Activated carbon (AC) used for removal of organic m icropollutants in European wastewater treatment plants (WWTPs) is usually produced from n on-renewable resources that need to be transported over long distances. Utilising local re sidual biomass as a raw material may be advantageous in terms of sustainability. This study investigated the environmental and energy balances of using biowaste and biomass from landsca pe management for micropollutant removal at a commercial scale WWTP. Both residual b iomasses were processed using the integrated generation of solid fuel and biogas from biomass (IFBB) technique to obtain a press cake that was used as feedstock for AC production. The results showed a lower global warming potential (GWP) and cumulative energy demand in com parison to a fossil-based conventional AC. Differences in GWP between residual and fossil ACs were enhanced when the end-of-life incineration step was considered, and residual AC h ad a lower social risk associated with its production. Energy efficiency of AC production was substantially increased by utilising waste heat generated in the pyrolysis process of biochar and by using electricity generated in a combined heat and power plant using biogas from the methanation of IFBB press fluids. Converting residual biomass into activated carbon u sing IFBB and a state-of-the-art pyrolysis and activation unit along with energy recovery woul d aid WWTPs to become self-sufficient in terms of raw materials and improve the sustainabili ty of WWTPs. Keywords: Life Cycle Assessment (LCA), Biochar, Soc ial Risk Assessment, Global Warming Potential, Sust ainable Resource Management 1. Introduction Due to its adsorbent properties activated carbon (A C) can efficiently remove organic micropollutants ( OMPs) from wastewater (Çeçen and Akta ş, 2012; Wang and Wang, 2016). OMPs are found in sew age water treatment systems and one of the main sources are pharmaceuticals originating f rom households, livestock farming and hospital effl uents (Mansour et al., 2018). The detection of OMPs in surface waters indi cates that the present treatment at waste water tre atment plants (WWTPs) Journal XX (XXXX) XXXXXX Author et al 2 is insufficient to remove OMPs effectively (UBA, 20 15), hence legal regulations are already in force i n Switzerland for the implementation of a fourth treatment stage at WWTPs (FOEN, 2015). Similarly, European Union water fram ework directive (2000/60/EC) and the German federal water act (Wass erhaushaltsgesetz, WHG) put forward legal obligatio ns to maintain the quality of water (UBA, 2018). The usage of powdered activated carbon (PAC) at WWT Ps as an additional treatment step is an option to effectively remove micropollutants (Boehler et al., 2012). ACs used in European WWTPs are either manufactured from bitumi nous coal (Bayer et al., 2005), which is related to substantial environ mental and social impacts assosiated with the mater ial extraction, or from cocunut shells, which frequently originate from mon ocultures with adverse environmental impacts due to land use change and the use of chemical fertilizers and pestisides (Har temink, 2005). A large share of these materials are produced in Asian countries and need to be transported over long distances. Using local residual raw materials (e.g. green cut, wood chips, fruit stones) instead, which are plent ily available across Europe (Scarlat et al., 2019), may overcome the con straints mentioned above and provide a pathway to c onvert these residues into a valuable resource. The public enterprise Eig enbetrieb Umwelttechnik in Baden-Baden, Germany, op erates a municipal WWTP and biomass recycling center on the same site. Among other things, landscape management residues, hereinafter referred to as Baden-Baden Mixture (BM), and househ old biowaste (BW) consisting mainly of kitchen and garden waste are regularly delivered there (Table S1.1), which up to now have mainly been used for biogas and compost p roduction. BM and BW were pre-treated using the integrated generation of solid fuel and biogas from biomass (IFBB) techn ique, which mechanically separates biomasses into a solid and a liquid fraction (Hensgen et al., 2011; Richter et al., 2010; Wachendorf et al., 2009). The solid fraction is released from ash and harmful elements (e.g. N, S, Cl and K) and the liquid fraction produced can be co-digested to produce biogas that can be us ed as an energy source for the IFBB process. For ca rbonisation and steam activation a Pyreg A500 plant (Pyreg GmbH, Dörth, G ermany) was put into operation which can handle ca. 900 t dry matter (DM) per year. The aims of this study were to assess the enviromen tal impact and energy demand resulting (i) from the production of AC based on biomass mixture (BMC) and biowaste (BWC) a nd (ii) from the usage of these AC as an additional treatment step for micropollutant removal at a WWTP in comparison to c onventional activated carbon (CC). A social risk as sessment was conducted (iii) aiming at a deeper insight into the social effects during production of ACs. The produ ction of AC from residual materials is well studied, however, most studies on the related environmental impacts were based on la boratory scale data. Therefore, using life cycle assessment (LCA) of AC production and use at a commercial scale WWTP would increase the understanding of environmental and social impacts o f this technical approach. 2. Methods 2.1 Life cycle inventory (LCI), system boundaries a nd impact categories Samples of raw and IFBB pre-treated BM and BW mater ials used in this study were obtained from the WWTP Baden- Baden. Press cakes from IFBB process were converted into ACs using a Pyreka (Pyreg GmbH, Dörth, German y) laboratory scale pyrolysis and activation reactor. The NORIT S AE Super (Boston, Massachusetts, USA) was used as a reference CC for benchmarking BMC and BWC. For the LCI, material flo ws and process data were calculated using data obta ined from the WWTP. The datasets used for modelling the processes are elaborated in section S2, and compiling was do ne using OpenLCA 1.9.0. A “gate to grave” approach was used in this study c onsidering all processes involved from handling of the incoming raw material at the facility until the use of ACs produ ced. Incineration of ACs with heat recovery was con sidered as an end of life option, whereas disposal of ash generated during th e incineration was not considered. The maximum annu al production capacity of 900 t (dry weight) of BM and BW input material o f the pyrolysis and activation unit was used for ca lculation of mass and energy balances (Figure 1). 2.2 Activated carbon production and usage Raw BM and BW was pre-treated according to the IFBB technology approach, as described in details by (J oseph et al., 2018). Ensiled BM and fresh BW were mashed in warm water (40 °C) for 15 min. Subsequently, solid (pres s cake) and liquid fraction was mechanical separated by a screw press. While the liquid fraction was co-digested along wi th sewage sludge to generate biogas, the press cake was used for AC pro duction (Figure 1). Through the IFBB-process, miner al and ash content of BM and BW could be significantly reduced, which in turn improved the feedstock quality for AC producti on (Table 1). Energy and material balances for all the processes involve d in BM and BW production at the WWTP were calculat ed (Table S1.2-1.5). For AC production, BM and BW were pyrolyzed at 900 °C and physically activated by adding water vapour as oxidation agent during pyrolysis, using the Pyreka laboratory -scale reactor. This ensured a very good transferab illity of results to the full- scale Pyreg A500 reactor, which is used at the WWTP Baden-Baden (details in the Supplementary material ). Journal XX (XXXX) XXXXXX Author et al 3 Figure 1. System boundary diagram elaborating the BM and BW AC production at the WWTP. The amount of input and output material and energy required for the production of the annual capacity AC is indicated along with the diagram. Ash disposal was not considered in this study. At the full-scale PYREG A500 reactor, the thermal e nergy required for the pyrolysis process was genera ted by burning the resulting pyrolysis gases. In addition, surplus hea t energy from pyrolysis was recovered and used to s upply the processes in the pre-treatment steps (Figure S1.3). After its use fo r the removal of OMPs from wastewater, incineration of used ACs, further referred as spent ACs, was assumed to be the typica l end-of-life stage. The resulting heat was assumed to offset local heat requirements and considered in the LCI. Table 1. Characteristics of Baden-Baden mix (BM) and Biowas te (BW) based materials at different steps in the p roduction process of activated carbon. Unit BM BW Raw material DM content % FM 36.3 34.7 Ash % DM 10.8 24.2 Volatile solids % DM 89.3 75.8 C concentration % DM 46.3 36.5 DM flow into press cake % 75.8 59.3 DM flow into press fluid % 24.2 40.7 Press cake DM content % FM 55.3 52.7 Ash % DM 6.5 10.8 Volatile solids % DM 93.6 89.2 C concentration % DM 47.9 43.1 Lower heating value MJ DM -1 18.0 16.1 DM content after drying %FM 85.2 85.2 Activated carbon Ash % DM 70.7 62.5 Volatile solids % DM 29.3 37.5 C concentration % DM 27.9 37.9 Lower heating value MJ DM -1 16.7 16.7 Press fluid DM content % FM 1.8 2.9 Ash % DM 23.2 54.6 Volatile solids % DM 76.8 45.4 C concentration % DM 41.5 26.0 Methane yield L n CH 4 kg -1 VS 275.0 395.0 Journal XX (XXXX) XXXXXX Author et al 4 2.3 Functional unit OMPs pose an ecotoxicological risk in aquatic envir onments (Oldenkamp et al., 2019). This risk was cal culated according to the USEtox model (Rosenbaum et al., 2008) for se lected pollutants, which were measured in our study (Table S2.1). By using the investigated ACs for wastewater treatment , OMPs were removed and, thus, the ecotoxicological risk in receiving water bodies was reduced. The factors defined in th e USEtox model for OMPs were multiplied with their concentration in wastewater to calculate the mean ecotoxicity value for all OMPs. Based on these values, the functional unit for the LCA was fixed to the required quantity of AC to reduce the aquatic freshwater exotoxicity potential to a defin ed level in terms of comparative toxic unit, ecotoxicity (CTU e L-1) per 1000 m 3 of wastewater (Figure S2.2) (Rosenbaum et al., 200 8). Performance of AC in terms of OMP removal from wast ewater was measured in laboratory experiments. For this, concentrations between 1.75 to 7 mg L -1 of AC were added to the wastewater to evaluate the influence of the dosage on the reduction rate of OMPs. The ecotoxicity reduction a t a dosage of 5 mg L -1 of CC was used as a benchmark. The required amount of BMC and BC to achieve the same ecotoxicity reduc tion was calculated used for the LCA (Figure S2.2). Based on these data, multiplication factors for the LCA were determined by dividing the amounts of BM and BW AC through the 5 mg L -1 of CC (Figure S2.1). 2.4 Social risk identification The social risks involved in the production chain w ere identified based on the country specific risk a ssessments adapted from (Saling et al., 2020) considering all main for eground processes. The processes were classified in to risk levels according to the intensity of social risk in the country of p roduction, using 11 social impact categories from V erisk Maplecroft™ (Table S4.1). If the country of production was known, the country’s social risks for 11 chosen social impact categories were obtained from Verisk Maplecroft™. These categories were weig hted according to the severity of the social issues as estimated by INEF (2015) (Table S4.1). Depending on the data availabi lity, social risk scores on country level or mean v alues for the 10 largest producer countries, weighted by production volume w ere used for the LCA. 3. Results 3.1 Ecotoxicity Removal efficiency of OMPs from wastewater was high er for CC compared to BMC and BWC (Table S2.2 and S 2.3). A dosage of 5 mg L -1, CC was able to attain a calculated ecotoxicity of approximately 55.000 CTU e L-1. To achieve the same ecotoxicity level, a dosage of 9.1 mg L -1 of BWC was needed, which resulted in a 1.82-fold d emand of BWC compared to CC. In contrast, performance of BMC was significantly b etter (6.1 mg L -1) resulting in an equivalent of 1.22 times CC (Figu re S2.2). Based on an annual amount of wastewater of around 9 ,510,000 m 3 to be treated and a dosage of 5 mg L -1 of CC, this resulted in a demand of approximately 48 t of CC. Considerin g the factors of 1.22 and 1.82 for BMC and BWC resp ectively, 87 t and 58 t per year was needed to achieve the same ecotox icity level. 3.2 Global warming potential The balance of GWP for BMC and BWC production (5.3 and 4.5 kg CO 2 eq. FU -1 respectively) was significantly lower compared to CC (20 kg CO 2 eq. FU -1) (Figure 2). Considering an incineration of the sp ent ACs and the recovered energy obtained from it, GWP for BMC and BWC was reduced t o 2.7 and 1.3 kg CO 2 eq. FU -1 respectively, whereas this resulted in an increased GWP of 29 kg CO 2 eq. FU -1 for CC. Heat usage for press cake drying had the h ighest GWP for BMC and BWC (7.1 and 4.5 kg CO 2 eq. FU -1 respectively). For CC, heat usage for pyrolysis an d activation was the highest contributor (14 kg CO 2 eq. FU -1). For the production of the required 58 t of BMC a nd 87 t of BWC, 345 and 515 t DM of raw material ne eded to be processed, respectively. Hence, the GWP associat ed with electricity usage in the facility was highe r for BWC as for BMC. Similarly, GWP related to material handling of pres s fluid digestates depended on the quantity that wa s transported and applied on agricultural land, leading to higher GWP for BWC , as 6,232 t FM was handled compared to only 4,689 t FM for BMC. Mining and transportation of hard coal for further processing led to substantial material handling rel ated GWP for CC. However, GWP related to infrastructure usage was negligible when compared to the overall GWP for all 3 ACs. The total usage related GWP was highest for CC, closely followed by BWC and much lower for BMC. In contrast to CC, whose production only generated GWP emissions, the production of BMC and BWC also e nabled GWP savings. Heat recovered from the activation unit sa ved 3.4 kg CO 2 eq. FU -1 for BMC compared to 5.1 kg CO 2 eq. FU -1 for BWC. Electricity recovered from combustion of press flui d-derived methane was higher for BWC and saved 5.7 kg CO 2 eq. FU -1 compared to 1.7 kg CO 2 eq. FU -1 for BMC. Application of digestate on agricultural fields replaced the need for conventional mineral fertilisers, thereby saving 3.3 (for BMC) a nd 1.9 (for BWC) kg CO 2 eq. FU -1, respectively. Energy recovered from incineration of spent AC resulted in GWP savings of 5.8 and 7.1 kg CO 2 eq. FU -1 for BMC and BWC respectively. But incineration also caused additional emissions of 3. 2 (BMC) and 3.9 (BWC) kg CO 2 eq. FU -1 from heat usage for de-watering Journal XX (XXXX) XXXXXX Author et al 5 and drying of spent AC sludge. For CC there was a s aving of 4 kg CO 2 eq. FU -1 including incineration, however, the benefits of incineration were overshadowed by additional emi ssions of 12 kg CO 2 eq. FU -1 resulting from the release of CO 2 during incineration and 2 kg CO 2 eq. FU -1 for de-watering and drying. Overall, GWP savings we re highest for BWC, followed by BMC and CC. Figure 2. Global warming potential (GWP) generated during th e production and usage of activated carbon derived from Baden- Baden mix (BM), Biowaste (BW) and Conventional coal (CC). Each section of the horizontal bars indicate s the amount of GWP emitted or saved by the respective parameters a s indicated in the legend. (+I) denotes the inclusi on of the end of life incineration stage and related effects on the GWP w ere visualised accordingly, whereas (-I) indicates the case without considering incineration. The negative values indic ate GWP savings and the positive values indicate GW P emissions, therefore a lower value would mean the option has a lower GWP . 3.3 Cumulative energy demand (CED) Energy needed for operating the processes was estim ated in terms of CED and classified into renewable and non-renewable CED according to the source of energy. Energy used to produce CC was almost completely based on non-re newable sources, whereas biomass was the primary source of energy fo r producing BMC and BWC (5% and 11% share of non-re newable CED for BMC and BWC, respectively). Total use of non-re newable energy was substantially higher for CC (269 MJ FU -1) than for BMC and BWC (54 and 83 MJ FU -1 respectively) (Figure 3). Due to energy recovery and material replacement, th e energy balance for BMC and BWC further improved t o 24 and 34 MJ FU -1 respectively, and the highest non-renewable CED re sulted from fuel usage for transportation of digest ate (almost 60% of the total non-renewable CED both for BMC and BWC). While incineration as an end of life option improve d the energy balance of all three ACs, the combustion of the loaded BMC and BWC even generated excess energy. Contrarily, t he balance for renewable CED was higher for both BMC and BWC (291 and 222 MJ FU -1 respectively) compared to 150 MJ FU -1 for CC (Table S3.3). Heat usage for press cake drying repr esented almost 75 % of the renewable CED from both BMC and BWC and it increased to 81 % when the heat required for dry ing AC sludge was considered. Though overall CED wa s higher for BMW and BWC, almost 90 % of it was based on renewable b iomass feedstock. Journal XX (XXXX) XXXXXX Author et al 6 Figure 3. Cumulative energy demand (CED) of non-renewable so urces generated during the production and usage of activated carbon derived from Baden-Baden mix (BM), Biowaste (BW) and Conventional coal (CC). Each section of th e horizontal bars indicates the amount of non- renewable CED used or saved by the respective parameters as indicated in the legend. (+I) denotes the inclusion of the end of life incineration stage and related effects on the GWP were visualised acc ordingly, whereas (-I) indicates the case without considering incineration . The negative values indicate the usage of energy and the positive values indicate the savings therefore a higher positive va lue would mean the option saves more energy than it uses. 3.4 Additional environmental impacts The acidification potential associated with the pro duction and usage of bio-based AC was significantly higher than for CC (Table S3.1). This was mainly caused by the SO x and NO x emissions from the activation process, which accou nted for more than 60% of the total acidification potential. Whil e NH 3 emissions from the press fluid digestate applicati on on the field had a substantial contribution of 26 and 33 % for BMC and BWC, respectively. Almost 42 % of the acidificatio n potential of CC was related to electricity and fuel demands during coal mining. Further emissions associated with heat and electricity usage during drying, pyrolysis and activation amounted to 53 % f or CC. BWC had the highest eutrophication potential follow ed by BMC and CC. NH 3 emissions from digestate application accounted for the largest share of total emissions both for BMC and BWC (38 % and 50 % respectively). For CC, nitrogen (N)- based emissions for heat and electricity generation were the highest contributors. Including the incin eration of loaded ACs reduced the acidification and eutrophication potent ial of all ACs (Table S3.1 and S3.2). Environmental and energy balances were also examine d at a dosage of 20 mg AC L-1, which is the highest standard dosage used at WWTPs, and it was found that this increased GWP balances of BMC, BWC and CC by a factor of 3.2 , 3.0 and 3.55, respectively. Considering incineration resulted in a lower GWP for BMC and BWC (6 and -4 kg CO 2 FU -1) but CC had a higher GWP (107 kg CO 2 FU -1) (Figure S3.1). 3.5 Social life cycle assessment Producing bio-based ACs at the WWTP was found to re duce social risks substantially compared to the con ventionally produced AC, as 83 % of all processes at Baden-Bade n were categorised as “low risk” in contrast to onl y 14 % for the production of CC (Figure S4.2). This was mainly due to the fac t that processes involved in the extraction of bitu minous coal take place in countries which usually perform poorly for almost a ll social impact categories (Table S4.2). While pro cessing, production and usage of CC took place at locations with somewhat l ower social risks, the associated use of diesel and natural gas resulted in very high to high social risks, respectively. Journal XX (XXXX) XXXXXX Author et al 7 4. Discussion 4.1 Ecotoxicity Performance of BMC and BWC in terms of OMP removal from wastewater was significantly lower compared to CC (Table S2.2 and S2.3). This may be explained by the signif icantly lower specific surface area and adsorption capacity of bio-based ACs. Hence, higher dosages of BMC and BWC were need ed to achieve the same level of ecotoxicity reducti on. Removal of OMPs from effluent water mainly depends on the adso rption characteristics of ACs used, such as specifi c surface area (Mailler et al., 2016). These characteristics are highly aff ected by the properties of raw materials (Erdo ğan et al., 2017) and the activation process (Bergna et al., 2019). In particular the in fluence of the raw materials on adsorption of OMP w ere not clear, although the process conditions for BMC and BWC production w ere comparable (Mailler et al., 2016). However, it is remarkable that the removal of OMP through BMC was higher than for BWC though the specific surface area for BMC was lo wer. The high ash and low carbon content for BMC and BWC may have affected the adsorption capacity, as these propert ies have an influence on the pore structure (Anisuzzaman et al., 2015). T he ash content of the raw biomass was significantly reduced by the IFBB pre-treatment, which improved the quality of the AC s (Hensgen and Wachendorf, 2018; László et al., 199 7). Further improvements in biomass processing, pyrolysis and a ctivation can be expected to increase the adsorptio n capacity of activated carbons. This, in turn, would result in a further r eduction of GWP, CED and other environmental impact s through BMC and BWC production and usage. 4.2 Global warming potential The combination of raw biomass collection, processi ng into ACs and their subsequent utilization at the investigated WWTP is unique. To facilitate the comparison with other studies, all life cycle stages including production , usage and end of life were examined individually and compared with similar stu dies. Emissions during the production stage exhibited max imum contributions to GWP in the life cycle of all ACs. To allow a comparison with other studies at this stage, the GW P generated by producing 1 kg of the respective AC was calculated and used instead of the functional unit. The production of 1 kg BMC and BWC resulted in emissions of 0.43 and 0.20 kg CO 2 eq. respectively, which was significantly lower than th ose reported for AC from olive waste cake (11.1 kg CO 2 eq. kg -1) (Hjaila et al., 2013). GWP for ACs produced from conventional raw materials was 1.15 kg CO 2 eq. kg -1 for coconut shell (Kim et al., 2019), 8.4 to 11.1 kg CO 2 eq. kg -1 for bituminous coal (Bayer et al., 2005; Gabarrell et al., 2012) and 2. 45 kg CO 2 eq. kg -1 for regenerated AC (ROFA CARBON, 2018), which is notably higher compar ed to BMC and BWC. In our study, GWP for CC production from bituminous coal (3.41 kg CO 2 eq. kg -1) was significantly lower than the range reported b y (Bayer et al., 2005) and (Gabarrell et al., 2012), which can be explaine d by the fact, that CO 2 emissions associated with removal of fossil carbon was not considered in the production stage for this study, but rather included in the end of life stag e as CO 2 emitted during incineration. The low GWP in our study can be attributed to the e fficient pyrolysis and activation process, where th e combustion of pyrolysis gas provided more heat than needed. While pyrolysis was reported to account for 30-47 % of t he total GWP (Gu et al., 2018; Hjaila et al., 2013), it was only 7% in the present study. Further, electricity generation from the IFBB press fluid reduced the energy demand for production and, conse quently, the GWP. The delivery of residual raw materials to the facil ity prevented CO 2 emissions related to transportation and extraction . Conversion factors from raw material to AC were sli ghtly different for BM (11 %) and BW (7 %) which wa s due to different physico-chemical properties, such as dry matter and carbon content, as well as differing mass flows du ring mechanical separation. With approximately 24 % the conversion efficiency from press cake to AC was equal for BM a nd BW. However, the transfer of BM dry matter into the press cake ( 76 %) was considerably higher than for BW (59 %), w hich resulted in a lower contribution of AC production to the entire GWP for BM, as less raw material had to be processed. On t he other side, the higher mass flow of BW dry matter into the press fluid inc reased GWP savings from electricity recovery. Thoug h the type of raw material affected the overall GWP, the impact of th e production process was much greater. Since the co mplete production and usage of BMC and BWC took place at the WWTP, materi al transport did not contribute to the GWP. However , emissions related to the transport of the reference CC were a lso neglectable (0.003 kg CO 2 eq. FU -1). Nevertheless, the GWP associated with the production did not consider the contrastin g performances of ACs in terms of OMP removal. To t his end an appropriate functional unit (i.e. the quantity of AC to achieve a defined ecotoxicity level in wastewater) was int roduced which allowed determining the amounts of ACs required and, thus, a true comparison between conventional and bio-base d ACs. Energy recovery through incineration of the spent B MC and BWC reduced the overall GWP for these ACs, w hereas the combustion of CC resulted in an increased GWP, as t he resulting CO 2 emissions originated from fossil sources. Hence, u sing AC produced from local residual biomasses has an ad vantage during incineration which is a commonly use d end of life step for ACs. OMP removal using ozonation treatment usua lly has a lower GWP (18 kg CO 2 eq. 1000 m-3) (Kounina and Wencki, Journal XX (XXXX) XXXXXX Author et al 8 2015) compared to CC treatment (20 kg CO 2 eq. 1000 m -3) (Baresel et al., 2019), however, the figures for BMC and BWC of 4 and 5 kg CO 2 eq. 1000 m-3 respectively in our study were much lower in comparison to ozona tion treatment. 4.3 Cumulative energy demand (CED) The energy balance of AC production and use is deci sive for the profitability of the concept (Blumenst ein et al., 2012). The CED for the production of BMC and BWC (30 and 15 MJ kg -1respectively) was notably lower than for CC (53 MJ kg-1). Our data differ widely from results of other studies, w hich found CED values between 158 and 241 MJ kg -1 for woodchip and coal- based AC respectively (Bayer et al., 2005; Gu et al ., 2018). The lower CED for BMC and BWC production in our study was mainly caused by the efficient pyrolysis and activa tion processes as well as by recovered electricity from anaerobic digestion of press liquids. Methane yield from the press fluid was considerably higher for BWC (395 L n CH 4 kg-1 VS) compared to BMC (264 L n CH 4 kg-1 VS). However, methane content and mass flow of vol atile solids into the press fluid were highly influ enced by material properties, such as dry matter content and IFBB pro cess parameters (Hensgen et al., 2011; Richter et a l., 2009). Even if the overall CED was lower for production of BMC and BWC , the share of non-renewable CED was higher than fo r CC. As these energy sources are based on depletable resources (A rvidsson and Svanström, 2016), additional research is needed to improve the sustainability of alternative AC production. Am ong others, this may include the substitution of di esel for transportation and processing through biofuels. 4.4 Additional environmental impacts Emissions of NO x, SO x and NH 3 damages land ecosystems by affecting plant health a nd change in soil pH (Rosenbaum et al., 2018). Though the emission of NO x and SO x from the activation process were the largest contr ibutors towards acidification, they were still below the German limit of 350 mg m-3 (Bundesministerium für Umwelt, Naturschutz und Reak torsicherheit, 2002). Mass flow of N during mechanical separation into press cake and fluid caused N emissions in the following stages, i.e. NO x emissions during pyrolysis of the press cake and a ctivation of the resulting biochar (Zhan et al., 20 18) as well as NO x and NH 3 emissions from the field application of digestates originating from the anaerobic digestion of press fluids (Eggleston, 2006) . Recent research showed that the flow of elements du ring mechanical separation was influenced by the ra w material properties, process parameters (Richter et al., 201 1) and scale of the conversion process (Joseph et a l., 2018). The modification of these factors would affect the flow of elements and, thus, reduce the acidification and eutrophicat ion potential. Further reduction is possible through the use of precision application techniques for digestates (Nicholson et al., 2018) and advanced flue gas cleaning systems (Turconi et al., 2013). 4.5 Social risk assessment For estimating the social risks associated with AC production, available average values on country lev el were used, as data from producers were not existent or not representat ive (Saling et al., 2020). Our results indicate tha t risks could be reduced when all processes take place in countries with a l ow risk level or when materials from countries with a lower associated social risk were used. To this end, materials need to be l abelled appropriately. The local production and use of AC from residual bi omass in industrialized countries like Germany, wou ld affect economic activities in the current production countries. Pos sible losses of employment could, however, be offse t by investments in alternative employment opportunities for people wor king in fossil-based sectors (Sparkes, 2008). Due t o a lack of revenue, this may lead to a further deterioration of the social s ituation in the developing countries. Regardless of this, socially responsible investments would help to improve the social condit ions and certification of socially compatible produ ction processes may contribute to a higher demand of products with lowe r social impact (Goedkoop et al., 2018). Neverthele ss, the conducted social risk assessment for AC production contained global and, therefore, inaccurate data. Hence, only social hotspots in the production process on country level could be identi fied (Saling et al., 2020). However, our findings m ay serve as a starting point for further investigations on the exact sourc es of social risk at a sector and company level. 5. Conclusion Using ACs produced from local residual biomasses fo r the removal of OMPs at a large-scale WWTP was fou nd to have lower environmental impacts in terms of GWP and CED when compared with conventional ACs. The novel con cept was found to be flexible in terms of raw material, though bot h BM and BW used in this study had distinctive prop erties. The IFBB step significantly reduced the ash content of feedstocks for AC production and the generated press fluids f urther improved energy Journal XX (XXXX) XXXXXX Author et al 9 generation and material replacement. Thus, IFBB was a crucial step in utilising residual biomass for A C production. On-site production and use of the AC enabled a significant reduction of monetary and environmental costs compa red to systems based on conventional ACs. The lower OMP adsorption perfo rmance of BMC and BWC could be overcome by using hi gher dosages of AC without hazarding GWP and CED advantages. Per formance of BMC and BWC is expected to be increased by further process optimization, which would render the novel system attractive for WWTPs that currently use conv entionally produced AC. Acknowledgement The authors would like to acknowledge the European Union for financing the Re-Direct project (NWE 294) through the INTERREG European regional development fund. We would also l ike to thank O. Hermann from Eigenbetrieb Umwelttec hnik Baden-Baden for providing the material and information needed for t he study. We are grateful to the support of K. Sten chly, T. Astor, J. Wijesingha, T. Fricke, R. Graß and A. Gerke.