Systematic Review of Dairy Processing Sludge and Secondary STRUBIAS Products Used in Agriculture

Worldwide dairy processing plants produce high volumes of dairy processing sludge (DPS), which can be converted into secondary derivatives such as struvite, biochar and ash (collectively termed STRUBIAS). All of these products have high fertilizer equivalent values (FEV), but future certification as phosphorus (P)-fertilizers in the European Union will mean they need to adhere to new technical regulations for fertilizing materials i.e., content limits pertaining to heavy metals (Cd, Cu, Hg, Ni, Pb, and Zn), synthetic organic compounds and pathogens. This systematic review presents the current state of knowledge about these bio-based fertilizers and identifies knowledge gaps. In addition, a review and calculation of greenhouse gas emissions from a range of concept dairy sludge management and production systems for STRUBIAS products [i.e., biochar from pyrolysis and hydrochar from hydrothermal carbonization (HTC)] is presented. Results from the initial review showed that DPS composition depends on product type and treatment processes at a given processing plant, which leads to varied nutrient, heavy metal and carbon contents. These products are all typically high in nutrients and carbon, but low in heavy metals. Further work needs to concentrate on examining their pathogenic microorganism and emerging contaminant contents, in addition to conducting an economic assessment of production and end-user costs related to chemical fertilizer equivalents. With respect to STRUBIAS products, contaminants not present in the raw DPS may need further treatment before being land applied in agriculture e.g., heated producing ashes, hydrochar, or biochar. An examination of these products from an environmental perspective shows that their water quality footprint could be minimized using application rates based on P incorporation of these products into nutrient management planning and application by incorporation into the soil. Results from the concept system showed that elimination of methane emissions was possible, along with a reduction in nitrous oxide. Less carbon (C) is transferred to agricultural fields where DPS is processed into biochar and hydrochar, but due to high recalcitrance, the C in this form is retained much longer in the soil, and therefore STRUBIAS products represent a more stable and long-term option to increase soil C stocks and sequestration.

Worldwide dairy processing plants produce high volumes of dairy processing sludge (DPS), which can be converted into secondary derivatives such as struvite, biochar and ash (collectively termed STRUBIAS). All of these products have high fertilizer equivalent values (FEV), but future certification as phosphorus (P)-fertilizers in the European Union will mean they need to adhere to new technical regulations for fertilizing materials i.e., content limits pertaining to heavy metals (Cd, Cu, Hg, Ni, Pb, and Zn), synthetic organic compounds and pathogens. This systematic review presents the current state of knowledge about these bio-based fertilizers and identifies knowledge gaps. In addition, a review and calculation of greenhouse gas emissions from a range of concept dairy sludge management and production systems for STRUBIAS products [i.e., biochar from pyrolysis and hydrochar from hydrothermal carbonization (HTC)] is presented. Results from the initial review showed that DPS composition depends on product type and treatment processes at a given processing plant, which leads to varied nutrient, heavy metal and carbon contents. These products are all typically high in nutrients and carbon, but low in heavy metals. Further work needs to concentrate on examining their pathogenic microorganism and emerging contaminant contents, in addition to conducting an economic assessment of production and end-user costs related to chemical fertilizer equivalents. With respect to STRUBIAS products, contaminants not present in the raw DPS may need further treatment before being land applied in agriculture e.g., heated producing ashes, hydrochar, or biochar. An examination of these products from an environmental perspective shows that their water quality footprint could be minimized using application rates based on P incorporation of these products into nutrient management planning and application by incorporation into the soil. Results from the concept system showed that elimination of methane emissions was possible, along

INTRODUCTION
Mineral phosphorus (P) is a listed European Union (EU) critical raw material due to its importance in food production (European Commission, 2017;Espinoza et al., 2020). As agriculture is the largest consumer of mined P in Europe (1.1 million tons in 2015; Eurostat, 2020), security of supply may be challenging because the source of non-renewable rock phosphate is in geopolitically sensitive regions (Cordell et al., 2009). The dairy processing sector produces P-rich dairy processing sludge (DPS) which, when used directly or in derived secondary products such as STRUBIAS (STRUvite, BIOchar, AShes), may reduce the dependence on mined rock P (Figure 1). The European dairy processing industry processed about 144.6 million tons or 140.4 billion liters of domestic milk in 2020 (Table 1), about 46% more than the USA which is the second largest milk producing country in the world (Agriland, 2020). It is estimated that dairy food processing wastewater treatment can generate up to 20 kg (mean 17.45 kg m −3 ) DPS per m 3 of milk processed (Ashekuzzaman et al., 2019b), which resulted in 2.45 million tons of DPS (wet weight) across the EU in 2020 ( Table 1).
Phosphorus in DPS can be recycled to fields, creating a circular economy, and can be used as a replacement for mineral P fertilizer produced from mined rock phosphate (Mayer et al., 2016). Similar to other organic fertilizers, land application only occurs in growing seasons (Sommer and Knudsen, 2021), meaning that storage is required for extended periods, which increases the cost of management. Storage and land application of DPS is a source of the greenhouses gases (GHGs), methane (CH 4 ) and nitrous oxide (N 2 O) (Smith et al., preparation) 1 , and when applied to soil may result in incidental losses of nutrients along surface or near surface pathways (Fenton et al., 2017;Shi et al., 2021a). As production becomes more centralized on a smaller number of larger farms in some EU member states (e.g., Denmark or France), the land bank opportunities for DPS application may become limited depending on the land use change and farmer willingness to accept DPS. Due to the nutrient content value and related transport costs, a land bank radius of about 10 km around a processing plant pertains. This may cause an oversupply to land areas near the processing plant and an increased risk of P losses to water (i.e., critical source areas). In addition to centralization of production, future application rates of fertilizer will be limited by P application, and not by nitrogen (N) rates. Presently, the limits imposed on N application rates have led to significant over-application of P (i.e., P:N ratio higher than the ratio needed by crops) and increased potential for eutrophication in regions with a high proportion of sludge and slurry production (Lu et al., 2012). As an example, in Brittany, France, the issue of P and eutrophication is of major importance due to the local high permeability bedrock (granite, shale and sandstone) and agricultural (high level of livestock density) context, which prevents localized organic fertilizer land application. This means there is not always a match between DPS and suitable land availability (e.g., P-deficient soils) in the local area near the dairy processing plant (Le Noë et al., 2018). Such conditions have created a market for the transport and application of DPS e.g., the total cost of spreading DPS (including transport, analyses, spreading and monitoring) is between 20 and 30 e per ton in France (Laperche, 2014). Any new regulations pertaining to the dairy industry will incentivise producers to reduce the volume and weight of DPS transported over long distances to end users. The post processing of DPS to produce secondary STRUBIAS products is a potential solution in this regard.
The main objective of this systematic review is to collate information that will help give DPS and their secondary products certification as P-fertilizers in accordance with technical proposals for new fertilizing materials under the Fertilizing Products Regulation (European Commission, 2003Huygens et al., 2019). The purpose of this systematic review is to contribute to decision making about the most sustainable FIGURE 1 | Global P resources and consumption of P and associated P usage streams in the EU in tons per year. Based on Ott and Rechberger (2012), Scholz and Wellmer (2013), and Schoumans et al. (2015).  Walstra et al. (2005) was used for mass to volume conversion; c Median P and N concentration of 35.9 and 48.9 g kg −1 (dry weight) and median dry matter content of 14.6% were used estimate total P and N content, respectively, data source: Ashekuzzaman et al. (2019a).
transformation of these "wastes" into high value bio-fertilizers used by both traditional and organic farmers.

Search Criteria
All search results were evaluated using the PRISMA statement (Page et al., 2021). In a systematic review, information was collated to answer the following two research questions: (1) how DPS and STRUBIAS production on in-situ and ex-situ treatment and processing units affect product characteristic, fertilizer replacement value (FEV) and P dynamics in soil, risks of GHG emissions and pathogen and heavy metal pollution?
(2) how to combine this information to develop production systems necessary to certify DPS and STRUBIAS products as

Screening of Search Results
Duplicates were removed manually and abstracts were screened by two screeners against the target research questions. Exclusion criteria, described below, were developed and selected and cross-checking was performed on these excluded articles.
If disagreements between the two screeners occurred, a third screener adjudicated. Full-text review was independently conducted by three reviewers and reasons for exclusion were annotated and tracked (e.g., "data pertaining to a different waste other than DPS or STRUBIAS"). The primary reasons for excluding papers were: (i) articles completely un-related to search questions; (ii) general knowledge papers; and (iii) papers that did not follow the basic criteria of scientific research (e.g., experimental design with sufficient replication). Articles clearly meeting the inclusion criteria were obtained for full-text review unless unavailable. These included articles related to the search inquiry, providing that scientific laboratory experiments or field studies had a minimum number of replicas and a negative control. Articles were not considered further when their title and abstract clearly indicated that the study did not meet the inclusion criteria (see Supplementary Figure 1). The studies included in tables consist of those that were reviewed in detail.

Greenhouse Gas Emission
Some data needed for the holistic review were not available in the literature and needed to be developed outside of the systematic review within the present study. Greenhouse gas emission from DPS sludge and secondary STRUBIAS products was calculated using a whole systems approach. Herein, methods presented calculate CH 4 , N 2 O, and ammonia (NH 3 ) emissions from sludge and secondary STRUBIAS products from production to field application. Ammonia emission is included in the calculation due to risk of N 2 O emission from NH 3 deposition to land. In this analysis, various conceptual scenarios (from production to field application) are compared with the aim to find potential scenarios that can minimize or eliminate emissions. Assumptions for management of sludge and STRUBIAS products for a Danish dairy production site are used in the scenarios as follows: each month 1/12 of 1 ton of annual produced standard sludge is transferred to a store and the temperature in the stored sludge is similar to the monthly average air temperature of Denmark. The tanks with stored sludge are emptied at the start of April and subsequently land applied to fields in April. In the scenarios where sludge is processed to biochar or hydrochar, the same monthly amount is treated and the products are land applied in April. The annual amount of DPS treated is 1 ton.

Calculation of CH 4 Emission
Methane emissions from stored liquid sludge is calculated with the CH 4 emission model as used in the Integrated Farm Systems Model (IFSM) for manure management in beef and dairy production systems (Chianese et al., 2009). This has been used previously to assess the impact of GHG reduction strategies in agriculture (Rotz and Hafner, 2011;Dutreuil et al., 2014). Equation 1 is used to calculate CH 4 emission from anaerobic stored livestock liquid manure and digestate from biogas plants if concentration volatile solids (VS) and air temperature is known (Baral et al., 2018): where F CH4 is CH 4 emission rate (g CH 4 kg −1 VS day −1 ), A is the pre-exponential factor of 31.2 g CH 4 kg VS −1 h −1 , Ea the apparent activation energy set to 81 kJ mol −1  giving the temperature response of CH 4 production. R is the gas constant (8.314 J mol −1 K −1 ) and T the temperature (K). The degradable fraction V SD is 56% of VS in the sludge (Smith et al., preparation) 1 . This equation and the parameters presented here is used to calculate CH 4 emission from stored DPS.

Emission of NH 3 and N 2 O
In the Danish GHG emission inventory (Nielsen et al., 2018), it is assumed that 0.5% of total N is emitted in form of N 2 O from stored liquid manure and sludge, and 1% of total-N is emitted from slurry and sludge applied to soil [based on the standard emission factors given by IPCC (2019)]. Another assumption is that N 2 O emissions from soil N are unaffected by any application of biochar or hydrochar applied to soil. When struvite is applied, it is assumed that 1% of the ammonium (NH + 4 ) is emitted as N 2 O as is the case for mineral fertilizer N applied to soil (IPCC, 2019).
During storage of the sludge 34% of the proteins are transformed to total ammonium nitrogen (TAN = NH 3 + NH + 4 ) (Mottet et al., 2010) and may be emitted from the storage. Emission of NH 3 from stores with a cover of PVC roof are set to 2.6% of TAN i.e., similar to emissions from covered stored liquid manure (Hansen et al., 2008) and of sludge injected into black soil to 2% of TAN (Olesen et al., 2020). The negative charge of biochar will contribute to a negligible NH 3 emission from soils to which hydro-biochar is applied (Chu et al., 2019), biochar from pyrolysis do not contain TAN and the NH + 4 in struvite is assumed not to volatilize due to the low pH (Sommer et al., 2004). At deposition on land or water, a fraction of NH 3 will be transformed to N 2 O and be emitted to the atmosphere. This indirect N 2 O emission is estimated using Equation 2 (IPCC, 2019): where F N2O is given in kg N 2 O, F NH3 is given in kg NH 3 -N emitted, 0.01 is a default factor given by the IPCC for calculation of the climate warming effect of emitted NH 3 , and 44/28 is to calculate from concentration given in 2 * N g mol −1 to 2 * N + O g mol −1 (IPCC, 2006).

Carbon Sequestration
During storage of untreated sludge, a fraction of C is lost in form of carbon dioxide (CO 2 ) and CH 4. This emission is calculated assuming that 2.6 kg VS is lost in the form of CO 2 and CH 4 for each kg of CH 4 produced. The sludge is heated in hydrothermal carbonization due to the oxidation of C, and in this process about 30% of C is lost in the form of CO 2 , and about 70% of C in treated biomass is retained in the hydrochar (Kambo and Dutta, 2015). In the pyrolysis process, more CO 2 is lost and between 25 and 35% (avg. 30%) of solids (40% C) are retained in the biochar (Kambo and Dutta, 2015). In the calculations used herein, the C retention of field applied sludge-C was set to 25%, which is between 12% of manure C input in the longer term (avg. 18 years; Maillard and Angers, 2014) and 35% for a 20-year period by scaling results from a study of transformation of C in digestate applied to soil (Thomsen et al., 2013). The longer retention time of C in sludge than in animal slurry was due to sludge organic matter (OM) from a wastewater treatment plant that has been transformed to a more stable form of C. It is calculated that C concentration in VS is 517 g C kg VS −1 , and this estimate is used to calculate C concentration in sludge, where OM is measured as VS.
It has been calculated that between 90 and 97% of the C in biochar from pyrolysis of a range of different biomasses will still remain in the soil after 100 years (Lehmann and Joseph, 2015;Wang et al., 2016). No study of the recalcitrance of biochar from pyrolysis of dairy sludge was found in the literature, and a conservative/cautious estimate is that 90% of the C in biochar from pyrolysis (PC) of dairy sludge is recalcitrant.
Hydrochar from hydrothermal carbonization (HTC) is produced at a lower temperature than when producing biochar by pyrolysis, and C component in the hydrochar has been oxidized less than pyrolysis products, i.e., less oxygen is added during the process. Our assumption is that C in hydrochar is less recalcitrant than biochar produced by pyrolysing dried biomass, and that only 50% C in hydrochar is recalcitrant. This assumption is supported by the study of Malghani et al. (2013), who stated that "although both HTC and PC chars were produced from the same feedstock, PC chars had markedly higher potential for carbon sequestration than HTC."

Treatment Options at Dairy Processing Plants and Volumes and Composition of DPS
The dairy processing industry generates a large volume of waste, which is high in OM. Discharge licensing, which is site specific, aims to prevent a reduction in surface water oxygen (O 2 ) concentrations and the onset of surface water eutrophication (Neal and Heathwaite, 2005). To achieve such discharge thresholds, a chain of treatment before discharge at dairy processing plants is needed (Figure 2). In fact, most DPS applied to land in the EU is first treated, composted or incinerated before being applied to land (Figure 2). Research now focusses on converting DPS into secondary STRUBIAS products (Figure 2).

Dairy Biological Wastewater Treatment
A schematic of typical dairy wastewater treatment processes and DPS generation is shown in Figure 2. Pre-treatment removes particulate matter and retains fats that could interfere with subsequent treatment. The pre-treatments and physio-chemical treatments most frequently used on dairy processing plants are buffer tanks, neutralization, sieving, flotation and degreasing (Droste and Gehr, 2018). Flocculation (coagulation) is the most simple and economical pre-treatment method, and reduces water turbidity by reducing particulate substances and fats, which could interfere with subsequent treatments (Carvalho et al., 2013). Pretreatment by flocculation enhanced by the presence of lactic acid bacteria (Lactobacillus plantarum), which ferment the lactose and produce lactic acid. The acidity of this compound precipitates the milk proteins, and thus significantly reduces the chemical oxygen demand (COD). This reduction increases with addition of flocculants such as chitosan or carboxymethyl cellulose (CMC) to this solution. The COD removed varied by 65-78% for CMC and 49-82% for chitosan (Dyrset et al., 1999).
In the secondary treatment, sludge undergoes aerobic treatment that is cost efficient and controllable (Kolev Slavov, 2017). This treatment step can reduce more than 90% of COD (Carta-Escobar et al., 2004;Kushwaha et al., 2011;Carvalho et al., 2013). In aeration ponds, the aerobic pathway uses microorganisms contained in biological reactors, which in the presence of O 2 , degrade the OM suspended in the sludge.
The soluble organic components transformed and emitted mainly in the form of CO 2 and NH 3 , while the insoluble pollution fraction, including the microorganisms, is recovered by separation and form "activated" sludge (Sustarsic, 2009). A large number of aerobic treatment types are being used e.g., rotating biological contractors (RBCs), sequencing batch reactors (SBRs), or membrane reactors (MBRs) (Goli et al., 2019). In particular, SBRs are effective at varying loading capacities (Kolev Slavov, 2017). In this step, a single tank used for filling, aeration, settlement and effluent withdrawal, and recycling of solids. The sludge produced during primary and secondary treatment may be treated anaerobically in biogas reactors. During this process, OM is transformed by microorganisms to CH 4 and CO 2 . This process produces small volumes of sludge (<0.05 kg of dry matter (DM) per kg of COD eliminated) (Omil et al., 2003). The most common anaerobic reactors are up-flow anaerobic sludge blanket (UASB), completely stirred tank reactors (CSTRs) and membrane anaerobic reactor systems (MARSs). Finally, where needed, the N, P, micropollutants or pathogenic microorganisms in the wastewater can be further reduced using a finishing treatment e.g., for P content reduction alternation between aerobic/anaerobic conditions or P coagulation and precipitation by ferric chloride or aluminum sulfate can be deployed (Rivas et al., 2010). After these treatments, the particles settle in a clarification pond. The supernatant is discharged, while the precipitated sludge is dewatered after addition of polymers to promote flocculation and separation between water and suspended matter (dewatering). This is then thickened (thickening) during its passage through dewatering grids. This sludge is stored in tanks or treated further and then used as a bio-based fertilizer (Huygens et al., 2019).

Wastewater Volumes and Contents
The volume of wastewater produced at processing plants can be high and is product-dependent. For example, 1 liter of processed milk can produce up to 10 L of effluent (Lateef et al., 2013). In the cheese manufacturing industry, the whey is the main pollutant discharged to water and soil. This is mainly due to its high carbohydrate content (4-5%) of which lactose is the main constituent (Kolev Slavov, 2017).
Consequently, DPS produced from cheese production, called Cheese Whey Wastewater (CWW), contains high concentrations of organic components contributing to a high COD and biological oxygen demand (BOD) (Ahmad et al., 2019). The COD of CWW is higher (0.79-77.3 g L −1 ) than COD of milk plant effluent (0.183-10 g L −1 ) (Carvalho et al., 2013) or milk-treated condensate wastewater (<0.001 g L −1 ), which is the water obtained during the concentration and evaporation processes of milk and its by-products (Bourbon and Huet, 2018). The CWW contains some milk or milk by-products, oils and greases, and cleaning water containing sterilizing agents, acid and alkaline detergent (Carvalho et al., 2013;Ahmad et al., 2019). As a result, the concentration of inorganic compound may be heterogeneous,   (Danalewich et al., 1998;Omil et al., 2003;Demirel et al., 2005;Byrne, 2011;Carvalho et al., 2013;Karadag et al., 2015;Kolev Slavov, 2017;Verma and Singh, 2017;Goli et al., 2019;Ferreira et al., 2021;Sivaprakasam and Balaji, 2021). with ranges observed for P from 8 to 510 mg L −1 and N from 14 to 1,462 mg L −1 (Demirel et al., 2005). The amount and composition of DPS is determined by the composition of milk and use of additives, the dairy production line and cleaning and disinfectant products used for cleaning. In addition, wastewater treatment, varies between plants (Figure 2), and depends on the type of dairy processing involved (Rico Gutiérrez et al., 1991;Karadag et al., 2015;Ahmad et al., 2019;Shi et al., 2021a). Recently, physical and chemical characteristics of 63 DPS samples (9 dairy processing plants in Ireland) were reported (Ashekuzzaman et al., 2019a). The main DPS types included in the study were bio-chemically treated activated sludge leading to aluminum-precipitated sludge (Al-DPS) and iron-precipitated sludge (Fe-DPS) depending the dosing of alum or ferric salt to remove P, and to lime-stabilized calcium-precipitated sludge (Ca-DPS) generated after dissolved air floatation (DAF). In some processing plants, mixing of DPS generated from aerobic biological wastewater treatment and DAF processes occurs before land application or further disposal processing. A few of the examined plants have anaerobic digesters (AD) that produced   AD sludge. That concentration of components varies between sludge categories and treatments and differences in composition was highest for N, P and K (Ashekuzzaman et al., 2019a). This difference is higher between activated and DAF sludge types than those in the combined sludge (sludge mixed from both activated and DAF process). The addition of chemicals such as Fe-, Al-, or Ca-based coagulants during the removal of P from the effluent may explain these results. In addition, the N concentration was lower in DAF sludge compared to AD sludge in contrast to the increased P concentration (Ashekuzzaman et al., 2019a). The addition of lime during the DAF process explains this result by an increase of the pH, which promotes the volatilization of N in the form of NH 3 . On the other hand, anaerobic degradation during the AD process increases the TAN concentration in the sludge due to organic N mineralization. Tables 2, 3 illustrate the physicochemical characteristics of dairy processing effluents and DPS reported in the literature. These characteristics highlight the variability in dairy effluent compositions related to the type of bio-products processed (i.e., milk, cheese, yogurt, and butter) (Omil et al., 2003;Carvalho et al., 2013;Karadag et al., 2015), and the wastewater treatment processes used (Britz et al., 2006;Ashekuzzaman et al., 2019a).
In France, information and data pertaining to the agronomic benefits and risks of applying organic waste to agricultural soils was collated by Houot et al. (2014) and used to re-evaluate the EU Sewage Sludge Directive (86/278) (CEC, 1986). This work emphasized the need for sludge used in agriculture to have P recycling as a main priority, and that this use must not be a risk to the environment or to human health due to their contents of heavy metals, organic trace compounds, pathogenic microorganisms and pharmaceutical compounds. To avoid some of these concerns, the EU Council Directive 86/278/EEC set limits for the content of heavy metals (Cd, Cu, Hg, Ni, Pb, and Zn) (CEC, 1986), and individual European countries have set limits for synthetic organic compounds and pathogens (Hudcová et al., 2019).
In a comprehensive study across nine Irish dairy plants, the concentration of heavy metals (i.e., Cr, Cu, Ni, Pb, and Zn) was examined in all major DPS types with lowest concentrations found in DAF sludge and highest in AD sludge (Ashekuzzaman et al., 2019a). Overall, the heavy metal concentrations across all tested DPS samples were significantly lower than limits set by the EU for avoiding accumulation in agricultural soil to which sludge is applied (CEC, 2008) and the levels were below those of livestock manure (Sommer et al., 2013), and composts (Bernal et al., 2017). The results of the Irish study are in line with the current knowledge on heavy metals content of DPS ( Table 4) and indicates that heavy metal concentrations will not be a limiting factor for legal and safe application rate of DPS to agricultural soils (Ashekuzzaman et al., 2019a;Shi et al., 2021b). The concentration varies between dairies and this is due to the diversity of the milk bio-products and the various possible steps in the treatment of the effluent. It is important to have knowledge pertaining to the heavy metal content of DPS and DPS-derived STRUBIAS products before land application, because farmers and society must be assured that the heavy metal content is lower (in soil and plants) than the limits given for use before making The values are collected from peer reviewed articles where data were given in as ranges  final decisions and rules of use of the waste as a fertilizer (Shi et al., 2021b). Organic Trace Compounds (OTCs) are chemical products (hydrocarbons and their derivatives, degradation products, solvents, etc.) present in organic waste or derived due to degradation of the organic compounds by the microorganisms in sewage treatment plants or in the soil. They often accumulate by biomagnification and bioaccumulation in biological organisms and cause irreversible damage to biological systems. They are directly or indirectly toxic to humans and animals (such as endocrine disruption and tumor initiation) (Barret et al., 2012). In Europe, there are proposals for limits on polycyclic aromatic hydrocarbons (PAHs) contents in municipal sewage sludge applied to land (CEC, 2010). Depending on the countryspecific regulations, the type of OTCs and the threshold limits differ. Where organic waste is land applied, the following three PAHs i.e., fluoranthene 5, benzo(b)fluoranthene 2.5, benzo(a)pyrene 2 and 7 PolyChloroBiphenyls (PCBs) (PCB 28,52,101,118,138,153,180) are considered good indicators of compound resistant to biodegradation, and according to French regulation must be below concentration limits (Hudcová et al., 2019). In contrast, German authorities do not regulate PAHs, while the threshold for the sum of nine PAHs is more stringent in Denmark than for the sum of three PAHs in France. Companies that handle wastes must be aware that the concentration of 16 PAHs depends on the type of sludge (Boruszko, 2017). Research has shown that anaerobic fermentation and post-flotation may reduce the content of PAHs up to seven times its original concentration (e.g., 689 µg kg −1 of DM) (Boruszko, 2017), with most of the reduction of hydrocarbons taking place in the final phase of fermentation. The concentrations of PAHs in DPS are low and do not exceed the amount allowed by the European Commission (CEC, 2009; Table 5). Therefore, their use in agriculture will not be limited by PAHs (Pérez et al., 2001;Boruszko, 2017), but it is recommended that these are investigated in DPS-derived STRUBIAS products.
DPS contains living microorganisms originating from the treated wastewater. There are pathogens (viruses, bacteria, fungi and parasites) derived from animal manure (Sobsey et al., 2006), and if dairy cows ingest grass from fields where dairy manure has been applied, then there is a risk for transfer of pathogens in infectious levels to the cows and subsequently to milk and therefore to wastes. The European Commission (CEC, 2000) has in its third draft of the Working Document on Sludge, proposed the following thresholds for a range of bacteria and worms for sludge to be recycled to soil: (1) E. coli <5 × 10 5 colony forming units (CFU) per gram (wet weight) of conventional treated sludge; (2) for advanced sanitized sludge E. coli must be below 1 × 10 3 CFU g −1 wet colony of treated sludge; (3) Salmonella Senftenberg W775 in sludge spiked with this microorganism must be reduced 99.99%; (4) no content of Ascaris ova (5) a sample of 1 g DM of the treated sludge must not contain more than 3 × 10 3 spores of Clostridium perfringens, and (6) a sample of 50 g (wet weight) of the treated sludge must not contain Salmonella spp.
There are few studies reported about reduction of pathogens in raw DPS. Laboratory studies have quantified reductions of microbial infectivity (inactivation) in animal organic wastes under controlled temperature conditions and the samples have been stored aerobically or anaerobically (Sobsey et al., 2006). For example, the high initial level of pathogens (Enterobacteriacea, fecal coli forms and E. coli) in dairy slurry was higher than in sludge from two urban wastewater treatment plants where anaerobic digestion was followed by mechanical dehydration (one treatment also received a heat-dried process). The concentrations in treated dairy slurry were 5.1 × 10 7 CFU g −1 of DM of Enterobacteriacea; 4.4 × 10 7 CFU g −1 of DM of fecal coli forms, and 4 × 10 6 CFU g −1 of DM of E. coli and all are below the limits set by the regulation. Concerning the persistence of these pathogens in the soil, then after a 80-day trial across soil/sludge treatments, the populations of fecal coliforms and E. coli decreased considerably or were not detectable (Estrada et al., 2004). The study of Ravva et al. (2006) also found that the pathogenic strain E. coli O157:H7 introduced in water from on farm dairy waste lagoons, failed to establish and proliferate in dairy wastewater microcosms with or without circulating aerators. On the other hand, high concentrations of Listeria are found in manure and sewage sludge and have survived in topsoil between 12 and 182 days (Sobsey et al., 2006). If these are present in DPS, then additional sludge treatments such as anaerobic digestion, hygienization by adding lime, or composting will reduce the concentration of pathogens to allowable values.
Dairy processing wastes (even after treatment at source) may contain harmful substances, which need testing and quantification across all the DPS and STRUBIAS types. The substances in focus should be antimicrobial drugs, hormones, pesticides, emerging contaminants, pathogens, disinfectants, persistent organic pollutant residues, microplastics, and nanoparticles in DPS or DPS-derived STRUBIAS (Shi et al., 2021a).

DPS-Derived STRUBIAS Production
Another strategy that is being deployed to manage DPS is to further process these raw products into other more usable and stable forms. Struvite (magnesium ammonium phosphate hexahydrate, MgNH 4 PO 4 , 6H 2 O) is widely used in agriculture due to its N and P content, which is in a form that efficiently can be used by plants (Adam et al., 2009). Phosphorus can exist as particulate and dissolved species in both organic and inorganic forms. The inorganic P species is mainly in orthophosphate form, which is plant available and important for soil fertility but can be readily lost to the environment (Frossard et al., 1996). However, the chemical composition of struvite obtained from DPS is not always consistent with pure struvite equivalents (Hall et al., 2020). Metal impurities such as Al, Fe, Ca, and small amounts of heavy metals can precipitate along with the struvite and could pose problems later for crops and soil when land applied.
The term "char-based materials" is used here to replace "biochar" in the STRUBIAS acronym as they have different terms depending on the technology. Char-based materials, obtained from the thermochemical conversion of biomass in an oxygen-depleted atmosphere, are porous and carbonaceous, and are more stable and C-rich and less toxic than the feedstock (Kambo and Dutta, 2015;Atallah et al., 2020). The significance of thermochemical treatment lies in overcoming the structural inferiority of biomass, which enhances the chances of energy and resource recovery from waste (Kambo and Dutta, 2015). There are many functions of char-based materials including, but not limited to, energy production, agriculture, C sequestration, wastewater treatment, and biorefinery (Kambo and Dutta, 2015). The utility of a specific char-based material for any particular application depends on its inherent properties. Feedstock, pre-treatment method, and temperature are all important (Amoah-Antwi et al., 2020). However, thermochemical treatments increase the risk of producing chars with other highly toxic compounds produced from high-temperature reactions such as PAHs, PCBs, dioxins, furans, and PCDD/Fs (Kambo and Dutta, 2015;Amoah-Antwi et al., 2020). Heavy metals present in the feedstock are most likely to remain and concentrate in the chars (Shackley et al., 2010).
Ashes are characterized as fly ash or bottom ash, or a combination formed through the incineration of biowastes by oxidation (Huygens et al., 2019). Ash normally contains valuable plant macronutrients such as K, P, S, Ca, and Mg (Haraldsen et al., 2011;Knapp and Insam, 2011;Brod et al., 2012). In addition, they contain large amounts of P (13.7-25.7% P 2 O 5 ), which are comparable to commercial superphosphate (Xu et al., 2012). Obstacles to the use of ash as a fertilizer or soil amendment could be their heavy metal contents (Franz, 2008;Herzel et al., 2016).

DPS and DPS-Derived STRUBIAS as Fertilizers
DPS and DPS-derived STRUBIAS products are used or research is underway to ascertain their potential as bio-based fertilizers in agriculture (Shi et al., 2021a). DPS is typically stored off site until applied to land in spring, whereas STRUBIAS products can be processed and stored until needed. Many knowledge gaps still exist pertaining to their respective effects on P dynamics once spread onto soils and their FEV. It should be noted that FEV is used herein (Shi et al., 2021a), but can be often known in the literature as Mineral Fertilizer Equivalence (MFE; Delin, 2012) or Mineral Fertilizer Replacement Value (MFRV; Schröder et al., 2007).

Phosphorus Dynamics in Agricultural Soils
In cropped agricultural systems, P applied to soil with fertilizers can be utilized by crops, absorbed by soil minerals (González Jiménez et al., 2019), or lost along surface (runoff) or subsurface (leaching and loss along natural or artificial lateral transport or deeper recharge to groundwater) pathways to surface waters (Murnane et al., 2016). Soil P transformation passes through several interconnected pools. These are the soluble P pool, which is considered to be immediately available for plants; labile or weakly adsorbed P, insoluble P chemically bound with Ca ions in calcareous and alkaline soils or occluded by Fe and Al oxides in acidic soils, P strongly adsorbed by hydrous oxides of Fe and Al, and insoluble organic P within soil organic matter (Stevenson and Cole, 1999;Bennett and Carpenter, 2002). Figure 3 presents a simplified diagram reflecting P soil cycling and interactions between these pools. Briefly, the soil P cycle consists of the following processes: weathering and precipitation, mineralization and immobilization, adsorption and desorption, and P losses through surface or near surface runoff and subsurface leaching with eventual recharge to groundwater (the proportions of which are dependent on soil/subsoil/bedrock permeability and chemistry).
Mineralization and immobilization of P are part of the organic P cycle. Mineralization is a process of transformation of organic P to soluble H 2 PO − 4 or HPO 2− 4 . Mineralization of organic P slowly releases soluble P, which is crucial during the growing season as it provides a continuous supply of P to crops. Mineralization of organic P in soil occurs through breakdown of organic bonds, which is driven by the release of enzymes produced by plants and soil microflora. Organic P mineralization is driven by phosphatase enzyme activity in soil, which mainly occurs during the growing season when soil temperature ranges between 18 and 40 • C (Prasad et al., 2016). Phosphatases synthesis are believed to be driven by P availability and enhances under P limiting conditions . However, in some cases application of mineral P (Paredes et al., 2011), and organic amendments of soil (Parham et al., 2002) can increase phosphatases activity. Some other factors which have an impact on phosphatases production are P availability and availability of other soil nutrients (Marklein and Houlton, 2012), soil moisture, pH, and availability of other soil nutrients, and energy supply (Acosta-Martínez and Waldrip, 2014; Prasad et al., 2016).
Precipitation and dissolution and desorption and absorption are part of inorganic P cycle. The direction of P transfer between inorganic P soil pools though precipitation and dissolution can be either reversible or irreversible, and can be impacted by a number of factors, including geochemical soil composition and soil pH. For instance, in acidic soils P precipitation occurs in the presence of Fe, Al, and Mg, and soluble P in such soil can be limited, while in alkaline soils precipitation primarily occurs through reactions involving Ca 2+ compounds (Prasad et al., 2014). Adsorption, or fixation, binds soluble P compounds to soil particles, whereas desorption releases P which is bound with soil minerals to soil solution, thereby increasing the soluble P pool. Unlike precipitation, this process is reversible, and P does not involve permanent change in chemical and structural changes in P-containing compounds and soil minerals.
The consideration of the aforementioned P fluxes in agricultural soil and recycled DPS composition is essential for developing guidelines of alternative P fertilizers. Specifically products derived from chemically treated dairy effluents treated with lime, ferric sulfate or aluminum chloride may contain elements, which can limit P release into available P pool such as Ca, Fe, and Al (Ashekuzzaman et al., 2019a). An inherent soil pH range optimal for P fertilizers to remain in soluble pool is between 6 and 7.5. Decreasing soil pH can lead to soluble P fixation by Fe and Al oxides. While fixation can be a limiting factor for soil P availability for crops (Daly et al., 2015;Prasad et al., 2016), fixation of P by minerals present in the soil is also a limiting factor. To ensure sustainable use of the P source and avoid P losses into the environment, such best practice should be followed (Science Communication Unit University of the West of England Bristol, 2013; Arenas-Montaño et al., 2021).

Fertilizer Equivalent Value of DPS and DPS-Derived STRUBIAS
The FEV defined as the equivalent application rate of an inorganic fertilizer achieved by an organic waste to achieve the same crop yield or nutrient uptake (Brod et al., 2012). The efficiency of most bio-based organic fertilizers is lower than inorganic fertilizers because of their slow nutrient release rates (Chen, 2006). The FEV of an organic fertilizer can both provide a quantitative estimate of the amount of efficient nutrients in biobased fertilizer and estimate of the actual value when compared with a chemical equivalent. This information, which is currently lacking, would give growers accurate information to help with nutrient management planning on farms.
Two methods used to assess the FEV of bio-based fertilizers such as DPS or DPS-derived STRUBIAS products, are pot or field-scale studies, which include different fertilizer rates, crops, and soils. The most common method is to compare yields or nutrient uptake results from DPS or DPS-derived STRUBIAS treatments with uptake from commercial mineral fertilizers as used with other organic fertilizers e.g., Lalor et al. (2011) examined the FEV of dairy cattle slurry. Typically, data fitted to linear, quadratic, or cubic polynomial regressions, creates a relationship equation. For example, Figure 4 illustrates a fitted polynomial function, describing crop yield or nutrient uptake corresponding to different mineral fertilizer application rates. This is the method used to determine the corresponding mineral fertilizer rate (×1) to any crop yield or nutrient uptake by a bio-based application. The mineral fertilizer rate, ×1, expressed as a percentage of total nutrient applied from that bio-based treatment and estimates the FEV. Alternatively, calculation of FEV by the apparent nutrient recovery method without the need of a response curve is used. There is, however, a difference between apparent N or P recovery (ANR or APR) and N-P FEV. The first is the N or P fraction taken up by the test crop of total applied nutrients and the second is the ratio of the apparent N and P recovery of bio-based fertilizer and that of mineral fertilizer at the same rate (Cavalli et al., 2016;Sigurnjak et al., 2019). They are determined as follows using Equations 3-6: NFEV (%) = ANR DPS TREATMENT ANR Mineral N TREATMENT × 100 (4) APR = P uptake TREATMENT − P uptake CONTROL total P applied TREARMENT (5) The most comprehensive grassland study on the FEV of DPS, conducted by Ashekuzzaman et al. (2021a,b), examined two main types of DPS. The first is aluminum or iron-precipitated activated sludge (Al-or Fe-DPS) and the second is a limestabilized calcium-precipitated sludge (Ca-DPS). At field scale, an assessment of N and P availability for crop yield and uptake in comparison to reference mineral fertilizers over one seasonal year was undertaken. Ashekuzzaman et al. (2021b) found N-FEV of 22-25, 54, and 8%, respectively, for Ca-DPS, Fe-DPS, and Al-DPS. They indicated that N-FEV varied between activated and lime treated DPS types, as affected by wastewater and sludge treatment processes and storage. The different treatments affect the proportion of mineral and organic N in the DPS, and thus the available N pool in amended soil. With regards to P availability, the results of Ashekuzzaman et al. (2021a) show that first-year cumulative P availability (over the four harvests) differs significantly between Al-and Ca-DPS where Al-DPS P-FEV was 109% compared to mineral P (applied at 40 kg P ha −1 ) and Ca-DPS P-FEV was only 31%. Their findings show that mineral P fertilizer was a better starter fertilizer that at application provided more readily available P for plant uptake than either Al-DPS or Ca-DPS, as they observed 50 and 16% P-FEV for the two DPS, respectively, in the first harvest. Although the Al concentration (1,122 mmol kg −1 ) in Al-DPS did not limit firstyear P bioavailability, the initial nature of P fractions, and their biological and bio-chemical mineralisation processes, might be the reason of lower P availability for immediate uptake by plant. For Ca-DPS, high Ca content (Ca/P molar ratio 1.86) and alkaline pH in Ca-DPS was likely to be associated with formation of low soluble Ca-P compounds and low P availability. Future studies on the aspect of P composition and mineralisation process in DPS would help to realize and correlate P uptake efficiency.
Other literature pertaining to the FEV of DPS, and especially DPS-derived STRUBIAS products is still limited. Many factors FIGURE 4 | Illustration of a FEV idealized response curve, where "a" is the intercept (crop yield or nutrients uptake at 0 kg ha −1 of mineral fertilizer); "b, c, and d" are the linear, quadratic, and cubic coefficients, respectively. affect the calculation of FEV such as the treatment processes used to produce a DPS or DPS STUBIAS type (crop type, fertilizer application rate, duration of experiment and scale of experiment (pot vs. field) (Brod et al., 2012;Cerný et al., 2012).
The dosing of Al or Fe salts used to capture P in sewage wastewater treatment plant affect P-FEV, because high concentrations (more than 2,800 mmol kg −1 ) of either or both may significantly reduce P bioavailability (Khiari et al., 2020). The P fertilizer effects of 14 different bio-based fertilizers had been tested in pot experiments with ryegrass (Delin, 2016). At the first cut, the P-FEV of Fe-and Al-precipitated sewage sludge were 37 and 33%, respectively (Delin, 2016). Falk Øgaard and Brod (2016) found that P-FEV of 11 sewage sludges treated with Al and/or Fe salts varied significantly between sludges, but was low for all sludges in a pot experiment with ryegrass. It was lowest at first cut, where it ranged from 2 to 24% (Falk Øgaard and Brod, 2016). Both studies indicated that sludge derived from a treatment with Fe had a higher P-FEV than when coagulation with Al salts occurred. This is due to a higher solubility of Fe phosphate compared to Al phosphate. The amount of Fe used is also important and sludge with a Fe:P ratio at 1:6 contains more plant available P than sludge with a higher ratio (e.g., Fe:P ratio of 9:8) (Kahiluoto et al., 2015). Calcium is another element that has an effect on the P availability. High dosing of Ca in the wastewater with a Ca:P ratio of 2:1 reduce the P-FEV due to formation of Ca-P compounds such as hydroxylapatite, which has a low solubility, an effect shown when using the sludge to produce compost and biochar (Nest et al., 2021). As mentioned above, solubility of the P crystals are affected by pH and liming increases the plantavailable P in sludge produced from the wastewater treated by Al or Fe salts (Krogstad et al., 2005;Montgomery et al., 2005;Bøen and Haraldsen, 2013).
When processing DPS to a STRUBIAS product, the FEV will change. This is influenced by the untreated DPS physiochemical characteristics and the processing methods and parameters used. For example, the P-FEV in ash produced from incineration is low when wood is used (P-FEV, 8-16%), but gets higher using chicken manure (P-FEV = 13-39%) (Yusiharni et al., 2007), and highest when incineration of biogas residue is used (MFE = 76-99%; Kuligowski et al., 2010). The plant availability of P in thermochemical products such as ash and biochar depends on the temperature during combustion/pyrolysis, and is halved by increasing the incineration temperature from 400 to 700 • C, which is due to hydroxyapatite formation (Thygesen et al., 2011). To increase the amount of P in STRUBIAS products, the use of flocculants or biological processes to increase P availability in raw DPS whilst avoiding high temperatures during the production of STRUBIAS could be implemented.
For N-FEV, the proportion of ammonium N (NH + 4 -N) to the total N content and the C/N ratio of the DPS or DPSderived STRUBIAS products are the most important factors for the FEV (Sommer et al., 2013;Webb et al., 2013). Ammonium is immediately available for the crop and is often a growth-limiting factor (Brod et al., 2012;Gómez-Muñoz et al., 2017). When fertilizer rates increased, the N-FEV of meat and bone meal (MBM) and composted fish sludge (CFS) decreased from 76-65% to 67-53%, respectively (Brod et al., 2012;Gómez-Muñoz et al., 2017). This is consistent with crop response trials, where increasing amounts of N are applied (e.g., Brod et al., 2012, see Figure 4). In that example, two industrial composts (i.e., neutral and acid Dynea composts) had only N-FEV values ranging from 7 to 30%, as they contained low amounts of NH + 4 and the N-mineralization rate was low. Acidification of the compost increased the ANR and N-FEV of Dynea composts compared to untreated compost (Brod et al., 2012). This is due to a reduced NH 3 emission during composting, as is seen when acidifying stored pig and cattle slurry. In long-term studies adding human sewage sludge (i.e., biosolids) to silage maize, the FEV was 55% at low application rates and 64% at high application rates (Cerný et al., 2012). This result, compared with the finding by Brod et al. (2012), implied that in short-term fertilizer application, doubling the application rate might not have a higher ANR and FEV, but in the long-term application, an increase of the rate increase ANR and FEV. The reason can be that higher application rate in short-term studies leads to emission of easily available N (NH 3 emission, denitrification) and that reduces FEV, while in long-term studies there still is this immediate loss but organic N increases in the soil and this will lead to higher amounts of N mineralised with time. Gómez-Muñoz et al. (2017) found in the long-term experiment that continuous application of agricultural and urban wastes improved soil quality, and longterm N availability correlates with the accumulation of N and C in soil. That study reported the ANR and FEV in the final year (2013) had generally increased compared to those in the first year of the study (2003), except for composted household waste and cattle deep litter. The effect of C:N ratio was documented for biochar produced by the pyrolysis of eucalyptus wood, as ANR values increased from 28-40% with increasing C:N ratios (2-4.9). Therefore, as new DPS-derived STRUBIAS products are emerging, there needs to be a test phase before their use in agriculture. This should involve short to long-term pot and field trials across crop and soil types to investigate P dynamics in soil and their N-P FEV values.

Potential Environmental Losses
During the storage and land application of DPS, there may be the risk of nutrient loss or emissions to waters (surface and subsurface) and/or the atmosphere, respectively.

Potential Losses From DPS/STRUBIAS to Waters
As with all fertilizers, there is an associated risk of pollutants loss to waters (surface and subsurface pathways) (Sørensen and Jensen, 2013) (Figure 3). DPS contains high levels of P and other constituents such as C, N, Na and Cl that can alter soil composition and runoff behavior Haynes, 2010, 2011). The application timing and method of DPS are both important factors to control to minimize pollutant losses to waters. Two recent studies have examined nutrient losses from DPS in field soil experiments. The first micro-plot lab study applied several DPS types to a grassland soil in Ireland and investigated the potential losses on P and N in runoff using simulated overland flow after 48 h of DPS application (Ashekuzzaman et al., 2020). That study found that the soluble P loss was highest for Ca-DPS (5.7 mg L −1 ) followed by Al-DPS (0.8 mg L −1 ) and Fe-DPS (0.15 mg L −1 ). In addition, P losses from DPS, including Ca-P-rich DPS, are much lower when compared to cattle slurry (7.0 mg L −1 ). With regard to N, that study observed dominant N losses were NH 4 -N (nitrate (NO − 3 ) losses were negligible) in the runoff pathway with concentrations ranging from 2.6 to 3.3 mg L −1 . Such concentrations are significantly lower than equivalent studies that focused on dairy cattle slurry (17.4 mg L −1 ). The availability of N in organic wastes can be predicted from their C:N ratio , and DPS has a C:N ratio of ∼6 (Ashekuzzaman et al., 2019a), which is comparable with human sewage sludge. According to Delin et al. (2012), this implies that around 50% of the N content is easily available, as also found for sewage sludge (Petersen et al., 2003). The second field study examined P accumulation in soil and potential losses in surface runoff and leaching (multi-depth) at seven sites in New Zealand (Lizarralde et al., 2021). Results showed that after the long-term application of DPS (based on N content), high amounts of P in the soil at least to 30-cm depths accumulated. The level of accumulation varied across soils and was due to the history of wastewater application, the capacity of the soils to sorb P and the land use and system management.
Organic fertilizers such as DPS delivered and applied on arable land (e.g., winter cereals) can be an effective component of any nutrient management plan. For practical reasons, DPS is often applied in autumn before sowing a winter cereal like winter wheat. However, under free draining soils (loamy sand and sandy loam soils with a yearly drainage surplus of 300-400 mm) and wet and cool North-European conditions, extra NO − 3 leaching losses, equivalent to 20-30% of total N, can be expected after application of organic fertilizers with similar N availability to winter wheat in autumn (Sørensen and Rubaek, 2012). Under conditions with less surplus precipitation or application to crops with a large capacity for N uptake in autumn, less NO − 3 leaching by application in autumn are expected. By waste application in spring, NO − 3 leaching is significantly lower (Sørensen and Rubaek, 2012) and NO − 3 leaching is often proportional to total N application (De Notaris et al., 2018;Pedersen et al., 2021) and thus nearly similar for organic N and mineral N. The total N applied with organic fertilizers is higher than with mineral fertilizers to obtain the same fertilizer value and thereby crop yield. This also implies that NO − 3 leaching is higher by application of organic wastes in spring compared to mineral N fertilization, but higher leaching losses can be prevented by use of cover crops (Pedersen et al., 2021). Pedersen et al. (2021) found extra NO − 3 leaching equivalent to 8% of the N input in the first year and 4% in the second year after application for both mineral and organic N applied to a loamy sand and a sandy loam soil in spring.
In contrast to N, soluble P is strongly bound to soil implying that very low leaching losses of P occur after application of organic fertilizers. However, if DPS is applied directly to soil and not incorporated or injected, there is risk of incidental P losses (0.3-7.6% of total input) by surface runoff (Ashekuzzaman et al., 2020) and by leaching through macropores in soil (Sørensen and Jensen, 2013). Such losses can both occur by transport in soluble form (e.g., dissolved reactive phosphorus) and in the form of particle and colloid-bound P. Christiansen et al. (2020) found large variation in water-extractable P (0.1-9 % of total P) in various sludge types and therefore the risk of soluble P loss is also variable. A reduced risk of P losses along surface runoff and subsurface macropore leaching pathways by incorporation of DPS into soil or by injection is possible (Sørensen and Jensen, 2013). When soils are loaded with excessive amounts of P over a longer period, the soil is saturated with P and P leaching to drains is significantly increased (Heckrath et al., 1995). By precipitation of P and N in struvite, nutrients become concentrated like in mineral fertilizers and can be stored and applied as for mineral equivalents. This means that its application can occur following best practice for precision farming i.e., right time, right place, right amount, right method, and right product.
After pyrolysis of sludge for biochar production, most of the organic N is lost. The availability and fate of this N is not well investigated e.g., Christiansen et al. (2020) found that a biochar derived from a mixture of human sewage sludge and straw contained 5% of total P in water-extractable form and most of the P content was soluble in a weak acid (citric acid). Therefore, a part of the P in sludge-based biochar solubilises in soil. Weak biochar binding on clay minerals and its low density can lead to environmental losses to waters. In addition, translocation of biochar due to hydrological connectivity is observed. For instance, Rumpel et al. (2006) showed that biochar accumulates at the bottom of slopes within the landscape and such losses are important to quantify as they can be delivered to surface water (Major et al., 2010). Therefore, biochar needs to be incorporated into soil to avoid loss of P in surface runoff either in soluble or in particulate form. This is also the case in grasslands, where a significant reduction in P losses may occur where injection rather than surface application of manure is practiced (Uusi-Kämppä and Heinonen-Tanski, 2008). This precision farming application method where available could be a DPS application method that minimizes incidental losses of pollutants in runoff during rainfall events.

Potential Losses From DPS/STRUBIAS to the Atmosphere
To date, there have not been many studies that have measured or calculated the accumulated emissions of GHG and NH 3 from the production, storage or land application of these products ( Figure 5). Therefore, DPS or STRUBIAS emissions of CH 4 , N 2 O, and NH 3 from production until after field application are calculated using a combination of information about emission from the products or by using similar products as a proxy. Herein, such risks and mitigation for each management step i.e., from production, to processing of the sludge, to storage of the sludge and sludge products, and to application of DPS and secondary STRUBIAS products is considered.
The DPS is usually stored anaerobically until application to soil. During this phase CH 4 and NH 3 may be emitted; little N 2 O is emitted during the storage phase as the waste tends to not have a surface crust where nitrification-denitrification may take place (Baral et al., 2018). However, application to soil emits N 2 O (Scott et al., 2000;Yoshida et al., 2015). A fraction of the C in sludge applied to soil will contribute to C storage. The emitted NH 3 can contribute to N 2 O emission after deposition to land or water. Transforming sludge into biochar, hydrochar, or ash will cause an emission of CO 2 , but this treatment will eliminate CH 4 emission and may affect N 2 O emission when applied to soil. The CO 2 emitted during treatment of the sludge is part of the circulation of C between the atmosphere, plants, intake by dairy cows, and recycling of the waste and therefore considered climate warming neutral. In recent field studies, biochar has not increased N 2 O emissions from "fertilized soil" when applied to fields (Clough and Condron, 2010;Taghizadeh-Toosi et al., 2011;Liao et al., 2020;Thers et al., 2020). There is no emission of N 2 O during storage of biochar, as it has been shown in compost studies that biochar reduces N 2 O production and emission (Shakoor et al., 2021). In several meta-analyses it has been shown that N 2 O emissions from soil decreases after the addition of biochar (Cayuela et al., 2014;Sri Shalini et al., 2020). Reasons for reduced N 2 O emission from soil treated with biochar could be improved soil aeration, increased soil pH, enhanced N immobilization, and possible toxic effect induced by biochar organic compounds (polycyclic aromatic hydrocarbons) on nitrifier and denitrifier communities (Taghizadeh-Toosi et al., 2011;Cayuela et al., 2014;Harter et al., 2014). In contrast, some studies show increased N 2 O emission from soil with biochar, which is attributed to an increased soil water content in the presence of biochar favoring denitrification, or the release of biochar embodied-N (Lorenz and Lal, 2014).
The N content of biochar or hydrochar is not high due to the transformation of N during initial feedstock thermolysis (Majumder et al., 2019). Although N content in biochar or hydrochar is low, the application of biochar materials into the soils can affect the soil N cycle. Biochar and hydrochar have been shown to adsorb NH + 4 on biochar particles and reduce NH 3 volatilization; however, the increased NH 3 volatilization, observed from some soil treated with hydrochar, is possibly due to the reduced ability to absorb NH + 4 associated with greater hydrophobicity of hydrochar (Clough and Condron, 2010;Taghizadeh-Toosi et al., 2012;Subedi et al., 2015).
Carbon in DPS and in DPS-derived STRUBIAS products added to soil will contribute to soil C storage, the sequestering potential or mean residence time (MRT) being related to the rate of transformation of the added carbon to CO 2 (Tian et al., 2009). It has been shown that CO 2 fluxes were suppressed when biochar was added to fertilized soils (Wang et al., 2016), which may be due to reduced enzymatic activity and the precipitation of CO 2 onto the biochar surface (Case et al., 2014). Ethylene, which is frequently present in biochar, can sometimes inhibit the transformation of C in soil (Spokas et al., 2010). However, if there is labile C input in biochar or hydrochar, it can result in positive priming effects (He et al., 2017) and increased CO 2 emissions, although part of the CO 2 may have originated from carbonate formed during pyrolysis (Kuzyakov et al., 2009). Pyrolysis and gasification materials have been assessed by many (e.g., Lal, 2009;Beesley et al., 2011;Wu et al., 2017) to increase soil organic C content and to improve overall soil health.
The present literature review shows that STRUBIAS production potentially can reduce GHG emission from all sites of the sludge management chain. An analysis of the emission of GHG from sludge is stored until it was applied to soil or alternatively processed, chars stored and then applied to soil was carried out to provide insight in the potential total GHG reduction due to production of biochar and hydrochar ( Figure 5). In calculations carried out using the model outlined in material and methods, the total GHG increase in the atmosphere due to sludge managed traditionally is 359 kg CO 2eqv ( Table 6). In contrast to sludge management, CH 4 emission during storage of STRUBIAS products is avoided, and N 2 O emission from biochar is negligible and from hydrochar reduced to 1/3 of the emission from sludge, because 55% of the N in sludge is recycled to the wastewater plant. Due to the recalcitrant nature of C in STRUBIAS products, more C is sequestered when these are applied to fields. Within the boundary of the sludge managing system, the emission from standard sludge management is 359 kg CO 2eqv (an increase in Sludge composition is dry matter 170 g kg −1 , ammonium-N 0.94% in DM, total-N 35% in DM, total C 39.0% in DM. Negative numbers means reduction in GHG in the atmosphere and positive increase GHG concentrations. CO 2 in the atmosphere). Producing hydrochar reduces GHG in the atmosphere corresponding to −30 kg CO 2eqv and biochar production to −92 kg CO 2eqv . Avoiding CH 4 and N 2 O emission from the sludge are the most important factors to reduce the climate warming potential of dairy waste management.

Effects on Soil Microorganisms
Soil microorganisms play a pivotal role in nutrient cycling in agricultural ecosystems. Their activities enhance the availability of essential nutrients for crop growth and contributes to an improvement of soil properties such as OM content and water retention capacity. The application of organic residues like DPS modifies soil microbial communities significantly. These modifications are highly variable and depend strongly on the composition of the bio-based residues applied.

Microbial Analysis
The methods used to assess soil microbial processes divide into abundance, diversity and activities. Determination of soil microbial biomass in soil by chloroform fumigation is a common indicator applied after the application of different types of sludge (Charlton et al., 2016). Other techniques that can provide information on microbial biomass are phospholipid fatty acid assays (PLFAs) and the quantification of total DNA. These two techniques are also powerful methods to determine changes in the diversity of soil microorganisms. DNA-based techniques such as 16S and 18S gene quantification, metagenomics and metabarcoding, provide very insightful information on the community composition (Abdelfattah et al., 2018;Bünemann et al., 2018;Bastida et al., 2019). PLFAs allow distinguishing between bacteria and fungi, and further distinctions between bacterial groups such as Gram + and Gram − bacteria (Frostegård and Bååth, 1996). Other commonly used methods to determine changes in microbial metabolic diversity are community-level physiological profiling (CLPP) assays such as Biolog Ecoplates (Liu et al., 2017). The latter technique provides profiles of potential degradation of different complex chemical C substrates, which assess the ability of soil microbial communities to degrade natural soil constituents (Siebielec et al., 2018). Lastly, enzymatic activities are the most common techniques to study soil microbial activities after the application of organic residues. The most common enzymes measured are phosphatases, β-glucosidases, dehydrogenases and ureases (Table 7). Enzymes are substratespecific and can be associated with different nutrient cycles in the soil (Burns et al., 2013).

Effects of DPS Application on Soil Microbial Communities
Little information is available and from a small number of research teams on the direct effects of dairy sludge on soil biological properties, and the studies do not always come to the same conclusion (Table 7). Reported increases of microbiological indexes and improvements in soil nutrient cycling after the application of DPS appear in the literature. Soil enzymatic activities (Frac and Jezierska-Tys, 2011;Frac et al., 2012;Oszust et al., 2015) and soil microbial diversity as revealed by CLPPs analyses (Frac et al., 2012;Oszust et al., 2015) were reported to increase after the application of dairy sludge. Gryta et al. (2014) show that the effect of heavy metal contamination due to DPS application caused a significantly lower (−61.54 %) C degradation as shown by a Biolog Ecoplate assay and DPS is a source of heavy metal contamination in soils (López-Mosquera et al., 2000). Heavy metal concentrations below guideline values found in studies by Ashekuzzaman et al. (2020) and Shi et al. (2021a) point to the importance of regulated application rates (based on P) but also recognizing heavy metal application which varies DPS types. As wastewater treatment is improved, heavy metal concentrations should be reduced further e.g., biological P, removal strives to replace the need for metal coagulants. Similar results were found when assessing other types of sludge such as sewage sludge, which has a high content of C and contributes to increased soil microbial biomass and activities (Torsvik et al., 1998;Abaye et al., 2005). Yet, similar to the results of Gryta et al. (2014) on dairy sludge, sewage sludge effects on soil microorganisms depend also on the amount of pollutants. Sewage sludge may contain large concentrations of pollutants such as heavy metals and pathogens that often can decrease soil microbial indicators (Torsvik et al., 1998;Charlton et al., 2016;Major et al., 2020; Table 7). Although concentrations of heavy metals and other pollutants may be low in DPS compared to sewage sludge, there is a need for long-term field trials, especially pertaining to bioaccumulation of heavy metals in soil (and crops) as this has the potential to damage soil microbial communities. Gryta et al. (2014) examined DPS heavy metal content on soil biology. Their results showed that high concentrations of heavy metals led to a decrease in most soil biological descriptive variables, associated with damage of the cell membrane, mitochondria or DNA (Dar, 1996;Tchounwou et al., 2012). Torsvik et al. (1998) compared the application of contaminated sewage sludge with unpolluted equivalents to examine microbial diversity. Using a technique that uses "total number of genomes, " as equivalent to the E. coli genome, their results indicated that the application of polluted sludge had negative impacts on soil biodiversity, causing a reduction of up to 6.5 times less biodiversity. Previous experiments also compared polluted vs. unpolluted sewage sludge with similar conclusions (Brookes and McGrath, 1984;Fliesbach et al., 1994). Other studies using Biolog (CLPPs) assessments (Banerjee et al., 1997) and PLFA  Gryta et al., 2014;6: Brookes and McGrath, 1984;7: Charlton et al., 2016;8: Banerjee et al., 1997;9: Fliesbach et al., 1994;10: Abaye et al., 2005;11: Roy et al., 2019;12: Fernández et al., 2009;13: Jorge-Mardomingo et al., 2013;14: Torsvik et al., 1998;15: Dar, 1996;16: Houben et al., 2019;17: Bastida et al., 2019;18: Nicolás et al., 2014;19: Kizilkaya and Bayrakli, 2005;20: Kunito et al., 2001;21: Markowicz et al., 2021;22: Siebielec et al., 2018;23: Liu et al., 2017;24: Singh et al., 2014).
analyses have shown similar results  Table 7). Similarly, several studies show the depressed activities of soil microorganisms by reduced enzymatic activities after polluted sludge applications to soil (Dar, 1996;Kunito et al., 2001;Kizilkaya and Bayrakli, 2005;Speir et al., 2007;Fernández et al., 2009;Markowicz et al., 2021; Table 7). Another risk derived from the application of DPS is the introduction of pathogens and organic pollutants. Research from sewage sludge shows the presence of human and animal pathogens such as E. coli, Listeria, Clostridium perfringens, Enterococcus, or Salmonella (Brochier et al., 2012). These pathogens may survive on plant tissues, in soils and in hydroponic systems to which sludge is applied (Brochier et al., 2012;Kyere et al., 2019). Native soil microbial communities are known to decrease the survival of potential pathogens (Xing et al., 2020), but specific strains of bacteria, such as E. coli O104:H4, may survive in soil for over a year after its inoculation (Knödler et al., 2016). Moreover, sewage and DPS introduce antibiotic resistant genes to soils, causing the development of antibiotic resistant bacteria with severe implications on human and environmental health (Rizzo et al., 2013;Dungan et al., 2018;Urra et al., 2019). Nevertheless, even if pathogens are present in DPS, their concentration is believed to be 10-15 times lower than in sewage sludge . As a consequence, the risks of introducing pathogens and other organic pollutants after the application of DPS are smaller than other residues such as sewage sludge.
Study results with respect to the effect of sludge applications in agriculture and their effects on soil communities vary, but for dairy derived sludge most studies show increases in microbial functional diversity (CLPP) and activities ( Table 7). Activities of dehydrogenase (Jezierska-Tys and Frac, 2009;Frac et al., 2012;Oszust et al., 2015), acid phosphatase (Frac and Jezierska-Tys, 2011), urease and protease (Frac et al., 2012;Oszust et al., 2015) have been reported to increase after the application of dairy sludge ( Table 7). CLPP experiments have also revealed increased degradation of C substrates (Frac et al., 2012;Oszust et al., 2015).
When sewage sludge is not heavily polluted, similar trends can be observed ( Table 7). Enzymatic activities have been shown to increase with the application of sewage sludge (Dar, 1996;Jorge-Mardomingo et al., 2013;Siebielec et al., 2018;Roy et al., 2019), soil microbial biomass (Charlton et al., 2016) and diversity indexes has been reported to improve based on Biolog assessments (Liu et al., 2017). DNA approaches have also revealed little effect on soil microbial diversity and bacterial antibioticresistance after the application of sludge (Rutgersson et al., 2020).
The application of unpolluted sludge to soils might be very beneficial in improving soil health and fertility. Sludge contains large concentrations of easily decomposable C, readily available N, P, and other essential nutrients for plant and microbial growth (Krogstad et al., 2005;Singh and Agrawal, 2008;Peltre et al., 2011). The recalcitrant components of sludge release at a slow rate by specific microbial groups such as P solubilising microorganisms (PSM) (Clarholm, 1985;Khan et al., 2009;Kuypers et al., 2018). The relative distribution of the different nutrient pools in sewage sludge is highly dependent on the production, treatments and origin of the sludge (Singh and Agrawal, 2008).
The high C content constitutes the most significant attribute of sludge that can affect the development and growth of soil microbial communities, because C is the most limiting element for bacterial and fungal growth in soils (Demoling et al., 2007;Hobbie and Hobbie, 2013). Boosting the C content in soil leads to a concatenated stimulus in the cycling of other nutrients such as N or P (Demoling et al., 2007). Field studies have also confirmed the positive effects of sludge application on both N and P cycles (Hallin et al., 2009;Frac and Jezierska-Tys, 2011;Houben et al., 2019). The application of large amounts of C is associated with a significant growth in soil microbial biomass (Charlton et al., 2016;Houben et al., 2019). Yet, this positive effect of sewage sludge on soil microbial communities has been reported to lead to significant nutrient immobilization by soil biota (Smith and Tibbett, 2004;Gómez-Muñoz et al., 2017). It is expected that sewage sludge with a high C:N ratio (>15) might lead to N immobilization by soil biota (Gómez-Muñoz et al., 2017). The same applies for the immobilization of P, application of organic materials with a high C:P ratio might lead to its immobilization (Zhang et al., 2018). Whenever sludge is applied as a bio-based fertilizer, these aspects should be considered. However, in the long term, the application of C-rich materials such as sewage sludge should improve soils from an agronomic and environmental point of view. Building up C content in soils would improve nutrient cycling, providing a slower release maintained over time and lower losses that might contaminate soils and water bodies (Gómez-Muñoz et al., 2017;Zhang et al., 2018).

CONCLUSION
This review collated information that will help give DPS and their secondary products certification as P-fertilizers in accordance with technical proposals for new fertilizing materials under forthcoming EU Fertilizing Product Regulations. It presents the current state of knowledge pertaining to dairy processing sludge and STRUBIAS bio-based fertilizers and identifies knowledge gaps and potential solutions to minimize environmental losses to soil, water and air. STRUBIAS products have a high P concentration compared to that of sludge, and a well-defined fertilizer efficiency of the P applied to fields. To achieve high P fertilizer efficiency, dairy wastewater treatment must aim at producing sludge with soluble P-components and avoid Al coagulation and P insolubility. In STRUBIAS production, conditions producing less soluble P should be avoided, i.e., high temperatures. The benefits of STRUBIAS production is a reduction of transport cost of P due to a high P concentration. In the development of production units, it is important that STRUBIAS products can be applied with traditional mineral fertilizer application machinery. Heavy metal concentration of known products are below the limits set for the use of these as fertilizers, and the risk of disease spreading and negative effects on microbial activity in soil is low. Producing STRUBIAS products eliminates GHG emissions from management of the sludge from dairies. A goal of STRUBIAS production could be recycling of plant nutrients and C to organic farms, thereby providing a sustainable circular economy. More information is needed to carry out an economic analysis (e.g., a cost comparison across dairy sludge management, production, and management systems for STRUBIAS products and mineral fertilizer), which should include a value chain analysis of the whole system.

DATA AVAILABILITY STATEMENT
The original contributions presented in the study are included in the article/Supplementary Material, further inquiries can be directed to the corresponding author/s.