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a Soil Plant and Ecological Sciences Division, P.O. Box 84, Lincoln Univ., Canterbury, New Zealand
b Dep. of Agronomy and Soils, 236 Funchess Hall, Auburn University, AL 36849-5412
c CSIRO Plant Industry, G.P.O. Box 1600, Canberra, ACT, Australia
d Animal and Food Sciences Division, P.O. Box 84, Lincoln Univ., Canterbury, New Zealand
* Corresponding author (SvenG.Sommer{at}agrsci.dk)
Received for publication June 19, 2001.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: IPCC, Intergovernmental Panel on Climate Change • TAN, total ammoniacal nitrogen (NH3 + NH4)
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Applying manure or slurries to agricultural land can lead to ground water contamination by nitrate (NO3) after nitrification of the ammonium nitrogen (NH4–N) present, and emission of NH3 (European Centre for Ecotoxicology and Toxicology of Chemicals, 1994), CH4 (Chadwick and Pain, 1997) and N2O (Jarvis et al., 1994), all of which contribute to climate change. Methane and N2O are greenhouse gases that contribute directly to climate change (Intergovernmental Panel on Climate Change, 1996). Ammonia, after deposition on land surfaces and water bodies and nitrification, acts as a secondary source of N2O (Mosier et al., 1998), and may also decrease the capacity of soils to absorb CH4 and act as a sink for this gas (Mosier et al., 1991, 1996). However, the magnitude of these direct effects and interactions is not known with certainty.
The Intergovernmental Panel on Climate Change (IPCC) has been coordinating the development of inventory methodologies for greenhouse gases (Intergovernmental Panel on Climate Change, 1997) and this has revealed the lack of information required to define appropriate factors for CH4 and N2O emission from different sources. For example, there is considerable doubt as to which emission factor should be used for calculating N2O emission from redeposited NH3 and the various organic forms of N used in agriculture (Mosier et al., 1998). In addition, the Intergovernmental Panel on Climate Change (1997) protocol assumes there is no CH4 emission from animal slurry applied in the field, whereas Chadwick and Pain (1997) and Sommer et al. (1996) have shown in laboratory studies that CH4 is emitted following pig and dairy slurry application to soil.
Developing strategies to reduce CH4 and N2O emissions from agricultural land requires an understanding of the production, emission, and consumption of these gases. In particular, studies are required on the interactions between these gases because it is of concern that strategies to reduce emission of one gas may increase emission of the other (Intergovernmental Panel on Climate Change, 1996). Consequently, it is necessary to measure emission of all three gases simultaneously in the system under study. This paper reports a study on NH3, CH4, and N2O emission from a pasture under conditions recommended by the Intergovernmental Panel on Climate Change (1997), after pig (Sus scrofa) slurry application, to assist in defining suitable factors to calculate CH4 and N2O emissions. Micrometeorological methods are available for accurately measuring NH3 volatilization in the field (Denmead et al., 1974; Denmead, 1983; Wilson et al., 1983; Schjoerring et al., 1992; Sherlock et al., 1995; Wood et al., 2000). Various chamber procedures for NH3 have also been implemented in the field, but these can introduce experimental artifacts that alter the dynamics and extent of NH3 volatilization (Freney et al., 1983; Black et al., 1985). Ammonia is very reactive with water compared with CH4 and N2O (Liss and Slater, 1974). The elevated NH3 concentration in a static chamber will, consequently, reduce NH3 emission from the soil covered by the chamber. The NH3 emission, therefore, was measured with a micrometeorological mass balance technique (Wilson et al., 1983) that does not affect the NH3 concentration above the soil amended with slurry. In contrast with NH3, static chambers are the preferred current method for measuring surface fluxes of both CH4 and N2O since these gases are less reactive with water and are much less affected by increases in chamber headspace concentrations (Mosier et al., 1991). Also in contrast to NH3, it is generally very difficult to detect the small concentration gradients of N2O and CH4 above the soil surface when these gases are being emitted. This is necessary for implementing micrometeorological techniques. Nevertheless, when comparisons between methods have been possible, little if any difference is observed between N2O emissions measured with a micrometeorological technique and static chambers (Fowler et al., 1997). Consequently, we decided to use static chambers to measure N2O and CH4 fluxes in this study.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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Ammonia Emission
Measurements of NH3 emission from the slurry-treated plot commenced immediately after slurry application on 9 August and continued for 10 d until NH3 emission returned to background levels. The fluxes were determined by the mass balance micrometeorological method described in Sherlock et al. (1995). The mass balance method equates the average surface flux density of NH3 from the circular plot amended with pig slurry to the difference between the integrated horizontal flux at a known downwind distance and that at the upwind edge (Wilson et al., 1983). With passive Leuning samplers (Leuning et al., 1985) mounted at several heights z (m), on masts placed at the upwind edge and at the center of the plot (radius 20 m), the net horizontal flux (F, g m-2 s-1) is derived from:
 | [1] |
is the mean horizontal flux (u is wind speed, m s-1 and c is NH3 concentration, g m-3) measured by each sampler at the downwind (dw) or upwind (uw) edge of the treated area. According to Leuning et al. (1985), is derived from:
 | [2] |
As a backup, NH3 flux was also determined with a second design of passive sampler (Ferm tubes) consisting of two parallel sampler units, each of which had three glass tubes (inner diameter 7 mm and lengths of 100, 100, and 23 mm) joined by silicone tubing (Schjoerring et al., 1992). The interior surface of the two 100-mm tubes was coated with oxalic acid for approximately 70 mm of the tube length. Glued to the end of the 23-mm tube was a stainless steel nozzle with a central hole of 1-mm diameter.
After exposure, the tubes were disconnected and the two 100-mm tubes analyzed separately. The coating was dissolved in 0.003 L deionized water and the NH4–N content determined by the indophenol blue colorimetric method with a flow injection analyzer (Tecator [Höganäs, Sweden] FIAstar).
The Ferm tube flux samplers were mounted on four masts at 0.2, 0.6, 1.2, 2.4, and 3.6 m above ground level, positioned at 0, 90, 180, and 270° angles around the circumference of the circular plot. The samplers continuously integrate the horizontal ammonia flux at the various heights. After analysis of the ammonia content in the tubes facing the NH3 source (exposed tubes) and surroundings (background tubes), the net horizontal flux can be calculated for each measurement height. Knowing the net horizontal ammonia flux, the vertical flux from the experimental plot was obtained by applying mass balance equations.
The average horizontal NH3 flux (Fhm, µg NH3–N m-2 s-1) through two glass tubes facing in the same direction at each measuring height (h) on each mast (m) either away from (background tubes) or toward (exposed tubes) the NH3 source was calculated by the following equation:
 | [3] |
The horizontal net flux (Fnet,h, µg NH3–N m-2 s-1) at each height was calculated by the equation (Schjoerring et al., 1992):
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The vertical NH3 flux from the plot (Fv, µg NH3–N m-2 s-1) was calculated by stepwise summation of the horizontal net flux over the height intervals represented by the flux samplers (Schjoerring et al., 1992):
 | [5] |
h is the height interval represented by the flux samplers and x (20 m) is the diameter of the experimental plot.
Methane and Nitrous Oxide Emissions
Measurements of CH4 and N2O emission with a closed chamber technique (Hutchinson and Mosier, 1981) commenced immediately following slurry application on 9 August and continued until 7 November. Emission was measured from microplots formed by pushing steel cylinders (24-cm diameter) into the soil. Five cylinders were placed in the slurry-treated plot and four cylinders were located on untreated soil upwind of the plot. Cylindrical gas-tight lids, having a headspace height of 10 cm and fitted with rubber septa for sampling, were attached to the steel cylinders during gas emission measurements, but were left open between measurements. Changes in headspace CH4 and N2O concentrations were used for calculating gas fluxes. Gas samples were collected with 50-mL syringes at 0, 10, and 20 min after lid closure. Samples were collected twice a day during the first 5 d, twice every second day during the period 5 to 11 d, once a day during the period 11 to 60 d, and once every second day during 60 to 90 d. Emission was measured at 1000 and 1600 h when two measurements were taken per day and at 1200 h on the other measurement days.
The gas samples were injected via a 10-port sampling valve into a carrier stream of N2, and a four port-switching valve directed the carrier gas stream to either of two gas chromatographs for CH4 or N2O determination. Methane was determined on an SRI (Torrance, CA) instrument equipped with a flame ionization detector. Nitrous oxide was determined on a Varian (Walnut Creek, CA) Aerograph Series 2800 equipped with a 63Ni electron-capture detector (Pye-Unicam, Cambridge, UK) and a stainless steel column (4-m length, 3-mm internal diameter) packed with Porapak Q (80/100 mesh) (Alltech Associates, Deerfield, IL). Detector and column temperatures were 350 and 20°C, respectively.
In general, CH4 and N2O fluxes were calculated with the logarithmic equation described by Hutchinson and Mosier (1981). At times when gas fluxes were small, they were determined with a linear equation.
Soil and Manure Analysis
Pig slurry was sampled from the slurry spreader and stored at 4°C until analysis. Pig slurry dry matter was determined gravimetrically, slurry density by weighing an aliquot of 20 mL in a graduated flask, total N and total C with a mass spectrometer (Tracermass; Europa Scientific, Crewe, UK), and NH4 and NO3 on a flow injection analyzer (Tecator FIAstar).
Soil samples (0–5 cm) were collected immediately before slurry application, three times per day for the first 5 d after application, once per day from 6 to 10 d, three times per week during 10 to 20 d, twice per week during 20 to 60 d, and once every second week during 60 to 90 d. Five replicate samples were taken on each occasion. Soil samples were well mixed in a field laboratory at the site and stored at temperatures less than 4°C until analysis. The pH of soil and canopy surfaces was measured with a portable pH meter and a flat surface pH electrode. Soil NO3, TAN, and volatile fatty acids were extracted with 2 M KCl (2 KCl to 1 soil, w/w). These suspensions were shaken for 30 min and filtered through Whatman (Maidstone, UK) #42 filter paper. Ten milliliters of each sample was used to determine NO3–N and TAN, with standard colorimetric techniques (Keeney and Nelson, 1982), on a FIAstar 5010 analyzer (Tecator). Five milliliters of the extract was used to determine volatile fatty acids. To each aliquot, 0.5 mL of 0.3 M oxalic acid and 0.5 mL of pivalic acid (25 000 mg L-1 in 0.3 M oxalic acid) as an internal standard were added. The mixture was centrifuged at 16 000 rpm for 10 min, the supernatant was passed through a 0.45-µm filter, and the resulting filtrates were stored at -18°C until analysis. The C2–C5 fatty acids in the filtrates were determined on a Shimadzu (Kyoto, Japan) GC-7A gas chromatograph equipped with a flame ionization detector (140°C) and a glass column (1.8-m length, 4-mm internal diameter) packed with Tenax GC (60/80) and FAL-M (Buchem BV, Apeldoorn, the Netherlands). Injector and oven temperatures were maintained at 210 and 136°C, respectively. The carrier gas was N2 adjusted to a flow rate of 60 mL min-1.
Calculations and Statistics
It has been shown that NH3 loss from both flooded and nonflooded fields is determined by two main variables, the equilibrium NH3 concentration and the wind speed (Freney et al., 1985; De Datta et al., 1989; Sherlock et al., 1995). Ammonia is lost from solution at the soil's surface when the NH3 gas concentration in equilibrium with the solution is greater than that of the atmosphere. Increasing wind speed increases the volatilization rate by promoting more rapid NH3 transport away from the surface. Immediately after applying animal slurry, urine, or fertilizer, the equilibrium NH3 concentration is much greater than that in the ambient atmosphere, thus the latter can be ignored and the relationship between the main variables is described by:
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 | [7] |
 | [8] |
 | [9] |
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The effect of environment and soil composition on N2O emission was analyzed by means of linear regression analysis (Stepwise; SAS Institute, 1989).
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RESULTS |
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Pig slurry application decreased the NO3 concentration in the 0- to 5-cm soil layer from 2.9 to 0.2 mg N kg-1 soil. There was little change in the NO3 concentration of the top 5 cm of soil until Day 10 when approximately 3 mm rain fell. After 20 mm rain fell on Day 20, the NO3 concentration in the surface soil decreased from 4 to 1.2 mg N kg-1 soil and then increased rapidly to approximately 11.4 mg N kg-1 soil (Fig. 1B and 2B). The NO3 concentration fell again after rain on Days 30 and 50 and returned to approximately 12 mg N kg-1 soil on each occasion. It finally decreased to the background value around Day 90 (Fig. 2B).
The concentration of volatile fatty acids in the soil solution declined from approximately 15 mM kg-1 soil just after slurry application to 4 mM kg-1 soil within 1.5 d (Fig. 2C). It then slowly decreased over the next 3 d to 0.8 mM kg-1 soil, which was not significantly different from the background concentration of 0.6 mM kg-1 soil.
Ammonia Emission
In the first 2.5-h period after slurry application, NH3 was lost at a rate of 4.7 kg N ha-1 h-1. In the next 4-h period, NH3 was lost at about half the initial rate (2.2 kg N ha-1 h-1) and overnight the rate dropped to 0.26 kg N ha-1 h-1 (Fig. 3A)
. After Day 1, NH3 volatilization rates declined markedly to average daytime and nighttime rates over the next 2 d of 0.57 and 0.11 kg N ha-1 h-1, respectively (Fig. 3A). Ammonia was lost at faster rates in the morning session (900–1300 h) than in the afternoon session (1300–1700 h) and the overnight rates were always low (Fig. 3A). As a result of the fast rates of loss, 24.8 kg N ha-1 was lost on Day 1 (Fig. 3B; equivalent to 6.8% of the applied N or 9.8% of the TAN). During the next two days, 7.8 and 4.8 kg N ha-1 were lost, and thereafter until Day 9, NH3 was lost at the rate of approximately 2.5 kg N ha-1 d-1. Little NH3 was volatilized after Day 10. The total NH3 emission was 57 kg N ha-1 (Table 2), corresponding to 15.5% of the total nitrogen applied to the field in pig slurry or 22.5% of its original TAN content. It can be seen from Fig. 3B that the results for NH3 loss obtained with the Ferm tubes (Schjoerring et al., 1992) were similar to those obtained with the passive samplers of Leuning et al. (1985).
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 | [11] |
Methane
Methane emission commenced at the very fast rate of 39.33 g C ha-1 h-1 immediately after pig slurry application and decreased to 10 g C ha-1 h-1 within 6 h (Fig. 4A)
. The CH4 emission rate then decreased, but emission continued at a lower rate for approximately 7 d. Thereafter, CH4 flux measurements indicated net uptake of atmospheric CH4 by the treated plot (Table 2). The control plot absorbed CH4 from the atmosphere at the rate of 9.1 µg C m-2 h-1 throughout the study period.
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 | [12] |
Nitrous Oxide
At all times emissions from the slurry-treated plot were significantly greater than those from the untreated control plots (approximately 0.08 g N ha-1 h-1; Fig. 5)
. Nitrous oxide emissions from the slurry-amended plot were low (1 g N ha-1 h-1) for the first 14 d after application, when the NO3 concentration in the 0- to 5-cm surface soil layer was low. In the 11 d that followed, the NO3 concentration in the surface soil increased (Fig. 2B), and N2O emissions increased to a peak of 7.5 g N ha-1 h-1. The emissions then declined to approximately 1 g N ha-1 h-1 after 20 d of rain-free weather (Fig. 1B). Nitrous oxide fluxes then increased in response to rain to create a second peak flux of 15.8 g N ha-1 h-1, 67 d after slurry application. The subsequent decline in daily N2O flux over the next 23 d to background values coincided with low rainfall and low mineral N in the soil layer (Fig. 1B and 2B).
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DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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The fastest rate of NH3 loss occurred immediately after application, and the loss rate then decreased quickly with time (Fig. 3A). In contrast, when urea or urine is added to the soil surface, little NH3 is lost initially and the maximum rate of loss does not occur until two or three days after application, and thereafter the loss rates decline slowly (Sherlock et al., 1995). The differing patterns at the beginning seem to be due to the composition of the material added. With fertilizer or urine, urea has to be hydrolyzed to produce NH3 and a source of alkalinity before any loss could occur, whereas the slurry contained a high TAN concentration and had a high pH, and thus NH3 could be lost immediately.
As noted above (Eq. [11]), a very strong relationship was found between the vertical flux density of NH3 and the product of [NH3]gas and wind speed, following pig slurry application to a pasture soil. In studies on NH3 losses following urea and urine application to unsaturated soils (Sherlock et al., 1995) the coefficients of inclination (slopes) of the comparable relationships were 0.69 x 10-4 after urea application and 0.9 x 10-4 after urine application. The slope in this current study was higher due to the different approach used to sample the soil, that is, in the study of Sherlock et al. (1995) soil was sampled by scraping off the top few millimeters of the soil surface, while in this current study soil samples were taken from the 0- to 5-cm depth. The TAN concentration will be lower when taking a 0- to 5-cm sample than in a sample from the topsoil of slurry- or urea-amended soil, and therefore the [NH3]gas x u product is lower.
The reasons for the low NH3 emission rates after the initial fast rate are uncertain. In the case of urea or urine the slow decline in loss rates may be due to acidification of the soil surface as a result of NH3 loss, leaching of urea below the soil surface by rainfall, oxidation by nitrifying organisms, reaction with the cation exchange complex, immobilization, and uptake by plants. While many of these factors would have operated to reduce NH3 loss after slurry application, none of them seems to be responsible for the rapid decline in NH3 loss rates on Day 1. The pH of the surface soil actually increased slightly during the first 3 d (Fig. 2A), there was little rainfall to leach the slurry into the soil during the first 20 d (Fig. 1B), and little NO3 was produced until after Day 20. The soil has considerable cation exchange capacity (15.4 cmol [Na+] kg-1), and NH4 would be adsorbed on the cation exchange complex in the surface soil as the slurry infiltrated. Immobilization would be expected to occur, especially as organic carbon was added in the slurry (Comfort et al., 1988; Kirchmann and Lundvall, 1993), but considerable NH4 remained in the soil until Day 40, and only 22.5% of the TAN was lost as NH3. Nitrogen uptake by plants should have been slow and small because the surface cover was small and growth slow because of the low soil and air temperatures. However, Thompson et al. (1990) studied NH3 loss from slurries in the absence of soil and concluded that the decline in NH3 loss rates was due to the formation of a crust on the slurry surface. With time, the crust becomes thicker and the slurry more viscous, thereby increasing resistances to diffusion within the slurry. The TAN in the crust surface can then become depleted due to NH3 volatilization (Sommer and Olesen, 2000).
The changes in the physical nature of the slurry would also restrict oxygen diffusion and ensure that the slurry remained anaerobic for up to 20 d, thus allowing CH4 formation in situ, and limiting NH4 nitrification (Fig. 2B). The addition of carbon in the slurry would stimulate microbial activity in the soil, rapidly consume any available oxygen present (Firestone, 1982), and guarantee that the slurry-treated surface soil remained anaerobic. The CH4 emission from the treated pasture at a rate of 39.6 g C ha-1 h-1 immediately after slurry application indicates that this CH4 was formed in the slurry pit prior to application, as there would not have been sufficient time for its formation in soil (Fig. 4A). It is noteworthy that this initial peak CH4 emission is not high in relation to the likely CO2 emissions. For example, Dendooven et al. (1998) measured average CO2 emission rates of 23 kg CO2–C ha-1 d-1 (960 g CO2–C ha-1 h-1) during the 28 d following pig slurry application (40 000 kg ha-1) to a bare soil. Further, small amounts of CH4 originating from this initial source may have diffused out from the viscous layer over the next few hours. However, the strong correlation between CH4 emission and volatile fatty acids (Fig. 4B) suggests that the CH4 emissions over the next 7 d were more likely to have arisen in the anaerobic soil layer from the metabolism of volatile fatty acids supplied in the slurry. The duration of CH4 emission was longer than that observed previously (Sommer et al., 1996). The low temperatures may have caused this during the experiment, as it is known that CH4 production is strongly related to temperature (Zeikus and Winfrey, 1976).
As discussed above, slurry addition appeared to restrict oxygen diffusion and the surface soil remained anaerobic for a lengthy period. While this limited nitrification for the first 14 d after application, it did not completely prevent it, as some NO3 was formed (Fig. 2B) and some N2O was emitted (Fig. 5). The likely sequence of reactions responsible for the small NO3 and N2O production seems to be the same as those operating in flooded soils, namely (i) NH4 in the slurry diffused to a zone in the soil containing residual oxygen, (ii) this NH4 was oxidized to NO3 by nitrifying organisms, (iii) the NO3 formed diffused back to the anaerobic zone, (iv) denitrification occurred with N2O and N2 production, and (v) the gaseous products diffused through the anaerobic layer to the atmosphere. During the diffusion through the anaerobic layer, N2O may have been further reduced to N2, resulting in little N2O emission (Patrick, 1982; Petersen et al., 1992).
The initial peak in N2O emission occurred during a period of active nitrification (Fig. 2B) and may have resulted from nitrification as well as by denitrification of the resulting NO3 (Jarvis et al., 1994; Lessard et al., 1996). However, the second peak occurred when NH4 concentrations were low (Fig. 2B), and NO3 concentrations (Fig. 2B) and soil moisture contents (Fig. 1D) were high, and thus appears to be mainly the result of denitrification. Lessard et al. (1996) have shown that both nitrification and denitrification may be sources of N2O within soils amended with manure N but that the relative contribution of the two processes could not be quantified. In this study the pool of NH4 was significantly higher than the NO3 pool during 0 to approximately 30 d after slurry application (Fig. 2B). Consequently, NH4 oxidation may have been the most prominent source of N2O emission in this period (Jarvis et al., 1994). After 30 d the NH4 pool was nearly depleted while the pool of NO3 was significantly higher. Thus, NO3 may have been the major N2O source from 30 d until the conclusion of the experiment at 90 d.
The N2O emission was slightly greater at 1600 than at 1000 h during days when emissions were measured twice daily, indicating a diurnal variation. Therefore, estimates of accumulated N2O emissions may be influenced by the sampling times selected, for example, estimated emissions would be higher from measurements performed at 1600 than at 1000 h (Brumme and Beese, 1992; Granli and Bøckman, 1994; Thornton et al., 1996). The diurnal variation was largely related to warming of soil during the day. We therefore assumed that N2O fluxes measured around noon would represent the average diurnal flux, and statistical analysis was performed with N2O emission rates and average daily climate data. Nitrous oxide emission was correlated (P < 0.15%) to the NO3 concentration in the topsoil (Table 3), confirming that most of the N2O emitted originated from NO3. Cumulative net precipitation and rain the day before a measurement had a significant positive effect on N2O fluxes (P < 0.01). This is consistent with stimulation of NO3 denitrification due to water saturation, which induces anoxic conditions in the soil. Rain on the day of measurement had a negative significant effect (P < 0.05) on N2O emission. Rain causes an immediate increase in soil water content to above field capacity in the topsoil and air-filled porosity will be very low under this condition. Consequently, N2O gas has to be transported by slow diffusion through water instead of a faster combination of diffusion and convection in soil air. This change in transport mechanism reduces the N2O flux until the soil water is again less than or equal to field capacity. Air temperature and incident solar radiation has a positive significant effect on N2O emission (Table 3). The increase with air temperature and incident solar radiation may be due to a reduced solubility of N2O in soil solution at increasing temperatures or an increase in N2O production with temperature (Müller et al., 1997).
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It is considered that animal slurries applied to soil do not constitute an important source of CH4 (Intergovernmental Panel on Climate Change, 1996), but very few studies appear to have been made on the amounts lost from pig slurry after surface application to field soils. Chadwick and Pain (1997) investigated the effect of soil and slurry type on CH4 emissions under laboratory conditions and showed that most of the CH4 emitted was derived from the slurry, that emissions declined to background levels after 48 h, and that more CH4 was emitted from pig slurry than cattle slurry when applied to a clay soil. In another laboratory study, Sommer et al. (1996) found that 30 g were emitted from a silt loam after applying the equivalent of 50 000 kg slurry per hectare. In that study the bulk of the CH4 (90%) was emitted within 4 h of slurry application. In the current study, the net CH4 emission was 0.07% of the carbon applied, and 46% of the total was lost within 6 h.
Simple integration and extrapolation of the daily emissions gives a total emission of 7.6 kg N2O-N ha-1 during the course of this study (2.1% of slurry N applied). Petersen (1999) reported accumulated N2O emissions between 0.14 and 0.64% of total N during a period of 6 to 8 weeks between slurry application and when crop uptake had depleted the inorganic N pools in the soil. Higher emissions reported in this study may be due to lower crop uptake of the mineral N applied in the manure. Based on data from long-term experiments with a variety of mineral and organic fertilizers, Mosier et al. (1998) recommend that only one factor should be used for calculating the N2O emission from different fertilizer types:
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Interactions
Aerobic soils constitute one of the most important sinks for atmospheric CH4 (Intergovernmental Panel on Climate Change, 1996). However, the capacity of soils to absorb CH4 and oxidize it to CO2 (Knowles, 1993) is reduced by agricultural practices that affect N cycling. This occurs because of the close relationships between pathways for NH3 and CH4 oxidation (Steudler et al., 1989; Mosier et al., 1991). There have been a number of studies on the effects of fertilizer N and cultivation on CH4 uptake, and the effect is greater in some soils than in others and in general NH4 addition has a greater effect than NO3 addition (e.g., Adamsen and King, 1993; Delgardo et al., 1996; Hütsch et al., 1994; Mosier et al., 1991, 1997; Willison et al., 1995). However, there have been few studies on the effect of slurry application on CH4 uptake (e.g., Hansen et al., 1993).
Two effects would be expected, a direct effect of the TAN from the slurry on the treated area and an indirect effect due to deposition of volatilized NH3 on the surrounding countryside. When NH3 is emitted into the atmosphere, some is absorbed by vegetation, some is dissolved in atmospheric water, converted to aerosols, and transported long distances (1000 + km), and some is deposited nearby (Ferm, 1998). Model estimates indicate that about 50% of the emitted NH3 is deposited on land or water surfaces within 50 km of the source (Ferm, 1998). When NH3 or NH4 is deposited back onto soil the capacity of the soil to take up atmospheric CH4 may be affected.
After 12 d when CH4 contained in the slurry had been emitted and production in the soil had ceased, atmospheric CH4 was absorbed by the treated area. During the next 78 d, 207 (SE ± 16) g CH4–C ha-1 (9.7 µg C m-2 h-1) were absorbed by the treated area. Thus, the slurry treatment did not reduce the capacity of the soil to absorb CH4. In the control plots, CH4 absorption during the same time period amounted to 194 (SE ± 9) g CH4–C ha-1 (9.1 µg C m-2 h-1). Thus, the addition of the slurry had essentially no effect on CH4 uptake. The CH4 uptake in the slurry-treated area was similar to that on the unfertilized area of Hansen et al. (1993) in Norway (9.7 µg C m-2 h-1) and almost twice that in the area they had treated with cattle slurry (5.9 µg C m-2 h-1).
Total ammoniacal nitrogen added in the slurry would be expected to contribute to the burden of N2O in the atmosphere in two ways, directly by nitrification and denitrification on the treated area, and indirectly by metabolism of deposited NH3 on the surrounding countryside (Mosier et al., 1998). As noted above, the direct emission was 7.6 kg N ha-1, and the indirect effect is 0.6 kg N ha-1 (calculated from the emission factor proposed by Intergovernmental Panel on Climate Change [1997] and Mosier et al. [1998], namely 0.01 times the amount of NH3 emitted, 57 kg N ha-1), giving a total of 8.2 kg N2O-N ha-1.
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CONCLUSIONS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
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R.Z. Khan (deceased), from 1995 to 1998 Ph.D. student at Soil Plant and Ecological Sciences Division, P.O. Box 84, Lincoln Univ., Canterbury, New Zealand. 
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 D. Ginting and B. Eghball Nitrous Oxide Emission from No-Till Irrigated Corn: Temporal Fluctuation and Wheel Traffic Effects Soil Sci. Soc. Am. J., May 6, 2005; 69(3): 915 - 924. [Abstract] [Full Text] [PDF]
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G. L. Velthof, J. A. Nelemans, O. Oenema, and P. J. Kuikman Gaseous Nitrogen and Carbon Losses from Pig Manure Derived from Different Diets J. Environ. Qual., March 1, 2005; 34(2): 698 - 706. [Abstract] [Full Text] [PDF]
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T. J. Clough, F. M. Kelliher, R. R. Sherlock, and C. D. Ford Lime and Soil Moisture Effects on Nitrous Oxide Emissions from a Urine Patch Soil Sci. Soc. Am. J., September 1, 2004; 68(5): 1600 - 1609. [Abstract] [Full Text] [PDF]
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 C. A. Rotz Management to reduce nitrogen losses in animal production J Anim Sci, January 1, 2004; 82(13_suppl): E119 - 137. [Abstract] [Full Text] [PDF]
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