Organo-mineral fertilizers (OMFs) are a mixture of materials of biological origin with one or more mineral fertilizers (Smith et al., 2020). This allows the use of BBFs rich in organic C as the organic matrix, such as compost. The use of OMF has been reported to have some advantages over mineral fertilizers alone, as addition of organic C can reduce N losses and increase N use efficiency (Richards et al., 1993; Antille et al., 2014; Florio et al., 2016). Similarly, OMFs increase plant P use efficiency by causing a prolonged release of plant available P in soil (Antille et al., 2013).

The enhanced nutrient use efficiency can be linked to various processes. For instance, there is an electrostatic attraction of phosphates to organic materials, reducing the P mobility in soil that could be otherwise fixed to clay minerals (Gwenzi et al., 2018; Luo et al., 2021). The chemical interactions between the organic matrix and mineral fertilizer can result in the formation of insoluble compounds that precipitate out of the soil solution (Mazeika et al., 2016; Carneiro et al., 2021; Luo et al., 2021). Additionally, the organic material acts as a physical barrier between the mineral fertilizer granule and the soil solution, thereby decreasing the solubility of the fertilizer granule (Limwikran et al., 2018). Lastly, the stimulation of soil microbial activity leads to the immobilization of nutrients into the microbial biomass and a subsequent gradual N release that reduces leaching (Richards et al., 1993; Mandal et al., 2007).

The FPR establishes that at least 3% of the total mass of an OMF must be composed of organic C if it is a liquid OMF, and at least 7.5% on a solid OMF, while the sum of macronutrients needs to be 6% or 8%, respectively (EC, 2019). This broad definition allows the production of OMFs with varying characteristics.

One type of OMF is where the organic matrix is used as the main source of one or more nutrients and mineral fertilizer is added to produce a specific nutrient ratio that will depend on specific soil and plant nutrient requirements (Rady, 2012; Antille et al., 2013; Anetor and Omueti, 2014). This type of solid OMF, in addition to supplying macro and micronutrients, can be used as a soil amendment to improve physical properties due to its large organic C content (Babalola et al., 2007). On the flip side, OMFs rich in organic C require the application of large quantities to the soil - in the order of tons per hectare- to adequately fulfill crop nutrient requirements (Antille et al., 2013; Mazeika et al., 2016). Such large required application volumes could result impractical.

A second group of solid OMFs has a small proportion of organic C from a material of biological origin. This material acts mainly as a protecting matrix for the mineral nutrients and not as a relevant nutrient source. Low organic C OMFs are designed to produce a slow release of mineral nutrients into the soil solution increasing the nutrient use efficiency (Richards et al., 1993). However, although the use of low organic C OMFs has the advantage of increasing the efficiency of the mineral fertilizer – and therefore reducing its application volume – they have the disadvantage of negligibly contributing with organic C to the soil.

There is a significant tradeoff between OMFs with high and low organic C. In those with low organic C, biowastes are not a significant source of nutrients. Therefore, their potential contribution to biowaste recycling is limited. If the mineral nutrients in the OMF are not from biological origin, it is advisable not to consider them a bio-based fertilizer (ESPP, 2023). Despite this limitation, biowastes in low organic C OMF play a crucial role in enhancing the nutrient use efficiency of mineral fertilizers, which is a significant benefit.

Various studies highlight the importance of fertilizer P placement for enhancing the fertilizer value (Quinn et al., 2020; Freiling et al., 2022). Applying mineral P fertilizers by placement close to the seeds maximizes crop P uptake in the early growth stages (Grant et al., 2001; Grant and Flaten, 2019). Another strategy is subsurface band application (10 cm to 20 cm depth). This is mainly adopted in tropical countries with highly sorbing P soils that are susceptible to long drought periods. Subsurface band application creates P-rich bands with high available P contents, that are less susceptible to the soil surface drying (Meyer et al., 2023). However, the placement of P-rich fertilizers, as would be the case with some BBFs, is not as effective as the placement of mineral P fertilizers. Two possible explanations are the lower P solubility in BBFs compared to mineral fertilizers; and nutrient imbalances, leading to the over application of elements that may be toxic to the plant (Lemming et al., 2016).

Placement of sewage sludge increased root proliferation in the placement zone but did not enhance plant growth and total P uptake (Lemming et al., 2016). It significantly reduced P uptake from the soil compared to a treatment where the fertilizer was completely mixed with the soil. Regarding sewage sludge ash, Lemming et al. (2016) observed that placement did not attract root growth to the placement zone and significantly reduced both P uptake and plant growth. Building up on that, Sica et al. (2023) suggested that biowastes may have sufficient soluble P to attract roots to the placement zone, however, available P may not be sufficient to sustain plant growth for longer periods. Therefore, treatments to increase the P solubility of bio-based fertilizers are needed, along with placement strategies.

Fertilization with organic fertilizers may lead to the overapplication of organic N in the placement zone. Accumulated organic N is mineralized over time, releasing high amounts of NH₃ into the soil (Chaves et al., 2004). Plants can suffer NH₃ toxicity symptoms, such as inhibiting root growth at concentrations above 100 mg·kg^-1 (Nkebiwe et al., 2016). To reduce NH₃ toxicity effects, placement should be done more than 10 cm from the seeds, and/or days before sowing, allowing NH₃ to be nitrified thus lowering its concentration over time, while the plant is starting to develop (Delin et al., 2018; Baral et al., 2021). Acidification may also reduce the negative effects of NH₃ toxicity (Pedersen et al., 2017).

Recycling P reduces phosphate rock mining, thereby avoiding the associated environmental impacts such as landscape degradation, contamination of water bodies, and emission of GHGs associated with transport (Higgins, 2001; Fayiga and Nwoke, 2016). Moreover, using phosphate rock and derivatives for fertilization can result in the contamination of water bodies and soils with heavy metals and hazardous elements like cadmium, uranium and arsenic (Fayiga and Nwoke, 2016). Similarly, reducing the reliance on synthetic N contributes to lowering net GHG emissions and energy consumption, as Haber-Bosch synthesis of NH₃ is one of the major fossil energy consuming processes worldwide (Osorio-Tejada et al., 2022; Gao and Cabrera Serrenho, 2023).

Beyond the substitution of mineral fertilizers, bio-based fertilizers offer additional environmental benefits. For example, by offering a slower release of nutrients compared to mineral fertilizers, nutrient leaching is reduced (Mandal et al., 2007). Bio-based fertilizers containing organic C enhance soil structure by increasing the organic matter content which also may be beneficial for microbial communities (Mayer et al., 2022). In soils in a Danish long-term field experiment, soils treated with compost significantly increased their organic C content compared to the mineral NPK treatment by up to 3% Moreover, they significantly reduced their bulk density from 1.6 kg·l^-1 down to 1.2 kg·l^-1, and improved their soil structure, thus requiring less energy consumption for tillage (Peltre et al., 2015).

Due to non-optimized nutrient ratios relative to crop demand in many BBFs compared to mineral fertilizers, fertilization with BBF and organic fertilizers could lead to overapplication of nutrients. An overview of environmental impacts associated with BBFs is available in Table 4. For instance, fertilization with manures can lead to overapplication of P when aiming to meet crop N demand (Sharpley, 1996). Similar problems arise when using digestates (Kadam et al., 2022), sewage sludge and other biowastes (Deeks et al., 2013; Lemming et al., 2019). Mitigation strategies encompass minimizing N losses during production, storage, and application of the BBF, thus maximizing the N content and its use efficiency, for which Ndegwa et al. (2008) and Pedersen and Hafner (2023) offer extensive reviews. The N to P ratio of some BBFs such as those derived from animal manures can be optimized by separating N from P through solid–liquid separation, or through segregation at-source of feces and urine before any processing step (Hjorth et al., 2009; Vu et al., 2016). Another strategy is using an organo-mineral combination, that provides the desired N to P ratio and corrects for the nutrient imbalance (Deeks et al., 2013).

The presence of contaminants could be a problem for some BBFs. Sewage sludges can contain cadmium, lead, and other heavy metals that may accumulate in soils after fertilization (Lu et al., 2012). In a study in Chile, significant amounts of microplastics accumulated in the soil after a decade of fertilization with sewage sludge (Corradini et al., 2019). The presence of persistent organic compounds like polychlorinated biphenyls, polycyclic aromatic hydrocarbons, traces of pharmaceutical compounds and hormones in sewage sludges, is discussed by Kanteraki et al. (2022). An overview of pollutant removal techniques for sewage sludge is provided by Geng et al. (2020). Olive pomaces contain phytotoxic phenolic compounds, lipids, and organic acids. Composting pomaces can help mitigate their toxicity by degrading some of the harmful compounds and increasing their fertilizer value (Muscolo et al., 2019; Ameziane et al., 2020).

Heavy metals are of potential concern in digestates and composts depending on the local input materials (Kupper et al., 2014; Kadam et al., 2022). Therefore, there is a pressing need for thorough monitoring and management of these biowastes. A breakthrough for the simplification of monitoring on a local scale was the obtention of conversion factors to predict the concentration of heavy metals in Belgian manure-derived digestates (de Castro et al., 2023). These conversion factors were based on process parameters of the AD processes, allowing the prediction of the concentrations of aluminum, chromium, cupper, iron, manganese and zinc based on the dry matter and the biodegradable fraction content of digestates.

Emissions of GHGs can result from both the production of BBFs and their soil application. In the production stage, mineral precipitates and thermally treated biowastes specifically raise concern as they require energy-intensive treatments. Thermal treatments and separation processes such as ammonia stripping, membrane electrodialysis, ion-exchange and struvite precipitation may emit less C to the atmosphere compared to traditional mineral fertilizer synthesis. However, their operational processes also demand substantial energy consumptions (Meneses-Quelal and Velázquez-Martí, 2020; Kar et al., 2023). Therefore, the source of the utilized energy will influence the environmental impacts of such treatments.

During soil application, NH₃ emissions from manure digestates are often higher than from untreated manure (Holly et al., 2017; Emmerling et al., 2020). Furthermore, using ammonium sulphate to target sulfur fertilization has been linked with increased and overlooked NH₃ emissions in neutral to alkaline soils (Powlson and Dawson, 2022). This not only emits NH₃, contributing to undesired N deposition in natural ecosystems, but also results in losses of recovered N and the energy invested in its recovery. Likewise, fertilization with sewage sludge causes emissions of N₂O and CH₄ (Alvarez-Gaitan et al., 2016).

Implementing mitigation strategies is essential to minimize emissions from BBFs. These strategies include selecting appropriate application timing and methods, managing soils to maximize C retention, and using stabilizing additives like nitrification inhibitors (Severin et al., 2016; Tariq et al., 2022). For instance, injecting digestates into the soil reduces NH₃ emissions compared to surface applications (Hou et al., 2015). Similarly, applying slurry to grasslands during spring rather than autumn is linked to less N₂O emissions (Maris et al., 2021). Soil management practices such as minimizing or eliminating tillage and implementing crop rotations and leys, maximize the amount of C retained in agricultural soils (Jarecki et al., 2003). Acidifying digestates with, e.g., sulfuric acid can significantly reduce NH₃ emissions (Pedersen and Nyord, 2023).

Single mitigation strategies can potentially result in pollution swapping. Therefore, understanding the interaction of all processes involved and the impact of specific local conditions is necessary. A comprehensive and integrated approach such as a life cycle assessment (LCA) is optimal to assess the potential environmental impacts of fertilization with BBFs (Jensen et al., 2020; Egas et al., 2023).

For example, Styles et al. (2018) used a LCA to assess BBF production from the liquid fraction of digestate of food waste in Sweden, applying NH₃ stripping and struvite precipitation. The BBF production and field application was compared to the conventional management of digestate’s liquid fraction, including storage and field application. They concluded that producing BBF from liquid digestate results in significant environmental benefits due to the avoidance of CH₄, N₂O and NH₃ emissions compared to the conventional management of liquid digestate. Moreover, application of that BBF enhanced the substitution of synthetic fertilizer due to the targeted use of nutrients. Another LCA on digestate utilization compared four alternative BBF production scenarios in relation to mineral fertilizer production (Alengebawy et al., 2022). This study included two technologies for nutrient extraction from the solid fraction and two more for the liquid fraction. Results showed that in all scenarios, BBFs constituted environmentally beneficial alternatives compared to mineral fertilizers, provided the digestate was pretreated to remove pollutants and pathogens.

Several studies have applied LCA to assess different aspects of fertilization with sewage sludge (Yoshida et al., 2018; Ding et al., 2021). For example, Yoshida et al. (2018) assessed the long-term impacts after field application of sewage sludge by using emission factors calculated by Bruun et al. (2016). Emission factors for sewage sludges were calculated considering sewage sludges with different properties, applied to three soil types, using three precipitation regimes and varying application amounts. This approach enabled the use of region-specific emission factors in the LCA (Yoshida et al., 2018). Normalizing the LCA results to yearly per capita emissions showed that human toxicity and ecotoxicity impacts were of greatest concern, largely due to the zinc and copper content in sewage sludge (Yoshida et al., 2018).

Ultimately, regional and local conditions may determine the sustainability and feasibility of using one BBF over another. Walling and Vaneeckhaute (2020) reviewed emission factors on organic and inorganic fertilizer production and use. This study recommended that emission factors should be estimated based on case-specific data due to the high variation in emissions depending on the composition of the fertilizer and the impact of local conditions like soil type and climate. In the LCA study of Beyers et al. (2022), the environmental impact of pig slurry acidification was assessed for the climatic, agronomic, and legislative conditions of Denmark, Spain, and the Netherlands. Slurry acidification reduced the environmental impacts related to emissions of GHG and NH₃. However, the acquisition of energy and materials for the acidification process led to increased off-farm impacts in some categories, including fossil resource depletion and human toxicity. Furthermore, the effectiveness of acidification to reduce environmental impacts varied between countries due to differences in legislative requirements and energy sources. Thus, specific regional conditions (soil, climate, legislation, and farming practices) are crucial for the overall environmental sustainability of fertilization with specific BBFs. The PLCI 2.0 model has been developed to account for regional differences in LCAs of P-containing BBFs applied in European regions (Rydgård et al., 2024). This model incorporates factors such as regional soil P concentrations, soil erosion rates and distribution of crop types. Furthermore, it enables the modeling of the impact of different fertilization practices on P losses, harvesting of P in crops and the substitution of mineral P fertilizer.

In the BBF market, primary stakeholders include livestock farmers with a surplus of manure for export/processing, and crop farmers who are potential end-users of BBFs (Jensen et al., 2017; Kurniawati, et al., 2023a). Other important players are food and pharmaceutical industries, waste management companies, recycling fertilizer companies and farmers with crop residues on the supplier side, as well as garden owners and horticultural producers on the end-user’s side (Jensen et al., 2017; Venegas et al., 2021). Governments, public institutions (local/national/EU), civil society, non-governmental organizations (NGO), the food industry, investors, media, and scientists are also stakeholders in the European BBF market (Nedelciu et al., 2019).

The influence of primary stakeholders (farmers) on the supply and demand of BBFs varies. While livestock farmers are incentivized to utilize surplus manure; crop farmers exert more power in accepting or declining recycled products based on their perceptions and preferences (Case et al., 2017). Thus, defining and addressing stakeholder requirements and preferences is crucial for BBF adoption.

The acceptance of BBFs among farmers in Europe is mainly influenced by four attributes: known nutrient contents, organic matter contents, cost and ease of application (Egan et al., 2022). Negative perceptions often arise from uncertainty regarding the N, P and K contents in BBFs (Tur-Cardona et al., 2018; Egan et al., 2022), contrasting with the precision offered by mineral fertilizers. Farmers are aware of the uncertainties in nutrient contents in organic fertilizers, and they prefer BBFs with nutrient ratios that fit crop demands (Egan et al., 2022). Therefore, reliable and known amounts and ratios of nutrients are essential for facilitating adoption.

A strength of BBFs is their organic matter content, and their perceived capacity to enhance soil structure, improve soil productivity, and increase water retention capacity of soils (Case et al., 2017; Gwara et al., 2021). These benefits are well-recognized by farmers (Egan et al., 2022). However, these advantages alone are not enough to completely substitute mineral fertilizers with BBFs and need to be accompanied by the other three attributes.

The cost of BBFs and their fertilizing properties relative to mineral fertilizers play a pivotal role in social acceptance. Another advantage of BBFs is their perceived low-cost and high nutrient content (Case et al., 2017). Farmers are more likely to adopt BBFs if they are competitively priced (Egan et al., 2022). Logistical costs and perceived higher overall expenses may deter adoption, necessitating a cost at least half that of mineral fertilizers for widespread acceptance (Tur-Cardona et al., 2018). Despite potential cost differences, the ecological co-benefits of BBFs mentioned before, could sway farmers toward their adoption with proper awareness (Egan et al., 2022). We discuss such potential in the next section.

The ease of application and the practicality of BBFs also play a major role. The form of the BBF (solid, semi-solid, liquid, or granulated) significantly influences farmer preferences and acceptance (Tur-Cardona et al., 2018; Egan et al., 2022). Solid and semi-solid forms are favored over liquids due to ease of application with existing farm machinery and improved forms, such as pellets, can enhance acceptance and willingness to pay (Hills et al., 2021). Thus, farmers prefer BBFs to be granular like mineral fertilizers.

Farmers may be reluctant to utilize BBFs due to the potential negative perception from clients and consumers of their products (Simha et al., 2017). In a Polish survey, an important source of negative perception stemmed from concerns about potential health risks associated with fertilization using nutrient-rich biowastes (Smol, 2021). Consumer acceptance of products fertilized with BBFs is greater for ornamental plants than for horticultural crops for consumption (Segrè Cohen et al., 2020). This indicates that final consumers of produce fertilized with BBFs may hold negative perceptions. Media and NGOs also influence consumer opinion and their perceptions of BBF products (Jensen et al., 2017) Even if farmers are willing to use recycled nutrients, the reluctance of final consumers and the food industry to consume products fertilized with BBFs may prevent them from doing so (Barquet et al., 2020; McConville et al., 2023). Therefore, it is important to raise awareness among produce consumers and inform them about the environmental advantages of BBFs.

Although precise estimations of the European BBF market size are not available, it is evident that the BBF market is smaller compared to the conventional mineral fertilizer market. Approximately half of the EU’s fertilizer inputs for P come from mineral P, and about two-thirds of N fertilizers are produced through the Haber-Bosch process (Schoumans et al., 2015; Einarsson et al., 2021). Both studies recognize animal manures as the most significant alternative sources of P and N aside from mineral fertilizers. Therefore, the BBF market is still in early stages, requiring substantial changes in product availability, quality, legislation, competition dynamics, and stakeholder perceptions (Kvakkestad et al., 2023).

A favorable landscape is currently emerging for the development of the BBF market. Legislative developments, as outlined in section 2, have established a legal framework for BBF products in the European market, providing a conducive environment for market growth. However, a crucial question arises: where should the focus be directed? In the previous section, we outlined five criteria important for BBF adoption. Additionally, literature has emphasized that BBFs should resemble mineral fertilizers in consistent supply, allowing for a seamless transition for farmers accustomed to using commercially available mineral fertilizers (Gregson et al., 2015; Case et al., 2017; Buysse and Cardona, 2020; Kvakkestad et al., 2023).

Buysse and Cardona (2020) suggested that rising prices of mineral fertilizers would create an opportunity for the BBF industry to develop, given the higher production costs of BBFs. The fertilizer market has experienced significant volatility (Figure 3) during the period from 2021 to 2023, particularly due to rising geopolitical tensions and the war in Ukraine, leading to several increases in mineral fertilizer prices (AGRIDATA, 2023). The price increase in late 2021 and 2022 created a new reality for BBF producers, making more nutrient recovery technologies economically feasible (Hermann and Hermann, 2021). Since early 2023, fertilizer prices have been decreasing and stabilizing. At the time of writing this article, mineral fertilizer prices have returned at the levels of mid-2021. However, despite the prices being lower than in 2022, they remain higher than previous pre-2021 levels. Given the ongoing geopolitical pressures on the European Union, it is unlikely that the prices will decrease further in the foreseeable future (Alexander et al., 2022; Brownlie et al., 2023; Rabbi et al., 2023).

The pricing and competitive position of BBF compared to mineral fertilizers are influenced by many factors: demand, availability of adequate BBFs, logistics, local legislation, and regional fertilization practices (Case et al., 2017; Kvakkestad et al., 2023). Previous studies recommended pricing of BBF significantly lower than mineral fertilizers (Case et al., 2017; Tur-Cardona et al., 2018). However, Moshkin et al. (2023) considered the willingness-to-pay and perceptions of potential users in an EU farmer survey and suggested that BBFs should be priced equivalently to mineral fertilizers. While willingness-to-pay may still be lower for BBFs compared to mineral fertilizers, marketing strategies could help justify higher prices and contribute to the development of the BBF industry. Moreover, branding of BBFs plays an important role in their adoption. For example, using the term “biosolids” instead of “treated sewage sludge” positively impacts consumer attitudes and acceptance (Lu et al., 2012).