e., anaerobic) conditions. Table 4 provides a monthly summary of the potential denitrification that could occur based on 100% availability of OC in lysimeter leachate solutions.
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Nitrogen Leaching Losses
Inorganic N leaching losses (NO-3–N + NH+4–N) during the 4-yr period totaled 0.62 (SE = ±0.09) kg ha-1 from the prairie and 179 (SE = ±16.0) and 201 (SE = ±44.8) kg ha-1 from the fertilized chisel-plowed and no-tillage maize agroecosystems, respectively (Fig. 5)
. Average nitrate N fraction of total leached inorganic N was 0.63 for the prairie site and 0.95 for both agroecosystems (Fig. 5). Considering that the fertilized agroecosystems received 190 kg N ha-1 as fertilizer every year for 4 yr, the leaching losses recorded represent 24 and 26% of the total commercial fertilizer applied during the study to the chisel-plowed and no-tillage agroecosystem, respectively.
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Carbon Leaching Losses
Cumulative total dissolved C leaching losses were greater for the fertilized maize agroecosystems than for the prairie, but OC leaching losses were comparable for all ecosystems (Fig. 7) . Mean TC leaching losses were 119 (SE = ±4.0), 435 (SE = ±141), and 502 (SE = ±118) kg ha-1 for the prairie, no-tillage, and chisel-plowed ecosystems. Mean OC leaching losses were 68 (SE = ±11), 127 (SE = ±33), and 174 (SE = ±16) kg ha-1 for the prairie, no-tillage, and chisel-plowed ecosystems, respectively (Fig. 7). Although IC concentrations fluctuated more than OC concentrations, approximately one-third of the total dissolved C was organic.
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CONCLUSIONS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The silt-loam to silty-clay-loam soils of this study are well drained and the opportunity for saturated soil conditions to limit the supply of O2 for sufficient lengths of time to allow denitrification to occur is generally unlikely. However, at depth in the agroecosystems, sufficient nitrate N and DOC concentrations existed to support anaerobic denitrifiers and, conceivably, subsoil denitrification during the winter and spring months when nearly saturated soil conditions exist for some time. Nonetheless, the potential for denitrification was limited by the supply of DOC in the agroecosystems and limited by the supply of nitrate N in the prairie. Based on available OC and nitrate N, the maximum contribution of denitrification below the root zone in the restored prairie and agroecosystems was less than 25% of the total amount of leached nitrate N, and probably much smaller because of the presence of O2 and the likelihood that not all the DOC was available.
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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REFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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soil storage), and evapotranspiration (ET) between 1 April and 3 November
