PFAS, PCBs, PCDD/Fs, PAHs and extractable organic fluorine in bio-based fertilizers, amended soils and plants: Exposure assessment and temporal trends Nicolas Estoppey a,*, Emma R. Knight b,c, Ian J. Allan b, Kuria Ndungu b, Gøril Aasen Slinde a, Jan Thomas Rundberget b, Kari Ylivainio d, Alicia Hernandez-Mora e,f, Erlend Sørmo a, Hans Peter H. Arp a,g, Gerard Cornelissen a,h a Norwegian Geotechnical Institute (NGI), P.O. Box. 3930, Ullevål Stadion, N-0806 Oslo, Norway b The Norwegian Institute for Water Research (NIVA), Økernveien 94, 0579 Oslo, Norway c Queensland Alliance for Environmental Health Sciences, The University of Queensland, 20 Cornwall Street, Woolloongabba, Queensland, Australia d Natural Resources Institute Finland (LUKE), Tietotie 4, 31600 Jokioinen, Finland e University of Natural Resources and Life Sciences (BOKU), Konrad Lorenz-Straße 24, 3430 Tulln an der Donau, Austria f AGRANA Research & Innovation Center (ARIC), Reitherstrasse 21-23, 3430 Tulln an der Donau, Austria g Norwegian University of Science and Technology (NTNU), 7024 Trondheim, Norway h Norwegian University of Life Sciences (NMBU), 1432 Ås, Norway H I G H L I G H T S G R A P H I C A L A B S T R A C T  Persistent organic contaminants in bio- based fertilizers, amended soils and plants  Concentrations below thresholds, except in a pyrolyzed sewage sludge  High EOF concentrations in sewage sludge and chicken manure BBFs  Decreased concentrations of legacy contaminants in sewage sludge over time  Low long-term risks for soils and low human exposure through cereal consumption A R T I C L E I N F O Editor: Fang Wang Keywords: Circular economy Agriculture Soil pollution Accumulation A B S T R A C T Bio-based fertilizers (BBFs) produced from organic waste contribute to closed-loop nutrient cycles and circular agriculture. However, persistent organic contaminants, such as per- and poly-fluoroalkyl substances (PFAS), polychlorobiphenyls (PCBs), polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs), as well as poly- aromatic hydrocarbons (PAHs) can be present in organic waste or be formed during valorization processes. Consequently, these hazardous substances may be introduced into agricultural soils and the food chain via BBFs. This study assessed the exposure of 84 target substances and extractable organic fluorine (EOF) in 19 BBFs produced from different types of waste, including agricultural and food industrial waste, sewage sludge, and * Corresponding author. E-mail address: nicolas.estoppey@ngi.no (N. Estoppey). Contents lists available at ScienceDirect Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv https://doi.org/10.1016/j.scitotenv.2024.177347 Received 11 August 2024; Received in revised form 29 October 2024; Accepted 31 October 2024 Science of the Total Environment 957 (2024) 177347 Available online 16 November 2024 0048-9697/© 2024 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license ( http://creativecommons.org/licenses/by/4.0/ ). Sewage sludge Literature review biowaste, and through various types of valorization methods, including hygienization at low temperatures (<150 C) as well as pyrolysis and incineration at elevated temperatures (150–900 C). The concentrations in BBFs (ΣPFOS & PFOA: <30 μg kg 1, Σ6PCBs: <15 μg kg 1, Σ11PAHs: <3 mg kg 1, Σ17PCDD/Fs: <4 ng TEQ kg 1) were found to be below the strictest thresholds used in individual EU countries, with only one exception (pyrolyzed sewage sludge, Σ11PAHs: 5.9 mg kg 1). Five BBFs produced from sewage sludge or chicken manure contained high concentrations of EOF (>140 μg kg 1), so monitoring of more PFAS is recommended. The calculated expected concentrations in soils after one BBF application (e.g. PFOS: <0.05 μg kg 1) fell below background contamination levels (PFOS: 2.7 μg kg 1) elsewhere in the literature. This was confirmed by the analysis of BBF-amended soils from field experiments (Finland and Austria). Studies on target legacy contami- nants in sewage sludge were reviewed, indicating a general decreasing trend in concentration with an apparent half-life ranging from 4 (PFOS) to 9 (PCDD/Fs) years. Modelled cumulative concentrations of the target con- taminants in agricultural soils indicated low long-term risks. Concentrations estimated and analyzed in cereal grains were low, indicating that exposure by cereal consumption is well below tolerable daily intakes. 1. Introduction The global population is expected to reach 9.7 billion by 2050 (UNDESA, 2022). Meeting food demands requires increasing agriculture production, and fertilizers play an essential role in this effort (Alexandratos and Bruinsma, 2012; Byrnes and Bumb, 1998). Conven- tional limited (phosphorus) and energy-intensive (nitrogen) inorganic fertilizers are not sustainable and ill-suited to meet this challenge (Cordell et al., 2009; Erisman et al., 2008). Bio-based fertilizers (BBFs) are products obtained by recycling nutrient-rich side streams, and constitute a valid alternative to conventional fertilizers (Babcock-Jack- son et al., 2023; Chojnacka et al., 2020; Svanback et al., 2019). BBFs can be derived from biomaterials of various origins (industry, agriculture, society) and many of them have a similar agronomical effectiveness to that of conventional fertilizers (Kurniawati et al., 2023; Sigurnjak et al., 2019, 2016; Vaneeckhaute et al., 2013; Wester-Larsen et al., 2024, 2022). Promoting the use of BBFs can thus significantly reduce depen- dence on conventional inorganic fertilizers, supporting the transition to a circular economy (Chojnacka et al., 2020; Svanback et al., 2019). However, to allow for a safe circular economy, it is essential to avoid recirculating persistent organic contaminants that can be present in the organic waste and that may accumulate in the environment and food chain. Historical applications of sewage sludges have been reported to contaminate agricultural land, groundwater, and adjacent bodies of water with per- and poly-fluoroalkyl substances (PFAS), polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs) and poly- chlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) (Johnson, 2022; Lindstrom et al., 2011; Mackiewicz-Walec and Krzebietke, 2020; Pepper et al., 2021a; Rohler et al., 2021; Washington et al., 2010a; Weber et al., 2018a). Concerns about the presence of these contaminants in BBFs have been raised because exposure to them is associated with a variety of health issues including various cancers and effects on the immune, nervous, and endocrine systems (Faroon et al., 2003; Fenton et al., 2021; Steenland et al., 2020; Van den Berg et al., 2006). PFAS are a group of thousands of compounds that are composed of a fluorinated alkyl chain and a polar head group that give them surfactant- like properties and extreme chemical and thermal stability (Buck et al., 2011; Dickman and Aga, 2022). They have been used in a wide range of products including fire-fighting foams, non-stick cookware, and fast- food containers (Glüge et al., 2020; Prevedouros et al., 2006). The growing health concerns associated with PFAS have been reflected in the drastic decrease of the tolerable intake set by the European Food Safety Authority (EFSA), i.e., from 150 and 1500 ng kg 1 body weight (kg 1 bw 1) per day for perfluorooctanoic acid (PFOA) and per- fluorooctanesulfonic acid (PFOS), respectively, in 2013, to 4.4 ng kg 1 bw 1 per week for the sum of PFOS, PFOA, perfluorononanoic acid (PFNA) and perfluorohexanesulfonic acid (PFHxS) in 2020 (Schrenk et al., 2020). Some legacy PFAS have been regulated and their use restricted (e.g., PFOA, PFOS, and PFHxS are listed in the Stockholm Convention on persistent organic pollutants (POPs)), but these have often been substituted by other, probably equally problematic, PFAS such as 6:2 chlorinated polyfluoroalkyl ether sulfonic acid (F-53B), hexafluoropropylene oxide dimer acid (GenX) or dodecafluoro-3H-4,8- dioxanonanoic acid (ADONA) (Munoz et al., 2019). Due to the very large number of different PFAS, it has been shown that the usual 30 to 60 target compounds account for only a small fraction of all fluorinated compounds present in various samples including human serum (Aro et al., 2021b) and sewage sludges (Aro et al., 2021a; Spaan et al., 2023). For these reasons, a proposal for a general restriction of all PFAS has been submitted to the European Chemical Agency (ECHA) in 2023. PCBs are also a group of synthetic compounds. They have been used exten- sively as insulating fluid in capacitors and transformers, as well as plasticizers in building materials such as paints and sealants (Erickson and Kaley, 2011; Reddy et al., 2019). Their ability to biomagnify and the extreme toxicity of the dioxin-like congeners led to a general ban in the 1980s. However, because of their persistence, they are still ubiquitous in the environment. PAHs and PCDD/Fs - contrary to PFAS and PCBs - are not manufactured intentionally but are mostly formed as unintended by- products from incomplete combustion and, for PAHs, can be naturally present in coal derivates and petroleum. Humans are mostly exposed to PAHs and PCCD/Fs via inhalation and diet (ATSDR, 1995; Marques and Domingo, 2019). PFAS, PCBs, PAHs and PCCD/Fs end up in organic waste (e.g., wastewater, manure, municipal organic waste) through various pro- cesses including excretion, laundry washing, food or green waste, improper sorting, and release from industrial point sources producing, using or treating these compounds (Andersen et al., 2008; Bolan et al., 2021; Gottschall et al., 2017; O'Connor et al., 2022; Thakali et al., 2022; Thompson et al., 2022). Due to their persistence, conditions encoun- tered in many valorization methods (e.g., drying, composting and anaerobic digestion processes) result in little (< 50 %) (Patureau and Trably, 2006; Siebielska and Sidełko, 2015) to no reduction in the amounts of persistent substances (Brandli et al., 2007, 2005; Lakhdar et al., 2009; Lazzari et al., 1999). In contrast, thermal conversion of organic waste by pyrolysis or incineration has been shown to provide good removal efficiency. Only small percentages of PFAS, PCBs and PCDD/Fs initially present in the waste have been measured in chars produced by different pyrolysis systems (500–800 C in anoxic condi- tions) (Kundu et al., 2021; McNamara et al., 2023; Sørmo et al., 2024, 2023), and a significant reduction was reported after combustion in a full-scale incinerator (850–1100 C in oxic conditions) (Bjorklund et al., 2023; Loganathan et al., 2007). Depending on the operating conditions (and the types of feedstocks), by-products such as PAHs and PCDD/Fs can form during thermal conversion, with pyrolytic conditions favour- ing the formation of PAHs and PCDDFs (Altarawneh et al., 2009; Chagger et al., 2000; Sørmo et al., 2024; Wang et al., 2017). Concen- trations in BBFs are thus expected to depend on the waste origin and the valorization process (especially the temperature), but not in the same way for all persistent organic substances. Studies investigating PFAS in diverse BBFs are rare (Kim Lazcano et al., 2020), and none have focused simultaneously on PFAS, PCBs, PAHs and PCDD/Fs in the same BBFs. Moreover, limit values of these substances in BBFs have only been N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 2 defined at country levels but not in EU regulations. Directive 86/278/ EEC (biosolids) and Regulation 2019/1009 (fertilizing products) regu- late heavy metals but not organic pollutants. Concentrations of regu- lated/legacy contaminants such as PFOS, PCBs, PCDDFs, and PAHs are expected to decrease with time (e.g., an apparent half-life of 10 for PCBs and 12 years for PCDD/Fs was reported for the 1993–2012 period in sewage sludges (Zennegg et al., 2013)) but reported temporal trends in organic waste are scarce (Gewurtz et al., 2024; Ulrich et al., 2016; Zennegg et al., 2013). The use of contaminated BBFs could constitute a risk for human and ecosystem health because plants grown in amended soils have been shown to accumulate POPs and other substances, thus representing a pathway for the trophic transfer into higher organisms, including humans (Blaine et al., 2013; Olowoyo and Mugivhisa, 2019; Pullagurala et al., 2018; Rorat et al., 2019; Wang et al., 2020). The total amount of persistent substances transferred to agricultural soils depends on BBF application rates and the pollutant concentrations in the BBF. The fraction of persistent substances available for plant uptake depends on contaminant physicochemical properties (e.g., hydrophobicity, water solubility, potential biodegradation), soil characteristics (e.g., amount and type of organic matter) and environmental conditions (e.g., pre- cipitation and temperature) (Reid et al., 2000). For example, the amounts of bioavailable PCBs, PAHs or PCDD/Fs are lower in soils rich in organic matter and black carbon because of the strong sorption of the contaminants to these matrices (Cornelissen et al., 2005; Ortega-Calvo et al., 2015) while the amounts of short-chain PFAS rapidly decrease in rainy conditions because of their high solubility (McLachlan et al., 2019; Stahl et al., 2013a; Weidemann et al., 2022). The transfer of bioavailable substances from soils to crops, and their translocation within plants vary significantly among contaminants, plant species and plant parts (Collins et al., 2006; Ghisi et al., 2019; Lesmeister et al., 2021). More hydro- phobic substances are expected to accumulate significantly in plant roots (Collins et al., 2006; Duarte-Davidson and Jones, 1996) although some of them have also been shown to translocate to other (edible) parts of the plants (Sun et al., 2019). Smaller and more hydrophilic persistent substances – such as short-chain PFAS – are more likely to be trans- located to shoots (Adu et al., 2023; Krippner et al., 2014; Lesmeister et al., 2021). For PFAS, this is of particular concern because short-chain PFAS tend to be used as substitutes for long-chain PFAS, and an increasing volume of recent studies has detected PFAS in edible parts of plants (Bao et al., 2020; Brendel et al., 2018). Very few studies have assessed human exposure to POPs and other persistent substances through BBF-amended plants, representing a knowledge gap for the large-scale application of such fertilizers in the EU. Furthermore, esti- mations of long-term build-up of persistent substances in soils as a result of repetitive BBF amendments are lacking. To address these collective concerns related to persistent organic substance contamination in BBF-amended soils and plants, this study assessed, for the first time, the exposure of 84 persistent organic sub- stances from 4 classes – i.e. 7 PCBs, 16 PAHs, 17 PCDD/Fs, 44 PFAS – in 19 BBFs produced by different methods (including incineration, pyrol- ysis, and hygienization at lower temperatures) and from various waste materials (including sewage sludge, biowaste, and agricultural and food industry waste). Another novel element was the quantification of extractable organic fluorine (EOF) in BBFs to account for non-targeted fluorinated compounds. Estimated contents of the studied contami- nants in soils and plants amended with BBFs were verified by analyzing BBF-amended soils. Literature providing concentrations of legacy con- taminants in sewage sludges was reviewed to provide temporal trends and model the build-up of these compounds in BBF-amended soils. The specific aims of the study were to (i) evaluate the impact of valorization methods and waste origins on the target pollutants in BBFs, (ii) assess the compliance of concentrations in BBFs with existing national threshold values, (iii) predict the concentrations in agricultural soils after one BBF application and over time, (iv) assess the risk for con- sumers by comparing expected concentrations in plants with tolerable intakes, (v) validate the estimates by measuring real samples from field trials, and (vi) review the literature on legacy PFOS, PCBs, PCDD/Fs and PAHs in sewage sludge to determine temporal change rates and assess the long-term risk for amended soils. Addressing these aims will aid the assessment of whether the use of BBFs can be considered a safe circular economic alternative, and whether specific waste types and valorization techniques require further development. 2. Materials and methods 2.1. Selection of bio-based fertilizers Nineteen BBFs were selected to cover the main categories of waste valorization methods, i.e., hygienization techniques at temperature < 150 C such as composting, biogasification and drying (12 BBFs), py- rolysis (3 BBFs), incineration (3 BBFs), and crystallisation (1 BBF) (see Table 1 and Supporting Information SI.1). BBFs produced by hygieni- zation at low temperature (<150 C) – hereafter referred to as “hygie- nization” - dominated the sample set because of their generally higher fertilizing values (see SI.1), especially in terms of N supply, which is the foremost limiting nutrient for crop yield (Fageria and Baligar, 2005). Crystallisation is expected to provide BBFs with low concentrations of contaminants (de Boer et al., 2018; Dong et al., 2023; Ronteltap et al., 2007), and for this reason only one struvite material was tested. The 19 selected BBFs also covered the three main categories of waste origin, i.e., green waste and livestock residues from the agricultural and food in- dustries (11 AgriFoodInduWaste-BBFs), sewage sludge (6 SewSludge- BBFs) and biowaste (2 Biowaste-BBFs) (Table 1). A stronger focus was placed on the two first categories because the resulting BBFs – often in form of pellets – are more suitable for large-scale commercialization, compared to the digestates or composts that are usually generated from the treatment of biowaste. Information on the nutrient content, dry matter and application rates of the tested BBFs can be found in the Supporting Information (SI.1). Analyses of 7 PCBs, 16 PAHs, 17 PCDD/ Fs, and 44 PFAS were first conducted on one replicate of each BBF to assess the contamination levels. Then, triplicate analyses of PCBs and PAHs as well as PFAS and EOF were conducted for the nine and thirteen most contaminated BBFs, respectively. 2.2. Preparation of BBFs Samples were ground and homogenized using an agate mortar and pestle, which was cleaned 3 times with methanol (VWR) and pentane (Merck) between each sample. Procedural blanks were done by grinding a certified reference material made of a clean loamy soil (CLNLOAM6, Merck, Norway), at least every 10 samples. Samples were stored in the freezer in glass containers (PCBs, PAHs and PCDD/Fs) or polypropylene (PP) tubes (PFAS and EOF) until analysis. 2.3. Target PCBs, PAHs and PCDD/Fs Samples were freeze-dried using method DIN 38414-22 (2000-09). The quantification of seven PCBs (PCBs 28, 52, 101, 118, 138, 153 and 180, additional information in SI.2) and 17 PCDD/Fs (full list in SI.2) was conducted based on the method DIN EN 16190 (2019–10). Briefly, after the addition of 13C-labeled standards (1 for each congener) to 5 g of sample, extraction was completed by accelerated solvent extraction (ASE, 180 C, 140 bar, 10 min) with toluene. Then, the extracts were purified by column chromatography using mixed silica columns (acid, neutral, basic) and aluminium oxide column, before being measured by gas chromatography coupled to high-resolution mass spectrometry (GC/ HRMS). Results for PCDD/Fs were expressed as WHO toxic equivalent (WHO-TEQ) by multiplying the concentrations of each compound by its corresponding WHO-05 toxic equivalency factor (TEF) (SI.2). The quantification of 16 PAHs (full list in SI.2)) – was achieved following method DIN EN 16181 (2019–08). After the addition of seven N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 3 deuterated standards (SI.2), an extraction was conducted by ASE with toluene. The extracts were purified using deactivated silica columns, dimethylformamide clean-up, and aluminium oxide columns. The 16 PAHs were quantified by GC coupled to tandemmass spectrometry (MS/ MS). Limits of quantification for PCBs, PAHs, and PCDD/Fs are given in SI.2. 2.4. Target PFAS Extraction of target PFAS was performed according to a method adapted from Knight et al. (2021) and Braunig et al. (2019). Freeze- dried samples (1.5–2 g) were placed into 15 mL PP tubes and an inter- nal standard mixture was added (0.03 mL of 0.1 ppm). After a minimum of 30 min, 5 mL of methanol/NH3 solution (99:1) was added to the samples before samples were shaken on a side-to-side shaker for 30 min, then sonicated for 20 min and then centrifugated for 10 min at 2700 rpm. A second extraction was performed with 3 mL of methanol/NH3 solution, after which the supernatants were combined and concentrated to 1 mL under a gentle nitrogen (N2) flow at 40 C and acidified with 0.01 mL of acetic acid. The extracts were passed through pre- conditioned (1 mL of methanol) Bond Elut Carbon cartridges (100 mg, Agilent) and collected in 1.5 mL PP vials. The extract tubes were rinsed with 0.5 mL and passed through the same cartridge and collected into the same 1.5 mL vial. Samples were concentrated to 1 mL under N2 at 40 C, and a 0.2 mL aliquot was transferred into a SpinX centrifuge tube (Merck), buffered (0.1 mL of 5.2 mM ammonium acetate in ultra-pure water), centrifugated (1 min, 10′000 x g), transferred to 1.5 mL PP vials and stored at 20 C. A total of 44 PFAS compounds were quantified using liquid chro- matography (UPLC, Acquity Ultra Performance HPLC system) coupled with quadrupole time-of-flight mass spectrometry (QToF-HRMS, Xevo G2-S instrument) from Waters (Milford, MA, U.S). Analysis included 13 perfluoroalkyl carboxylic acids (PFCAs), 10 perfluorinated alkyl sul- phonic acids (PFSA), six perfluorooctane sulfonamido substances (pre- FOS), four fluorotelomer sulfonic acids (FTSA), three polyfluoroalkyl phosphate diesters (diPAP) and eight other PFAS of interest, e.g., sub- stitutes of PFOS or PFOA (the whole list of target compounds is provided in Supplementary Information SI.3). Separation was carried out on a Acquity BEH C8 reversed phase column (100  2.1 mm, 1.8 μm, Wa- ters). Acetonitrile and water with 5.2 mM NH4OAc were used as mobile phases for chromatographic separation, using a flow rate of 0.5 mL min 1. Negative ion electrospray was used as ionization source. The analytical method parameters are detailed in Supplementary Informa- tion SI.3. Three procedural blanks were subjected to the whole analysis procedure, using a certified reference material (Clean loam soil CLNLOAM6, Supelco) as matrix. In all of the procedural blanks, con- centrations of target PFAS were below the analytical limit of quantifi- cation (LOQ) of the method; perfluorobutanesulfonic acid (PFBS) was found at a very low concentration (0.12 μg kg 1) in one of the three blanks (SI.3). When replicate(s) presented a value three fluorinated carbon atoms were acceptable in all matrices (mean recoveries in each group >77 %, see Supplementary Information SI.3). For PFAS with three fluorinated carbon atoms (i.e., PFBA and PFPrS), chromatographic quality criteria were not met. Thus an alternative extraction method based on acetonitrile - was used for these compounds (Langberg et al., 2020). Briefly, the samples (1.5–2 g) were extracted twice with aceto- nitrile (8 ‡ 6 mL) both times using an ultrasonic bath (30 min) and shaking (30 min) and then concentrated under N2. The quantification by LC-QToF-HRMS was performed using the same parameters described above; the recoveries were 104 % for PFBA and 115 % for PFPrS. Table 1 Valorization methods and waste origins of the 19 selected BBFs of European origin. Acronyma Valorization method Waste origin Category Short description Categoryb Short description CGO Crystallisation Struvite precipitation Sewage Sludge Wastewater supernatant EPH Incineration >850 C, granulating Agriculture & food industry Sunflower husk ash PLA Incineration >850 C Agriculture & food industry Poultry litter ash ADC Incineration 900–950 C Sewage Sludge Calcined phosphate from sewage sludge ashc CRA Pyrolysis HTC, 190 C Agriculture & food industry Sludge from juice-making industry MBC Pyrolysis 300–450 C Agriculture & food industry Chicken manure BAG Pyrolysis 650 C Sewage sludge Sewage sludge BA1 Hygienization Fermentation & distillation Agriculture & food industry Wheat and maize MO14 Hygienization Pelletising Agriculture & food industry Vegetable by- products from food industry & animal proteins BIO Hygienization Pelletising Agriculture & food industry Meat and bones, apatite, vinasse, chicken manure, K2SO4 OPU Hygienization Pelletising Agriculture & food industry Chicken manure FEK Hygienization Drying and pressing Agriculture & food industry Chicken manure OG2 Hygienization Hydrolysis Agriculture & food industry Horn meal (pig bristles) ECO Hygienization Pelletising Agriculture & food industry Blood and feather meal RAN Hygienization Drying & granulating Sewage sludge Sewage sludge and biowaste PRV Hygienization Biogasification & hygienisation Sewage sludge Sewage sludge and biowaste NNP Hygienization Infrared drying Sewage sludge Sewage sludge and industrial sludge VERMI Hygienization Biogasification & vermicomposting Biowaste Biowaste and manure PLP Hygienization Composting Biowaste Biowaste, peat and wood chips a For confidentiality reasons, acronyms are used instead of full names. b The waste origin category refers to the characterizing raw material; some BBFs contained mixtures of wastes as they were real-world, commercially available (or in development) products. c Calcination is a thermal treatment whereby the substrate is exposed to very high temperatures (usually >800 C) without melting under restricted supply of ambient oxygen, generally for the purpose of removing impurities or volatile substances. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 4 2.5. Extractable organic fluorine (EOF) Non-target extractable organic fluorine (EOF) analysis was con- ducted on separate portions of all samples (1–5 g) by using the same extraction and clean-up procedures described above for the targeted PFAS analysis without the addition of internal standards. The EOF in the extracts was measured as inorganic fluoride (ions) on a Combustion Ion Chromatography (CIC) system. The CIC system consisted of an Analytik Jena combustion unit (Jena, Germany) coupled with a 920 Absorber Module and a 930 Compact Ion Chromatography (IC) Flex from Met- rohm (Herisau, Switzerland). Aliquots of 0.1 mL were combusted in quartz boats at 1050 C (several boats were needed as co-extracted alkali metals devitrified the quartz) and the combusted fluorine was absorbed in deionized water and transferred to the IC. Measured fluoride peak areas were corrected by subtracting the peak area of empty boat blanks that were injected immediately before or after the extract. Concentra- tions of EOF were determined from the area of the fluorine peak using PFOA (perfluorooctanoic acid) standards in methanol (R2 > 0.99). Calibration quality control samples were run throughout the worklist and were within 15 % of the nominal concentration. The detection limit (calculated as blank mean of five procedural blanks plus three times their standard deviation) was 20.8 μg kg 1. Quality control sam- ples included three procedural extraction blanks (certified reference material, CLNLOAM6, Supelco) and multiple methanol blanks analyzed together with the BBF extracts. It was verified that inorganic fluorine was not co-extracted with EOF using a sodium fluoride spike in the sample replicates. Organofluorine recovery was performed using PFOA spiked in samples replicates (BIO, ECO, RAN). The recoveries deter- mined from the spiked samples with PFOA ranged from 83 % for SewSludge-BBFs to 7–12 % for AgriFoodInduWaste-BBFs. The low re- coveries observed for the latter were slightly improved to 21–28 %when working with acetonitrile extraction. To obtain a fluorine mass balance, the concentrations of the measured target PFAS (CPFAS) were converted into fluoride concentra- tions and summed up (CF ) using Eqs. (1) and (2) (Aro et al., 2021a): %F ˆ nF MF MPFAS (1) CF ˆ X CPFAS %F (2) where %F is the mass fraction of fluorine in PFAS, nF is the number of fluorine atoms in a PFAS molecule,MF is the atomic mass of fluorine and MPFAS is the molecular mass of PFAS. 2.6. Impact of waste origins and valorization methods, and compliance assessment The comparisons of PCB, PAH, PCDD/F, PFAS, EOF concentrations between groups (or sub-groups) of BBFs were conducted using one-way ANOVA (pˆ 0.05) followed by a post-hoc Tukey's test (ሠ0.05) using R 4.3.2 software. Limit values of the target pollutants in fertilizers have not been defined in EU regulation. The strictest values used in individual EU countries were used for compliance assessment, i.e., 200 μg kg 1 for P 6PCBs (Luxemburg, PCB 118 not included), 3 mg kg 1 for P 11PAH (Denmark, naphthalene (NAP), acenaphtylene (ACY), anthracene (ANT), benzo[a]anthracene (BaA), chrysene (CHR), and dibenz(a,h) anthracene (DBahA) not included), 20 ng TEQ kg 1 for P 17PCDD/Fs (Luxemburg, same compounds as those studied in the present study), and 100 μg kg 1 forPPFOS and PFOA (Germany) (Collivignarelli et al., 2019a; Hall et al., 2020). Note that for PCBs and PAHs, some countries have set thresholds for individual compounds that, in certain situations, may be the limiting values (100 μg kg 1 for individual PCB in Germany and Croatia, 1 mg kg 1 for benzo[a]pyrene (BaP) in Germany, see SI.4). 2.7. Concentration in BBF-amended soils and plants Concentrations expected in BBF-amended soils were calculated for a worst-case scenario using maximum allowed application rates of fertil- izers. The EU Nitrate Directive 91/676/EEC allows a maximum of 170 kg ha 1 y 1 as manure-based N, with N generally representing the limiting factor for application rates (Amery and Schoumans, 2014; Collivignarelli et al., 2019b). Also applying the 170 kg N ha 1 y 1 threshold to non-manure-based BBFs (see N content in BBFs in Supple- mentary Information SI.1), maximum allowed application rates were calculated to range from 1.1 (OG2, N content of 15 %) to 13.1 t ha 1 y 1 (VERMI, N content of 1.3 %). As BBFs with a very low N content ( 1 %) are primarily intended to supply P, applying the limit for N was deemed irrelevant. Instead in those cases, an application rate of 50 kg P ha 1 y 1 was used as a basis for the calculations (upper end of the range of allowed rates in several EU countries; Amery and Schoumans (2014)), resulting in application rates between 0.6 (ADC, P content of 8.1 %) and 1.0 t ha–1 y–1 (PLA and BAG, P content of 5.2 %). The mass of persistent organic substances entering agricultural soils was determined by multiplying these rates with the concentrations measured in BBFs. Then, expected concentrations in (initially not contaminated) soils after one BBF application were calculated by dividing the mass of contaminants entering the soils by the amount of soil (3′900’000 kg ha 1 for soil with a density of 1.3 g cm 3), assuming that most of the studied pollutants accumulated in the top 30-cm surface layer (see discussion in Section 3.4 regarding more mobile PFAS) (Di Guardo et al., 2020; Wellmitz et al., 2023). The concentrations expected in plants were calculated based on the calculated concentrations in soils and literature-based factors (BAF). Literature-based soil porewater partition coefficient (Kd) values were used to discuss the sorption of these contaminants to soil. Kd and BAF can be found in SI.5. To verify the calculated concentration in soils and plants, these es- timates were compared to values measured in samples from two field trials conducted with a selection of the studied BBFs in 2021. Soils and maize amended with the BBFs EPH and OPU were obtained from a field trial in Langenlebarn, Austria (48.32093, 16.10166), and soils and barley amended with PLP and OPU from a field trial in Jokioinen, Finland (60.863839; 23.521162); see Supplementary Information SI.6 for additional information about the sites. At both sites, each of the studied BBFs or control (no BBF application) treatments were replicated four times (complete randomized block design) and application rates were based on total P rates of 30 kg ha 1. The size of each plot was 5 m  10m (Jokioinen) or 6 m 9 m (Langenlebarn). Both maize and barley were grown up to maturity and harvested. Composite soil samples after harvest from the topsoil layer were collected from each plot. The soil and grain samples were dried at 40 and 60 C, respectively. Composite soil and grain samples from the four replicates of the specific BBF and con- trol treatments were analyzed for PCBs and PFAS in triplicate using the same method as for BBFs (Sections 2.2 to 2.4). The comparisons of concentrations of these contaminants between amended and non- amended soils were done using a Student's t-test (p ˆ 0.05). 2.8. Temporal trends Studies reporting ‘PFOS’, ‘PCB’, ‘PDCDD/F', and ‘PAH’ concentra- tions in ‘sewage sludge’ and ‘biosolid(s)’ were investigated using these keywords in Google Scholar. This selection (37 studies for PFOS, 13 for PCBs, 11 for PCDD/Fs, 17 for PAHs, see SI.9 and SI.10) is not claimed to be fully exhaustive. However, it is expected to be representative of the last two decades of PFOS data (most recently regulated compound), and three decades of PCB, PCDD/F and PAH data (compounds banned or regulated for a longer time). Individual sample years were used when contaminant concentrations were reported annually; an average value was used for studies mentioning sampling campaigns conducted over a range of years without providing concentrations for individual years (see details in SI). In those cases where mean and median concentrations N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 5 of contaminants were not directly reported in the studies, they were calculated from individual concentrations. 3. Results and discussion 3.1. Impact of waste origins and valorization methods The average concentrations measured in the 19 BBFs ranged from 85 %) was not explained by the 44 PFAS targeted in the present study (Fig. 3). 3.1.1. Hygienization Significant differences were observed between the BBFs, depending on the valorization methods and waste origins. Among the BBFs that underwent a hygienization at low temperature (<150 C), the average concentration was significantly lower in AgriFoodInduWaste-BBFs than in SewSludge-BBFs, by a factor of 59 for P 7PCBs, 5.4 for P 16PAHs, 10 for P 17PCDD/Fs, 5.4 for Σ44PFAS and 2.2 for EOF. Biowaste-BBF presented intermediate concentrations for all substances, neither of which were significantly different from those in either AgriFoodInduWaste-BBFs or SewSludge-BBFs, except PCB concentra- tions, which were significantly higher in Biowaste-BBFs (SI.8). The overall elevated concentrations observed in SewSludge-BBFs can prob- ably be explained by the fact that the majority of PCBs, PAHs, PCDD/Fs and longer-chain PFAS present in wastewater are distributed in sludge because of their high sorption potential (Sinclair and Kannan, 2006; Tian et al., 2012; Urbaniak et al., 2017; Zhang et al., 2019). For PCBs, the higher average contamination level in biowaste-BBFs was caused by the elevated concentrations measured in one specific BBF (PLP), which could be explained by random organic contaminant impurities that typically occur in some biowaste streams (Amlinger et al., 2004). 3.1.2. Pyrolysis and calcination The contamination levels in BBFs obtained by pyrolysis (CRA, MBC and BAG) were quite variable due to the variety of pyrolysis methods utilized. The relatively low temperature used to produce CRA (190 C, hydrothermal carbonization) and MBC (300–450 C) did not effectively remove all persistent organic substances (> 6 μg kg 1 of P7PCBs and Σ44PFAS in CRA, > 5 μg kg 1 of Σ44 PFAS in MBC). The deployment of higher pyrolysis temperatures >600 C usually achieves this goal (Sørmo et al., 2024, 2023), as exemplified by BAG (650 C) for PCBs, PCDD/Fs and PFAS. However, this does not guarantee the absence of unintentionally formed compounds such as PAHs (23.2 mg kg 1 Σ16 PAHs in BAG) (Dai et al., 2014; Sørmo et al., 2024). The importance of using high temperatures to remove the target persistent organic sub- stances was further confirmed by the results obtained for BBFs produced by incineration at temperatures >850 C (PLA, ADC, EPH). These BBFs contained concentrations close to or below LOQ. The same three currently studied BBFs (PLA, ADC, EPH) were also shown to be free of pesticides and pharmaceuticals in a previous study (Dong et al., 2023). The removal efficiency of organic contaminants by high-temperature incineration is known to be very high (Bjorklund et al., 2023; Logana- than et al., 2007). 3.1.3. Crystallisation The BBF obtained by crystallisation (CGO) was almost free of tar- geted PCBs, PAHs, PCDD/Fs and PFAS. Previous studies have shown that Fig. 1. Concentrations of P 7 PCBs (PCBs 28, 52, 101, 118, 138, 153 and 180), P 16 PAHs (NAP, ACE, ACY, FLE, PHE, ANT, FLU, PYE, BaA, CHR, BbF, BkF, BaP, IDP, BghiP, DBahA) and P 17 PCDD/Fs (see entire list in SI.2) in 19 BBFs produced by different methods and from various waste materials. The concentrations of compounds regulated by the strictest thresholds (in Luxembourg for PCBs and PCDD/Fs, in Denmark for PAHs) are represented as dark blue bars, the concentrations of substances not included in the thresholds (i.e., PCB 118, and PAHs NAP, ACY, ANT, BaA, CHR, DBahA) are represented as light blue bars. See Table 1 and SI.1 for details about the valorization methods and waste origins of BBFs. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 6 Fig. 2. Concentrations of P 44 PFAS, including 13 PFCAs, 10 PFSAs, 13 PFAA precursors (6 preFOS, 4 FTSAs, 3 diPAPs), and 8 other PFAS (see entire list in SI.3), and extractable organic fluorine in 19 BBFs produced by different methods from various waste materials. The concentrations of the regulated PFOS and PFOA (threshold of 100 μg kg 1 in some German regions) are represented as dark blue bars, the concentrations of the other 42 PFAS are represented as light blue bars. Note the different x-axis scale for target PFAS and EOF. See Table 1 and SI.1 for details about the valorization methods and waste origins of BBFs. Fig. 3. Fraction of organic fluorine explained by the 44 PFAS targeted in the study, i.e., 5 short-chain perfluoroalkyl acids (short-chain PFAAs), 18 long-chain perfluoroalkyl acids (long-chain PFAAs), 13 PFAA precursors (6 preFOS, 4 FTSA, 3 diPAP), and 8 other PFAS (see entire list in SI.3) in 13 BBFs produced with different methods from various waste materials (see details about BBFs in Table 1 and SI.1). N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 7 struvite precipitation produces fertilizers and amended plants with low concentrations of organic micropollutants (de Boer et al., 2018; Ron- teltap et al., 2007). In the CGO struvite, pesticides and pharmaceuticals were also found to be < LOQ in a previous study (Dong et al., 2023). 3.1.4. Valorization comparison Considering all BBFs, no statistically significant differences in pollutant concentrations were found between the three main groups of valorization methods, i.e., hygienization, pyrolysis, and incineration (crystallisation was not considered for statistical analysis as only one such BBF was investigated). This can be explained by the high variability caused by the waste origin in the hygienization group and by the higher PCB or PAH concentrations in some samples of the pyrolysis and incineration groups. When considering the SewSludge-BBFs only (overall highest contaminated BBFs and triplicates available for most of BBFs), concentrations of PCBs (ANOVA, F(2,12) ˆ 54.1, p < 0.01) and PFAS (ANOVA, F(1,10) ˆ 5.16, p ˆ 0.042) in BBFs that went through hygienization were significantly higher than those in BBFs obtained through pyrolysis (Tukey, PCBs: p < 0.01, PFAS: p ˆ 0.041) and incin- eration (PCBs: Tukey, p ˆ 0.011). This underscores the potential of high temperature thermal processes (i.e., pyrolysis and incineration) for persistent organic substance removal. In contrast, concentrations of PAHs in pyrolyzed products were statistically significantly higher ((ANOVA, F(2,12) ˆ 52.9, p < 0.01) than in BBFs obtained through hygienization and incineration (Tukey, p < 0.01 for both comparisons), confirming the earlier trade-off described between organohalogen pollutant removal and PAH generation (Sørmo et al., 2024). However, it should be noted that the amount and bioavailability of PAHs produced is very dependent on the pyrolysis technology (Hale et al., 2012). 3.2. Compliance assessment All measured concentrations of PCBs, PAHs, PCDD/Fs and PFAS in BBFs were well below the limit values set in individual EU countries except in one case (PAHs in the pyrolyzed product BAG) (Figs. 1 and 2). High concentrations of PAHs in pyrolyzed products (such as in BAG) are often explained by uneven heat distribution and vapor trapping during pyrolysis or cool zones in the post-pyrolysis area, and can often be drastically reduced by modifying the pyrolysis unit design (Buss et al., 2022). Post-treatment of pyrolyzed products at moderated temperature (100–300 C) can be used to thermally desorb PAHs (Kołtowski and Oleszczuk, 2015). Moreover, the bioavailability of PAHs in pyrolyzed products has been shown to be generally low (Hale et al., 2012). The present results are thus encouraging for the application of the studied BBFs as an alternative to conventional inorganic fertilizers. However, it is important to target – and regulate – more PFAS in the future, since a large fraction of total organic fluorine (>85 %) was not explained by the target PFAS in this study. This total organic fluorine can originate from non-target precursors that are known to be predominant in some organic waste such as sewage sludge (e.g., fluorotelomer alcohols, see section 3.1.1). This can also be explained by unconventional PFAS and non- PFAS organofluoride substances widely used in products such as phar- maceuticals (Spaan et al., 2023), batteries (Guelfo et al., 2024), or pesticides (Lasee et al., 2022) that could contaminate various types of organic waste. 3.3. Comparisons to current trends in contamination of organic waste 3.3.1. PFAS and EOF Overall, PFAS patterns in SewSludge-BBFs were dominated by PFOS (0.8–14.7 μg kg 1), perfluorohexanoic acid (PFHxA, < LOQ – 5.4 μg kg 1), ethylperfluorooctane sulfonamidoacetic acid (EtFOSAA, 1.3–2.9 μg kg 1), methylperfluorooctane sulfonamidoacetic acid (MeFOSAA, 0.6–2.1 μg kg 1), 12:2 fluorotelomer sulfonate (12:2 FTS, 0.7–1.6 μg kg 1), 10:2 FTS (0.4–1.5 μg kg 1) and PFOA (< LOQ – 2.4 μg kg 1). In raw sludges, the predominance of PFOS and its precursors MeFOSAA and EtFOSAA (as well as some other long-chain PFAS such as PFOA) has been reported by many studies (Higgins et al., 2005; Schultz et al., 2006; Sepulvado et al., 2011). PFAA precursors (entire list in SI.3) represented a significant fraction of target PFAS in SewSludge-BBFs (25–86 %) (Fig. 3). In contrast to recent studies that reported high concentrations of fluorotelomer phosphate diesters (diPAPs) in raw sludges (> 50 % of the targeted PFAS) (Aro et al., 2021a; Thompson et al., 2023), the concen- trations of these precursors were < LOQ in the three SewSludge-BBFs. An explanation could be that most of the diPAPs had already been transformed to fluorotelomer alcohols (FTOHs, not measured in this study) and perfluorocarboxylic acids (PFCAs) (Butt et al., 2014; D'eon and Mabury, 2007; Lee et al., 2010a; Yoo et al., 2010). This assumption is supported when considering that mostly even chain-length PFCAs (PFHxA, PFOA, PFDA) were quantified in PRV and RAN, which is consistent with the biological production of PFCAs from fluorotelomer- based compounds (Lee et al., 2010a). Moreover, PFHxA presented the second-highest concentrations of all the targeted PFAS (up to 5.4 μg kg 1). Unlike diPAPs, FTSs – which are also precursors of PFCAs – were detected in SewSludge-BBFs. The presence of FTSs could be explained by the slower degradation of these compounds (Wang et al., 2011; Zhang et al., 2016) and/or by their potentially initially higher concentrations (compared to diPAPs). The relatively high EOF concentrations in SewSludge-BBFs (320  133 μg kg 1) – which were consistent with levels reported in sludges (Aro et al., 2021a) – support that PFAA pre- cursors and other non-targeted PFAS, such as the above-mentioned FTOHs, can be present in high concentrations in these matrices. 3.3.2. PFAS temporal trends Although the degradation of PFAA precursors is expected to extend in time the occurrence of PFAA in organic waste, a decrease of regulated PFAS has been reported in sludges (Fredriksson et al., 2022; Gewurtz et al., 2024; Ulrich et al., 2016) and sludge based-products (Kim Lazcano et al., 2020). Between 2000 and 2010, average concentrations of PFOS in sludges were often reported to reach 100 μg kg 1 (Becker et al., 2008; Higgins et al., 2005; Schultz et al., 2006; Sepulvado et al., 2011) whereas average concentrations are currently getting closer to 10 μg kg 1 (Aro et al., 2021a; Eriksson et al., 2017; Fredriksson et al., 2022; Sørmo et al., 2023; Ulrich et al., 2016). Considering the concentrations of PFOS re- ported in the literature (Table 2), it appears that in sewage sludge in general, concentrations of PFOS have decreased by about 50 % every four years in Europe (or every six years in America), whether using the mean (Fig. 4) or median (SI.9) values. These data represent diffusely polluted municipal sewage sludges that – to the best of our knowledge - were not affected by major PFAS hotspots (see information about wastewater treatment plants and potential local PFAS sources in Table 2). A similar decrease has been reported in Europe by time-trend monitoring campaigns conducted by Fredriksson et al. (2022) and Ulrich et al. (2016); the same trend seems to apply to other long-chain PFAS (e. g., PFOA) and precursors because these PFAS have been replaced by industry with short-chain PFASs (Fredriksson et al., 2022). Recently, in Canada, Gewurtz et al. (2024) reported a decrease of regulated PFAS concentrations in sewage sludges, except for PFOS, for which a slower response to regulations/phase-outs seems to occur in this country. In Biowaste-BBFs, PFOA was present at higher concentrations than PFOS, and short-chain PFAS were predominant (Fig. 3), in line with previous studies and in contrast to profiles measured in SewSludge-BFFs (Choi et al., 2019a; Stahl et al., 2018; Thakali et al., 2022). Another main difference with SewSludge-BBFs was that PFAA precursors were not detected in Biowaste-BBFs. These two differences in profile could potentially be explained by the higher uptake of short-chain PFAS by plants which constitute a large fraction of the biowaste (Ghisi et al., 2019; Stahl et al., 2018); in addition short-chain PFAS likely accumulate less in sludge than long-chain ones in a wastewater treatment plant. However, there are also potentially many different sources of contami- nation from other inputs to biowaste streams, for example the high presence of packaging plastics in the two Biowaste-BBFs (Estoppey et al., N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 8 Table 2 PFOS concentrations in sewage sludges reported by European and American studies. WWTP: wastewater treatment plant. Study Sample year Median (μg kg 1) Mean (μg kg 1) Number of WWTPs (and country) Potential local PFAS sources Bossi et al. (2008) 2004 – 18.4 Six WWTPs (Denmark). Becker et al. (2008) 2006 100 100 One WWTP (Germany). 2/3 of wastewater came from commercial and industrial sources including breweries, food, plastics and tobacco industries. Zhang et al. (2010) 2008 213 333 Three WWTPs (Switzerland). Navarro et al. (2011) 2006 28.3 63.9 Twenty WWTPs (Spain). Llorca et al. (2011) 2010 73.5 84.2 One WWTP (Spain). Esparza et al. (2011) 2009 39.5 40.5 Four WWTPs (The Netherlands). Sun et al. (2011) 2008 78.0 139 Twenty WWTPs (Switzerland). The sewage sludges from three WWTPs presented PFOS levels ca. 8 higher than sewage sludges from the 17 other WWTPs. These three WWTPs were probably impacted by local wastewater sources (incl. Chromium electroplating and surface finishing industries). Gomez-Canela et al. (2012) 2012 1.91 1.94 Fifteen WWTPs (Spain (12) and Germany (3)). The three (Spanish) sewage sludges presenting the highest PFOS concentrations were produced byWWTPs treating wastewater from industrial sectors (with vehicle, textile, and chemical industries). Arvaniti et al. (2012) 2009 4.3 4.3 Two WWTPs (Greece). One received 80 % domestic wastewater and 20 % industrial wastewater, the other only domestic wastewater. Stasinakis et al. (2013) 2011 6.5 7.3 One WWTP (Greece). Martínez-Moral and Tena (2013) 2013 1.38 1.32 Different WWTPs (Spain). Perkola and Sainio (2013) 2010 63.0 63.0 One WWTP (Finland). Campo et al. (2014) 2010 51.7 229.1 Sixteen WWTPs 16 (Spain). 2011 0.01 38.0 Filipovic and Berger (2015) 2013 2.9 4.2 Three WWTPs (Sweden). Alder and van der Voet (2015) 2011 75 177.1 PFAS-related industrial and commercial activities in the catchment of 35 of the 45WWTPs: metal plating industries (22), fire brigade training sites and foam suppliers (8), textile/textile finishing industries (9), landfill leachates (5), paper manufactures (5), packaging supplier (1), airports (2). Ulrich et al. (2016) 2008 – 48.0 Several WWTPs (Germany). 2009 – 21.5 2010 – 18.0 2011 – 23.0 2012 – 24.0 2013 – 15.5 Navarro et al. (2016) 2011 8.2 14.8 Sixteen WWTPs (Spain). Zacs and Bartkevics (2016) 2015 0.16 0.27 WWTPs from the Baltic area. Eriksson et al. (2017) 2012 4.7 5.2 Three WWTPs (Sweden) The three WWTPs received domestic wastewater and water from hospitals. One WWTP received wastewater from textile and chemical industries. 2014 2.7 3.2 2015 3.0 2.4 Stahl et al. (2018) 2013 7.3 23.1 Different sewage treatment plants (Gemrnay). Abril et al. (2020) 2018 20.1 20.4 Ten sewage treatment plants (Spain). Aro et al. (2021a) 2017 3.9 4.9 Ten WWTPs (Finland, Sweden, Denmark, Norway, Faroe Islands). Fredriksson et al. (2022) 2004 27.5 27.5 Two WWTPs (Sweden). 2005 30.0 30.0 2007 26.0 26.0 2008 27.5 27.5 2009 27.5 27.5 2010 20.0 20.0 2011 15.0 15.0 2012 19.8 19.8 2013 10.1 10.1 2014 13.0 13.0 2015 12.0 12.0 2016 6.7 6.7 Sørmo et al. (2023) 2021 24.5 25.0 Three WWTPs (Norway). Higgins et al. (2005) 2001 77.3 400 Eight WWTPs (USA). All the WWTPs received at least 50 % domestic waste. One WWTP received papermill effluent; it however presented some of the lowest PFOS concentrations. Schultz et al. (2006) 2004 – 100 One WWTP (USA). Sinclair and Kannan (2006) 2004 30 31 Two WWTPs (USA). One WWTP influenced by domestic and commercial discharge, the other WWTP had an additional industrial discharge. Only small difference in PFOS concentrations between sludges (mean: 37 vs. 25 μg kg 1, median: 28 vs. 32 μg kg 1) Loganathan et al. (2007) 2005 61.0 61.7 Two WWTPs (USA). D’eon et al. (2009) 2002 41.0 95.2 Six WWTPs (Canada). Sepulvado et al. (2011) 2005 145 143 Biosolids (USA). No information about WWTPs (biosolids obtained from Metropolitan Water Reclamation District of Greater Chicago). Venkatesan and Halden (2013) 2001 – 403 Ninety-four WWTPs (USA). (continued on next page) N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 9 2024). These plastics were shown to be correlated with high contents of short-chain PFAS in composts (Choi et al., 2019b). Total concentrations of target Σ44PFAS (6.7  3.5 μg kg 1) were similar with recent data reported in literature (Bolan et al., 2021; Choi et al., 2019a; Kim Lazcano et al., 2020; O'Connor et al., 2022; Sivaram et al., 2022; Stahl et al., 2018; Thakali et al., 2022). The relatively low EOF concentrations (48  7 μg kg 1, μg F kg 1) in Biowaste-BBFs indicate that large amounts of precursors and long-chain PFAS are not expected in Biowaste-BBFs. In AgriFoodInduWaste-BBFs, the low concentrations of Σ44PFAS in (2.8  2.8 μg kg 1) were in accordance with results from Munoz et al. (2022) who reported median Σ44PFAS concentrations of 0.66 μg kg 1 for pig slurry, poultry manure and dairy cattle. Interestingly, EOF concentra- tions were relatively high in chicken manure-based FEK and OPU (434 and 143 μg kg 1, respectively). Soils on which (free-ranging) chicken are grown are often enriched in organic carbon through the build-up of feed waste and manure; these soils can sorb PFAS strongly and be an important sink, exposing the chickens through digestion of contami- nated soil particles and intake of soil organisms (Lasters et al., 2022). The closeness to point pollution sources cannot be excluded, which would amplify the contamination of chickens. Concentrations of PFAS in BBFs produced by pyrolysis and inciner- ation were > 5 μg kg 1 at an operating temperature < 500 C (CRA, MBC), and < 0.3 μg kg 1 for combustion conducted at >600 C (BAG, PLA, ADC, EPH). This is consistent with recent studies reporting PFAS removals >90 % when pyrolysis temperatures were > 500 C (Kundu et al., 2021; McNamara et al., 2023; Sørmo et al., 2023) and very low concentrations in the bottom ashes after incineration (Bjorklund et al., 2023; Loganathan et al., 2007). EOF concentrations were higher in CRA produced at 190 C (107 μg F kg–1) than in BAG produced at 650 C (25 μg F kg 1), indicating a beneficial effect of high temperature to reduce PFAS concentrations. PFAS removal occurring during thermal processes does not mean that PFASs are fully destroyed. At pyrolysis temperatures between 500 and 800 C, Sørmo et al. (2023) showed that moderate amounts of (shorter chain) PFAS were present in the flue gas (< 3 % of the total PFAS-mass in the waste) and high amounts of (longer chain) PFAS were expected to be concentrated in condensation oils, requiring treatment at higher temperature to be destroyed (McNamara et al., 2023). In contrast, incineration temperatures (> 800 C) were shown to allow for mineralization of PFAS (Gehrmann et al., 2024; RIVM, 2021) although small amounts of PFAS were measured in flue gas (4.0–5.6 ng m 3) (Bjorklund et al., 2023). 3.3.3. PCBs and PCDD/Fs In SewSludge-BBFs, PCB profiles were dominated by the hexa- chlorinated PCBs 138 and 153 (> 35–46 % ofP7PCBs), as reported by previous studies on raw sewage sludges (Antolín-Rodríguez et al., 2016; Urbaniak et al., 2017; Zennegg et al., 2013). PCDD/F profiles were dominated by OCDD (73 %) and HpCDD (9 %), similar to what was found in sewage sludges (contribution of around 80 % for these two congeners, (Elskens et al., 2013; Zennegg et al., 2013) and in back- ground air in Europe (contribution of >60 %) (Degrendele et al., 2020). Atmospheric deposition and wash-off by rain into combined sewer sys- tems is one of the main sources of these compounds (Zennegg et al., 2013). Concentrations of P 7PCBs (6.5  1.1 μg kg 1) andP7PCDD/Fs (2.4  1.6 ng TEQ kg 1) in SewSludge-BBFs were at the low end of the range of concentrations reported in sewage sludges since the 1980s (Fig. 5). This underscores the drastic decrease of PCB and PCDD/F contamination in sewage sludge over the last decades due to the phasing-out of PCBs and the introduction of PCDD/F restrictions (UNEP, 2001). Since the wash-off of atmospheric deposition into combined sewer systems is a primary source of PCBs and PCDD/Fs in sewage sludge (Zennegg et al., 2013), the observed decreasing trend in this organic waste is most likely due to the reduction of emissions of these compounds into the air (EEA, 2024). Fig. 5 shows the projected Table 2 (continued ) Study Sample year Median (μg kg 1) Mean (μg kg 1) Number of WWTPs (and country) Potential local PFAS sources Guerra et al. (2014) 2010 13 534a or 29.6 Fifteen WWTPs (Canada). Sludges from one WWTP receiving industrial wastewater contained very high PFOS concentrations (13′100 and 2099 μg kg–1). These concentrations impact very much the mean value: 534.2 μg kg 1 (considered) vs. 29.6 μg kg 1 (not considered) Armstrong et al. (2016) 2005 1.0 1.1 One WWTP (USA). 2006 50.4 51.7 2007 28.8 27.3 2008 21.1 21.3 2009 19.0 19.3 2010 18.8 15.6 2011 13.5 15.6 2012 12.6 12.1 2013 17.0 21.2 Gottschall et al. (2017) 2008 7.2 7.2 One WWTP (Canada). The WWTP processed domestic, commercial and industrial wastewater Kim Lazcano et al. (2020) 2018 9.9 10.1 Four WWTPs (USA). PFOS concentrations in sludges prior studied treatment process have been used to determine the mean and median values. Letcher et al. (2020) 2017 5.7 10.9 Twenty WWTPs (Canada). Kim Lazcano et al. (2020) 2014 10.3 18.2 Eleven commercially available bio-based products (USA). Gewurtz et al. (2024) 2009 22.8 1072a or 45.8 Twenty-seven WWTPs (Canada). Sludges from one WWTP presented high PFOS concentrations in 2009 (7617.5 μg kg 1); no significant industrial source was known. These concentrations impact very much the mean value: 1072 μg kg 1 (considered) vs. 45.8 μg kg 1 (not considered).2010 8.3 10.5 2011 14.6 12.8 2013 14.7 37.7 2014 3.54 62.3 2015 6.0 8.4 2016 9.4 11.0 2018 12.1 15.6 2019 9.9 17.7 2021 7.9 10.3 a Value not considered in the determination of the temporal trend (Fig. 4) because the contribution of a local hotspot was suspected. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 10 evolution of PCB and PCDD/F concentration in sewage sludge-based fertilizers, using (i) the apparent half-life reported by Zennegg et al. (2013) (10 years for PCBs, 12 years for PCDD/Fs) and (ii) the apparent half-life found when fitting literature data to sludges sampled between 1990 and 2023 (6 years for PCBs, 9 years for PCDD/Fs) (Abad et al., 2005; Alcock and Jones, 1993; Blanchard et al., 2004; Eljarrat et al., 2003, 1999; Elskens et al., 2013; Fijalkowski et al., 2017; Fuentes et al., 2007; Katsoyiannis and Samara, 2004; Kaya et al., 2015; Martínez et al., 2007; McGrath et al., 2000; Roskosch and Heidecke, 2018; Sava et al., 2024; Sørmo et al., 2024; Stevens et al., 2001, 2003; Zennegg et al., 2013). In Biowaste-BBFs, concentrations of P 7PCBs (11.3  3.8 μg kg 1) were at the low end of the range of concentrations in composts or digestates from organic household and green wastes reported between 2000 and 2020 (7.9–63 μg kg 1) (Antolín-Rodríguez et al., 2016; Barcauskait _e, 2019; Benísek et al., 2015; Brandli et al., 2007, 2005; Govasmark et al., 2011; Hellstrom et al., 2011; O'Connor et al., 2022; Visniausk _e et al., 2018). For P 17PCDD/Fs, concentrations in biowaste- BBFs (1.5 1.3 ng TEQ kg 1) were not significantly different from those measured in SewSludge-BBFs, supporting findings from Elskens et al. (2013). The obtained concentrations are slightly lower than those measured in composts in the 2000s (8.5–9.5 ngTEQ kg 1) (Brandli et al., 2005) and in the 2010s (4.1  1.5 ngTEQ kg 1), most probably because of the general decrease of PCDD/Fs emissions. AgriFoodInduWaste-BBFs have been poorly investigated in terms of PCB and PCDD/F contamination. Low PCB concentrations have been reported in manure (0.99 μg kg 1, or 0.15 ng TEQ kg 1) or in manure compost (2.7 μg kg 1) (Barcauskait _e, 2019; Elskens et al., 2013; Ng et al., 2008) which is in accordance with manure-based BBFs included in this study (< 0.5 μg kg 1 in FEK, OPU, BIO). Concentrations of P 17PCDD/Fs in AgriFoodInduWaste (0.19 ng TEQ kg 1) were in accordance with the most recent concentrations reported in manure (0.15 ng TEQ kg 1) (Elskens et al., 2013) and slightly lower than those reported for manure in the 2000s (0.1–4 ng TEQ kg 1) (Ng et al., 2008; Stevens and Jones, 2003; Welsch-Pausch and McLachlan, 1998). The presence of PCBs and PCDD/Fs in manure can be explained by the fact that livestock farming accumulates these substances from soils (contaminated from past releases or emissions from buildings and con- struction works) and that significant fractions of these contaminants are expected to leave animals in manure (Weber et al., 2018a, 2018b; Welsch-Pausch and McLachlan, 1998). This is especially true for free- range chickens that take up more soil than other farm animals per body weight (Weber et al., 2018a, 2018b). The PCB concentrations < LOQ in animal (OG2, ECO) and plant (MO14, BA1) based BBFs are in accordance with low PCB levels being observed in meat, cereals, vege- tables and fruits (0.011–2.26 μg kg 1) (Esposito et al., 2017; Schwind et al., 2009; Zhang et al., 2008). Elskens et al. (2013) showed that dioxin-like PCB concentrations in plant and animal-based fertilizers (0.05–0.06 ng TEQ kg 1) were significantly below those in sewage sludges and composts (1.6–2.2 ng TEQ kg 1), confirming that PCB levels in AgriFoodInduWaste-BBFs are at least 10 times lower than in SewSludge-BBFs and Biowaste-BBFs. In BBFs produced by pyrolysis, concentrations of P 7PCBs ranged from < LOQ, when obtained by dry pyrolysis (> 300 C), to concen- trations similar to common feedstock (6.8 μg kg 1), when obtained by wet pyrolysis (< 200 C). This is in line with previous studies that report a 1–2 order magnitude reduction of P7PCB concentrations with dry pyrolysis of sewage sludges (from 7.6 to 20.7 μg kg 1 (Sørmo et al., 2024) or 274.2 μg kg 1 (Mosko et al., 2021) in sludges to <0.25–1.7 μg kg 1 or < LOQ - 26.6 μg kg 1 in pyrolyzed products, respectively) and no significant reduction (or generation) of PCBs by wet pyrolysis (HTC) (Brookman et al., 2018; Tasca et al., 2022). Concentrations of P 7PCDD/ Fs were very low in most of the BBFs (<0.005 ng TEQ kg 1) and slightly higher in BAG obtained by dry pyrolysis at 600 C (0.3 ng TEQ kg 1). Conversion of PCBs into PCDD/Fs in BBFs is unlikely due to the low levels of PCBs. These results show that processes involving thermal treatment do not generally generate by-products, but require close monitoring of operating conditions to avoid uneven heat distribution or Fig. 4. Mean PFOS concentrations in sewage sludge reported in the literature from Europe (open black circles, 23 studies) and America (open grey squares, 14 studies) and measured in SewSludge-BBFs (filled black circle, this study). The solid black and grey lines are the fits of European and American data and represent decreases of PFOS concentration with an apparent half-life of 4 and 6 years, respectively; dashed black lines show a 95 % confidence interval for European values. All concentrations are given in Table 2 and graphs with median values are presented in SI.9. The samples for the current study were taken in 2020. Note the logarithmic y-axis. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 11 cool zones in the post-pyrolysis area, especially when substantial chlo- rine sources are present in the feedstocks (Altarawneh et al., 2009; Buss et al., 2022; Chagger et al., 2000; Wang et al., 2017). 3.3.4. PAHs In SewSludge-BBFs, PAH profiles were dominated by phenanthrene (16 %), fluoranthene (15 %), and pyrene (18 %), consistent with earlier results (Chen et al., 2019). Concentrations of P 16PAHs in SewSludge- BBFs (1.41  1.28 mg kg 1) were in line with concentrations reported in sewage sludges throughout Europe from 2010 (close to 1 mg kg 1) and about an order of magnitude lower than those measured in the beginning of the century (see SI.10) (Alhafez et al., 2013; Baran and Oleszczuk, 2003; Berset and Holzer, 1999; Boruszko, 2017; Busetti et al., 2006; Perez et al., 2001; Roskosch and Heidecke, 2018; Sørmo et al., 2024; Stevens et al., 2003; Suciu et al., 2015; Villar et al., 2006). Thus, similarly to the other substances, a general decrease in PAH concentrations seems to occur with a half-life of <10 years (SI-10). Furthermore, the hygienization process used to produce the studied BBFs (i.e., drying and/or anaerobic digestions) is expected to decrease the amount of PAHs; a reduction of >60 % was indeed reported when applying aerobic stabilization (Trably et al., 2005; Włodarczyk-Makuła et al., 2021). Concentrations of P 16PAHs in Biowaste-BBFs (0.86  0.3 mg kg 1) were similar to those measured in composts from organic household waste and green waste (1.7–1.9 mg kg 1) (Brandli et al., 2005; Farrell and Jones, 2009). Low concentrations in Biowaste-BBFs can be explained by the degradation of some PAHs in composts, espe- cially low molecular weight PAHs (up to 90 % reduction) (Brandli et al., 2007; Houot et al., 2012). The low concentrations of P 16PAHs in AgriFoodInduWaste-BBFs (0.26  0.36 mg kg 1) are in good agreement with the very low concentrations reported in cereals (< 0.001 mg kg 1) (Einolghozati et al., 2022) or manure (< 0.5 mg kg 1) (Mackiewicz- Walec and Krzebietke, 2020). Fig. 5. Concentrations of P PCBs 28, 52, 101, 138, 153, 180 and 118 when measured - (upper panel) and of P 17 PCCD/Fs (lower panel) in sewage sludge-based fertilizers reported in the literature and measured in SewSludge-BBFs (this study). The solid grey line is the decrease of concentrations in sewage sludge reported by Zennegg et al. (2013) with an apparent half-life of 10 years for PCBs and 13 years for PCCD/Fs. The solid black line is the fit of all data presented and shows a decrease of concentrations with an apparent half-life of 6 years for PCBs and 9 years for PCDD/Fs; dashed black lines show a 95 % confidence interval. All numerical data can be found in Supplementary Information SI.10. Note the logarithmic y-axis. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 12 In BBFs produced by pyrolysis, low concentrations of P 16PAHs in two of the pyrolyzed products (< 1 mg kg 1 for CRA and MBC) and one high concentration exceeding existing thresholds (23.2 mg kg 1 for BAG), were consistent with the values reported byWang et al. (2017) for 102 pyrolyzed products, i.e., < 1.5 mg kg 1 for about 90 % of the py- rolyzed products, and up to 100 mg kg 1 in 10 % of the cases. In ashes, the concentrations of P 16PAHs measured in BBFs (< 0.5 mg kg 1) were comparable to those reported in bottom/bed ashes from municipal waste/biomass incineration (< 1 mg kg 1 in most of the cases) (Enell et al., 2008; Masto et al., 2015). 3.4. Expected and measured concentrations of PCBs and PFASs in soil 3.4.1. Contamination after one BBF application Expected concentrations of the persistent organic substances in amended soils (initially not contaminated) - after one application –were < 0.033 μg kg 1 forP7PCBs, < 0.009 ng TEQ kg 1 for Σ17PCDD/Fs, < 0.004 mg kg 1 for Σ15PAHs, < 0.04 μg kg 1 for Σ44 PFAS (Table 3). All these concentrations were at the low end of the ranges of background concentrations in soils; PFOS and PFOA were used for the comparison with literature since data for these two congeners are the most widely reported (Table 3). These results indicate that one BBF application does not constitute a significant contribution to the contamination of agri- cultural soils by target substances. Measurements conducted on soil samples from field trials revealed modest but statistically significant (t-test, p < 0.05) increases in PCB concentrations between the control soils and the BBF-amended soils. Indeed, increases of 0.12 and 0.50 μg kg 1 (P7PCBs) were measured in PLP-amended soils (Finland) and EPH amended-soil (Austria), respec- tively (Supplementary Information SI.11). The increase in PCB concen- trations in PLP-amended soils was slightly higher than the expected increase (0.12 vs. 0.032 μg kg 1) and can probably be explained by the variability of PCB concentrations in the BBFs and soils. In contrast, the measured PCB concentration increase in EPH-amended soil was much higher than expected (0.40 vs. 0.0018 μg kg 1). It appeared that some PCBs (e.g. PCB 138 and PCB153, responsible for half of the increase in EPH-amended soil) were < LOQ in EPH, indicating that the increase of PCB concentrations in the amended soil had most probably another origin than the BBF itself. An explanation could be that 9 months separated the sampling of the pre-trial and amended soils; PCB contamination through the atmosphere or from other plots might be the cause of this increase in Finland. Another explanation could be the heterogeneity and uncertainty for measurements at such low levels. In any case, this measured difference (0.40 μg kg 1) was at the low end of background concentrations in soils (0.01–58 μg kg 1). Regarding PFAS, PFOS was the only quantifiable congener in soils from the field trials. The concentrations in OPU-amended soils were not higher than in controls or pre-trial soils (Supplementary Information SI.11). In Finland, the concentration of PFOSwas higher in the control (3.1 μg kg 1) than in the OPU-amended soil (0.6 μg kg 1), most probably due to soil heterogeneity. 3.4.2. Prediction of pollutant accumulation in amended soils Persistent organic substances have been shown to accumulate in soils amended with recycled fertilizers (Sepulvado et al., 2011; Umlauf et al., 2011; Washington et al., 2010a, b; Weber et al., 2018a, b). A linear in- crease with increasing biosolid loading rate was for example reported by Sepulvado et al. (2011). Therefore, the prediction of contamination in amended soils should take into account the expected persistent organic substance concentrations in BBFs in the future (Eq. (3)) as well as the cumulative concentrations over time (Eq. (4)). As discussed in Section 3.3, the concentrations of legacy contaminants in SewSludge-BBFs are expected to decrease with an apparent half-life of <10 years. Scenarios with half-life values of 10, 15 and 20 years were used for conservative assessment (worst-case scenarios). For contaminant concentrations in BBFs, Eq. (3) was used: Ccont;BBF;tˆn ˆ Ccont;BBF;tˆ0*2 t=x (3) where Ccont,BBF,tˆn is the contaminant concentration in SewSludge-BBFs in n years, Ccont,BBF,t ˆ0 is the average concentration of contaminant in SewSludge-BBFs at tˆ 0 (concentrations given in Figs. 1 and 2), and x is the half-life of contaminants in BBFs (10, 15 and 20 years for the pre- sented scenarios). For contaminant concentrations in soils, Eq. (4) was used: Ccont;soil;tˆn ˆ Ccont;soil;tˆ0‡ Xn tˆ0 mBBF msoil Ccont;BBF;tˆn (4) where Ccont,soil,tˆn is the accumulated contaminants in amended soil after n year, Ccont,soil,tˆ0 is the initial concentration in soils at t ˆ 0 (e.g., median values given in Table 3), mBBF is the average mass of BBF applied per hectare annually (4.4 tons; average application rate for SewSludge- BBFs), and msoil is the average mass of soil amended per hectare (3900 tons; 30 cm of soil with density of 1.3 g cm 3). Additional information about the model can be found in SI.12. For PFAS, the predictions are complicated by the potential leaching of some PFAS and the transformation of precursors (Holly et al., 2024; Pepper et al., 2021b; Ye et al., 2024). While short-chain PFAS (CnF2n‡1COOH, n  6, and CnF2n‡1SO3H, n  5) are readily mobilized from soils, long-chain product PFAS such as PFOS and its precursors remain mostly in the upper soil horizon because the desorption can be relatively slow (Gellrich et al., 2012; Gnesda et al., 2022; Lee et al., 2010b; Maizel et al., 2021; Sepulvado et al., 2011; Stahl et al., 2013b; Washington et al., 2010b). Therefore, predictions should consider an accumulation of the long-chain PFAS in the top layer of the soils whereas the short-chain will reach the groundwater more rapidly. Regarding the transformation of precursors, previous studies reported that concentra- tions of MeFOSAA and EtFOSAAmeasured in amended soils were shown to be much lower than those calculated from the concentration in sludges, suggesting transformation of these precursors (Pepper et al., 2021b; Sepulvado et al., 2011). Thus the predictions for PFOS were based on all precursors (FOSA, EtFOSAA and MeFOSSA, (Kolanczyk Table 3 Expected concentrations in soils amended with SewSludge-BBF, Biowaste-BBFs, AgriFoodInduWaste-BBFs, as well as BBFs obtained through pyrolysis and inciner- ation, when maximum allowed application rates are used (0.6–13.1 t 1 ha 1 y 1, see SI.1 for individual application rates). The amount of soil in the top layer (30 cm) was determined using a density of 1.3 g cm–3 (i.e., 3′900’000 kg ha 1). Hygienization Pyrolysis Incineration Background concentration in soils a SewSludge-BBFs Biowaste-BBFs AgriFoodInduWaste-BBFs P 15PAHs (mg kg 1) < 0.004 < 0.004 < 0.001 < 0.002 < 0.0003 0.004–7.3 (mean: 0.37) P 7PCBs (μg kg 1) < 0.011 < 0.033 < 0.001 < 0.005 < 0.003 0.01–58 (mean: 2.9) P 44PFAS (μg kg 1) < 0.036 < 0.022 < 0.013 < 0.008 < LOQ PFOS (μg kg 1) < 0.016 < 0.002 < 0.00008 < 0.0002 < LOQ 0.02–162 (median: 2.7) PFOA (μg kg 1) < 0.003 < 0.003 < LOQ < LOQ < LOQ 0.02–124 (median: 2.7) P 17PCDD/Fs (ng TEQ kg 1) < 0.004 < 0.009 < 0.003 < 0.0002 < 0.0000003 0.5–28.9 (mean: 3.18) a Background concentrations are fromMeijer et al. (2003) for PCBs, Vives et al. (2008) for PCDD/Fs, Nam et al. (2008) for PAHs, Brusseau et al. (2020) for PFOS and PFOA. For PCFF/Fs values reported are for Italy only, but studies conducted in other countries fell in this range (Environment Agency, 2009). N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 13 et al., 2023)) fully transforming to PFOS, in combination with an apparent half-life of 10, 15 or 20 years for both PFOS-precursors and PFOS in BBFs. For all scenarios, the concentration of PFOSwould reach a plateau at a value below 0.3 μg kg 1 in (initially non-contaminated) soils amended with SewSludge-BBFs (Fig. 6). Although these numbers are subject to large uncertainty due to the factors described above, it is apparent that those concentrations are well below the median back- ground concentrations of PFOS in soils (2.7 μg kg 1, Table 3) and the limit values set by some countries (the strictest being 0.8 μg kg 1 in The Netherlands (Hall et al., 2020)). When applying SewSludge-BBFs on soils already contaminated by PFOS (e.g., median background concen- tration given in the literature), the increase due to BBF application is expected to be <11 % (SI.12). Thus, the application of BBFs is not ex- pected to constitute a risk for contaminating the topsoil layer with PFAS. However, short-chain PFAS in BBFs could potentially constitute a risk for groundwater contamination. For PCBs, PCDD/Fs and PAHs, conservatively assuming there is no degradation and transfer of these persistent organic substances in soils (Umlauf et al., 2011), the concentration of P 7PCBs, P 17PCDD/Fs and P 16PAHs would reach plateau values of 0.25 μg kg 1, 0.1 ng TEQ kg 1 and 0.3 mg kg 1, respectively, in (initially non-contaminated) soils amended with SewSludge-BBFs (Fig. 6). Those concentrations are well below the mean background concentrations of PCB, PCDD/Fs and PAHs in soils (2.9 μg kg 1, 3.18 ng TEQ kg 1 and 0.37 mg kg 1, respectively). In cases of existing soil contamination by PCBs, PCDD/Fs and PAHs, (e. g., median background concentration given in the literature), the in- crease due to BBF application is <8 % for PCBs, <3 % for PCDD/Fs and <12 % for PAHs (SI.12). Fig. 6. Expected concentrations of legacy substances in the top layer (30 cm) of initially non-contaminated soils amended with SewSludge-BBFs at an average maximum allowed application rate (4.4 t y 1 ha 1) using apparent decreases target substances in BBFs with a half-life of 10, 15 and 20 years (see Figs. 4 and 5, and discussion in Section 3.3). For conservative assessment, it was assumed that all quantified PFOS precursors transform to PFOS, and these contaminants do not significantly transfer in soils, nor degrade. The amount of soil in the top layer (30 cm) was determined using a density of 1.3 g cm 3 (i.e., 3′900’000 kg ha 1). Equations and results with initially contaminated soils are given in SI.12. N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 14 This indicates that the application of BBFs is not expected to constitute a soil or crop contamination risk for PCBs, PCDD/Fs and PAHs in the long term. Previous studies also stated that the application of organic fertilizers containing similar levels of PCDD/Fs and PCBs to the studied BBFs does not significantly affect the soil since the contribution from fertilizers is usually very low compared to that of atmospheric depositions (Amlinger et al., 2004; Elskens et al., 2013; Stevens et al., 2003; Timmermann et al., 2003; Umlauf et al., 2011). 3.5. Uptake by plants and risk for humans Organic pollutant bioavailability in soils and uptake by edible plant parts are usually low for all but the most mobile substances (Clarke et al., 2010). PFAS are the studied compounds that present the lowest sorption to soil (and thus the highest bioavailability), with soil-water partitioning coefficient (Kd) values around 0.3 L kg 1 for PFBA, 1 L kg 1 for PFOA, 8 L kg 1 for PFOS in common soils and under ambient pH conditions (SI.5) (Nguyen et al., 2020). For cereal grains, some of the (maximum) bio- accumulation factor (BAF) values reported are 0.48 for PFBA, 0.16 for PFOA and 0.06 for PFOS (SI.5) (Doucette et al., 2018; Krippner et al., 2014; Stahl et al., 2009; Wen et al., 2014). PFBA, PFOA and PFOS concentrations expected in BBF-amended grains were calculated using these BAF values (Table 4). The data show that consumption of food from BBF-amended plots would only contribute minimally to tolerable daily intake (TDI), confirming results from previous studies on sludges (Gottschall et al., 2017). The other substances studied herein present much stronger sorption to soil particles. Kd values - calculated from organic carbon to water partitioning coefficient, Koc (log Koc ˆ 0.00028 ‡ (0.983 x log Kow) (Di Toro, 1985)), at an organic carbon content of 3 % - range between 200 and 120′000 L kg–1 for PAHs, 11′000 and 520′000 L kg–1 for PCBs, and 30′000 and 2′200’000 L kg–1 for PCDD/Fs (SI.5). Therefore, reported BAF for these compounds are very low, i.e., < 1 for PAHs, < 0.005 for PCBs and < 0.0005 for PCDD/Fs (see SI.5), the main uptake for these compounds originating from the atmosphere rather than the soil (Nizzetto et al., 2008; Su et al., 2007). Concentrations of PAHs, PCBs and PCDD/Fs in plants were based on conservative BAF values (1, 0.005 and 0.0005, respectively). The consumption of cereals from BBF-amended plots only minimally contributed to tolerable intake levels (Table 4). Low uptake in crops due to BBF amendment was confirmed by all measured PCB and PFAS concentrations in barley and maize grains being < LOQ, except PCB 28 which was quantified at a very low con- centration (0.11 μg kg 1) in barley grains from Finland. The presence of this PCB can be explained by an uptake from the atmosphere which has been shown to be a more important pathway for such compounds (Collins et al., 2006; Nizzetto et al., 2008). This study assessed the risk of BBFs for cereals. However, BAFs for PFOS for other edible plants could be higher. Values ranging between 0.1 and 1.67 were reported for lettuce, and between 0.07 and 0.7 for radish roots (Blaine et al., 2013). For these vegetables, in the worst case, the concentrations of PFOS could be up to 30 times higher than the PFOS concentrations presented in Table 1 for cereals. Therefore, the use of tested SewSludge-BBFs for vegetables should not be recommended without additional treatment (e.g., thermal treatment). More experi- mental data on plant uptake would strengthen the conclusion, and are recommended for future studies. 3.6. Further discussion of processes impacting the availability and uptake of contaminants In this study, the risk of BBFs was assessed using a conservative approach (i.e., assessment of the worst-case scenario) to ensure that contaminants do not enter the food chain at levels of potential health concern. Therefore, for the long-term assessment, we assumed no degradation and migration of the studied legacy contaminants (PCBs, PCDD/Fs, PAHs, PFOS) as well as a full transformation of all measured PFOS precursors (FOSA, MeFOSAA and EtFOSAA) to PFOS. In addition, the highest literature-based BAF values were used to assess the uptake of contaminants in cereals. Hereafter, we discussed to which extent contaminant concentrations would change if these conservative as- sumptions were not met. 3.6.1. Degradation Degradation rates of PCBs, PCDD/Fs and PFOS in soils are very low, with half-lives ranging from a few years to hundreds of years (Campanella et al., 2002; Dickman and Aga, 2022; Sinkkonen and Paasivirta, 2000; Terzaghi et al., 2021; Umlauf et al., 2011). Heavier PAHs are also very persistent (e.g., half-life of many years), whereas lower molecular weight PAHs are more prone to biodegradation (e.g., half-life of the three-ring molecule phenanthrene is a few months) (Shuttleworth and Cerniglia, 1995). If degradation occurs, contaminant concentrations in soils reach maximum values after times that depend on the half-life and the scenario studied in section 3.4.2. (see graphs in S.12). These maximum concentrations are about 46 % of the plateau values reported in Section 3.4.2 if the degradation of compounds occurs with a half-life of 20 years, 32 % of the plateau value with a degradation half-life of 10 years, and 23 % of the plateau value with degradation half-life of 5 years. 3.6.2. Migration Themigration of contaminants in (saturated) soils mostly depends on contaminant sorption (Kd values), soil physical properties (bulk density ρb and porosity θ), and environmental conditions (precipitations). The retardation factor Rf (i.e., the estimate of how much slower a contami- nant moves compared to water) is given by Rfˆ 1‡ (Kd ρb/θ). For PAHs, PCBs, and PCDD/Fs, Kd values range between 200 and 2′000’000 L kg–1 Table 4 Expected concentrations in grains of cereals grown on BBF-amended soils, when considering maximum allowed application rates (0.6 to 13.1 t 1 ha 1 y 1, see SI.1 for individual application rates) and BAF values of 0.48 for PFBA, 0.11 for PFOA and 0.06 for PFOS (Bizkarguenaga et al., 2016; Blaine et al., 2013; Ghisi et al., 2019; Krippner et al., 2014; Lesmeister et al., 2021; Stahl et al., 2009; Wen et al., 2014) as well as conservative BAF values of 1 for PAHs, 0.005 for PCBs and 0.0005 for PCDD/Fs (Kacalkova and Tlustos, 2011; Paraíba et al., 2010; Strek et al., 1981). Hygienization Pyrolysis Incineration Tolerable intakea SewSludge-BBFs Biowaste- BBFs AgriFoodInduWaste- BBFs P 15PAHs (μg kg 1) 0.3–3.7 1.4–3.0 0.01–1 0.1–1.8 < LOQ – 0.3 0.3 μg kg 1 b.w.d 1 for BaP (US EPA) P 7PCBs (ng kg 1) < 0.02 < 0.02 < 0.005 < 0.009 < 0.001 20 ng kg 1 b.w. d 1 (US EPA) P 17PCDD/Fs (pg TEQ kg 1) < 0.002 < 0.005 < 0.002 < 0.00006 < 0.0000001 2 pg TEQ kg 1 b.w. week 1 for PCDD/Fs& dioxin like-PCBs (EFSA) PFBA (ng kg 1) < LOQ < LOQ 6.7 < LOQ 5.4 < LOQ – 3.5 < LOQ – PFOS (ng kg 1) 0.04–0.90 0.086–0.093 < LOQ 0.004 < LOQ – 0.0073 < LOQ 4.4 ng kg 1 b.w. week 1 for P PFOS, PFOA, PFNA, PFHxS PFOA (ng kg 1) < LOQ 0.333 0.219–0.374 < LOQ < LOQ < LOQ a Updated from Popli et al. (2022). N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 15 for an organic carbon content of 3 % (field trials: 2–3.1 %). Assuming a bulk density of 1.3 g cm 3 and a soil porosity of 50 % (field trials: clay- rich soils), the retardation factor ranges between 500 and >5  109. If the annual precipitation (field trials: about 750 mm year 1) fully in- filtrates into the soil and water moves downward, the migration of contaminants is expected to range between 0.15 μm year 1 (for Rf ˆ 5 109) and 1.5 mm year 1 (for Rfˆ 500). Therefore, assuming no colloidal transport, 200 to 2 109 years are needed for PCBs, PCDD/Fs, and PAHs to migrate from the top layer (30 cm), confirming that the effect of migration is negligible for these compounds. For PFOS, the average Kd value is about 8 L kg 1 at pH 7.2 (field trials: pH 6.2–6.9) (Nguyen et al., 2020). Using the above-mentioned approach, it can be estimated that 9 years would be required for PFOS to migrate from the top layer (30 cm) with precipitation of 750 mm year 1, confirmed by Sørmo et al. (2024) who showed that 90 % leaching of PFOS from a 1 m soil layer would occur after 15 years (pH 6.2, precipitation 750 mm). The effect of such leaching on PFOS concentrations in soil is similar to a decrease with a half-life of about 10 years as shown in SI.13. In these conditions, con- centrations that reach 23 % of the plateau values reported in Section 3.4.2 can be expected. If the top layer is not saturated, the proportion of PFOS retained is expected to be higher because the retention in the vadose zone is stronger due to high adsorption at the air-water and solid- water interfaces (Gnesda et al., 2022). 3.6.3. Precursor transformation The degradation of PFOSprecursors in soilswas shown to occurwith a half-life in the range of weeks (Lange, 2001; Mejia Avenda~no and Liu, 2015; Rhoads et al., 2008; Zabaleta et al., 2018). PFOS yields reported in literature widely vary, from 3% (Mejia Avenda~no and Liu, 2015) to 30% and more (Zabaleta et al., 2018; Zhao et al., 2016). A PFOS yield of 30 % would lead to PFOS concentrations that equal 72% of the concentrations presented in Section 3.4.2. 3.6.4. Plant uptake Available BAF values for cereal grains are very scarce (see SI.5) (Doucette et al., 2018; Krippner et al., 2014; Stahl et al., 2009; Wen et al., 2014). BAF values for wheat grains (0.054–0.062) are the highest BAFs reported in the literature; they were used in Section 3.5 as con- servative values. Using the lower BAF values of oat (0.004–0.017) lead to concentrations expected in grains that are 6 % to 28 % of the values provided in Table 4. In summary, the degradation rates of PCBs, PCDD/Fs, and PFOS in soils are very low, with half-lives ranging from a few years to hundreds of years, while lower molecular weight PAHs degrade faster. Migration of these contaminants in saturated soils is minimal due to high sorption, with PFOS taking some dozens of years to migrate 30 cm. Plant uptake of these contaminants is generally low. 4. Conclusion To allow a safe circular economy through the use of bio-based fer- tilizers (BBFs), it is essential to assess the exposure of persistent organic substances in BBFs and the risk they constitute to the environment and food chain. Regarding the 19 studied BBFs - produced with different methods and from various waste materials - PCDD/F, PCB, PAH and PFAS concentrations were below the strictest limit values used in indi- vidual EU countries, except in one case. The present results thus repre- sent a positive incentive for the implementation of the studied BBFs as alternatives to conventional inorganic fertilizers. BBFs produced from agricultural and food industry waste through hygienization processes were shown to be particularly promising. They contained extremely low concentrations of target substances, and the absence of thermochemical treatments retains nutrient solubility and fertilizer value of the product. However, attention must be paid to the relatively high concentrations of EOF in chicken-manure based BBFs, and our results suggest that future monitoring studies could target more PFAS, especially PFAA precursors to ensure the safe use of these BBFs. In addition, the EOF concentrations in BBFs produced from sewage sludge were relatively high; therefore, additional PFAS should be included in the quantification methods as new knowledge increases about PFAA precursors, and fluorinated compounds in general. The use of pyrolysis and incineration to produce BBFs should probably be regarded as the preferred option for the valo- rization of organic waste containing elevated persistent organic sub- stance concentrations. Expected concentrations of the target substances in soils, even upon maximal allowed BBF application rates, were at the low end of the background soil concentration ranges, indicating that the application of the selected BBFs does not constitute a substantial contribution to the contamination of agricultural soils by these legacy pollutants. This was confirmed by PCB and PFAS analyses conducted on soil samples from BBF field trials. In addition, the consumption of food from BBF-amended plots would only contribute minimally to reaching tolerable intake thresholds, as evidenced by modelling and measure- ments. Though these results are promising for BBFs considered in this study, it should be kept in mind that contaminated BBFs can reach the market, e.g., in case they are produced near a contaminant hotspot or made from a contaminated waste stream such as paper or sewage sludge. Therefore, close monitoring of contaminant concentrations in BBFs is critical. The BBFs included in this article were in part selected because of their advanced commercialization stage. This study indicates that long- term use of commercial BBFs does not represent a risk for agricultural soils in terms of contamination with PCBs, PCDD/Fs, PAHs and target PFAS. However, it is still recommended to monitor for these and other hazardous substances in BBFs to ensure that the concentrations measured are in line with those of the BBFs reported in this study. To capture the effects of as many bioavailable contaminants as possible - including their potential interactions and pollutants that may not be detected by chemical analysis alone – it is also recommended to use ecotoxicological assays (Albert and Bloem, 2023). To reduce the risk that non-monitored contaminants potentially present in BBFs enter the food chain via BBF-amended plants, an avenue for further investigation could be the addition of sorbents, such as bio- char, to the organic waste presenting the highest risk (e.g., manure or sewage sludge) (Konczak and Oleszczuk, 2018; Stefaniuk and Oleszczuk, 2016). The high sorption of contaminants to the sorbents would signif- icantly reduce the availability of contaminants to plants (and other soil organisms). In addition, the plants would benefit from the reduced nutrient leaching in the amended soil (Knowles et al., 2011; Sarkhot et al., 2012). CRediT authorship contribution statement Nicolas Estoppey: Writing – original draft, Project administration, Methodology, Investigation, Conceptualization. Emma R. Knight: Writing – review & editing, Methodology, Investigation. Ian J. Allan: Writing – review & editing, Methodology, Investigation, Conceptuali- zation. Kuria Ndungu: Writing – review & editing, Methodology, Investigation. Gøril Aasen Slinde: Project administration, Methodol- ogy, Investigation, Conceptualization. Jan Thomas Rundberget: Methodology, Investigation, Conceptualization. Kari Ylivainio:Writing – review & editing, Project administration, Methodology, Investigation. Alicia Hernandez-Mora: Writing – review & editing, Methodology, Investigation. Erlend Sørmo:Writing – review & editing, Methodology, Conceptualization. Hans Peter H. Arp: Writing – review & editing, Methodology, Funding acquisition, Conceptualization. Gerard Corne- lissen: Writing – review & editing, Methodology, Funding acquisition, Conceptualization. Declaration of competing interest The authors declare the following financial interests/personal re- lationships which may be considered as potential competing interests: N. Estoppey et al. Science of the Total Environment 957 (2024) 177347 16 Nicolas Estoppey reports financial support was provided by EU Horizon 2020 research and innovation program under grant agreement 818,309 (LEX4BIO). Emma R. Knight reports financial support was provided by EU Horizon 2020 research and innovation program, through the project ZeroPM (101036756). Ian Allan reports financial support was provided by EU Horizon 2020 research and innovation program, through the project ZeroPM (101036756). Goril Aasen Slinde reports financial sup- port was provided by EU Horizon 2020 research and innovation pro- gram under grant agreement 818,309 (LEX4BIO). Kari Ylivainio reports financial support was provided by EU Horizon 2020 research and innovation program under grant agreement 818,309 (LEX4BIO). Alicia Hernandez-Mora reports financial support was provided by EU Horizon 2020 research and innovation program under grant agreement 818,309 (LEX4BIO). Hans Peter H. Arp reports financial support was provided by EU Horizon 2020 research and innovation program, through the project ZeroPM (101036756). Hans Peter H. Arp reports financial support was provided by Research Council of Norway, through the Miljøforsk project SLUDGEFFECT (NFR 302371). Gerard Cornelissen reports financial support was provided by EU Horizon 2020 research and innovation program under grant agreement 818,309 (LEX4BIO). If there are other authors, they declare that they have no known competing financial in- terests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgment This work received funding from the European Union's Horizon 2020 research and innovation program under grant agreement 818309 (LEX4BIO), with further support from ZeroPM (101036756, H.P.H.A, E. R.K, I.J.A) and the Research Council of Norway, through the Miljøforsk project SLUDGEFFECT (NFR 302371, H.P.H.A). 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