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Journal of Environmental Quality 31:524-532 (2002)
© 2002 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America

TECHNICAL REPORT
Atmospheric Pollutants and Trace Gases

Nitrous Oxide and Ammonia Fluxes in a Soybean Field Irrigated with Swine Effluent

R. R. Sharpe* and L. A. Harper

Southern Piedmont Conservation Research Center, USDA-ARS, 1420 Experiment Station Road, Watkinsville, GA 30677

* Corresponding author (rsharpe{at}arches.uga.edu)

Received for publication February 2, 2001.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 REFERENCES
 
In the United States, swine (Sus scrofa) operations produce more than 14 Tg of manure each year. About 30% of this manure is stored in anaerobic lagoons before application to land. While land application of manure supplies nutrients for crop production, it may lead to gaseous emissions of ammonia (NH3) and nitrous oxide (N2O). Our objectives were to quantify gaseous fluxes of NH3 and N2O from effluent applications under field conditions. Three applications of swine effluent were applied to soybean [Glycine max (L.) Merr. ‘Brim’] and gaseous fluxes were determined from gas concentration profiles and the flux-gradient gas transport technique. About 12% of ammonium (NH4–N) in the effluent was lost through drift or secondary volatilization of NH3 during irrigation. An additional 23% was volatilized within 48 h of application. Under conditions of low windspeed and with the wind blowing from the lagoon to the field, atmospheric concentrations of NH3 increased and the crop absorbed NH3 at the rate of 1.2 kg NH3 ha-1 d-1, which was 22 to 33% of the NH3 emitted from the lagoon during these periods. Nitrous oxide emissions were low before effluent applications (0.016 g N2O–N ha-1 d-1) and increased to 25 to 38 g N2O–N ha-1 d-1 after irrigation. Total N2O emissions during the measurement period were 4.1 kg N2O–N ha-1, which was about 1.5% of total N applied. The large losses of NH3 and N2O illustrate the difficulty of basing effluent irrigation schedules on N concentrations and that NH3 emissions can significantly contribute to N enrichment of the environment.

Abbreviations: DOY, day of year


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 REFERENCES
 
IN THE UNITED STATES, swine operations produce more than 14 Tg of manure each year (Sweeten, 1992) and much of the manure is generated in relatively small geographic areas. Land application is the preferred method of manure management but high concentrations of animals may lead to excessive applications. Application of organic waste materials on agricultural land has received considerable attention in recent years because of potential environmental problems such as water quality degradation, air pollution through N gas emissions, odors, and dispersal of pathogenic organisms (Dosch and Gutser, 1996; Edwards et al., 1996; Paul and Zebarth, 1997).

Gaseous emissions of N from waste applications generally occur through volatilization of NH3 and loss of N2 and N2O through nitrification–denitrification processes. These emissions are affected by waste characteristics, method of application, climatic conditions, and chemical and physical soil properties (Misselbrook et al., 1996; Lowrance et al., 1998). As much as 13% of applied NH4 can be volatilized during irrigation before the effluent reaches the soil, with an additional 10 to 70% lost within 48 h after irrigation (Sharpe and Harper, 1997; Schilke-Gartley and Sims, 1993). Large losses of NH3 can result in insufficient N availability for plant uptake or environmental degradation. Ammonia is the primary neutralizing agent for acid gases in the atmosphere and is a common component of atmospheric aerosols. Volatilization of NH3 may result in eutrophication of natural ecosystems. Natural ecosystems are thought to be net sinks for NH3 (Denmead et al., 1976; Hutchinson et al., 1972; Van Hove et al., 1987) and cropping systems have shown significant absorption capacity (Harper et al., 1987). Atmospheric NH3 from agricultural sources has been implicated in forest decline (McLeod et al., 1990; Nihlgard, 1985) and species changes in European heathlands (Van Hove et al., 1987).

Nitrous oxide is a radiatively active trace gas that contributes to global warming and to the destruction of atmospheric ozone when it is converted to nitric oxide (Crutzen, 1981). Anthropogenic sources account for about 41% of total N2O emissions, but the strength of individual sources is uncertain (Intergovernmental Panel on Climate Change, 1996). Consequently, better source strength estimates are required for all systems. Atmospheric concentrations of N2O are increasing at about 1.5 µg m-3 yr-1 and stabilization of N2O concentrations at current levels would involve reductions in anthropogenic emissions of more than 50% (Sanhueza and Zhou, 1996). The primary biogenic sources of N2O are nitrification and denitrification of soil N (Knowles, 1982; Poth and Focht, 1985). Nitrous oxide emissions are favored by low oxygen (O2) concentrations, high soil organic C, and NO3 (Payne, 1981).

Several studies have shown that land application of animal waste increases both NH3 (Lockyer et al., 1989; Whitehead and Raistrick, 1992) and N2O emissions (Cabrera et al., 1994; Stevens and Cornforth, 1974; Egginton and Smith, 1986). These studies dealt with cattle and poultry manure or slurry applications and there is little information available concerning NH3 and N2O emissions associated with application of waste lagoon effluent. Our objectives were to quantify gaseous losses of NH3 and N2O after effluent application with a traveling gun irrigation system and relate N2O emission rates to soil water content and soil temperature.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 REFERENCES
 
This research was conducted on a 1200 sow, farrow-to-finish swine farm located in the Coastal Plains of North Carolina. The waste management system was a "flush" type with recycled water from the lagoon used to flush the pits beneath slatted floors. There were two pits beneath each house. The pits were flushed in a 4-h cycle: one pit being was flushed and 4 h later a second pit was flushed. The waste from all the houses was flushed into a single 2.7-ha anaerobic lagoon. Effluent from the lagoon was applied to the field by a traveling big gun sprinkler system. The experimental site was an 11-ha field with sandy loam soils of the Norfolk (fine-loamy, kaolinitic, thermic Typic Kandiudult)–Rains (fine-loamy, siliceous, semiactive, thermic Typic Paleaquult)–Goldsboro (fine-loamy, siliceous, subactive, thermic Aquic Paleudult) series (Brandon, 1986). Soil characteristics are shown in Table 1. Soybean was planted in 18-cm rows on 17 July 1997 and the only N fertilizer applied was through three effluent applications during the growing season.


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  		Table 1. Soil properties prior to irrigation.

 
Effluent was applied through the irrigation system three times during the summer growing season. Quantities applied, total N, and inorganic N contents for each application are shown in Table 2. Effluent pH at time of applications was 8.0. Most of the N in the effluent was NH4–N with only small amounts of NO3–N and organic N. Effluent application was about twice as large in the first irrigation as in the last two. Nitrogen applied in the three irrigation events was 144.6, 59.1, and 70.8 kg N ha-1, respectively. Total applications during the study were 274.6 kg ha-1 of total N and 239.2 kg ha-1 of NH4–N. Initial effluent application was relatively late in the growing season (day of year [DOY] 237), thus there was limited opportunity for plant uptake.


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  		Table 2. Quantity of effluent, effluent concentrations, and total N, NH3–N, and NO3–N applied in each irrigation. DOY = day of year.

 
Soil N (0- to 0.1- and 0.1- to 0.2-m depths) measurements were made through the season. Soil samples were taken from 10 areas in the field and each sample consisted of four cores. Soil temperatures (0.05-, 0.10-, and 0.15-m depths) were measured every 15 min using copper-constantan thermocouples. Total plant N was measured using Kjeldahl digestion, and soil N (NH4 and NO3) was determined colorimetrically in a 2 M KCl soil extract. Losses of effluent NH3 during the irrigation events were calculated by collecting effluent samples in glass containers and comparing N concentrations in the catchment samples to samples collected directly from the lagoon close to the irrigation pipe.

Micrometeorological instrumentation was located near the center of the field to obtain a minimum fetch of 100:1 (upwind canopy distance to measurement height) in all directions for wind profile development. Micrometeorological data and atmospheric NH3 and N2O concentrations were determined throughout the season starting on DOY 200 (20 July) and ending at harvest on DOY 250. Windspeed (sensitive cup anemometers; Model 106-LED-DC, Thornthwaite & Assoc., Pittsgrove, NJ) and air temperature (aspirated thermocouples; Model ASPTC, Campbell Scientific, Logan, UT) profiles were measured at six heights (plant height plus 0.2, 0.4, 0.6, 0.8, 1.6, and 2.7 m). Atmospheric concentrations of NH3 and N2O were measured at plant height plus 0.6 and 1.6 m using tunable diode laser spectroscopy (TDL). The TDL (Model TGA100, Campbell Scientific) technique is based on infrared spectroscopy (Dias et al., 1996; Edwards et al., 1994). The diode laser is mounted in a liquid nitrogen cooled dewar and a heater in the dewar gives precise control of the laser in the 78- to 110-K region. The laser was operated in the IR spectral region between 3000 and 3025 cm-1. The sample and reference cells are 1.54 m and 0.05 m long, respectively. Both sample and reference detectors were Peltier-cooled mercury–cadmium–tellurium IR detectors (EG&G Judson, Montgomeryville, PA). The instrument had a short-term (sample period) total noise of about 10 ppbv. The TDL's electronics were integrated with a PC for software control of the digital signal processing, laser function, real-time display of laser-operating functions, and for data storage. Atmospheric gas concentrations were measured 10 times per second. Delta N2O concentrations (concentration differences between the two sampling heights) were calculated every minute, and delta NH3 concentrations were calculated every two minutes. Delta concentrations were averaged over 30-min periods for use in the flux gradient technique. Mylar balloons were used to transport references gases to the field to calibrate the laser spectrometers.

Ammonia and N2O gas flux densities above the canopy were determined during the measurement seasons using gas concentrations and the flux gradient gas transport technique using the momentum balance transport coefficient. The relationship for gas flux density is:

[1]
where F = gas flux densities (kggas ha-1 d-1), n = atmospheric gas concentration (µg gas m-3), z = gradient measurement height (cm), and Kmb = momentum balance transport coefficient. A negative F value indicates absorption and a positive value emissions from the field. The momentum coefficient (Kmb) is determined from the relationship:

[2]
where k = von Karman constant, µ = windspeed (cm s-1) at upper height (µ2) and lower height (µ1), zd {cong} 0 (cm) for a water surface, and {Psi} = stability correction factor (Dyer and Hicks, 1970). Errors associated with the flux gradient gas transport technique have been discussed by Harper (1988) and Denmead and Raupach (1993). Data were analyzed using the stepwise regression procedures of SAS (SAS Institute, 1991).


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
REFERENCES
 
Nitrous oxide flux measurements were taken during three periods prior to effluent applications. In general, N2O fluxes were small (<=0.016 kg ha-1 d-1) before effluent applications except for DOY 206.5 to 207.5 (Fig. 1A) , when average N2O flux was about 0.25 kg ha-1 d-1. The increased emissions of N2O during this period were probably due to saturated soil conditions caused by 60 mm of rain on DOY 204 and 205 (Fig. 1C). Assuming an N2O emission rate of 0.016 kg ha-1 d-1 during periods when fluxes were not measured, total emission would have been about 0.53 kg N ha-1 from planting until the first irrigation.



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  		Fig. 1. Nitrous oxide flux and rainfall from a soybean field before irrigation with swine waste effluent, day of year (DOY) 200 to 236.

 
Nitrous oxide flux rates were substantially greater after effluent application than before (Fig. 2A) . Average N2O emission rate on DOY 236 was about 0.01 kg N2O ha-1 d-1. After Irr 1 on DOY 237, average N2O emissions increased to 0.5 kg N2O ha-1 d-1 within 6 to 8 h. Peak emission rates, however, were not observed until after about 30 h (DOY 238). Emissions rates decreased after DOY 238 but remained elevated, relative to pre-irrigation rates, until Irr 2. Peak emission rates were observed within 8 h after Irr 2 and 3. Other studies have shown highly variable time courses for peak response following animal waste application. Sharpe and Harper (1997) and Whalen et al. (2000) reported maximum responses within several hours of swine effluent applications. Studies with cow and poultry manure have shown peak emissions from 1 to 7 d after application (Cabrera et al., 1994; Egginton and Smith, 1986; Lessard et al., 1996). Nitrous oxide is a product of both nitrification and denitrification of soil N. Highest rates of production would occur at low O2 concentrations, which limit use of O2 as an electron acceptor in the nitrification process (Klemedtsson et al., 1988) and inhibit reduction of N2O to N2 in the denitrification process (Focht, 1974). A rapid release of N2O would be expected from the application of a liquid organic fertilizer such as swine effluent. Swine effluent provides a large source of NH4 for nitrification, decreased O2 concentration by increasing soil water, and a soluble C source for microbial growth. Nitrous oxide emissions after the first two irrigations averaged about 0.58 and 0.78 kg ha-1 d-1. Emissions after Irr 3 were smaller, about 0.18 kg N2O ha-1 d-1. The lower response may have been due to a sharp decrease in temperature since soil water content was similar on DOY 242 and 246 (Fig. 1C). Soil temperature decreased about 5°C (Fig 2C) and air temperature decreased about 15°C (Fig. 3) from DOY 246 to 247. From DOY 237 through 250, total N2O–N emission was about 4.1 kg N ha-1 or about 1.5% of the applied N.



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  		Fig. 2. Time course for (A) N2O flux and irrigation (Irr) with effluent, (B) soil moisture and rainfall, and (C) soil temperature from a soybean field.

 


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  		Fig. 3. Time course for NH3 flux and irrigation (Irr) with effluent from day of year (DOY) 237 to 250.

 
Before irrigation there was little or no NH3 flux from the soil–crop system (Fig. 1B). There were, however, three short periods during which atmospheric NH3 was absorbed by the crop. Each period of absorption was associated with periods of high atmospheric NH3 concentrations. Two of these periods are shown in Fig. 4 and 5 . On DOY 212.5 to 213.5, the differential range in NH3 concentrations between the two measurement heights increased to 8 to 32 µg NH3 m-3 (Fig. 4B). During this time, NH3 absorption rate averaged about 1.17 kg NH3 ha-1 d-1 (Fig. 4A). Similarly, on DOY 218 to 219 (Fig. 5B), the delta NH3 concentration increased to a differential range of 15 to 25 µg m-3 with corresponding periods of short-term NH3 absorption by the crop (Fig. 5B). During these periods atmospheric concentrations were probably greater than the NH3 compensation point (Farquhar et al., 1980) inducing NH3 absorption. Numerous studies have shown significant absorption of atmospheric NH3 by plants (Hutchinson et al., 1972; Harper and Sharpe, 1995; Parton et al., 1988).



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  		Fig. 4. Time course for (A) NH3 flux, (B) differential (delta) NH3 concentrations at 60- and 160-cm heights, and (C) wind direction and windspeed before effluent irrigation from day of year (DOY) 210 to 213.

 


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  		Fig. 5. Time course for (A) NH3 flux, (B) differential (delta) NH3 concentrations at 60- and 160-cm heights, and (C) wind direction and windspeed before effluent irrigation from day of year (DOY) 217 to 221.

 
The experimental field was southwest of the 2.7-ha waste handling lagoon supplying the effluent for irrigation and the measurement site was located about 325 m from the lagoon. Using the standard compass convention of north being 0° and east being 90°, then wind directions of 0 to 60° would indicate wind blowing from the lagoon toward the measurement site. Thus, NH3 volatilized from the lagoon was blown toward the experimental site when winds were less than 60°. From DOY 210 through 220, there were several extended periods when the wind was blowing from the lagoon toward the experimental site (Fig. 4C and 5C). However, during most of these periods windspeed was greater than 100 cm s-1 and atmospheric NH3 concentrations were not affected (Fig. 4B and 5B) probably due to turbulence and mixing of NH3 from the lagoon with the bulk air mass. During periods of low windspeed and wind direction from the lagoon [DOY 212 to 213 (Fig. 4C); DOY 218 to 219 (Fig. 5C)], atmospheric concentration increased and the plants absorbed NH3, as indicated by the negative flux rates. This forced absorption of N can occur in natural systems such as a forest (Nihlgard, 1985). There is no indication for N saturation in leaves (Van der Eerden et al., 1992) and uptake is probably linearly related to atmospheric NH3 concentrations. During these periods about 1.2 kg NH3 ha-1 d-1 was absorbed by the crop. If the entire field (11 ha) absorbed NH3 at a similar rate, then a total of about 39.6 kg of NH3 would be absorbed over the 3-d period. The lagoon emitted 15 to 22 kg NH3 ha-1 d-1 during the summer of 1997 (Harper et al., 2002). Over a 3-d period the 2.7-ha lagoon would emit 120 to 178 kg NH3. Thus, 22 to 33% of the NH3 emitted from the lagoon was absorbed by the crop during these short-term periods.

Ammonia emissions were large immediately after irrigation with effluent, but decreasing to background levels within 24 to 48 h (Fig. 3). There was a large pulse of NH3 within 4 h after Irr 1, but the largest emissions were observed 24 to 30 h after irrigation. With Irr 2 and 3, the maximum emission rates were directly following irrigation. Previous research with waste effluent application has shown maximum emission rates directly following applications as in Irr 2 and 3 in this study (Sharpe and Harper, 1997; Smith et al., 1996). The slower maximum response with Irr 1 may have been due to the quantity of N applied and to the small quantity of precipitation on DOY 238. The larger quantity of applied N would have extended total time of NH3 volatilization and evaporation of the rain may have resulted in a pulse of NH3 emissions. Total N applied in Irr 2 and 3 and in the Sharpe and Harper (1997) and Smith et al. (1996) studies was less than half that applied with Irr 1. The larger maximum flux rates after Irr 1 and 3 (125 and 111 kg NH3 ha-1 d-1, respectively) than for Irr 2 (59 kg NH3 ha-1 d-1) were probably due to decreased windspeed. After Irr 1 and 3, windspeeds ranged from 150 to 200 cm s-1 while windspeeds after Irr 2 were 15 to 100 cm s-1. Previous research has shown that the NH3 flux is dependent of both air temp and windspeed (Fenn and Kissel, 1974; Bremner and Mulvaney, 1978).

During the irrigation events, about 12% of the NH4–N in the effluent was volatilized as NH3 before reaching the crop or soil surface. Total emissions after the three irrigations were 42, 11, and 11.1 kg NH3 ha-1, respectively. These losses ranged from 30% of applied NH4 for Irr 1 to 15% for Irr 3. More than twice as much effluent was applied during Irr 1 than Irr 2 or 3 (Table 1) and this may have resulted in pooling of effluent on the crop and soil surface allowing for greater NH3 volatilization. In total, about 35.4 and 1.5% of the effluent N was lost to the atmosphere as NH3–N and N2O–N, respectively.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 REFERENCES
 




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