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Published online 31 May 2006
Published in J Environ Qual 35:973-981 (2006)
DOI: 10.2134/jeq2005.0320
© 2006 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

3,4-Dimethylpyrazol Phosphate Effect on Nitrous Oxide, Nitric Oxide, Ammonia, and Carbon Dioxide Emissions from Grasslands

S. Menéndeza,*, P. Merinob, M. Pintob, C. González-Muruaa and J. M. Estavilloa

a Department of Plant Biology and Ecology, University of the Basque Country, Apdo. 644, E-48080 Bilbao, Bizkaia, Spain
b Basque Institute for Agricultural Research and Development, NEIKER. Bo Berreaga 1. E-48160 Derio, Bizkaia, Spain

* Corresponding author (gvbmevis{at}ehu.es)

Received for publication August 22, 2005.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Intensively managed grasslands are potentially a large source of NH3, N2O, and NO emissions because of the large input of nitrogen (N) in fertilizers. Addition of nitrification inhibitors (NI) to fertilizers maintains soil N in ammonium form. Consequently, N2O and NO losses are less likely to occur and the potential for N utilization is increased, and NH3 volatilization may be increased. In the present study, we evaluated the effectiveness of the nitrification inhibitor 3,4-dimethylpyrazol phosphate (DMPP) on NH3, N2O, NO, and CO2 emissions following the application of 97 kg N ha–1 as ammonium sulfate nitrate (ASN) and 97 kg NH4+–N ha–1 as cattle slurry to a mixed clover–ryegrass sward in the Basque Country (northern Spain). After slurry application, 16.0 and 0.7% of the NH4+–N applied was lost in the form of N2O and NO, respectively. The application of DMPP induced a decrease of 29 and 25% in N2O and NO emissions, respectively. After ASN application 4.6 and 2.8% of the N applied was lost as N2O and NO, respectively. The application of DMPP with ASN (as ENTEC 26; COMPO, Münster, Germany) unexpectedly did not significantly reduce N2O emissions, but induced a decrease of 44% in NO emissions. The amount of NH4+–N lost in the form of NH3 following slurry and slurry + DMPP applications was 7.8 and 11.0%, respectively, the increase induced by DMPP not being statistically significant. Levels of CO2 emissions were unaffected in all cases by the use of DMPP. We conclude that DMPP is an efficient nitrification inhibitor to be used to reduce N2O and NO emissions from grasslands.

Abbreviations: ASN, ammonium sulfate nitrate • DMPP, 3,4-dimethylpyrazol phosphate • NI, nitrification inhibitors • WFPS, water filled pore space


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
THE INEFFICIENT USE of nitrogen fertilizers can lead to large N losses from the soil–plant system which, apart from economic loss, can lead to undesirable ecological problems such as nitrate leaching or N gaseous losses. Gaseous losses of N can occur as ammonia (NH3) volatilization or as nitric oxide (NO), nitrous oxide (N2O), and nitrogen (N2) emissions.

Ammonia volatilization is a phenomenon involving both the chemical conversion of ammonium N to dissolved ammonia gas and the physical transport of the ammonia gas into the air, causing negative effects on atmospheric quality (Barthelmie and Pryor, 1998), soil acidification (Van der Eerden et al., 1998), and water eutrophication (Bobbink et al., 1992). Ammonia losses occur during animal housing, slurry storage, slurry application, grazing, fertilizer application, and to a lesser extent from crops (Jarvis and Pain, 1990). Among agricultural sources of NH3 volatilization, livestock manure is by far the most important, causing 80% of emissions in the world (Van Der Hoek, 1998). Therefore, minimizing NH3 losses from agricultural sources is essential for environmental protection.

Emissions of N2O and NO, which arise mainly from both the denitrification and nitrification microbial processes, constitute important sources of N loss from agricultural soils to the atmosphere. Chemical reactions seem to be important only for the production of NO, but not of N2O (Bremner et al., 1980; Nelson, 1982), and NO production by chemodenitrification may only be significant under acidic conditions (Van Cleemput and Baert, 1984). Microbial denitrification requires an anaerobic environment, whereas aerobic conditions are necessary for nitrification (Bremner and Blackmer, 1979). Nevertheless, both processes can take place simultaneously in the soil (Abbasi and Adams, 1998), since aerobic and anaerobic microsites can exist within the same soil aggregate (Kuenen and Robertson, 1994). Denitrification has been considered a prokaryotic process for more than a century and has been extensively studied in several bacteria (Zumft, 1997). Laughlin and Stevens (2002) have recently shown the potential for fungi to produce N2O by denitrification in grasslands. In both the nitrification and denitrification processes, NO is an intermediate which is not always emitted, since it can be further metabolized in soil. Although both nitrification (Dunfield and Knowles, 1999; Godde and Conrad, 2000) and denitrification (McKenney et al., 1982; Remde and Conrad, 1991; Schafer and Conrad, 1993) processes can consume NO, relative consumption by denitrification seems to be higher (Skiba et al., 1993). Nitrification is therefore believed to be the main source of NO (Anderson and Levine, 1986). Regarding environmental considerations, NO contributes to the formation of acid rain, while N2O is involved in global warming and contributes to the destruction of parts of the ozone layer; N2O has great importance as a greenhouse gas because it has a mean atmospheric residence time of more than 100 yr (Prather et al., 2001). Agricultural soils are considered to be a major source of N2O, and approximately 35% of the global annual N2O emission is attributed to agriculture (Isermann, 1994).

Nitrification inhibitors (NI) like 3,4-dimethylpyrazol phosphate (DMPP) are compounds that delay the bacterial oxidation of ammonia to nitrite in the soil (first step of nitrification) for a certain period of time by depressing the activity of Nitrosomonas bacteria in the soil. Application of NIs with several ammonium-based fertilizers has been shown to decrease N2O emissions (Bronson et al., 1992; Mosier 1994; De Klein et al., 1996). The NI DMPP is applied at very low rates compared to the recommended doses for other NIs, and DMPP application with ammonium sulfate nitrate (ASN) as ENTEC 26 has reduced N losses by nitrate leaching (Fettweis et al., 2001; Bañuls et al., 2001) and N2O emissions (Weiske et al., 2001; Macadam et al., 2003).

After NI application, N persists in the soil in ammonium form for a longer period of time, thus increasing the risk of enhanced ammonia volatilization. Few studies have reported the effect of NIs on NH3 volatilization. Prakasa Rao and Puttanna (1987) reported enhanced NH3 volatilization losses after dicyandiamide-treated urea application. On a global scale, CO2 is the most important greenhouse gas contributing to global warming (Lal and Kimble, 1995). Despite the fact that soils are an important source of CO2, only a few studies have determined NI effects on soil CO2 emissions. Weiske et al. (2001) suggested that DMPP may reduce soil CO2 emissions.

In the edaphoclimatic conditions of the Basque Country (northern Spain), there is a high potential for gaseous N losses in cut grassland soils, from both denitrification (Estavillo et al., 1994; Barton et al., 1999) and nitrification processes (Merino et al., 2001; Estavillo et al., 2002). In these conditions, we have already reported that DMPP is an efficient NI, reducing N2O losses both when applied with ASN 26% as ENTEC 26 (Macadam et al., 2003) and with cattle slurry (Merino et al., 2005). The aim of the present work was to evaluate the effect of DMPP on the simultaneous emissions of NH3, N2O, NO, and CO2 after the application of cattle slurry or ASN.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
This work was conducted in a cut grassland in the Basque Country (northern Spain) during the spring of 2003. The soil was a poorly drained clay loam (34% fine sand, 3% coarse sand, 34% silt, and 29% clay in the top 10 cm) with a pH (1:2 H2O) of 6.6. A typical permanent pasture [perennial ryegrass (Lolium perenne L. var. Herbus), 60%; intermediate ryegrass (Lolium x hybridum Hausskn. var. Texi), 32%; white clover (Trifolium repens L. var. Huia), 8%] was sown at a density of 40 kg seeds ha–1 in April 2001. The meteorological conditions of the location during the assay were the following: rainfall for the period of the study (from 19 May to 17 July) was 113 mm, soil temperature ranged between 12.8 and 21.5C°, and air temperature between 13.3 and 27.5°C.

Two kinds of fertilizer were applied: ASN 26% and cattle slurry. Slurry was obtained from a concrete storage pit on a dairy farm. Application rates were 97 kg N ha–1 for ASN 26% and 97 kg NH4+–N ha–1 (i.e., 181 kg total N ha–1) for the slurry. Nitrogen in ASN consisted of 7.5% nitric and 18.5% ammoniacal. The combination of ASN with DMPP was applied as available in the market, that is, ENTEC 26 granules, a product developed by BASF (Ludwigshafen, Germany) and commercialized by COMPO (Münster, Germany). A treatment with no fertilizer was also included as a control. Four treatments were examined as follows: control (C), ASN 26% (M), ENTEC 26 (M + DMPP), slurry (S), and slurry + DMPP (S + DMPP). A randomized complete block factorial design with four replicates was established, with each experimental plot measuring 4 x 3 m.

Ammonium sulfate nitrate 26% and ENTEC 26 were surface applied in granular form by hand. Slurry was surface applied. Total N in the slurry was analyzed by Kjeldahl digestion and automatic distillation–titration by a Kjeltec Auto 1035 analyzer (Tecator AB, Höganäs, Sweden). The amount of NH4+–N in the slurry was determined by direct distillation with MgO. Slurry was analyzed 2 d before application for dose calculations and again verified at the time of application. Characteristics of the slurry were: total N = 0.43% (w/fw); NH4+–N = 0.23% (w/fw), and C to N ratio = 11.22.

DMPP was added at a proportion of 0.8% of the NH4+–N applied (approximately 1 kg ha–1), diluted in a small volume of water. This solution was sprayed on slurry + DMPP plots just after slurry application. The rate of DMPP applied was equal to that incorporated in ENTEC 26.

Emission Measurements
Ammonia
Emissions of NH3 were measured for 5 d after fertilizer application. Measurements were made in the four plots per treatment at 1, 4, 6, 10, and 19 h from fertilizer application on the first day. On the second and third days, the frequency of measurements was decreased to three times per day, and sampling was further reduced to once per day on Days 4 and 5.

Ammonia emissions were measured using an open chamber technique (Hinz, 2005). Opaque PVC chambers with a volume of 6.75 L and an area of 0.0314 m2 were fitted tightly onto a frame which was inserted 3 cm into the soil. The chambers had a small hole on the top for air exhausting. The inside walls of the chambers were covered with polytetrafluoroethylene (PTFE) film to ensure minimum uptake of the soil-emitted NH3 by the walls. In each plot, one chamber was placed and repositioned daily to account for spatial variation. To remove the ammonia from the air entering the chambers, inlet air was taken from outside the plot area and pumped by an air compressor (1 L per min) through two 1-L glass flasks in series containing 500 mL of H3PO4 (10%) and 500 mL H2O, respectively. The air was then filtered through silicagel (CoCl2-free) to absorb excess water. Concentrations of NH3 were measured at the air inlet and outlet of the chamber using a photoacoustic infrared gas analyzer (Model 1302 Multi-Gas Monitor, detection limit 0.2 ppm; Brüel and Kjær, Nærum, Denmark) for approximately 5 min, once the steady-state value had been reached. Fluxes of NH3 were calculated from the concentration differences between inlet and outlet air, the air flow rate through the chamber, and the surface area covered by the chamber.

Nitrous Oxide
After fertilizers had been applied, N2O emissions were measured daily for 2 wk in the four plots per treatment. Measurements continued at a frequency of once or twice per week for 59 d. Sampling was always conducted in the morning (from 1000 to 1400 h). Soil temperature was measured at a depth of 10 cm using a soil thermometer (HI 935005; Hanna Instruments, Woonsocket, RI). Nitrous oxide emissions were measured using a closed air circulation technique in conjunction with a photoacoustic infrared gas analyzer (Model 1302 Multi-Gas Monitor, detection limit 0.03 ppm) for 40 min (Merino et al., 2001). The measurement chambers had the same characteristics as those used for NH3 measurements, except that they were completely closed chambers. Air from the headspace of the chamber was directed through 40-m PVC tubes via a multipoint sampler to the analyzer in a closed system. Fluxes were calculated from the linear concentration increase in the chamber headspace with time (r2 > 0.90).

Nitric Oxide
Nitric oxide emissions were measured just before N2O measurements using an open chamber technique as described by Harrison et al. (1995). The flux chambers had the same characteristics as those used for NH3 measurements. Inlet air was filtered through charcoal and purafil to remove ambient O3 from the air stream (thus eliminating reactions between O3 and NO within the chamber) and then pumped into the chamber via PTFE tubing at a rate of 1 L min–1. Concentrations of NO were measured at the air inlet and outlet of the chamber using an NO–NO2–NOx chemiluminescence analyzer (Model AC31M, detection limit 0.35 ppb; Environnement SA, Poissy, France). Fluxes of NO were calculated from the concentration differences between inlet and outlet air, the air flow rate through the chamber, and the surface area covered by the chamber.

Carbon Dioxide
Carbon dioxide emissions were measured simultaneously with N2O emissions using the closed air circulation technique in conjunction with a photoacoustic infrared gas analyzer (Model 1302 Multi-Gas Monitor, detection limit 3.4 ppm). Fluxes were calculated from the linear concentration increase in the chamber headspace with time.

Cumulative Losses
Cumulative gas emissions during the sampling period were estimated by averaging flux in two successive determinations, multiplying that average flux by the length of the period between the measurements, and adding that amount to the previous cumulative total.

The global warming effect of cumulative N2O emissions was estimated following the recommendations of the Intergovernmental Panel on Climate Change (1996), using a global warming potential (GWP) factor of 310.

Soil Water Content
On each day that gaseous emissions were measured, eight soil cores (0- to 10-cm depth and 2.5-cm diameter) were collected from each plot. Samples from each plot were combined and the gravimetric water content was determined. Soil water filled pore space (WFPS) (Aulakh et al., 1991) was calculated from the bulk density, assuming a soil particle density of 2.65 g cm–3.

Data Analysis
The LSD test was used for multiple comparisons of the instantaneous flux means, using the SPSS (2004) software. Differences between cumulative emissions in the different treatments were compared by ANOVA and the separation of means between treatments by Duncan test. Emissions of N2O followed a logarithmic distribution and log transformations of these emissions were used for statistical analyses. Significance for ANOVAs as well as LSD and Duncan tests was determined at p < 0.05.


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Emissions
Ammonia
Volatilization of NH3 from soil is a major N loss mechanism which reduces the efficiency of applied nitrogen. Denmead et al. (1976) measured NH3 volatilization rates of between 23 and 474.3 g NH3–N ha–1 h–1 from a grassland canopy without fertilization. In our experiment, maximum NH3 volatilization rates were lower than 15 g NH3 ha–1 h–1 in control, M, and M + DMPP treatments, and no statistical differences in instantaneous flux (Fig. 1) or cumulative emissions (Table 1) were observed between these treatments. Thus, the application of ASN 26%, either alone or as ENTEC 26, had no effect on NH3 volatilization.


Figure 1
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  		Fig. 1. Ammonia volatilization rates measured during the first 110 h following slurry and mineral fertilizer applications. ({blacktriangleup}) Control, ({blacksquare}) ammonium sulfate nitrate, ({square}) ENTEC, (•) slurry, and ({circ}) slurry with 3,4-dimethylpyrazol phosphate (DMPP). The vertical bars indicate LSD at 0.05 between treatments for each sampling time.

 

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  		Table 1. Cumulative emissions of the different gases up to Day 59 and CO2 equivalents (CO2e) for cumulative N2O emissions as well as total CO2e corresponding to N2O + CO2 following the recommendations of the Intergovernmental Panel on Climate Change (1996).

 
When slurry is spread on a field, low NH3 volatilization rates occur during the spreading process (Phillips et al., 1991). In contrast, high rates of volatilization occur during the first few hours subsequent to application (Mattila, 1998). This loss from surface-applied slurry is attributed to both the high initial concentration of total ammoniacal N and the high pH of surface soils amended with newly spread slurry (Sommer and Sherlock, 1996). In these experiments, the pH of the unamended soil was 6.6. Chantigny et al. (2004) described a pH increase of 1 to 3 units in the top 2 cm of soil following slurry application, resulting in high NH3 volatilization rates in the first 6 h after slurry application. This increase in soil pH following slurry addition may partly be attributed to the alkaline slurry pH (7.2 in our case) and to the dissociation of slurry carbonates (Génermont, 1996; Sommer and Sherlock, 1996), which occurs rapidly in non-alkaline soils such as ours (Rochette et al., 2000; Chantigny et al., 2001). It is known that the total ammoniacal N in the surface-applied slurry decreases rapidly due to volatilization, infiltration, and nitrification (van der Molen et al., 1990). In our experiment, maximum volatilization rates (250–270 g NH3–N ha–1 h–1) took place immediately after slurry application, decreasing to control levels 70 h after slurry application (Fig. 1). About 60% of total NH3 emissions occurred during the first 24 h. These values are consistent with those reported by Thompson et al. (1990) and Bussink et al. (1994) who observed that 45 and 60%, respectively, of the total measured NH3 emissions occurred within the first 24 h after application of cattle slurry.

Ammonia volatilization from surface-applied slurry increases with increasing temperature (Sommer et al., 1991) and decreases with increasing rainfall (Klarenbeek and Bruins, 1991). The moderate temperatures and dry weather conditions during our experiment could have favored NH3 losses in the slurry treatments. The maximum NH3 flux measured in these experiments was much lower than those reported by Vandré et al. (1997) (7.56 kg NH3–N ha–1 h–1) after the application of cattle slurry at 50 kg NH4+–N ha–1 under dry weather conditions. Stevens and Laughlin (1997) have reported cumulative NH3–N losses of up to 93% of the NH4+–N applied with surface application of cattle slurry, and Stevens and Logan (1987) described cumulative losses of 51% in a clayey soil. In our trial (clay loam soil) cumulative losses were less than 11% of the NH4+–N applied with the slurry (Table 1).

Higher NH3 volatilization losses might be expected after the application of NIs. Some authors have reported enhanced NH3 volatilization losses due to the application of NIs such as nitrapyrin or DCD together with urea (Smith and Chalk, 1978; Prakasa Rao and Puttanna, 1987). In our study, DMPP application induced higher NH3 losses when applied together with slurry, although not significantly at {alpha} = 0.05 (Table 1). This finding corroborates that of Wissemeier et al. (2001) who did not find differences in NH3 volatilization losses after ASN application with or without DMPP application.

Nitrous Oxide
It is known that N2O emissions are closely related to soil water content (Davidson, 1991). In our study, most N2O losses took place during the first 24 d after fertilizer application, when WFPS was over 60% (Fig. 2). The maximum flux (0.7–1.3 kg N2O–N ha–1 d–1, Fig. 2A) was observed when WFPS was close to field capacity, when both nitrification and denitrification derived N2O losses can occur simultaneously (Estavillo et al., 2002).


Figure 2
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  		Fig. 2. Nitrous oxide (A) and nitric oxide (B) emission rates and water filled pore space (WFPS) and soil temperature (C). ({blacktriangleup}) Control, ({blacksquare}) Ammonium sulfate nitrate, ({square}) ENTEC, (•) slurry, ({circ}) slurry with 3,4-dimethylpyrazol phosphate (DMPP), (+) soil temperature at the 10-cm depth, ({diamondsuit}) WFPS in the control and mineral fertilizer treatments (0- to 10-cm depth), and ({diamond}) WFPS in the slurry treatments (0- to 10-cm depth). The vertical bars indicate LSD at 0.05 between treatments for each sampling time.

 
The control treatment showed emission rates below 0.1 kg N2O–N ha–1 d–1 throughout the assay (Fig. 2A). Higher N2O emission rates were observed in slurry treatments than in mineral treatments (Fig. 2A), as also reported by Breitenbeck et al. (1980) and Cates and Keeney (1987), as well as by Merino et al. (2002) in a previous experiment in the same area. Slurry treatments had higher soil WFPS than the control and the mineral treatments during the study (Fig. 2C). Slurry also has a high organic carbon content, suggesting that the higher N2O emissions in the slurry treatments may be due to N2O produced during denitrification. Weier et al. (1993) found that denitrification rate was correlated with organic C supply. The readily oxidizable carbon present in the slurry may have produced an increase in O2 consumption (De Klein et al., 1996; Neeteson and Van Veen, 1987; Rice et al., 1988), and provided an energy source for denitrifying bacteria. In particular, the high respiratory activity of decomposing organic matter could cause anaerobic microsites ("hot spots") for denitrifiers (Christensen et al., 1990) which, together with the higher soil WFPS observed, would lead to increased denitrification rates in plots with slurry. Comfort et al. (1988) indicated that only 10 to 15 d were required for readily oxidizable carbon to be metabolized. In our case, 24 d passed before N2O flux from slurry treatments equaled those from mineral treatments (Fig. 2A). At times after Day 24, N2O emission rates from all fertilized plots were the same as the control treatment (Fig. 2A), due to the soil water content falling below 60% (Fig. 2C), and no further N2O emission was detected (Fig. 2A). Another reason for the higher emission rates in the slurry treatments could be the extra organic N applied with the slurry, which could be mineralized and serve as a potential source of ammonium or nitrate which could eventually form N2O. In previous studies in which the rate of N application in slurry and inorganic fertilizer was based on total N rather than mineral N, short-term N2O emissions (Tilsner et al., 2003) or total N losses by denitrification (Estavillo et al., 1994) were lower after the slurry application than after the mineral fertilizer application because the organic N in the slurry was not mineralized quickly. In fact, it has been described that even 8 mo after slurry application organic N in the slurry is still being mineralized (Díaz-Fierros et al., 1988).

Low cumulative N2O emissions were measured in treatments with mineral fertilizer (M and M + DMPP) which were not statistically higher than those in the control (Table 1). In the literature, percentages of fertilizer N lost as N2O from grasslands are in the range of 0 to 4% (Bouwman, 1990; Fowler et al., 1997). Our results show higher losses of 4.5 and 16% for ASN and slurry, respectively (Table 1). When cumulative N2O emissions were expressed as CO2 equivalents (Table 1), we observed that N2O warming effect ranged from 7% (control treatment) to 33% (S treatment) of the total CO2 warming effect corresponding to the sum of CO2 + N2O total equivalents.

The effect of DMPP on N2O emissions from slurry was observed during the first 7 d (Fig. 2A), when N2O flux was lower in the S + DMPP treatment than in the S treatment. Müller et al. (2002) verified that DMPP had no effect on the activity of denitrifying enzymes. The observed decrease in N2O emissions after DMPP application is presumably due to DMPP's effect as a nitrification inhibitor, reducing N2O emissions from nitrification by slowing down the nitrification process or by reducing the nitrate available for denitrification to N2O. In slurry treatments, DMPP reduced cumulative N2O emissions by 30% after 59 d (Table 1). In contrast, we did not find any significant effect of DMPP in mineral fertilizer treatments (Table 1). This finding was unexpected since ENTEC 26 has been reported to decrease N2O emissions. Weiske et al. (2001) reported an average 49% decrease in the amount of N2O released in a 3-yr field experiment comparing different crop rotation systems fertilized with ENTEC 26 in comparison to ASN 26%. In a previous study in our area, DMPP applied as ENTEC 26 also proved to be effective, reducing N2O losses by 58% compared to calcium ammonium nitrate 26% (Macadam et al., 2003). However, in the same area Merino et al. (2005) found a 48% decrease in DMPP efficiency in reducing N2O emissions after slurry application when the mean soil temperature was 16°C or above. In the present study, the mean soil temperature was 18°C, so this factor may be responsible, at least in part, for the decreased efficiency of DMPP in reducing N2O emissions. Application of DMPP with slurry significantly decreased N2O emissions (Fig. 2A, Table 1). The reasons underlying the low efficiency of ENTEC 26 for the reduction of N2O emissions in the current study remain to be determined.

Nitric Oxide
Fluxes of NO-N were around 10 times lower than those of N2O-N. In the control treatment, NO-N flux never exceeded 15 g NO-N ha–1d–1. The emissions were always higher in the treatments with inorganic fertilizer (M and M + DMPP) than in the treatments with slurry (S and S + DMPP) (Fig. 2B). The NO emissions were strongly related to soil water content. It is known that NO emissions are affected by changes in soil water content (Skiba et al., 1992) and soil texture and structure (Skiba et al., 1997). Maximum NO flux (107 g NO-N ha–1d–1) was observed 30 d after ASN application (Fig. 2B), when the WFPS was below 50% (Fig. 2C) and the main process for NO production was expected to be nitrification because of the well-aerated conditions. This is consistent with previous reports where maximum NO emission rates coincide with maximum nitrification rates (Pinto et al., 2004). Gut et al. (1999) reported maximum NO flux of 50 g NO-N ha–1 d–1 after ammonium nitrate application and 26 g NO-N ha–1 d–1 after slurry application in wheat fields at field capacity. These reported rates are in the same range as those observed in our experiment (Fig. 2B) during the first 24 d when WFPS was near to field capacity (i.e., 70%). The relatively high NO-N flux measured after Day 24, when WFPS decreased down to 35 to 50%, is similar to the values reported by Akiyama et al. (2000) when WFPS was 30 to 42%. These authors measured maximum emission rates of 72 g NO-N ha–1 d–1 from urea application and 180 g NO-N ha–1 d–1 from a mixture of ammonium sulfate and urea.

It is interesting to point out that in the treatments with slurry (S and S + DMPP) and in the control, the ratio Ln NO/N2O was lower than in the treatments with mineral fertilizer (M and M + DMPP) at WFPS values higher than 50% (Fig. 3). As CO2 emissions indicate (Table 1) and will be further discussed, there was a respiration decrease in the mineral treatments. So, the comparatively higher respiration rates in the control and slurry treatments could have led to a situation of comparatively less oxygen availability, especially at WFPS values higher than 50%, resulting in an increase in N2O emissions by denitrification relative to NO emissions.


Figure 3
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  		Fig. 3. Effect of water filled pore space (WFPS) on the NO to N2O ratio. ({blacktriangleup}) Control (C), ({blacksquare}) ammonium sulfate nitrate (M), ({square}) ENTEC (M + 3,4-dimethylpyrazol phosphate [DMPP]), (•) slurry (S), ({circ}) slurry with DMPP (S + DMPP).

 
Zerulla et al. (2001) and Irigoyen et al. (2003) described a drastic reduction in the efficiency of DMPP at high temperatures (20°C) in terms of the duration of NH4+ content in soil. Our measurements indicate that ENTEC 26 is still able to reduce NO emissions 56 d after fertilization (p < 0.05) (Fig. 2B) in the same temperature range (18–22°C). Thus, NO emissions can be a sensitive indicator of the duration of the effect of DMPP applied as ENTEC 26, whereas, as previously commented, it showed no effect in reducing N2O losses.

Application of DMPP significantly reduced cumulative NO emissions for both slurry and mineral fertilizer treatments (Table 1). Cumulative losses accounted for 0.7 kg NO-N ha–1 in S treatment and 2.8 kg NO-N ha–1 in M treatment, which is 0.7% of the NH4+–N applied in the slurry and 2.8% of the N applied as ASN (Table 1). Application of DMPP reduced NO emissions 25% in slurry treatments and 44% in mineral fertilizer treatments (Table 1).

Carbon Dioxide
Several authors have reported a positive correlation between soil WFPS and CO2 emission rates when WFPS is lower than 60%, and a negative correlation when WFPS is higher than 60% (Davidson et al., 1998; Kiese and Butterbach-Bahl, 2002). In our study, we did not find a correlation between CO2 emission rates and WFPS or soil temperature (r2 < 0.1), although there was a general trend indicating that the highest CO2 flux occurred when WFPS was near 70% during the first 20 d and the lowest emissions occurred when WFPS was at its minimum (around 35%) on Day 56 (Fig. 4). Application of organic manure in grassland has been shown to enhance soil respiration in comparison with a zero-N control or with an inorganic fertilizer application (Jones et al., 2005). We observed enhanced CO2 emissions during the first 4 d after the slurry application in both S and S + DMPP treatments (Fig. 4), although cumulative emissions after 59 d were not significantly different from the control (Table 1). On the other hand, M and M + DMPP treatments showed lower CO2 emissions than the control from Day 4 to 38 (Fig. 4) when WFPS was higher than 50%, which resulted in significantly lower cumulative emissions (Table 1). This result is consistent with that of Bowden (2000) who also showed that, in the short term, inorganic N fertilization reduced CO2 emissions in a forest soil. In our experiments, lower CO2 emissions in the mineral treatments were coincident with the higher Ln NO to N2O ratios in these treatments compared with the control and slurry treatments when WFPS was higher than 50%, as previously discussed.


Figure 4
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  		Fig. 4. Carbon dioxide emission rates. ({blacktriangleup}) Control, ({blacksquare}) ammonium sulfate nitrate, ({square}) ENTEC, (•) slurry, ({circ}) slurry with 3,4-dimethylpyrazol phosphate (DMPP). The vertical bars indicate LSD at 0.05 between treatments for each sampling time.

 
Weiske et al. (2001) reported a decrease in CO2 emissions following the use of NIs (DMPP and DCD) in field experiments with summer barley, maize, and winter wheat after fertilization with ASN 26%. These authors argued that this might be due to an effect of NIs on C mineralization. In contrast, we did not observe any significant difference in CO2 emissions with DMPP application in slurry or mineral fertilizer treatments (Fig. 4, Table 1). Consequently, the results of the present study do not support the idea of an effect of DMPP on C mineralization. Weiske et al. (2001) also measured N2O and CH4 emissions, and reported that DMPP reduced the global warming potential by 30% when expressed as CO2 equivalents. This occurred because both CO2 and N2O emissions were reduced by DMPP and because the soil acted as a sink rather than as a source of CH4. In our study, we did not observe a significant effect of DMPP on the total global warming potential when expressed as CO2 equivalents corresponding to the effect of N2O + CO2 (Table 1). But when we studied the contribution to the global warming of N2O and CO2 separately, we could observe that DMPP significantly reduced the contribution due to N2O in the slurry treatments.


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
The application of cattle slurry to grasslands induced high N2O and significant NO and NH3 emissions. Application of DMPP reduced N2O and NO losses by 30 and 25% respectively, without significantly increasing NH3 losses. The application of ASN induced higher NO emissions than those associated with slurry. DMPP applied as ENTEC 26 reduced these emissions by 44%, but surprisingly did not significantly reduce N2O emissions. Carbon dioxide emissions were unaffected by the use of DMPP when applied with ASN or with slurry.


   ACKNOWLEDGMENTS
 
This project was funded by the Spanish Commission of Science and Technology (MCyT project number AGL2003-06571-CO2-02) and by the University of the Basque Country (9/UPV00118.310-13533/2001 and UPV/Compo-Agri SL UE A02/A04). S. Menéndez held a grant from the Ministerio de Educación y Ciencia of the Spanish Government (FPU, Programa Nacional). We would like to express many thanks to Leire Careaga and Azucena González for technical assistance.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 




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