of the area together with the historically parcelled rural soil distribution and the current high industrial pressure has dramatically increased rural soil prices, which does not contribute to enhance farmland availability. Efficient use of nutrients is one of the major keys of sustainable agricultural production systems because inefficient nutrient use not only results in excessive and potentially harmful losses to the environment, it also negatively affects economic performance of production systems (Oenema and Pietrzak 2002).

This review focuses on N2O emission especially in the Basque Country as related to grassland soil, considering the effect of management and mitigation options.

Keywords: denitrification, grassland, mitigation, nitrous oxide iNTRODUCTION

Nitrogen applied in fertilizers and manures is not always used efficiently by crops. To this respect, soil N availability can have a major impact on N2O fluxes, thus, N fertilization increases N2O fluxes relative to unfertilized controls (Smith et al., 1998).

There is still a big uncertainty in estimated nitrogen balances at different scales. The gap between input and output of N can be up to 40%. Denitrification is suspected to be the most important process that can explain this gap (Erisman and Sutton, 2007). N2O is generated by the microbial transformation of nitrogen during nitrification and denitrification processes in soils and manures, and is often enhanced where available nitrogen exceeds plant requirements. Regarding environmental aspects, N2O is involved in global warming and contributes to the destruction of parts of the ozone layer. It has a mean atmospheric residence time of more than 100 years (Prather et al., 2001). N2O is mainly produced by denitrification, especially under wet conditions (Oenema et al., 2005). The climatic characteristics of the Basque Country provide the favourable conditions for the growth of grasslands and the development of livestock enterprises. The climate is Atlantic, with an average temperature of 14°C and an annual precipitation ranging between 1200 and more than 2000 mm (http://www.euskalmet.euskadi.net). The soils studied were typical of the region, and classified as dystric Gleysol for the top 10 cm (2.2% fine sand, 41.6% coarse sand, 30.4% loam, 25.7% clay for the top 10 cm). They are poorly drained soils, which affects soil water content and gas diffusivity in soil, promoting anaerobic conditions for denitrification. In fact, we found high rates of emissions on clay soils, suggesting that N2O emission was regulated by soil aeration (Gregorich et al., 2005).

Initial studies in the Basque Country of Spain (Estavillo et al., 1994 and 1996) carried out using soil incubations determined that denitrification was the main way of N losses in grassland, being the greatest losses in spring and autumn following fertilization events, with denitrification rates up to 2 kg N ha-1 day-1. These are seasons in which heavy rainfalls occur coinciding with fertilizer applications. Annual losses were about 10-15% of N applied, reaching in occasions 20-37% of the N applied (Estavillo et al., 1994).Values similar to this rate have also been observed in other studies carried out later in the area, with 1.56 kg N ha-1 day-1 (Estavillo et al., 2002). In fact, denitrification potential measurements carried out in different soil types (grassland, forest and crop) showed that grassland soils presented the highest denitrification potentials due to their high organic matter content (Virgel et al., 1995).

There was a linear relationship between potential denitrification in grassland and soil microbial carbon. To this respect, the importance of microbial N pool in grassland was evident: distribution of N showed that microbial biomass N content was 10 to 30 times larger than the mineral N and 1 to 5 times higher than that extracted by herbage, with values of 120, 6 and 28 kg N ha-1 respectively for the fifth cut carried out in autumn (Estavillo et al., 1997). From 1997 we have been monitoring N2O fluxes in the field after fertilization events considering different treatments: fertilisation type, use of nitrification inhibitors and land use change. Field measurements took place the day the fertilizer was applied, the day after the fertilizer was applied daily for two weeks and then approximately weekly until fluxes diminished. A common methodology based on manually operated static chambers was used for N2O flux measurements (Merino et al., 2001). To minimize the effects of diurnal variation, sampling was always carried out in the morning. Nitrous oxide emissions were measured using a closed air circulation technique in conjunction with a photoacoustic infrared gas analyzer (Bruel and Kjaer 1302 Multi-Gas Monitor). Fluxes were calculated from the linear concentration increase in the chamber headspace with time. Alternatively, gas measurements were carried out by gas chromatography equipped with an electron capture 63Ni detector. An intercomparison of N2O analyses was conducted for equipments calibration.

Cumulative N2O emissions during the sampling period were estimated by averaging the rate of loss between two successive determinations, multiplying that average rate by the length of the period between the measurements, and adding that amount to the previous cumulative total. Table 1 shows cumulative N2O values for the studies considered referred to the period of study and its conversion to CO2 equivalents.

Usually higher percentage of N losses from the N applied were found from mineral fertilisation, although this trend was not consistent in all the studies. As mentioned above, the highest N2O losses from mineral or organic fertilisation observed in our studies took place in spring and autumn. These losses coincided with high soil water contents; in fact, denitrification rates greater than 0.2 kg N ha-1day-1 occurred mainly after fertilization events if the soil was wet (soil water content > 35%) and soil nitrate was high (>7mg N kg-1 dry soil) (Estavillo et al., 1994). The difference between the different fertilizers would be that the increase in nitrate content in soil after mineral fertilization produced immediate losses of N by denitrification, whereas the nitrate content in soil after slurry application was lower. Thus, in the case of slurry fertilisation, losses occurred during longer periods of time, due to the mineralization of N between fertilization events. We have found that although nitrate availability determines the potential denitrification, the soil must be also we for denitrification to occur (Del Prado et al., 2006; Estavillo et al., 1994). Merino et al (2001) also observed that the interaction between nitrate and WFPS on N2O emissions seems to indicate that the nitrate effect varies according to different ranges in WFPS. When these variables were considered in a purpose oriented experiment at laboratory scale, (Del Prado et al., 2006), we observed that under wet conditions (WFPS>60%) the fertilised soils gave cumulative N2O emissions 3.5 times greater than the unfertilised ones; on drier conditions (WFPS<60%), this difference was a factor of about 1.5. Fertilised and unfertilised soils under wet conditions resulted in about 10 times greater N2O emissions than soils with the same treatments but under dry conditions.

Grassland soils in particular have a high potential for mineralization and subsequent nitrification and denitrification (Merino et al., 2001). The microbial processes nitrification/denitrification involved in the production of N2O have been studied by means of soil incubations with acetylene (Merino et al., 2001). In this sense, N2O derived from cattle slurry applied to a grassland soil came in a greater proportion from nitrification rather than from denitrification, while for the mineral fertilisation most N2O came from denitrification.

Table 1. Grassland derived N2O emission measured at field expressed as cumulative emissions and CO2 equivalents (CO2e) for N2O emissions following the recommendations of the Intergovernmental Panel on Climate Change (1996)



Mg CÜ2eha-1






Merino et al., 2001









Merino et al., 2001















Macadam et al., 2003















Pinto et al., 2004












Merino et al., 2005











Menendez et al., 2006









Menendez et al., 2008

S (solids)



S (liquid)



*Values are expressed as percentage of the N applied emitted, which in the case of slurry is calculated with respect to the ammonium fraction applied. '-'-'Tillage took place immediately before fertilisation. (2)Tillage took place three days before fertilization.

*Values are expressed as percentage of the N applied emitted, which in the case of slurry is calculated with respect to the ammonium fraction applied. '-'-'Tillage took place immediately before fertilisation. (2)Tillage took place three days before fertilization.

High nitrification activity has been observed in the soil under study from both kinds of fertilisers (Merino et el., 2001; Merino et al., 2002; Pinto et al., 2004; Merino et al., 2005; Menendez et al., 2006). Typically, a period of about 5 days following slurry and mineral fertilization was necessary to decrease soil ammonium content from fertilised treatments to the amount found in unfertilized treatments. This decrease might have been due to immobilization, nitrification or plan uptake. Probably all the processes were occurring simultaneously, but in most cases, nitrification was the main process involved as no favourable conditions for volatilization occurred (Merino et al., 2002) or a proportional increase in soil nitrate was accounted. Also, N2O fluxes from mineral or slurry fertilisation have occurred mainly in the first week after application. Similarly, Comfort et al (1990) found that most of the N2O emission occurred within the first 5 days following injection of manure into the soil when CO2 evolution was greatest. On a global scale, CO2 is the most important greenhouse gas contributing to global warming, although most of the contribution from agriculture to potential greenhouse warming does not originate from CO2 but from N2O and CH4 (Izaurralde et al., 1997). This was confirmed after the measurements carried out in our experiments, where CO2 emissions were evaluated in grassland soils. We observed enhanced CO2 emissions during the first 4 days after the slurry application, although cumulative emissions after 59 days were not significantly different from the control (Menéndez et al., 2006).

In a study carried out in a grassland soil managed by a commercial farm, (Merino et al., 2001), N2O emissions were greater if mineral fertilizer was used instead of slurry fertilizer, even if a higher amount of N was applied as slurry than as mineral fertiliser, with cumulative losses over a year of 7.9 and 5.9 kg Nha-1 respectively. Emissions were minimum two months alter fertilisation, with background levels for slurry and mineral fertiliser of 4 and 2 g N2O-N ha-1d-1 respectively. Also in this case the seasons with the highest emission of N from the system were spring and autumn, with 4.3 for mineral and 2.0 kg N ha-1 for slurry respectively. The highest losses of N2O by denitrification were found after the application of mineral fertilization, while slurry acted as a slow release fertilizer which supplies nitrate continuously at a small rate. This was the main difference observed between the slurry and the mineral fertilization with respect to N2O emission. The determination of the kind of fertilization that produced greater N2O losses resulted from the interaction of different factors, such as management (date of application, distribution of fertilisers, heterogenecity of cattle slurry), edaphoclimatic conditions. Flechard et al (2007) also observed from a study carried out most European climatological zones that N2O measurements were extremely variable in time and in space at each site, depending on weather and management practices. In summary, those factors influencing soil microbial activity in soil, which in the case of slurry fertilisation is important due to the pulses of emissions occurring following mineralization of soil organic N. This slow release fertiliser effect observed in slurry fertiliser was simulated when nitrification inhibitors were applied together with mineral fertiliser or with cattle slurry. Adapting to social and policy demands of environmentally sustainable dairy production, some projects have been already developed in the territory regarding to study how different fertilizers (cattle slurry or inorganic fertilizer) amended with commercial nitrification inhibitors or slurry electroflotation techniques affect on nitrous oxide and CO2 emissions on a temperate grassland. Nitrification inhibitors used in the field have been proposed as management alternatives to reduce both nitrate leaching and denitrification, providing greater N availability to the sward. It is likely that the use of nitrification inhibitors (NIs) will have increased potential in the long-term, although it currently shows limited potential (Smith et al., 2007). DCD (diciandyamide) and DMPP (3,4 dimethylpyrazol phosphate) are the nitrification inhibitors that have been evaluated in our conditions. They act by delaying the bacterial oxidation of ammonia to nitrite in the soil by depressing the activity of Nitrosomonas bacteria in the soil. These NIs have been known to successfully reduce N2O emissions from mineral fertilisers (MacTaggart et al., 1997; Linzmeier et al., 2001). The mitigation potential of these inhibitors applied with slurry has seldom been studied, having been evaluated by our group even in terms of phytotoxic effects in clover (Macadam et al., 2003). Nitrification inhibitors applied in grassland soils efficiently delayed nitrification rates, keeping soil mineral N content as ammonium for a longer time (Merino et al., 2001; Merino et al., 2002, Macadam et al., 2003, Merino et al., 2005). The maintenance of soil mineral N in the ammonium form after NIs application led to a reduction in cumulative N2O emissions, showing a percentage of reduction of about 40-60% with both inhibitors. Nevertheless, phytotoxic effects on white clover were observed with the application of DCD, which did not appear with the application of DMPP. When white clover was grown with DCD in the growth chamber, it showed the same visual symptoms of phytotoxicity as those observed previously in the field. They consisted mainly of chlorosis and further necrosis at the border of the leaves. DCD application caused a reduction in clover yield and a nutrient imbalance in leaves. These results led us to recommend the use of DMPP instead of DCD in grasslands as this inhibitor prevented N2O emissions in identical amount, but without causing damage to clover. As mentioned above, DMPP has been widely studied if applied with mineral fertilisation, but few studies can be found considering its effect with slurry fertiliser. In this sense, and due to the decreasing mineral fertiliser amounts used by farmers in grasslands parallel to the increasing amounts of slurry applications in our region, application of DMPP with slurry was evaluated at different times of the year: spring and autumn. As previously mentioned, at this time of the year, periods of rainfall frequently coincide with warm temperatures, which may affect DMPP performance as a nitrification inhibitor. We proved that DMPP effect is lower during spring owing to the slightly higher mean soil temperature in that season, leading to a possible faster degradation of DMPP. In our field study in spring, DMPP had no more effect after day 14, in comparison to autumn, when DMPP delayed ammonium nitrification during 58 days (Merino et al., 2005). With respect to N2O losses, 69% of N2O emissions were decreased 22 days after slurry application in autumn. With respect to to CO2, we found lower emissions from mineral fertilization than from control treatment and slurry fertilization, with 11.8, 15.9 and 15.3 Mg CO2 ha-1 respectively. No effect of the nitrification inhibitor DMPP applied with mineral or with slurry fertiliser was found. Yield and botanical composition of the grass were not affected by the use of DMPP.

Adapting industrial processes of manure treatment is another action that has been evaluated in the Basque Country as a feasible way to decrease gaseous N emissions (Menendez et al., 2008). To this respect, the electroflotation process, aimed to decrease the volume of slurries from intensive livestock farms was considered. The industrial process consists basically of an electrolysis of the slurry catalyzed by iron which leads to the flocculation of the solid particles, giving as a final result a solid and a liquid fraction. Thus, the solid and liquid fractions derived from the process were applied on a grassland soil to study the influence of N2O and CO2 emissions. As a result, the solid and liquid fractions of electroflotation can be considered useful products as fertilizers. Both products caused an increase in grassland yield with respect to the original untreated slurry, with grassland N extraction being even higher after the solid fraction application. Regarding environmental concerns, if applied under temperate conditions, they do not modify the risk of global warming, with N2O and CO2 emissions caused by their application being of the same magnitude as those caused by the application of the original untreated slurry.

Tillage is one of the management variables that may enhance or retard emissions of greenhouse gases from agriculture (Ugalde et al., 2007). Tillage influences interactions between soil structure and biota, which in turn influences the stability of nitrogen within the soil matrix. Tillage of grassland by farmers to grow maize, Zea mays (L), as a summer crop is a common practice in our region. Maize is sown immediately after tillage, and N fertiliser is applied at the same time to boost maize growth. As spring has been determined as a season when great losses of N2O may take place, we conducted a field experiment in which the immediate effects on N2O emission following tillage and mineral fertiliser application were studied (Pinto et al., 2004) at this time of the year. Also, production of N2O was studied at different soil depths. In both the ploughed and unploughed treatments, the 0-10 cm layer was the major contributing layer to gaseous N production (Estavillo et al., 2002). A period of 3 days was necessary for the ploughed soil to reach the concentrations of the available mineral N originated from organic matter mineralization. This may have occurred as a consequence of mineralization following the incorporation of organic matter into the soil, which increases soil nitrogen availability and provokes important differences in the production and emission of N2O (Table 1). As result, a recommended practice to reduce N loss from land use changes involving tillage would be to avoid immediate fertiliser addition as there is an extra N supply from mineralization of organic matter at this time.

Variability should have been reduced in this compilation by use of a common methodology, and by the fact that emissions are considered at the level of grassland sited in the same edaphoclimatic area, ensuring an improved estimation of N2O emissions in the situations under study. Nevertheless, the highest variation was found for slurry treatment, the treatment that has received higher attention in our studies due to the environmental and economic advantages derived from its use at farm level. Besides, the uncertainty in the estimates coming from its use as fertiliser have required a deeper consideration by scientists. A wide variation was observed (Table 1), with ranges of 1.4 to 8.5 % of applied N for slurry fertilisation for 20 days in autumn.

The interaction of soil, climate and management systems needs further evaluation. Policies that support better agricultural land and fertiliser management practices are needed. Prudent management of N inputs, considering the type of fertiliser, time of application, use of nitrification inhibitors and land use and technology for slurries minimization can be an effective strategy to minimize N2O emitted from grassland.

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