3.1 Supply and quality of soil organic matter

Return of crop residues to the soil is standard practice in both organic and non-organic arable production; however, because of the more diverse crop rotations in organic production systems, larger and more diverse inputs of herbage from legume-based green manures, leys, cover crops (CC) and intercrops are returned to soil than in the non-organic cropping systems commonly used in Europe (Gattinger et al., 2012). The N content in the crop residues of legume-based systems is typically higher than that from non-legume systems (Watson et al., 2002). Commonly used external sources of organic inputs in organic cropping systems are animal manures and slurries, composts or biogas residues, and organic fertilisers based on animal manure or municipal waste (Løes et al., 2017). The great diversity of N mineralisation patterns among the organic fertilisers and crop residues result in a large variation in how much N and C are rapidly degradable in the organic inputs.

Through the application of organic amendments and various crop residues from arable and forage crops, C and N are applied to soil, and the soil organic matter (SOM) content is often higher in organic than in non-organic arable crop rotations: Marinari et al. (2007) reported about 40 % more total organic C a short time after the application of organic matter compared with mineral fertiliser; Marriott and Wander (2006) documented concentrations of SOC that were about 14 % higher in organic than in non-organic systems; Gattinger et al. (2012) showed 3.5 Mg C more in SOC stocks in organic compared with non-organic production in a global meta-analysis; Aguilera et al. (2013) reported a SOC concentration 19 % higher in a meta-analysis from the Mediterranean; Hu et al. (2018a) showed 0.4 Mg C ha⁻¹ yr⁻¹ more SOC accumulated with an organic than a non-organic treatment at Foulum, but 0.4 Mg less C than in the non-organic treatment at Flakkebjerg in long-term field trials; and Pimentel et al. (2005) documented 15 % higher SOC concentrations in legume-based organic versus non-organic crop rotation in a long-term field trial. The quality of SOM differs between non-organic and the more diversified organic, arable crop rotation, with a higher share of labile SOM (Lynch, 2015) and thus easily degradable organic matter (C and N) in soils in organic crop rotations (Marriott and Wander, 2006; Marinari et al., 2007; Martyniuk et al., 2016). The higher content of degradable SOM in organic crop rotations is a valuable soil fertility asset as it provides a short-term pool for plant nutrient supply (Marriott and Wander, 2006; Martyniuk et al., 2016). SOM turnover rates vary with soil texture and climate: they are higher when organic carbon is less protected from decomposers (low clay content) and in warm climates with suitable moisture (Burke et al., 1989).

3.2 Soil biological activity

Soil microbes contribute directly to plant residue decomposition and to the mineralisation and turnover of SOM, in addition to earthworms and other soil fauna (Kuiper et al., 2013; Lubbers et al., 2013). The inflow of degradable organic matter provides substrate for soil organisms, and the application of organic matter increases the growth of microbial communities, their enzyme activities and the microbial diversity compared with an unfertilised control or soil fertilised with only mineral fertiliser (Anderson and Domsch, 1989; Marinari et al., 2007; Thangarajan et al., 2013), although such changes in the topsoil may be slow processes (Petersen et al., 2013). Accordingly, higher biological activity has commonly been found in arable soils managed organically compared with non-organically managed soils (Mäder et al., 2002; Gomiero, 2013; Hartmann et al., 2015; Lori et al., 2017). In their meta-analysis, Lori et al. (2017) found that organic systems had 32 % to 82 % greater microbial biomass C, microbial biomass N, total phospholipid fatty-acids, and dehydrogenase, urease and protease activities than conventional systems. They found that when both organic and non-organic systems included legumes, the organic system displayed a higher microbial N content than the non-organic counterpart. In cases where only the organic systems contained legumes, the difference in microbial N between the two systems was even more pronounced. The abundance of earthworms can be twice as high in organic than in non-organic systems (Pfiffner and Mäder., 1997; Filser et al., 1999; Hansen and Engelstad, 1999; Riley et al., 2008). More abundant earthworm populations are found when large amounts of animal manure or green manure are applied to soil (Hansen and Engelstad, 1999; Frøseth et al., 2014), when autumn ploughing is avoided (Pfiffner and Luka, 2007) and in the absence of tractor traffic (Hansen and Engelstad, 1999).

3.3 Soil N-dynamics within organic arable cropping systems

Crop N supply and SMN in organic farming relies largely on mineralisation of N in soil organic matter, N in organic amendments and crop residues and BNF of legume-based crops (Gattinger et al., 2012; Lorenz and Lal, 2016). A principal sketch for N dynamics in organic systems is given in Fig. 1 and an overview of performance of selected drivers of SMN in organic and non-organic crop rotations are given in Table 1.

Figure 1The principle of N dynamics in organic crop rotations.

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Table 1Performance of selected drivers of SMN in organic and non-organic cropping systems.

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Despite substantial inputs of N from biological nitrogen fixation (BNF; Kayser et al., 2010; Pandey et al., 2017) and from organic amendments, the N supply is often below optimum for plant growth in arable organic farming (Berry et al., 2002; Tuomisto et al., 2012). BNF may be a more important N source than N from organic manures and fertilisers in organic arable cropping systems (Pandey et al., 2018).

The major release of plant available N from added organic matter depends on the C:N ratio, the mineral N content and the degradability of the C of added plant residues manure, compost and other decomposed amendments. Bhogal et al. (2016) showed that for pig slurry and poultry layer manure with a C:N_org of 9–12:1, up to 70 % of the organic N was mineralised after five growing seasons, whereas in cattle slurry and straw-based farmyard manure with a C:N_org of 10–21:1, only 10 %–30 % of N was mineralised (net). For crop residues with a high C:N content there will be initial immobilisation and the net mineralisation may only start very late or after the growing season of the main crop (Li et al., 2015b). The mineralisation process of the more stable N can continue over years to decades. Simulation modelling has shown that even over a 20-year period, only 10 %–15 % of organic N in applied manure may be taken up by crops, with the rest being lost or retained in soil organic matter (Berntsen et al., 2007). The high microbial activity and high content of organic matter affect N cycling. Hu et al. (2018b) found a higher net N mineralisation of added organic matter in soils with a prehistory of CC use, indicating positive legacy effects of CC use, which was attributed to a greater microbial biomass N in systems with the use of CC.

The design of the rotation, as well as its management, influences the potentially mineralisable N (PMN) pool. Working in three different organic arable systems, Spargo et al. (2011) showed that the PMN pool amounted to 315 kg N ha⁻¹ on average. They showed that more diversified crop rotations resulted in larger PMN pools. Poudel et al. (2002) reported a 112 % and 56 % greater PMN pool in the organic system in comparison to the conventional and low-input systems, respectively. Moreover, they observed a slower and more continuous release of mineral N in the organic systems compared with the more rapid release of mineral N from synthetic N fertilisers applied in non-organic systems. Moyo et al. (2016) reported higher PMN in soils under wheat following a cut and mulched red clover ley compared with after a ley where the residues had been removed. This indicates the importance of total N-input to the soil. The importance of total N-input in plant residues for crop N uptake was also observed by Petersen et al. (2013).

SMN was found to be very low under grass–clover leys (Table S1, Watson et al., 1993; Nadeem et al., 2012; Brozyna et al., 2013; Frøseth et al., 2014; Krauss et al., 2017b), as grasses quickly take up soil NO₃ in the root zone (Brophy et al., 1987). Frøseth et al. (2014) observed low levels of SMN in a 1-year grass–clover ley, irrespective of whether the herbage was mulched or removed. After termination of a ley, the concentration of SMN usually increases (Table S1, Ball et al., 2007; Brozyna et al., 2013; Krauss et al., 2017b). Even in the year following the termination of a ley, the content of SMN can still be high (Hansen et al., 2007; Jończyk and Martyniuk, 2017). Kayser et al. (2010) pointed out that N provided by spring ploughing of both a 1-year grass–clover ley and a 3-year grassland ley resulted in high concentrations of SMN (0–90 cm, 61 kg N ha⁻¹ and 95 kg N ha⁻¹, respectively) in the following autumn after harvest of spring triticale. Much of this SMN may not have been available for the crops during spring and is likely to have been mineralised after the end of the growing season. In contrast, low concentrations of SMN were observed in topsoil and subsoil after spring ploughing of a 1-year grass–clover ley in four field trials in Norway, in late spring and in autumn after the harvest of a spring barley crop (Table S1) and in the following year (Frøseth et al., 2014; Frøseth, 2016).

4.1 Mechanisms for N₂O emissions

Many processes contribute to N₂O production in soils, but the dominant mechanisms for N₂O emissions from terrestrial agricultural soils are the microbial processes of nitrification, nitrifier denitrification (as a result of incomplete nitrification) and denitrification (Firestone and Davidson, 1989; Butterbach-Bahl et al., 2013). Nitrification and denitrification are both biological processes; thus, the same mechanisms will cause N₂O emissions in organic and non-organic farming systems. However, fertilisation, and crop and soil management practices differ substantially between these two systems (Sect. 3), and the relative importance of the relevant triggers consequently differ (Sect. 6).

Nitrification is the microbial oxidation of NH₃ to NO₂ and ultimately NO₃, where N₂O is produced as a by-product via some partially understood biotic and abiotic reactions of hydroxylamine (Anderson, 1964; Liu et al., 2017). Nitrifier denitrification occurs when NO₂ produced during nitrification is reduced to N₂O (by denitrifying organisms), instead of being oxidised to NO₃, under fluctuating oxic–anoxic conditions (Firestone and Davidson, 1989). Denitrification is the microbial anaerobic reduction of NO₃ via NO₂ to gaseous NO, N₂O and N₂, which are ultimately transported through the soil to the atmosphere. Denitrification is the main source of N₂O production in soils, as the N₂O yield potential of denitrification is much higher (1 %–100 %) than that of nitrification (0.1 %–1 %) (e.g. Andersson et al., 1993; Butterbach-Bahl et al., 2013).

The ratio between the gaseous products of denitrification depends on NO₃ availability, oxygen availability in the soil, the amount of easily decomposable carbon as an energy source, soil pH and the microbial community structure (Bakken et al., 2012). Oxygen availability depends on soil microbial activity and gas diffusivity, which, in turn, depends on soil moisture content, texture and density. Gas diffusivity is a promising predictor for N₂O fluxes from soils with varying bulk density as observed by Balaine et al. (2013), who found that the production of N₂O increased when the relative gas diffusivity was between 0.006 and 0.020 and the soil became anaerobic.

As explained above, the risk of N₂O emission increases as the soil carbon content increases (Li et al., 2005). N₂O and N₂ production correlates with total organic C, water soluble C and mineralisable C in soil, but the increased availability of C also decreases the ratio of N₂O:N₂ (Sahrawat and Keeney, 1986). Emissions from soil organic matter (background emission) will vary between years because of variations in temperature and precipitation (Brozyna et al., 2013; Hansen et al., 2014) and between different categories of crops due to different time windows with high SMN (Dobbie and Smith, 2003).

Organic amendments and plant residues that provide carbon that is easily decomposable by microbes may enhance microbial activity and deplete soil oxygen via enhanced soil respiration. In addition, degradable carbon is an energy source for denitrifying bacteria. In accordance with this, Köster et al. (2011) concluded that bacterial denitrification was the main process for producing N₂O during the first 3 weeks following the application of biogas residues, and high carbon availability was an important cause of this. Li et al. (2016) concluded that denitrification was the main cause of N₂O emission after the addition of legume-based residues. Several studies have shown higher rates of N loss through denitrification from soils treated with organic amendments such as manure, composts and plant residues when compared with unamended or mineral N treated soils (Thangarajan et al., 2013). In line with this, the incorporation of residues by tillage increases soil respiration and N₂O fluxes due to microbial stimulation (Krauss et al., 2017a).

Low soil pH inhibits the activity of the N₂O reductase enzyme and, thus, the N₂O:N₂ ratio increases (Liu et al., 2010). At higher soil pH, the denitrification rate is higher, but the N₂O:N₂ ratio is lower as a greater part is completely denitrified to N₂. At low temperatures, nitrous oxide reductase is hampered (Holtan-Hartwig et al., 2002), but, conversely, denitrification rates are also reduced (Butterbach-Bahl et al., 2013).

4.2 Legumes during active plant growth

In general, unfertilised legumes have small N₂O emissions during their growing period, particularly when grown in mixtures with non-legumes. Low N₂O emission are found during growth of grain legumes (Rochette and Janzen, 2005; Dusenbury et al., 2008; Pappa et al., 2011; Jensen et al., 2012; Jeuffroy et al., 2013), green manure crops and CC (Baggs et al., 2000b; Brozyna et al., 2013; Li et al., 2015a; Peyrard et al., 2016; Shelton et al., 2018) as well as for grass–clover leys (Baggs et al., 2000b; Ball et al., 2002; Nadeem et al., 2012; Brozyna et al., 2013; Krauss et al., 2017b; Skinner et al., 2019). This is consistent with low SMN concentrations during growth (Sect. 3) and negligible N₂O emissions associated with BNF by the legume–rhizobium symbioses (Rochette and Janzen, 2005; Carter and Ambus, 2006). However, following termination or senescence of the legume crops, reactive N released from dying roots and nodules may lead to enhanced N₂O emission (Rochette and Janzen, 2005).

Surface mulching of harvested herbage may theoretically enhance N₂O emissions due to mineral N released from the herbage. However, several studies have shown that mulching of grass–clover herbage on the growing ley only causes a slight increase in N₂O emissions (Möller and Stinner, 2009; Nadeem et al., 2012; Brozyna et al., 2013). None of these studies measured ammonia volatilisation from mulched herbage, which could have been a major loss of mineralised N corresponding to the findings of Larsson et al. (1998). Volatilised NH₃ will be redeposited elsewhere and may result in increased N₂O formation downstream (IPCC, 2006).

4.3 Crop residues

As outlined in Sect. 4.1, there is an enhanced risk of N₂O emissions from agricultural soils when easily degradable carbon and N are simultaneously available, and denitrification is probably the main source for this. Because legume-based crop residues also increase SMN (Sect. 3.3), increased N₂O emissions have been reported in field trials whether the residues are from grain legumes (Pappa et al., 2011; Jeuffroy et al., 2013), grass–clover (Baggs et al., 2000b; Ball et al., 2007; Nadeem et al., 2012; Brozyna et al., 2013; Skinner et al., 2019), intercropped clover (Pappa et al., 2011) or CCs (Baggs et al., 2000b; Peyrard et al., 2016; Pugesgaard et al., 2017). However, the increase of N₂O fluxes after the incorporation of crop residues and other plant material might be small and a negligible part of the total N₂O emissions (Peyrard et al., 2016; Pugesgaard et al., 2017; Shelton et al., 2018).

The C:N ratio of incorporated herbage does affect N₂O emissions, with higher emissions expected from herbage with a low C:N ratio (Chen et al., 2013). From this, one should expect higher N₂O emissions from legume residues than from cereals or grasses (e.g. Rochette and Janzen, 2005). However, Larsson et al. (1998) observed the same N₂O–N emission factor (EF; 1 % of applied N) from mulched alfalfa (C:N ratio of 11) as from mulched grass with a C:N ratio of 21, but a higher EF than from a mulched grass with a low N-content (C:N ratio of 36, EF =0.1 %). The N₂O fluxes might be high despite a high C:N ratio when the carbon source is easily degradable as observed with fodder radish by Li et al. (2015a) (Table S2).

N₂O emissions may also be associated with the previous incorporation of plant residues. In accordance with this, Skinner et al. (2019) observed enhanced N₂O fluxes after a maize crop succeeding a grass–clover ley. Measurements of N₂O fluxes shortly after the incorporation of plant material, or measurements in the following year, only tell part of the story. The enhanced content of various fractions of SOM derived from crop residues, ley and CC (Sect. 3.2) are likely to increase the long-term background emissions of N₂O (Sect. 4.1). In a 10-year-old field experiment with and without legume-rich CC in the crop rotation, Pugesgaard et al. (2017) concluded that crop residues were an important source of N₂O, and that mineralisable C, rather than N input, was the main driver for N₂O emission. Contrary to this, Peyrard et al. (2016) observed that although N₂O fluxes increased for a few days after incorporation of CC, the contribution of such events to cumulative N₂O emissions were negligible in a 3-year low-input field trial. In their study, however, the CC treatments started when the N₂O measurements started. More studies in long-term experiments with the continuous use of CC are needed to verify the actual impact of crop residues in a long-term perspective in various field situations, as the addition of plant material to soil also affects the soil structure, the soil biological activity and N turnover.

4.4 Organic fertilisers

Organic fertilisers vary widely in the content and types of N and C compounds causing large variations in N₂O emissions after application. Animal slurries have a higher content of NH₄–N and contain more easily degradable N and C than solid manures and composts; thus, they are stronger triggers for rapid N₂O emissions shortly after application (Charles et al., 2017, Sect. 3). In accordance with this, Krauss et al. (2017b) observed higher N₂O emissions shortly after the application of cattle slurry than after the application of composted solid cattle manure. Correspondingly, in a field experiment with spring barley fertilised with various organic slurries, Baral et al. (2017) observed the highest N₂O EF in the treatment with the highest application of organic matter, and thus the highest content of easily degradable C. Meijide et al. (2007) and Chantigny et al. (2007) found that the use of digested slurry, with a lower content of degradable C compared with untreated pig slurry, reduced soil N₂O emissions by 25 % and 50 %, respectively.

The effect of organic fertilisers depends on soil type and the content of SOC. Degradable C applied with organic fertilisers will trigger microbial respiration and denitrification in a soil with a low SOC content to a greater extent than it would in a soil with a high SOC content (Chantigny et al., 2010; Pelster et al., 2012), whereas the impact of easily available N is higher in a soil with a high SOC content (Petersen et al., 2008). Due to the higher SOC content and higher share of labile SOM in organic crop rotations compared with non-organic crop rotations (Sect. 3.1), the short-term effect of organic fertilisers with a high content of degradable C on N₂O emission are likely lower in organic than in non-organic crop rotations.

Also, the absence of synthetic fertilisers containing easily available N means that organic fertilisers are likely to have a smaller short-term impact on N₂O emissions in organic than in non-organic crop production. In a meta-analysis, Charles et al. (2017) found that the N₂O EF was higher when soils received organic amendments in combination with synthetic fertilisers. They found EFs for liquid manures + synthetic fertilisers of 2.14 % (±0.53), for composts + synthetic fertilisers they reported 0.37 % (±0.24), and the corresponding EFs for manure and compost were 1.12 % (±0.18) and 0.00 % (±0.17), respectively.

However, the long-term impact of manures is not included in these EFs. In contrast with the short-term fertiliser effect, a long-term fertilisation with organic fertilisers may enhance N₂O emissions through enhanced background emissions. Chang et al. (1998) observed that annual N₂O emissions increased with manure rate when different rates of solid feedlot manure and thus N application were applied for 21 years. Their manure rate was far greater than would be applied under organic farming conditions, but the possibility for enhanced background N₂O emissions after long-term input of organic matter via manuring and the application of crop residues should be considered. Krauss et al. (2017b) found that fertilisation with slurry and manure compost increased annual N₂O emissions during winter wheat after more than 10 years of differentiated management compared with sole slurry fertilisation (mean values in the period, 369 d, were 2.2 and 2.9 kg N₂O–N ha⁻¹, respectively). They related this to higher microbial biomass and SOC. Mean values for the upper 10 cm of soil were 28 and 30 Mg C ha⁻¹ for fertilisation with slurry and manure compost and sole slurry fertilisation, respectively.

4.5 Contribution of total N-input and high emission events to N₂O emissions

Skinner et al. (2014) concluded in a review that soil characteristics (soil N content) had a greater impact on N₂O emissions from organic production than the total N-input by fertilisation. Lack of correlation between N₂O emissions and N fertilisation in organic production corresponds to the more recent findings of Krauss et al. (2017b) and Pugesgaard et al. (2017). Pugesgaard et al. (2017) observed no significant correlation between N₂O emissions and N input in fertiliser/manure, for either annual N₂O emissions or spring emissions, but N₂O emissions were correlated with N input in residues from the previous main crop and CC. This agrees with the findings of Bouwman et al. (2002), Van Groenigen et al. (2010), Shcherbak et al. (2014) and Peyrard et al. (2016), who observed low EFs when N fertilisation was below optimum as commonly found in organic production systems.

As discussed earlier, N₂O emissions in organic crop rotations are driven by enhanced background emissions from the long-term input of organic matter and episodes with enhanced N₂O emission after the application of crop residues or organic fertiliser. However, this poses the question as to whether N₂O emissions are primarily driven by background emissions or by episodes with high N₂O fluxes. We used data from mainly organic field trials (Table S2) and one non-organic trial fertilised with organic fertilisers (Baral et al., 2017) to calculate the impact of the highest N₂O fluxes on the total N₂O emissions. In Frick (CH), Edinburgh (UK), Aberdeen (UK) and Ås (NO), the highest daily flux rates were 605, 211, 297 and 94 g N₂O–N ha⁻¹ d⁻¹, respectively (Table S2). Because of the high flux rates, we hypothesised that high flux events were responsible for a major part of the N₂O emissions from these systems. The single days with the highest fluxes correspond to 18 % (65 d measurement period) in Frick, 2 % in Edinburgh (161 d), 17 % in Aberdeen (38 d) and 2 % in Ås (218 d) of the cumulated N₂O emissions in the measurement periods. The five highest daily N₂O fluxes corresponded to 22, 7, 55 and 5 % of the N₂O emission in the measurement periods in these investigations, respectively. However, in field trials conducted on well-structured sandy loams at either Foulum or Flakkebjerg in Denmark (Table S2, 105 to 365 d), peak N₂O fluxes from 1 or 5 days only constituted from < 1 % to 8 % and 5 % to 14 % of the total emission in the periods, respectively. The highest daily flux rate in these trials was only 78 g N₂O–N ha⁻¹ d⁻¹. This was in a non-organic treatment heavily fertilised with cattle slurry and digested sewage sludge (476 kg total N ha⁻¹, Table S2; Baral et al., 2017). From this we reject the hypothesis that high flux events were responsible for a major part of the N₂O emissions from these systems, rather background emissions seemed to be the major N₂O source. However, a simple regression model (Eq. 1) showed that the average daily N₂O flux correlated positively with episodes with high N₂O fluxes. From this we can conclude that when the conditions for high N₂O fluxes are met for 1 or more days, there is a large chance of high total N₂O emissions in the period. The small peaks in the Danish field trials reveal that well-drained sandy soils promote rapid water infiltration and good gas-diffusivity and, in turn, low N₂O emissions.

We wanted to explore the impact of soil conditions on the highest N₂O flux peak. In a stepwise regression (Eq. 2), the content of clay, SOC, NO₃–N in soil and soil temperature had significant positive impacts on peak N₂O fluxes in the selected investigations shown in Table S2 (P_clay, P_SOC < 0.001, $P_{{\mathrm{NO}}_{\mathrm{3}}}$, P_temperature < 0.01, P_WFPS < 0.05). The content of NH₄–N in soil did not affect peak N₂O fluxes. These findings indicate that denitrification is the main cause of high N₂O-flux rates in these studies. To visualise the impact of different factors on the mean and the highest N₂O flux and the contribution of the five highest daily fluxes to the cumulated N₂O emissions in the measurement period, we grouped the investigations according to the percentage (%) of clay in soil, N added with organic fertilisers, SOC, soil pH, experimental period, type of crop and mean daily precipitation in the period (Fig. 2). Because of the lack of information, we were neither able to include total N-input nor soil porosity in the regression analyses and the box plots.

Figure 2Box plot (the lines in the boxes indicate the mean, the crosses indicate the median, and the whiskers indicate the upper and lower quartile) for (a) the mean N₂O flux in the period (g N₂O–N ha⁻¹ d⁻¹) and the highest daily flux rate in the same period (g N₂O–N ha⁻¹ d⁻¹), and (b) the percent (%) contribution of the five highest daily fluxes to cumulated N₂O emissions in the period. N denotes the number of sites, and n is the number of observations.

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In line with the findings of Skinner et al. (2014), we did not find a clear impact of N added with organic fertilisers; however, the soils with the lowest SOC content showed the lowest mean and lowest peak N₂O emissions. This supports the hypothesis that the SOC content of the soil, and thus the content of SOM, are important drivers for N₂O emissions. At high SOC contents other factors seem to be more important. The impact of the type of crop on N₂O emissions mainly follow the trends we have seen in other investigations (Sect. 5.2, 5.3): higher N₂O emissions when cereals succeed a short-term ley (grass–clover or clover), and mixtures with monocotyledons and legumes do not have higher N₂O emissions than cultures with only monocotyledons. The high residual effects of previous leys support the idea that background emissions are the main driving force for N₂O emission in organic crop rotations. The high N₂O emissions (up to 600 g N₂O–N ha⁻¹ d⁻¹) from one occasion in Switzerland with the incorporation of weeds in wet soil have a large impact on all box plots (Fig. 2). None of the other groupings in Fig. 2 showed any clear causal effect on N₂O emissions.

The low effect we observed of pH on N₂O emissions (Eq. 2, Fig. 2) is contradictory to enhanced N₂O emissions caused by a hampered N₂O reductase enzyme often observed in acid soils (Bakken et al., 2012). A reason for this could be that in these soils and with these management methods, other factors meant more for N₂O emissions than pH. Although denitrification was likely to have been the main cause of N₂O fluxes at the highest N₂O flux rates, the soil NO₃ concentrations are often low between the episodes with high N₂O flux rates, as observed by Chirinda et al. (2010), Nadeem et al. (2012), Brozyna et al. (2013), Li et al. (2015a) and Pappa et al. (2008).

4.6 Impacts of earthworms

Abundant earthworm populations in organic crop rotations (Sect. 3.3) are likely to influence N₂O fluxes as they significantly affect mineralisation and reduction of N compounds to N₂O and N₂ (Prieto, 2011). N₂O is emitted from intestinal microbes but is also released from nitrates emitted in body fluids in the earthworm gut as well as from casts, middens and burrows (Prieto, 2011). Conversely, earthworms improve soil porosity and aggregate stability (Bronick and Lal, 2005) and thus gas diffusivity and water infiltration in soils, which will reduce N₂O emissions. Epigeic species (living near the surface and feeding on surface litter) and anecic species (deep burrowing) are well known to enhance N₂O production, because they feed directly on decomposing herbage (Evers et al., 2010; Lubbers et al., 2011; Nebert et al., 2011). Endogeic earthworms that feed on SOM particles are most common in cultivated arable soils (Hansen and Engelstad, 1999), and they do not increase denitrification (Postma-Blaauw et al., 2006). There are too few published results to robustly predict the impact of earthworms in arable organic crop rotations on N₂O emissions as this will depend on local climatic and edaphic conditions.

5.1 Mechanisms for NO₃ leaching

NO₃ leaching is an abiotic process driven by diffusion and convection (e.g. Johnsson et al., 1987), where NO₃ is transported out of the root zone along with the downward water flow. In addition to soil water content, soil texture and structure are important in determining leaching rates. Fine textured soils have slower infiltration rates than coarse textured soils, and porous sandy soils are most vulnerable to leaching, also because these soils often have more shallow rooting depths than loamy soils (Askegaard et al., 2005). The impact of soil type was clearly demonstrated by Askegaard et al. (2011) who found that, depending on soil type (coarse sand > loamy sand > sandy loam) and precipitation, 20–100 kg N ha⁻¹ yr⁻¹ were leached on average for the crop rotations. In their study, the location on coarse sand had 200–300 mm more rainfall per year than the other locations. The leaching was considerably higher than for Swedish clay soils which showed a value of 20 kg N ha⁻¹ yr⁻¹ (Stenberg et al., 2012).

Due to its high mobility in soil, NO₃ can easily be lost from the agroecosystem by leaching during periods with high drainage rates. A well-developed active root system enhances NO₃ uptake, whereas a poor root system will not utilise all of the NO₃ within the soil profile (Dunbabin et al., 2003). NO₃ remaining in the soil after the growing season of crops, or subsequently mineralised, will greatly increase the risk of leaching loss. This could occur outside of the growing season, but also when there is poor crop establishment caused by unfavourable seedbed structure or from crop failure caused by diseases or pests (Stenberg et al., 2012). If crop failure coincides with rainy weather, the risk of severe NO₃ leaching is large. This was observed by Torstensson et al. (2006) and De Notaris et al. (2018) in organic farming systems, where potato growth was restricted due to early crop termination following disease outbreaks. Annual leaching was 75 kg N ha⁻¹ when potato was cultivated in the year after green manure compared with 98 kg N ha⁻¹ after pea/barley (Torstensson et al., 2006; Table S3), whereas De Notaris et al. (2018) observed substantially higher leaching rates. They measured 213 kg N ha⁻¹ yr⁻¹ leached when the potato followed green manure and 133 kg N ha⁻¹ yr⁻¹ after grain legumes in a year with early occurring potato late blight. This was substantially higher than in previous years (140 and 78 kg N ha⁻¹, respectively).

Extreme rainfall events and/or periods of drought can significantly affect leaching for a variety of reasons. A field experiment over 13 years in the UK showed that N leaching in winter from fertilised grass (non-organic) was highly correlated with the preceding summer's soil moisture deficit, with the highest losses following dry summers (Tyson et al., 1997). In this case, poor grass growth due to drought led to a build-up of NO₃ from unused fertiliser present in the autumn. Prolonged mineralisation of organic fertilisers or crop residues due to drought may also lead to a similar situation in organic farming systems. Tosti et al. (2016) found, under Mediterranean rain-fed conditions, that the risk of NO₃ leaching was mainly at the onset of drainage due to rainfall, i.e. at the initial stage of growth, and was typically variable among years depending on the timing of heavy rains. Thus, amendments applied at the pre-crop stage would be a risky practice for NO₃ leaching. Most N leaching studies in organic farming in Mediterranean environments focused on row and vegetables crops (e.g. Campanelli and Canali, 2012), because these systems are most demanding in N inputs and thus have higher N applications and potential leaching than in common arable crops.

Figure 3Lower and upper level of NO₃ leached (kg NO₃–N ha⁻¹ yr⁻¹) in various field investigations: with or without a cover crop (CC), different amounts of N applied, and with or without grass–clover ley. “s.” refers to summer, and “w.” refers to winter. The data used are given in Table S3.

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5.2 Legumes

In a crop rotation with a large contribution from BNF, some of the N inputs will be retained in crop residues and in particular in mulched green manures (Frøseth et al., 2014). In their review, Crews and Peoples (2005) found that when the N input was based on BNF, the proportion of the N retained in the soil was higher (58 % of legume N) than in the fertilised systems (31 % of fertiliser N). From this, it can be assumed that the risk of N release outside the growing season are high in rotations with legumes. However, in their meta-analyses of crop yield and N dynamics as influenced by CCs, Tonitto et al. (2006) concluded that NO₃ leaching was reduced by 40 % on average in legume-based systems relative to conventional fertiliser-based systems. The reason for this is probably the large difference in N input between legume-based systems relative to conventional fertiliser-based systems. The response of NO₃ leaching to N input in fertiliser, manure and residues may also differ between sites due to soil type and precipitation (Pandey et al., 2018).

5.3 Cover crops

CCs are grown between main crops to minimise NO₃ leaching. Many field trials in non-organic systems have shown reduced leaching using CCs (e.g. Rasse et al., 2000; Torstensson and Aronsson, 2000; Constantin et al., 2010; Valkama et al., 2015). This is also the case for organic crop rotations (Tonitto et al., 2006; Askegaard et al., 2011; Tosti et al., 2014, 2016; De Notaris et al., 2018). The reduction in NO₃ leaching can be substantial. Studies in Nordic countries report reductions of 50 %–60 % in N leaching (Torstensson and Aronsson, 2000, Fig. 3; Askegaard et al., 2011; De Notaris et al., 2018). If the cash crop fails, the effect of CCs on reduced leaching can be even higher. In a year when potato late blight caused crop failure in potato, the CCs reduced N leaching by 95 % when the potato succeeded a grain legume (from 133 to 6 kg N ha⁻¹ leached), and by 92 % when the potato succeeded a green manure ley (from 213 to 17 kg N ha⁻¹ leached; calculated from Table S3, De Notaris et al., 2018).

De Notaris et al. (2018) concluded that the use of CCs had a larger impact on leaching than a substantial variation in N surplus between alternative cropping systems. In three long-term field trials (13–17 years) in northern France, Constantin et al. (2010) observed that CCs were the most efficient management option for reducing NO₃ leaching (from 36 % to 62 %). Good establishment and growth of the CC is essential to obtain sufficient uptake of SMN and thus reduce NO₃ leaching. Stenberg et al. (1999) found no significant reduction in NO₃ leaching during winter from a ryegrass CC that was undersown in spring. They explained this by poor CC establishment. De Notaris et al. (2018) also observed occasions with a very small effect of CCs on NO₃ leaching. They related this to CC growth and identified threshold values in CC above-ground biomass determined in November, above which N leaching was reduced to a stable low level. NO₃ leaching from spring wheat averaged 15 kg N ha⁻¹ yr⁻¹ with CC biomass above 0.9 Mg ha⁻¹, and 41 kg N ha⁻¹ yr⁻¹ with CC biomass below 0.9 Mg ha⁻¹. In potatoes, the average N leaching was 11 and 80 kg N ha⁻¹ yr⁻¹ with CC biomass above and below 1.5 Mg ha⁻¹, respectively.

Including legumes in CC mixtures does not seem to reduce the ability of CCs to reduce NO₃ leaching (Tonitto et al., 2006; Tosti et al., 2014; De Notaris et al., 2018; Shelton et al., 2018). In a field trial with barley, hairy vetch and a 50:50 mixture of both species as CC, Tosti et al. (2014) found that, in all years, the barley–vetch mixture decreased N leaching to the same level as pure barley, both during its own growing cycle and after CC incorporation into the soil. De Notaris et al. (2018) concluded that the same degree of reduced NO₃ leaching was obtained with legume-based CCs as with non-legume CCs. The CC was either undersown in spring or after the harvest of the main crop. The undersown legumes were white clover and red clover, and winter vetch was used in the mixture sown after harvest. Shelton et al. (2018) found greater NO₃ leaching from intercropped hairy vetch than from simultaneously grown wheat and wheat–hairy vetch mixture. When the CC is a pure stand of legumes, the CC does not necessarily reduce N leaching (Tosti et al., 2014; Valkama et al., 2015; Shelton et al., 2018). Tosti et al. (2014) concluded that hairy vetch sown as a pure crop in autumn showed high BNF, but no NO₃ leaching mitigation effect compared with bare soil. Valkama et al. (2015) found in their meta-analysis of Nordic studies of undersown CCs that legumes (white and red clovers) in pure stand did not diminish the risk of NO₃ leaching.

5.4 Tillage

Tillage stimulates soil N mineralisation, at least in the short term. Tillage is also often associated with soil incorporation of plant residues that may lead to net N immobilisation or mineralisation, depending on the quality of the residues. Therefore, timing of tillage is crucial for the fate of the mineralised N, whether SMN becomes available for crop uptake or is subject to leaching. In general, incorporation should consider soil type, climate conditions and type of herbage (C:N ratio). Thorup-Kristensen and Dresbøll (2010) suggested the late incorporation of CCs in high-rainfall areas on sandy soils, and earlier in low-rainfall areas on NO₃ retentive soils. Field studies have shown rapid N mineralisation from N-rich plant material, even at low temperatures (Breland, 1994; Thorup-Kristensen and Dresbøll, 2010). Thus, spring incorporation has been recommended to increase N recovery by subsequent crops. However, under Scandinavian conditions, there may still be a deficit in crop-available N, even after spring incorporation of a green manure ley (Känkänen et al., 1998; Frøseth et al., 2014). Under such conditions, Torstensson and Aronsson (2000) suggest that late autumn incorporation of CCs, instead of spring incorporation, would be preferable with respect to N availability for the subsequent crop and would not substantially increase NO₃ leaching. Under Mediterranean climatic conditions, characterised by mild, rainy winters and warm to hot, dry summers, there is a risk of NO₃ leaching if residues are incorporated prior to the wet season. No studies were found that measured NO₃ leaching with respect to the timing of tillage in organic arable farming under these conditions.

In organic crop production, the timing of cultivation for mechanical control of perennial weeds may conflict with the aims of high N use efficiency (Melander et al., 2016). Askegaard et al. (2011) found, on sandy soils in Denmark, that the management of crop and soil during autumn was the main determinant of N leaching. Stubble harrowing in autumn for controlling perennial weeds, followed by bare soil during winter, led to an average of 25 kg N ha⁻¹ more leached than for soils left untouched with a cover of weeds/volunteers. NO₃ leaching increased with increasing number of autumn soil cultivations.

Reducing tillage intensity may also enhance the need for weed management, and thereby the risk of NO₃ leaching. In their meta-analysis comparing different reduced tillage intensities in organic farming, Cooper et al. (2016) found that the weed incidence was consistently higher, by about 50 %, when tillage intensity was reduced, although this did not always lower the yields. Compared with conventional tillage, reduced tillage may reduce NO₃ leaching, but this depends on the establishment and growth of the succeeding crop (Känkänen et al., 1998).

Bare fallow has traditionally been a method to control perennial weeds by repeated tillage of superficial or deeper top soil layers. In practical organic farming, a bare fallow can sometimes be used before or after the main crop if conditions have promoted perennial weeds. However, if carried out in the growing season, soil temperature and moisture conditions favour soil microbial activity and therefore the build-up of SMN, which greatly increases the risk of NO₃ leaching (Borgen et al., 2012).