Organic resource composition can be variable, both within and among fertilizer types. In Table 1 the average dry matter, N, P and organic matter (OM) contents of a selection of the main organic resources used in Dutch agriculture are shown. The N:P ratio of an organic resource is an important parameter for balanced fertilization and to avoid excess application of N or P (see Section 6.1). The N:P ratio is generally higher for cattle manures than for pig and poultry manures (Table 1). The OM:nutrient ratio should be high when organic resources are mainly applied as soil improver and not as a source of nutrients. Vegetable, garden, and fruit-derived (VGF) compost has a higher OM:nutrient ratio than manures and this ratio is higher for solid manures than for slurries. Cattle slurry has the lowest N and P content on fresh product basis. Meat and bone meal is relatively rich in P and has a lower N:P ratio than the other organic resources. The UK Nutrient Management Guide (AHDB, 2023) also presents the composition of a range of organic resources. For cattle slurry, pig slurry, poultry manure, solid cattle manure and solid pig manure, they report values of 2.6; 3.6; 28; 6, and 7 g N kg⁻¹ fresh product, and 0.52; 0.65; 7.4; 1.4; 2.6 g P kg⁻¹ fresh product, respectively. Most of these values are somewhat lower than the Dutch, but the number of samples and the uncertainty level of the UK data is not available. These differences indicate however that the Dutch data -although based on a vast number of samples- may not necessarily be representative for other countries, and that average manure composition can depend on the common livestock system and level of intensity in each country.

The MITERRA-EUROPE model was used to calculate the amount of carbon and nutrients that are applied on agricultural soils in the EU-27 + UK (Table 2). The results show that in total over 78,000 Gg of C is applied through organic resources annually in EU-27 + UK (Table 2). The total amount of organic resource applied to soil amounted 6,970 Gg N (on average 41 kg N ha⁻¹ agricultural land, based on 171.7 million ha arable and grassland), 1,593 Gg P (9 kg P ha⁻¹), and 78,294 Tg C (456 kg C ha⁻¹) in EU-27 + UK in 2017 (Table 2). The amount of N is somewhat less and that of P somewhat higher than the with MITERRA-EUROPE estimated amounts of applied mineral N and P fertilizers, i.e., on average 66 kg N and 7 kg P per ha agricultural land. Solid pig and cattle manure and cattle slurry are the most used organic resources. More than half of the C applied to soils by organic resources in the EU-27 + UK is applied as solid manures. Relatively high amounts of cattle and pig slurries are applied in countries in NW Europe such as Belgium, Denmark, Netherlands, and Germany, whereas solid manures applied are more common in France, Spain, and part of central Europe (Figure 1). The application rate of VGF compost is relatively high in Denmark, Germany, and the Netherlands and that of sewage sludge in Spain, Denmark, Germany and Italy (Figure 1).

Van Dijk et al. (2016) made an assessment of the P flows in Europe in 2005. Phosphorus excreted as manure by livestock was almost fully applied to agricultural land and amounted to about 1,749 Gg P. A relatively small part of bio-waste was composted and used in agriculture in 2005 (15 Gg P), most bio-waste was lost in landfills and incineration (212 Gg P). Part of the sewage sludge in Europe was used in agriculture (137 Gg P) while a larger part was lost by landfilling and incineration (234 Gg P). On average 35% of the produced sewage sludge is applied in agriculture in the EU (Huygens et al., 2022), with very large variations between countries (European Commission, 2022). From the late 1980s onwards, the risk of prion diseases has decreased the recycling of P from animal slaughter waste in Europe. A part (22% in 2022) of the total produced animal slaughter waste has since then been labeled as high risk material (category 1 in the regulations on Animal By-products; EU, 2009; Huygens et al., 2019; Dobbelaere, 2023). This part cannot be used as feed or fertilizer, and is landfilled after incineration or used in cement kilns. Therefore, only a relatively small portion of the waste of slaughtered animals was used as fertilizer (16 Gg P in 2005). The remaining part (281 Gg P) is lost by landfilling and incineration (Van Dijk et al., 2016), although these data could be outdated (Huygens et al., 2019). The total amount of P loss (727 Gg P) by landfilling and incineration of food waste (212 Gg P), sewage sludge (234 Gg P) and slaughtered animals (281 Gg P) in 2005, is high, i.e., more than half of the mineral P fertilizer use (Van Dijk et al., 2016). Phosphorus is a life-essential, irreplaceable element and the fossil P reserves are limited (USGS (U.S. Geological Survey), 2022). As Europe has no significant P mines, it is highly dependent on the import of phosphate ore to produce P fertilizers (De Ridder et al., 2012). Clearly, increased recycling of P from these waste streams can strongly reduce the use of mineral P fertilizers.

There is an ongoing trend in which less municipal waste is landfilled and more waste is recycled, composted and incinerated (Eurostat, 2023). The amount of food waste is currently estimated to be between 158 and 298 kg per person per year in Europe (Corrado and Sala, 2018), which is about 20% of the food produced (Stenmarck et al., 2016).

The availability of organic resources for use in agriculture as fertilizer or soil conditioner from food processing, sewage sludge, and municipal bio-wastes is expected to strongly increase over the coming decade, because of the ambitious targets of recycling and reuse of municipal waste of the EU Waste Framework Directive sets (EU, 2018). The collection and recycling of municipal bio-waste is projected to increase in order to reach the EU's objective to reuse and recycle 65% of the municipal waste by 2035 (ECN, 2022). In comparison, in 2022, 48% of the municipal waste generated in the EU was recycled of which 18% organically, through composting and anaerobic digestion (Eurostat, 2023). A large fraction of the municipal waste (42%) is still incinerated or landfilled which offers potential for increasing recycling (Eurostat, 2023). The P recycling potential from waste water remains largely untapped in the EU (Serrano-Gomez et al., 2023). Some countries in Europe, such as Germany and Switzerland, have implemented measures to make the technical recovery of P from sewage sludge mandatory (Hušek et al., 2023), to ensure that P is recycled back into agricultural production. In addition, P recovery from slaughterhouse waste has become obligatory in Switzerland¹ (ESPP (European Sustainable Phosphorous Platform), 2015). Further innovations in the development of waste water management, such as the use of urine source separation technologies, will increase the availability of nutrients for use in agriculture (Larsen et al., 2021). The circular economy action plan (European Commission, 2020b) has fostered many research initiatives, exploring a broad range of nutrient valorization pathways like the resource recovery from food-processing waste waters (Durkin et al., 2022). The overall recovery potential for N and P from currently untreated or unutilized waste streams, such as food chain waste, manure and waste water has been estimated to be 4.2 Tg of N and P in the EU, which represents about 35% of the currently used mineral fertilizers in EU agriculture (Chojnacka et al., 2019). The increased recycling of organic resources will decrease the dependence on mineral N and P fertilizers in the EU.

Data on the composition of organic resources are very important in order to increase nutrient use efficiency and to decrease the risk of emissions to the environment. Measurement techniques are developing, and new techniques become available to characterize soils, manures, crops, and emissions to the environment to improve nutrient management (Bühler et al., 2021; Horf et al., 2022; Reijneveld et al., 2022; Deru et al., 2023; Ge et al., 2023). These data can be used in decision support tools and fertilization recommendation to support farmers in management of nutrients (Section 7.2.1).

Analytical methods used to characterize fertilizing products have recently been standardized by the EU, as a result of the FPR that entered into force on 16 July 2022 (see CEN and ISO technical committees for the proposed and published methods).² These methods are used for samples with varying composition due to the heterogeneity of organic resources. Furthermore, Member States are allowed to have national regulations for fertilizing products based on national standards and different methods than prescribed in the FPR. Comparison of samples from similar organic resources (e.g. from cattle slurry) and developing datasets about the composition of organic resources therefore remains difficult.

New techniques to determine C, and nutrients in organic resources are near-infrared spectroscopy (NIRS), low-field nuclear magnetic resonance for (NMR) and pyrolysis (Py-GC/MS) (He et al., 2020). Moreover, sensor-based techniques can analyze the total and ammonium N-, P-, K content of organic resources on site or on-board for precision farming with fast access to data and at a low cost (Sørensen et al., 2015; Evangelista et al., 2021; Morvan et al., 2021; Feng et al., 2022; Thiessen et al., 2022; Horf et al., 2024). However, the accuracy of these techniques is often less than classical chemical methods, and depends strongly on the number of reference samples of a specific material (Derikx et al., 2021). Specific techniques for compost, for example to test the stability or biodegradability of compost, such as Oxitop^®, rottegrad (self heating test) (Binner et al., 2011), or the presence of plastic (Bläsing and Amelung, 2018), are rarely tested and used for other organic resources.

Besides analysis of their composition, properties related to decomposability of OM in organic resources are often characterized in incubation studies, in which organic resources are applied to soil. Examples are the determination of humification coefficients of organic resources (Section 4.2; Lashermes et al., 2009) and N mineralization. These incubation studies have similarities with pot and field studies to which they can be related (see Section 4.1), but also to measurements of plant-available nutrients such as citrate-extractable P (Delin, 2016), and Potential Mineralizable Nitrogen (PMN) (Westerik et al., 2023).

Many new analytical techniques are available, yet there are still many uncertainties involved in how the data obtained can be used to improve nutrient management. A combination of the aforementioned techniques can help to optimize the use of organic resources and decrease emissions to the environment. However, there are many challenges in the interpretation of the results and the handling of the large number of data generated with these techniques. Representative sampling of organic resources remains a challenge, even for these new analytical techniques.

The effects of organic resource application on crop growth, yield, and quality are generally obtained by harvesting, weighing, and analyzing crops in a field experiment. Spectrum sensing techniques (e.g., handheld scanners using NIRS), from which information of the status of a crop in a field or field experiment can be rapidly obtained, are currently available and rapidly developing (Burkart et al., 2018; Luo et al., 2022). Crop images from satellites or unmanned aerial vehicles such as drones can also be used to get insight into the development of crops at a larger (field or farm) scale. This type of information can be used to predict crop response to (organic) fertilizers and, where necessary, adjust fertilization using precision farming techniques (Maresma et al., 2016; Zhang et al., 2020), as well as improve crop growth models (Dlamini et al., 2023). Data on crop growth and development in combination with data on nutrient input and availability of organic resources can be used in decision support tools for farmers to decide if a top dressing of a fertilizer is required to obtain the optimal yield (Section 7.2.1).

Plant nutrient availability of organic resources can be quantified by their Nitrogen or Phosphorous Fertilizer Replacement Value (NFRV or PFRV, respectively), which specifies the amount of mineral fertilizer needed for a similar crop response as the organic resource (Schröder et al., 2007; Schils et al., 2020). Accurate FRVs are required to underpin fertilizer recommendations and maximize nutrient use efficiencies. FRVs are mostly applied to N, and NFRVs can be determined in a field or pot experiment in which the rates of organic and mineral fertilizers required to obtain equal crop N uptake are compared (Jensen, 2013; Westerik et al., 2023). NFRVs are commonly determined for the first year after application (i.e., short-term FRVs) and the variation in FRV is large, as is shown by two review papers in Table 3. Variation can be caused by variation in organic resources composition, time and method of application, climatic conditions, soil properties, reference fertilizer and test crop (Jensen, 2013). For example, the first year NFRV of cattle slurry can increase from 36 to 53% when injected instead of surface applied, most likely through a reduction in NH₃ volatilization (Schröder et al., 2007). NFRVs are expected to increase over time because of mineralization of residual organic N in the years after application (Gutser et al., 2005; Schröder et al., 2007; Hijbeek et al., 2018).

Since field experiments for the determination of FRVs are time intensive and costly, attempts are being made to predict FRVs based on easily obtainable chemical properties. Total N, C:N, mineral N and Potential Mineralizable Nitrogen (PMN) appear to be good predictors for NFRV in pot and field experiments, and citrate-extractable P can predict PFRVs (Stadler et al., 2006; Delin et al., 2012; Delin, 2016; De Notaris et al., 2018; Westerik et al., 2023). NFRVs generally increase with higher total and mineral N contents, lower C:N ratios and higher PMN (Tables 1, 3), but the strength of these correlations needs to be tested under a range of field conditions. This is especially relevant with the rapid emergence of new organic resources, and the need for product-specific recommendation rates and FRVs. These new products can in theory have improved plant nutrient availability, but adequate FRVs are required to underpin fertilizer recommendations and maximize the nutrient use efficiency of these products.

NFRVs of organic resources are often used in national fertilizer recommendations and decision support tools as well as action programs of the EU Nitrates Directive. Currently, there are striking differences of assumed NFRVs between European countries, therefore large differences exist in recommended or allowed rates of organic resources. For example, the NFRV of farm yard manure (FYM) is assumed to be 10% in the UK, whereas it is assumed to be 60% in Germany (Jordan-Meille et al., 2022). Such differences can result from alternative calculation methods, experimental setups and timeframes (Gutser et al., 2005; Jensen, 2013; Schils et al., 2020; Westerik et al., 2023). For example, some countries (Belgium, Switzerland, Italy and Germany) take into account long term effects (>1 year) of fertilizer application on yield, whereas other countries do not. There is a reasonable agreement between countries in the order of decreasing NFRV, i.e., pig slurry > cattle slurry > poultry manure > FYM (Webb et al., 2013). However, there are also differences in NFRV among countries which are, besides the calculation method and considered period, probably also due to differences in application time and method. Standardization of calculation methods should serve to improve N use efficiency of organic resources.

Besides their supply of N and P, organic resources also supply other nutrients and trace elements (Sager, 2007). Furthermore, repeated organic resources application can improve soil properties such as bulk density, aggregate stability, infiltration capacity and water retention, although these effects are largely dependent on soil and crop type (Fu et al., 2022). Organic resource application can therefore benefit soil fertility and improve yields beyond the supply of N and P (Hijbeek et al., 2018; Kok et al., 2023).

Although knowledge of NFRVs of different organic resources is rapidly emerging, insight in the chemical properties to explain and predict variability in NFRVs as well as standardization of determination methods of NFRVs is required for adequate fertilizer recommendations of new and existing products. Besides, data of supply of P, K and other macro- and micronutrients from organic resources to crops is still scarce.

The addition of organic resources influences soil organic carbon (SOC) by the addition of organic material and by increasing aggregate stability and the organic C content of aggregates (Yilmaz and Sönmez, 2017). Generally, the potential to sequester C in the soil increases with increasing C:N rate of organic resources (Hijbeek et al., 2019).

The decomposability of organic resources varies largely between and within different groups of organic resources (Lashermes et al., 2009; Levavasseur et al., 2022). Levavasseur et al. (2022) found that for example the mineralization of C from animal manures varied between 0 and 500 mg C per g added C for incubation studies lasting an equivalent of 1 year under field conditions. To predict C mineralization in the field after organic resource application, Levavasseur et al. (2022) recommend to calibrate models per product and not per fertilizing product group. In addition, they recommend to also include product characteristics such as the Van Soest analysis of fiber (Van Soest and Wine, 1968).

The humification coefficient (HC) is the percentage of applied OM that is not decomposed after 1 year of soil application, and is used as an indicator for OM stability of fertilizing products. HCs are used in fertilizer recommendations and in simple models for estimation of C mineralization after application of OM (Janssen, 1984; De Neve et al., 2003; Egene et al., 2021). The HC is highly variable within and among different types of organic resources (Table 4). Several factors may influence the HC, such as climatic factors, soil characteristics, characteristics of the input material, and anthropogenic activities. Additionally, the relationship with initial SOC content and HC values was, based on a mathematical method, found to be non-linear: a higher SOC content leads to lower HC values (Tan et al., 2014). The wide ranges reported in Table 4 imply that using a single HC per fertilizer type for calculations of C build up, may lead to significant over- or underestimations (of up to 30%).

Other parameters used to describe C storage potential or mineralization include the indicator of remaining organic carbon (IROC) (Lashermes et al., 2009), the humus production capacity (VDLUFA, 2021), and the cumulative C-CO₂ emissions over a chosen period of time (Mondini et al., 2017). All of these terms point toward decomposability trends with C from animal by-products including meat and bone meal being most decomposable and C from compost being least decomposable. For Lashermes et al. (2009), VDLUFA (2021) as well as Mondini et al. (2017), the fact remains that these decomposability terms vary strongly within product groups.

In short, commonly, animal by-products are most easily decomposed, followed by sewage sludge, manure and compost, but the variation in C mineralization is very high within the categories.

Ammonia emissions after application of manures or other organic resources can constitute a major N loss pathway from animal farms. Loss of NH₃ through volatilization reduces N use efficiency of the fertilizer, contributes to the formation of airborne particulate matter, and may induce acidification, eutrophication, and indirect N₂O emissions after deposition (Behera et al., 2013; Räbiger et al., 2020). Emission patterns differ for various fertilizer types. For liquid fertilizers, such as slurry and digestates, volatilization typically occurs during the first hours or days after application, whereas for solid manures, composts, and plant residues NH₃ emissions are typically lower and can be spread out over a longer timeframe, as the OM in the fertilizer breaks down. Generally, emissions are higher for liquid manures and slurries with high levels of available N, such as slurry digestates or liquid poultry manure, than for solid manures and composts (Table 5). In an elaborate screening of over 30 biobased fertilizers, Wester-Larsen et al. (2022) reported a high NH₃ volatilization potential for digestate products, and a low potential for solid products like composts.

Composition strongly affects the magnitude of NH₃ losses from organic resources, especially the total N and ammoniacal N contents and the fertilizer pH. Moreover, as NH₃ volatilization strongly depends on the chemical equilibrium between ammonium, dissolved NH₃, and gaseous NH₃, environmental factors such as temperature, wind speed, soil moisture, and contact surface between the fertilizer and air play a crucial role. Mitigation strategies can therefore be aimed either at altering the composition or chemical properties of the fertilizer, e.g., by acidifying manures (Wagner et al., 2021; Silva et al., 2022; Pedersen and Nyord, 2023) or treating them with plasma (i.e., acidifying slurry and adding NO₂ and NO₃ by N fixation) (Graves et al., 2019), or at ensuring that the environmental conditions are unconducive for NH₃ volatilization. The latter strategy includes applying liquid fertilizers shortly before rainfall events and applying water after slurry application (Webb et al., 2014a), by injecting or incorporating fertilizers into the soil (Thompson and Meisinger, 2002; Powell et al., 2011; Dell et al., 2012; Webb et al., 2014a; Wester-Larsen et al., 2022), or by covering the fertilizer with a foamlike substance to reduce air contact (Park et al., 2006; Lee et al., 2007; Bajagain et al., 2022). Generally, assuring quick infiltration of liquid organic resources into the soil and minimizing their exposure to air may reduce NH₃ volatilization after application. However, the aforementioned measures may in turn lead to pollution swapping to e.g., enhanced NO₃ losses or N₂O emissions (see also Section 7.1).

National inventories for reporting NH₃ emissions from agriculture often make use of emission factors that describe the cumulative emissions from a source as a fraction of total or ammoniacal N (Table 5). There are examples of national Tier 2 or Tier 3 emission calculation methodology that include emission factors for specific NH₃ abatement measures, e.g., by differentiating for various application methods in The Netherlands (Van der Zee et al., 2021), but generally national inventories do not include the option to account for these measures. Emission factors for new organic resources in combination with abatement measures are desirable for accurate calculation of NH₃ emissions and for providing farmers and governments with a means of accountability for management actions taken. Deriving these emission factors will be a major challenge, complicated by the increasing number of products and measures, and by the inherent difficulty of accurately determining NH₃ emissions. Recent efforts to synthesize information on NH₃ emission in databases with information from lab, field and model studies that describe emission data (Hafner et al., 2018; Beltran et al., 2021) can provide context for newly derived emission factors.

Application of organic resources to soil can result in N₂O emissions through several processes, of which the most important are nitrification—the oxidation of NH 4+ to NO 2- and NO 3- with N₂O as an intermediate product, and denitrification—the reduction of NO₃ to N₂ with N₂O as a product of incomplete denitrification (Chadwick et al., 2011). In the last decades many studies focused on N₂O emissions following manure application (Chadwick et al., 2011; Thorman et al., 2020). The emission factors of manures (i.e., the cumulative N₂O-N loss as a proportion of total N applied to soil via manure) can range from 0.1 to 3.3% (Chadwick et al., 2011; Thorman et al., 2020; Van der Weerden et al., 2021), reflecting differences in manure type and composition, soil type, management and climate (Table 6). The emergence of manure processing techniques leads to new organic resources (Section 6.2.1). The composition of these products varies strongly (Chadwick et al., 2011), and so do the N₂O emissions following application to soil. However, many countries rely on the IPCC tier 1 default N₂O emission factors, in which the type of organic resource, soil type and application method is ignored. This leads to large uncertainties in the quantification of N₂O emission at local to global scales (Tian et al., 2020). The default aggregated IPCC-2019 N₂O emission factors are 0.3–0.4% of N applied for manures and 1.0% for other organic resources (IPCC, 2019).

Incorporation of slurry into the soil has already been shown to be an effective method to reduce NH₃ emission (Section 5.1), however some studies have shown that soil injection increases the risk of N₂O emission, pointing to a risk of pollution swapping (Velthof and Mosquera, 2011). Others report similar N₂O emissions between broadcast and soil injection (Vallejo et al., 2005). The 4R nutrient management strategy (Section 6.1), including a lower application rate, splitting of N doses, and the application of nitrification inhibitors (Herr et al., 2020; Thorman et al., 2020) could balance the input and output of N and subsequently reduce the N₂O emission. Organic resource application should furthermore be timed to avoid application during wet conditions, actively support crop growth, and reduce excess N in the soil as it could result in large quantities of N lost by NO₃, which is in turn an indirect source of N₂O emission.

It is assumed that the risk of N₂O emissions from VGF compost, sewage sludge and meat- and bone meal are low, because of the generally lower mineral N content of these fertilizers.

Methane is produced by methanogenic bacteria under strict anaerobic conditions (Zeeman, 1991). Agricultural soils are generally well-aerated and are therefore not a source CH₄, but can rather act as a sink of CH₄ (Hansen et al., 2024). During storage of liquid manure, however, CH₄ may be produced from easily degradable C components in the manure, such as volatile fatty acids (Kupper et al., 2020). After application to soil, part of the CH₄ produced during storage and which is dissolved in the slurry can be released (Chadwick et al., 2011). However, this is not CH₄ produced in soils. This does not hold for wetland rice soils, which are a large source of CH₄. Application of organic resources to wetland rice may increase CH₄ emission (Hu et al., 2023). The surface area in Europe used for wetland rice production is extremely low, thus risk of high CH₄ emissions from agricultural soils in EU remains negligible.

It has been hypothesized that N from organic resources has a larger risk of being lost by leaching than mineral N due to the untimely mineralization of fertilizer-derived organic N outside the growing season. This idea has likely also been taken into account when setting the specific limit for animal manure of 170 kg N ha⁻¹ in the Nitrates Directive (EEC, 1991) to prevent nitrate leaching from agriculture. However, use of labeled N shows that only a few percent of leached N in 1 year originates from the N fertilizer applied in that year, both for mineral N and organic resource N (Powlson et al., 1986; Choi et al., 2004; Frick et al., 2022). In the long term, buildup of soil organic N by crop residues and organic organic resources is the most important source of N leaching (Goulding et al., 2000). Nitrate leaching increases with long-term N application (Blicher-Mathiesen et al., 2014; Wang et al., 2019), and wrong timing of application, for example the application use of pig manure in autumn (Shepherd and Newell-Price, 2016). A meta-analysis by Ren et al. (2022) shows lower N leaching at higher SOC contents. A meta-analysis by Wei et al. (2021) shows that substitution of a part of the applied mineral N by animal manure N results in limited yield loss and decreases N leaching on average with 30%, while yields do decrease at large substitution rates (Ren et al., 2022). Another study showed that organic inputs in addition to mineral N did not affect leaching and runoff, but also did not improve yield (Wei et al., 2021). Application of easily available C via organic amendments can decrease N leaching by enhancing denitrification, but this might also decrease overall nutrient use efficiency (Qin et al., 2017). A meta-analysis showed a 13% decrease of nitrate leaching due to biochar addition although the mechanisms behind this remain unclear (Borchard et al., 2019).

Phosphorus leaching from soil due to various P fertilizers is mainly driven by the P surplus in the long-term and not by fertilizer type or composition (Blake et al., 2003; Eichler-Löbermann et al., 2007; Messiga et al., 2015; Lemming et al., 2019). In the case of long-term use of certain fertilizers, such as calcium-rich poultry manure (Lehmann et al., 2005), iron-rich sewage sludges, or organic P rich sources, specific P forms can be formed in soil with different leaching susceptibility (Liu et al., 2020). A long term experiment with animal manure resulted in a lower P sorption to the soil in comparison to six other types of organic resources (composts, sewage sludges) (Nest et al., 2016), implying a higher risk of P leaching of manure than other organic resources. For runoff, risk assessment tools such as P indices often rely on landscape and management factors to estimate the risk of losses (Osmond et al., 2017; Ros et al., 2020). Distance to the closest waterways, slope, the use of cover crops and the timing and method of (organic) fertilizer application are examples. Additionally, weather conditions like heavy rainfall after application of manure increases risk of surface runoff of N and P toward surface water and these events may contribute to a large extent to annual N and P leaching from applied manure (Van der Salm et al., 2012).

Application of biodegradable waste material is challenged with avoidance of contamination of the soil, especially with “emerging contaminants”, such as microplastics, nanoparticles and pharmaceuticals (Ng et al., 2018; Kacprzak et al., 2022). Therefore, the research on waste processing to produce safe and sustainable organic resources is still ongoing (Kurniawati et al., 2023), while at the same time recent EU legislation has banned synthetic polymer microparticles from various products (REACH EU 1907/2006) and fertilizing products (EU 2019/1009) within a transitional period of 5 years.

In Table 7 an overview of heavy metals in selected organic resources is presented. The heavy metal content is generally very low in struvite and meat and bone meal (Möller and Schultheiss, 2015), and higher in manure, compost and sewage sludge (Dach and Starmans, 2005; Smith, 2009; Kupper et al., 2014). The contents of Cu and Zn in animal manures are generally relatively high compared to those of other heavy metals. Cu and Zn in animal manure often originate from additives in feed (Adamse et al., 2017). The composition of these additives is controlled in the EU since 1970 (70/524/EEC) by limit values for Cu and Zn specific for each animal type, which have been reduced since then (EU, 2009). However, the limit values for Cu and Zn in the FPR are sometimes still exceeded, making these manures unsuitable for marketing as EU fertilizer. In contrast to Cu and Zn, the Cd and Pb contents in animal manure are low in comparison to compost and sewage sludge (Table 7).

Levels of persistent organic contaminants, such as per- and polyfluoroalkyl substances (PFAS), polycyclic aromatic hydrocarbons (PAH), poly chlorinated biphenyls (PCB), dioxine-like PCBs (dl-PCB) and dioxines and dibenzofurans (PCDD/F), are normally very low in animal manure (Berset and Holzer, 1995; Munoz et al., 2021), but can be elevated in products such as sewage sludge (Munoz et al., 2021; Huygens et al., 2022) and compost (Brändli et al., 2007a,b; Saveyn and Eder, 2014; Huygens et al., 2020; Munoz et al., 2021). Limit values of organic contaminants have been set in the FPR for certain input categories (Component Material Categories: CMC) for the production of fertilizers (Table 8). There are concerns about the bioaccumulation and toxicity of pharmaceuticals from both manure and sewage sludge (Gworek et al., 2021). In some cases the use of pyridine herbicides that have a relative high toxicity for plants but not for grazing animals can result in herbicide levels in manure from grazing animals that are too high for use as a general fertilizer. This has resulted in proposed limit values for EU fertilizer products from animal manure for clopyralid and aminopyralid (Huygens, 2023). Some animals receive antibiotics that end up in animal manure. These may potentially leach from the soil or be harmful for soil biota (Rietra et al., 2023) if degradation is slow and sorption is low. Additionally, they may pose a threat to insects living in dung (Bruinenberg et al., 2023) and birds living on cattle farms (Buijs et al., 2022).

It has been estimated that 67 and 83% of sewage sludges (on the short- and long term, respectively) fails to comply with current limit values, if no pollution prevention is implemented (Aubain et al., 2002). For sewage sludge (ashes), techniques to extract phosphate are promoted by research (Canziani et al., 2023), which seems especially relevant for sewage sludge that does not comply with the limit values.

In summary, many types of contaminants exist and their prevalence and concentrations vary with the type of fertilizing product, e.g., levels heavy metals and pharmaceuticals are higher in manure and other contaminants are more concentrated in e.g., sewage sludge.

The 4R nutrient stewardship guidelines are a framework that describes practices for the application of fertilizers to increase nutrient use efficiency and reduce environmental footprints (Johnston and Bruulsema, 2014; Flis, 2017, 2018). The practices focus on fertilizer application using the Right source, the Right rate, at the Right time, in the Right place. Management of organic resources according to these principles has the potential to increase nutrient use efficiencies and mitigate emissions to the environment. According to a synthesis of several meta-analyses by Young et al. (2021), 4R strategies generally positively impacted crop yield, crop N uptake, and soil organic carbon (SOC), while reducing N₂O emissions, NH₃ emissions and N surplus. Replacing mineral fertilizers with organic resources increases the SOC content and results in reductions of N losses. Moreover, nutrient management generally has a larger effect on sustainability indicators compared to crop, soil and water measures (Young et al., 2021). The 4R nutrient stewardship guidelines can be based on a decision support tool approach using combinations of soil, crop, organic resource, and weather data (Section 7.2.1.) Decision tools can also include the risk of emissions and measures to mitigate these emissions, although currently most decision support tools only focus on fertilizer recommendations (Nicholson et al., 2020).

For animal manures and other organic resources, however, it is likely that proper 4R nutrient stewardship is more difficult to attain compared to mineral fertilizers. Animal manures, for instance, usually contain multiple nutrients in different ratios, which requires careful consideration of the right source to meet crop demand of certain nutrients without under- or overapplication of others. Herein lies an increased risk of pollution swapping, for instance between NH₃ volatilization, N leaching, and N₂O emissions, compared to mineral fertilizers.

Since nutrients are often (partly) in organic forms and mineralize over time, control of nutrient release from organic resources remains a challenge, as it is partly governed by external factors such as moisture and temperature. Repeated applications of organic resources increases N mineralization in following years (Schröder et al., 2007; Jensen, 2013; Bhogal et al., 2016; Sorensen et al., 2017). The right time of application is essential to limit the mineralization and loss of nutrients outside of the crop growing season. For manures however, the window of application may be limited by on-farm storage capacity. The right placement of organic resources (near the crop roots) can be more challenging since regulations or available techniques do not always allow this.

Proper implementation of 4R strategies in combination with new measurement and processing techniques have great potential to reduce losses and increase nutrient use efficiency for organic resources, especially if they partly replace mineral fertilizers. Despite the universality of the 4R principles, optimizing strategies for application of organic resources requires tailoring them to specific environmental and management contexts (climate, soil type, cropping system, etc.). Herein still lies a big challenge.

Structural measures such as the reduction or relocation of livestock and crop production are an option for improving nutrient and C management and to reduce emissions from agriculture. In comparison to technical and management measures, structural measures generally refer to more drastic changes in the agricultural system, e.g., through transitioning toward lower livestock numbers, relocating or shifting livestock types, or changing land uses and cropping systems. Structural measures will need to be taken if large scale implementation of technical and management measures as shown in Sections 6.1 and 6.2 are not sufficient to reach the environmental targets. Besides their relevance for reaching environmental targets, structural measures will impact the nutrient demand (e.g., for fertilizing cropland) and nutrient supply (e.g., from livestock manure), which need to be considered in the future management of organic resources. The examples of studies exploring reduction potentials at the national or European scale in Table 10 show large ranges but also a high reduction potential for structural measures.

A combination of measures have to be taken to increase nutrient use efficiency, improve soil quality and reduce gaseous emissions and leaching from organic resources so that the targets of the different environmental policies can be met. Combinations of measures may have synergies and strengthen their effects; however there is also a risk of trade-offs or pollution swapping. Different types of pollution swapping between measures can take place (Figure 3). Examples of the different types of pollution swapping risks are:

Clearly, avoiding pollution swapping and finding synergies in the implementation of combinations of measures is challenging. If sufficient information on organic resources is available, potential trade-offs can be taken into account. For example, Velthof and Mosquera (2011) calculated the total effect of application method on N₂O emissions, taking into account indirect emissions from NH₃ volatilization and the need for extra mineral fertilizers due to loss in NFRV after surface application. Based on their results and IPCC emission factors, they showed that, although injection of cattle slurry on grassland can lead to higher direct N₂O emissions than surface spreading, the indirect emissions from NH₃ volatilization and the additional mineral fertilizer need result in higher total N₂O emissions for surface application (Velthof and Mosquera, 2011). This effect was not the same for pig slurry on arable land due to the high emission factor for injection of pig slurry, which shows that trade-offs are often heavily context dependent. In-depth knowledge on the composition, agronomic value and risks on the emissions of organic resources to the environment is therefore needed to derive effective nutrient management strategies. This especially holds for new waste products from the circular economy and new products from manure treatment (Figure 2), as the agronomic and environmental consequences of the use of many of these products is not yet clear.

Fertilizer recommendations and decision support tools are important sources of information for farmers to increase nutrient use efficiency and decrease nutrient losses. Recent papers show that there is a large diversity in recommendations and tools between countries in the EU and that harmonization, shared learning, and a collective approach is recommended to tackle environmental problems (Nicholson et al., 2020; Higgins et al., 2023). The EU FAST tool (https://fastplatform.eu/) is developed for European Commission's DG Agriculture and Rural Development as nutrient management tool to support EU farmers, farm advisors, developers of digital solutions, and policy makers. This tool contains different modules, including a fertilizer module. Besides nutrient management tools, tools are also developed to calculate soil C sequestration and GHG emissions from farms (Arulnathan et al., 2020; Alexandropoulos et al., 2023), such as the Cool Farm Tool (Hillier et al., 2011) (https://coolfarm.org/). There is also an increasing interest of the food industry in these type of tools, as many industries have made commitments to decrease their GHG emissions. In the coming decade, farmers might be subject to an increasing level of accountability for the environmental footprint of their farm and stimulated to use aforementioned tools, by both governments and industry partners. This can give rise to new challenges because of the extra administration and record keeping these tools demand or because of management changes required to achieve set targets.

Fertilizer recommendations and decision support tools require information about the composition (Sections 2.1 and 3.1), the plant-availability of nutrients (Section 4.1) and degradability of OM of organic resources (Section 4.2). Ideally, analyses of the composition and nutrient availability of organic resources in combination with (real-time) data on soils, crop and weather are used, but at least standard values for a range of different organic resources have to be available for reliable recommendations (Nicholson et al., 2013). If tools are also used for estimating the risk of gaseous emissions and leaching, information on emission factors of NH₃ and GHG for organic resources and coefficients for calculation of leaching are required (Section 5). Emission factors can be derived from default values of the EMEP/EEA air pollutant emission inventory guidebook, and IPPC guidelines, or from country specific emission factors. Farmers should also be aware about the potential risks of contaminants present in the organic resources, especially when they use these fertilizers frequently (Section 5.5). The use of farm specific emission factors and coefficients would improve the accuracy of emission calculations, using for example process-based soils models (Heinen et al., 2020).

The rapid development of analytical techniques for composition of organic resources, soils, crops, weather, and emissions to the environment, will further improve farm specific nutrient management strategies and DSS-tools. However, there are large challenges regarding the interpretation and integration of the huge amount of data that will be derived from these new techniques. Ultimately, they have to lead to scientifically underpinned and clear recommendations to farmers about daily field operations.

Lastly, the implementation of measures, that for example promote manure treatment or change of animal housing systems, is challenged with financial and practical barriers. Building manure processing installations or new stables is expensive and requires permits; all in all, it usually takes several years before such a change is implemented. For the biogas sector the key barriers are the lack of financial incentives and policy frameworks that ensure profitable operation until at least the investment is paid back (Kampman et al., 2017).