Journal of Environmental Quality 31:1789-1794 (2002)
© 2002 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
TECHNICAL REPORTS
Atmospheric Pollutants and Trace Gases
Application Technique and Slurry Co-Fermentation Effects on Ammonia, Nitrous Oxide, and Methane Emissions after Spreading
I. Ammonia Volatilization
S. Wulf*,
M. Maeting and
J. Clemens
Institute of Plant Nutrition, University of Bonn, Karlrobert-Kreiten-Str. 13, 53115 Bonn, Germany
* Corresponding author (se.wulf{at}uni-bonn.de)
Received for publication October 29, 2001.
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ABSTRACT
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Ammonia emissions after spreading animal manure contribute a major share to N losses from agriculture. There is an increasing interest in anaerobic co-digestion of liquid manure with organic additives. This fermentation results in a change of physical and chemical parameters of the slurry. Among these are an increased pH and ammonium content, implying a higher risk of NH3 losses from fermentation products. To compare different application techniques and the effect of fermentation on NH3 volatilization, we used the standard comparison method and tested it for reliability. This method seems to be perfectly suited for experiments with a large number of treatments and replicates if prerequisites concerning the experimental layout are considered. We tested four different application techniques on arable and grassland sites. The more the substrate was incorporated into the soil or applied near the soil surface on the grassland site, the less NH3 was lost. Injection of the substrate reduced losses to less than 10% of applied NH+4 on both sites, whereas losses after splash plate application amounted to more than 30%. Trail shoe application on grassland performed as well as injection. Harrowing on arable land also reduced emissions efficiently, if harrowing occurred within the first 2 h after application. Emissions from trail hose–applied co-fermentation product were not greater than from unfermented slurry. Better infiltration of the less viscous substrate seemed to have compensated for the increased loss potential.
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INTRODUCTION
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AMMONIA (NH3) volatilization after spreading of liquid organic fertilizers is an important vector of N loss from agricultural systems. Up to 90% of NH+4–N applied with slurry can be lost through NH3 emissions (Horlacher and Marschner, 1990), substantially reducing the amount of plant-available N. In the early 20th century research was performed to improve the fertilizer effect of liquid waste. In addition to the use of slurry additives, application technique was also identified as a possible means to reduce NH3 losses (Blanck, 1918; Gerlach, 1918; Heck, 1931). However, with increasing use of mineral fertilizer, research interest in gaseous N losses decreased rapidly until NH3 became an environmental issue. Ammonia lost from agriculture eventually contributes to atmospheric N input into natural or near-natural ecosystems, not only promoting soil and surface water acidification, eutrophication, and forest dieback (Ellenberg, 1985; Fangmeier et al., 1994), but also causing N2O emissions, thus being an indirect greenhouse gas. Estimates show that up to 90% of anthropogenic NH3 emissions originate from agriculture (Buijsman et al., 1987) and most of this is assigned to animal husbandry, especially the spreading of animal manure. To reduce environmental hazards associated with application of animal manures to land it is necessary to know how much of the nitrogen added is plant available. This is the basic requirement to avoid either yield losses or overfertilization and consequently harmful gaseous (N2O) or leaching 
losses of nitrogen. Reducing NH3 losses during application is the only way to make animal manures an efficient fertilizer and to minimize NH3 deposition in other ecosystems, where it might cause nutrient imbalances or increased N2O losses.
There are already some publications on the effect of application techniques on NH3 emissions in field experiments, but most of them compare only two techniques (Dosch and Gutser, 1996; Rubaek et al., 1996; Ferm et al., 1999), or do not measure throughout the experiment, but monitor emissions for selected periods only and estimate the emissions for the periods in between (Malgeryd, 1998; Morken and Sakshaug, 1998; Weslien et al., 1998; Smith et al., 2000).
Co-fermentation is an upcoming technology, not only because it is a way of producing renewable energy (biogas), but also because organic wastes can thus be recycled into efficient fertilizers and it gives extra income to farmers. But while fermentation of slurry is of increasing relevance, information on NH3 emissions from fermented slurries is still scarce, especially concerning the effect of different application techniques. Rubaek et al. (1996) reported similar or even lower NH3 emissions from fermented substrates compared with raw slurry, whereas Kuhn (1998) postulates an increase of NH3 emissions through slurry fermentation.
Various methods for determining NH3 emissions in the field exist. The most common are micrometeorological methods (Ferguson et al., 1988; Generemont et al., 1998) or wind tunnels (Braschkat et al., 1993), the first involving large homogeneously fertilized areas, the latter rather expensive technical equipment. Because NH3 emissions are strongly influenced by highly variable environmental factors such as temperature, wind speed, and precipitation it is indispensable to test a series of application techniques simultaneously to determine useful mitigation strategies. This, however, can only be achieved with plot experiments and a measurement technique allowing concurrent measurement from a large number of treatments and replicates. Methods designed for this purpose are the dynamic chambers designed by Svensson (1994) and the standard comparison (SC) method described by Vandré and Kaupenjohann (1998). Both methods use passive diffusion samplers. The Svensson method involves the measurement of equilibrium NH3 concentrations inside and outside of dynamic chambers on every plot, followed by micrometeorological calculations. For the SC method only one passive sampler per plot is needed and emission rates are calculated by simple comparison of passive sampler concentrations on standard and treatment plots. As the SC method is simple and can easily be adjusted to large plot numbers, we used this method and tested it for reliability.
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MATERIALS AND METHODS
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Site Description and Experimental Setup
Both farms were situated in the western part of Germany near the city of Bonn. The grassland experiment (Heidgen) was started in April 1999 on a poorly drained Stagno–Gleyic Luvisol (FAO classification) at a grass sward height of approximately 10 cm with wet and cool weather at the beginning of the investigation. Mean daily air temperature was 9°C, mean volumetric soil moisture content was 40%, and precipitation during the 4 d of the experiment was 4.5 mm. The trial on arable land (Klein–Altendorf) was started on fallow land three weeks later on a well-drained Luvisol (FAO classification) when rather warm and dry weather prevailed, with 13°C air temperature, 18% soil moisture content, and no rainfall during the first 4 d after applying the slurry. Information on site characteristics is given in Table 1.
The two experiments were conducted in the same manner in a completely randomized design with four replications for each treatment. Plot size was 9 m2 with 8- or 12-m spacing between the plots to minimize the drift of NH3 from one plot to another. Ammonia measurements were conducted during the four days after spreading on each site, as it is known that most of the NH3 losses occur during the first 24 h after spreading and volatilization after 3 d is negligible (e.g., Weslien et al., 1998; Sommer and Jacobsen, 1999).
Slurries
The digested slurry used was a co-fermentation product. This substrate was produced through combined anaerobic fermentation of 70% dairy cow slurry together with 30% biowaste (organic household waste). Mean duration of fermentation was approximately 40 d under mesophile temperature conditions (42°C).
Cattle slurry was used as a reference with one application technique only. The slurry was the same as used for fermentation. Slurry and the fermentation product differed in their chemical parameters (Table 2). During fermentation methanogenic microorganisms producing CO2 and CH4 digest organic compounds. Nitrogen from this organic pool is transferred to inorganic nitrogen during this process. So the share of NH+4–N from total nitrogen rises and constituents that can be oxidized by chemical (COD) or biological processes (BOD5) as well as dry matter content are reduced.
Application Techniques
The substrates were applied at a rate of 30 m3 ha-1 by splash plate, trail hose, trail shoe, and injection. For injection a custom-made tractor-drawn device was used. The substrate was injected 10 cm deep into the soil; spacing between the injector tines was 30 cm. The other application techniques were simulated by hand. The trail hose treatment was applied with watering cans near the soil surface. For the trail shoe application the slurry was conducted through a hose from the watering can to the end of a tine attached to a helve that was drawn through the grass sward. The splash plate was constructed from a watering can with a circular plastic disk (diameter = 20 cm) attached at an angle of 120° at a distance of 10 cm from the sprout. On arable land, an additional trail shoe treatment was used with immediate shallow incorporation with a garden harrow with tines of 5 cm length attached to a helve.
Application of substrate on a single plot took about 10 min and gas flux measurement started immediately thereafter. The substrates were applied to the all plots within 90 min.
Ammonia Measurement
Ammonia volatilization was measured with the standard comparison method described by Vandré and Kaupenjohann (1998). This is an open method designed for plot experiments where NH3 is collected in passive samplers installed on each plot 15 cm above the soil surface. Emission rates are calculated by comparing passive sampler signals from the experimental plots with known emission rates from standard plots.
The samplers were constructed from rectangular 250-mL polyethylene flasks with windows cut into the sides to allow gas exchange. The windows were covered with a polyethylene net to prevent insects from entering the samplers. Forty milliliters of 0.05 M H2SO4 were filled into the samplers to collect NH3 and were replaced every 4 h. Ammonium concentrations in the passive sampler solution were analyzed colorimetrically with a continuous flow analyzer (Bran + Lübbe [Norderstedt, German] AA3) and gave mean signals over the sampling period.
On two standard plots of same size and exposition as the experimental plots, known amounts of NH3 from gas cylinders where fed into a carrier stream passing through a tube system (commercial polypropylene sewage tubing with plug connections, 50-mm i.d.) and was emitted through small holes (diameter = 0.5 mm) on the top of the tubes, creating a two-dimensional source by a network of point sources (four holes per square meter). Homogeneity of gas distribution through this system was tested by measuring the flux from each hole with a digital flow meter (ADM 1000; J&W Scientific, Folsom, CA) giving a coefficient of variation of 7%. As air flow over the plots is turbulent, perfect mixture of the gas can be assumed before reaching the passive sampler. Flux rates were adjusted to 62 µg NH3–N m-2 s-1 on one of the standard plots and 31 µg NH3–N m-2 s-1 on the other with two-stage pressure relief valves and flow meters with input needle valves. The flux was checked and readjusted every 2 h as temperature fluctuations during day and night caused slight changes of NH3 flux due to temperature effect on viscosity of the gas. Oscillation of the gas flow due to NH3 ebullition in the valves as described by Vandré and Kaupenjohann (1998) was avoided by using a mixture of 7.5% NH3 in synthetic air instead of gas cylinders with pure NH3.
Knowing the flux rate (QS) and ammonium concentration in the passive sampler solutions (SS) collected from the standard plots, a transfer factor (f) for each 4-h period could be determined (Eq. [1]) and allowed the calculation of flux rates on the experimental plots (QE) from passive sampler NH+4 concentrations (SE) (Eq. [2]). The atmospheric background concentration (SB) was measured with samplers placed in the area surrounding the experimental site:
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RESULTS AND DISCUSSION
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Measurement Methodology
As the standard comparison method to determine NH3 volatilization is a very new technique and only little information is available on weakness and strength of this method, some information on the reliability of the results obtained will be given.
Transfer Factors
To produce reliable results, the accurate determination of transfer factors for each sampling period is very important. Figure 1
shows the time course of the transfer factors over the measuring period in the grassland experiment. Variation of the factors with time is high and at some sampling dates there are great differences in the factors calculated from high and low emission standard plots. According to Vandré and Kaupenjohann (1998), the transfer factor is independent of NH3 emission rate but depends on surface characteristics and meteorological conditions. Therefore, the differences in factors calculated from the two standard plots most probably result from the abovementioned temperature effect on flow rates and reflect the difficulty of checking and readjusting the flow meter needle valves regularly. The high variability with time is due to changes in meteorological conditions, as those factors have a direct effect on the transfer of NH3 from the tube system to the passive sampler. Linear multiple regression showed that the most important factor is wind speed, explaining 60% of the fluctuations of transfer factors in the grassland experiment. Taking into account also air temperature and humidity, more than 74% of the variation with time could be explained. This makes meteorology a powerful tool to verify the transfer factors and to eliminate outliers before calculating mean transfer factors from the standard plots and emission rates from the experimental plots. After elimination of outliers the mean error of the transfer factors calculated from the two standard plots was 8%. On arable land no comparison of transfer factor and wind speed was possible, as the anemometer on the field failed to work. But it can be assumed that the determination of transfer factors was reliable, as the mean error of the transfer factors was only 10%.
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Fig. 1. Time course of transfer factors and wind speed during the measuring period in the grassland experiment.
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Plot Interaction
A prerequisite of the standard comparison method is a homogeneous wind field on the experimental site to assume the same micrometeorological conditions for standard and experimental plots and to use the transfer factor measured on standard plots for the calculation of emission rates from the experimental plots. Another important requirement is to minimize the drift of NH3 from one plot to another to get accurate signals for each single plot. Sufficient distance between the plots is necessary to achieve this. In the grassland experiment we used 8-m spacing as did Vandré and Kaupenjohann (1998), but mean variability between replicates with mean emission rates > 2.5 µg NH3–N m-2 s-2 was much higher with 69% compared with 24% determined by Vandré and Kaupenjohann (1998), indicating some spatial dependency of calculated emission rates. Figure 2
illustrates where the highest replicate for each treatment and sampling period was found most often. Highest replicates were found most often close to the two standard plots, indicating that spacing between plots was not large enough to prevent drift effects. This was most obvious next to the standard plots, because apart from the first sampling period, when emission rates from applied substrates were highest, source strength of NH3 from the standard plots was much greater than from the experimental plots.
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Fig. 2. Number of sampling periods with highest emissions measurements within the four replicates of the same treatment during the first 12 measurement periods. Each bar represents a plot; the arrangement of bars resembles the experimental layout. As an example for treatment arrangement, splash plate application is shaded grey.
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In the experiment on arable land, no spatial dependency of calculated emission rates could be found and mean variability of emission rates > 2.5 µg NH3–N m-2 s-2 was 30%. The reason for this might be that spacing between the plots was 12 m, but also different meteorological conditions might be responsible.
Despite the mentioned restrictions, the standard comparison method produced clear and statistically significant results. As all other methods for quantifying NH3 volatilization are subjected to strong limitations concerning number of treatments or replicates, this method so far seems to be the only sensible technique to cope with larger plot experiments designed to allow statistical analysis of the results. To further improve measurement accuracy it might be reasonable to include micrometeorological models into the calculation of transfer factors. Spacing between plots should be made as big as possible, taking into account that micrometeorological and soil conditions also should be homogeneous. In addition, it is advisable to reduce the source strength of standard plots after the first or second sampling date, because the corresponding emissions from the experimental plots also decrease rapidly.
Ammonia Volatilization
Ammonia emissions were highest immediately after application on both sites. In Fig. 3
the time course of emission rates on arable land is shown. Cumulated emissions after 24 h ranged from 88% of total emissions after splash plate application on both arable and grassland and 70% after harrowing on arable land. The following days, emissions decreased more and more until background NH3 concentration was reached after four days. A diurnal rhythm of emission was evident with higher emission during the day and lower emission at night (Fig. 3), as already described by Horlacher and Marschner (1990) and Vandré et al. (1997). This can be attributed to temperature effects on NH3 solubility and the equilibrium concentration of NH3 and NH+4 according to Henry's law (Hengnirun et al., 1999).
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Fig. 3. Ammonia emission rates from arable land after spreading co-fermented slurry with different application techniques. Data represent means and standard deviation (n = 4).
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The influence of application technique on NH3 volatilization was already evident at the first sampling date. The highest emission rates occurred immediately after splash plate application on both sites and trail hose application on arable land. Trail shoe application on grassland or harrowing on arable land and injection on both sites significantly reduced emissions during the first 4-h period after spreading the substrate. The differences in overall emissions can clearly be shown with cumulated emissions (Fig. 4)
. The more the substrate was incorporated into the soil or applied near the soil on the grassland site, the less NH3 was lost. Injection of the substrate reduced losses to about 10% of the applied NH+4–N on both sites, whereas losses after splash plate application amounted to more than 30%. These emission levels and reduction potential of injection correspond to mean NH3 losses measured by Smith et al. (2000), with 24 and 33% of NH+4–N for splash plate–applied slurry on moist arable and grassland, respectively. Mean emissions were reduced to 5 and 12% by injection. Nevertheless, overall emissions strongly depend on environmental conditions. Sommer et al. (1991) found a correlation of wind speed and temperature with NH3 volatilization. Reitz and Schürer (1999) report reduced NH3 losses after simulated precipitation, whereas according to Vandré et al. (1997), NH3 losses after spreading slurry on moist arable land might be increased due to impeded infiltration. Thus, emissions ranged from 22 to 76% after splash plate application and 10 to 39% after injection (Vandré et al., 1997).
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Fig. 4. Cumulated NH3 losses after spreading co-fermented slurry with different application techniques to arable (a) and grassland (b). Data represent means and standard deviation (n = 4).
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In our experiment spreading with trail hoses on arable land resulted in almost the same NH3 volatilization as splash plate because the fermented substrate was of low viscosity, having small dry mass content, and could not be applied in a defined band but spread over a greater surface area due to encrustation of the soil surface (Fig. 4a). On the grassland site, trail hose application performed slightly better (not significant), as some of the substrate filtrated through the grass and thus was less exposed to atmosphere and wind (Fig. 4b). When the substrate was applied directly to the soil surface with trail shoes on the grassland site, emissions were as low as after injection, most probably because resistance of diffusion through the grass was high and the substrate could infiltrate directly into the soil. On the arable site, shallow incorporation by harrowing immediately after trail hose application substantially reduced emissions, but was less efficient than injecting the substrate (Fig. 4a). The major part of this reduction can be assigned to emissions during the first 4 h after application, as 21% of applied NH+4–N was lost in this short period after splash plate application, 17% after trail hose application, and only 9% after harrowing. In our experiment, harrowing occurred less than 5 min after application, whereas in Germany, incorporation after spreading of slurry on bare soil in praxis is often realized more than 6 h after application when mitigating effects of harrowing are only small. Malgeryd (1998) found a reduction of 60% in NH3 emissions through harrowing surface-applied slurry 4 h after spreading. For fermented substrates as used in our experiments, this seems to be far too late as already about 60% of total losses from splash plate– or trail hose–applied NH+4–N occurred during the first 4 h after application. The exponential decrease of emission rates implies that most of these losses can be assigned to emissions within 1 or 2 h after application.
Comparing NH3 losses from slurry and co-ferment after trail hose application, the influence of physical and chemical properties of the substrates on N volatilization can be seen (Fig. 5)
. On both sites overall emissions did not differ significantly between fermented substrate and raw slurry, but especially during the grassland experiment there was a difference in the time course of emissions. Whereas higher NH3 volatilization occurred from the co-ferment immediately after application, emissions from the undigested slurry continued over a longer period, resulting in higher (not significant) overall NH3 losses from unfermented slurry, especially in the grassland experiment. The slurry had a lower NH+4 content and pH than the co-ferment, but much higher dry mass content so that it could be applied in defined bands, whereas the co-ferment was spread over a greater surface area. Due to the less defined application and higher NH+4 content, emissions immediately after spreading were higher from the co-ferment than from slurry. But while the co-ferment infiltrated much faster, the slurry band stayed on top of the grass or soil surface, retaining much of its moisture and emitting NH3 over a longer period. On grassland this effect tends to be more important, as the high dry matter content impedes filtration through the grass cover and a certain amount of airflow will also take place under the slurry band, being likely to promote volatilization. The effect of higher infiltration capacity of the co-ferment with low dry matter content can be compared with studies with separated slurries. Frost et al. (1990) and Vandré et al. (1997) report significant reduction in NH3 emissions when applying separated slurries with splash plate on grassland or trail shoe followed by a flexible harrow on arable land, respectively. Theoretical considerations imply that substrates of higher NH+4 content and pH should cause higher emissions than those of smaller NH+4 content and pH (Kuhn, 1998). In our study co-ferment had both higher pH and higher NH+4 content but NH3 volatilization after spreading was not higher than after slurry application. Increased loss potential due to chemical factors seems to be compensated by improved infiltration of the less viscous fermented substrate.
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Fig. 5. Cumulated NH3 losses after spreading co-fermented slurry and unfermented cattle slurry with trail hoses on arable (a) and grassland (b) sites.
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CONCLUSIONS
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Comparison of NH3 emission from a large number of treatments and replicates can be achieved with the standard comparison method described by Vandré and Kaupenjohann (1998). Drift of NH3 from one plot to another occured and increased the variability between replicates in the grassland experiment, where distances between plots were only 8 m. Nevertheless, these effects were small and still allowed distinction of significant treatment effects. Increasing the distance between plots can reduce the drift of NH3. Apart from this, flow from the standard plots should be reduced after the first day. Flow regulation might be improved with mass flow controllers instead of needle valves and flow meters.
This study comprises only one study on grassland and arable land under specific conditions. Nevertheless, from this limited number of measurements some conclusions can be made.
Ammonia emissions can effectively be reduced with appropriate application techniques. The more the substrate is incorporated into the soil, the less NH3 is lost, but it is questionable if emission reduction through injection compared with harrowing on arable land justifies higher costs of application. Grassland application with trail shoe might reduce NH3 volatilization as effectively as injection, but further experiments including different sward heights are needed to verify this effect.
As a consequence of the very high NH3 emissions immediately after spreading co-fermented slurry and the exponential decrease of emissions, incorporation on bare arable land should be performed within a very short period after application, best of all in one operation with spreading, for example, by mounting a harrow on or trailing it after the slurry-spreading device.
The higher potential for NH3 losses from fermented slurry due to pH and NH+4 content does not essentially induce higher volatilization, as the substrate can infiltrate much faster into the plant cover or soil because of smaller dry matter content and reduced viscosity.
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ACKNOWLEDGMENTS
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We are greatly indebted to Mr. B. Passmann for allowing us to set up the grassland experiment on his farm and Mr. H. Scheb-Wetzel (Institut für Betriebslehre) for providing the arable land site. Dr. M. Thelen (Institut für Landtechnik) helped us very much in organizing and applying the injection treatment. Above all we would like to thank Mrs. W. Schlüter for her valuable support in the field and laboratory. This project was funded by the Deutsche Bundesstiftung Umwelt (DBU).
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REFERENCES
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TOP
ABSTRACT
INTRODUCTION
