Adsorption Performance Assessment of Agro-Waste-Based Biochar for the Removal of Emerging Pollutants from Municipal WWTP Effluent

Academic Editor: Marek Kosmulski Received: 12 November 2025 Revised: 10 December 2025 Accepted: 11 December 2025 Published: 17 December 2025 Citation: Luki´c, D.; Vasi´c, V.; Živanˇcev, J.; Anti´c, I.; Pani´c, S.; Petronijevi´c, M.; Ðuriši´c-Mladenovi´c, N. Adsorption Performance Assessment of Agro-Waste-Based Biochar for the Removal of Emerging Pollutants from Municipal WWTP Effluent. Molecules 2025,30, 4803. https://doi.org/10.3390/ molecules30244803 Copyright: © 2025 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https://creativecommons.org/ licenses/by/4.0/). Article Adsorption Performance Assessment of Agro-Waste-Based Biochar for the Removal of Emerging Pollutants from Municipal WWTP Effluent Dragana Luki´c * , Vesna Vasi´c , Jelena Živanˇcev, Igor Anti´c , Sanja Pani´c , Mirjana Petronijevi´c and Nataša Ðuriši´c-Mladenovi´c University of Novi Sad, Faculty of Technology Novi Sad, Blvd. Cara Lazara 1, 21000 Novi Sad, Serbia; [email protected] (V.V.); [email protected] (J.Ž.); [email protected] (I.A.); [email protected] (S.P.); [email protected] (M.P.); [email protected] (N.Ð.-M.) *Correspondence: [email protected] Abstract Wastewater treatment plants (WWTPs) have been identified as the major sources of contaminants of emerging concern (CECs) in water bodies, as they are not designed to remove organic micropollutants efficiently. Consequently, many technologies have been explored for WWTP upgrading, including activated carbon adsorption. However, the high production cost and environmental challenges associated with activated carbon production limit its application in industrial settings. Therefore, a wide range of alternative materials has been investigated as potential replacements. In this study, biochar produced from waste raspberry biomass was evaluated as an adsorbent for the removal of pharmaceuticals and pesticides quantified in the secondary effluent of municipal WWTP. The results showed that the biochar efficiently removed almost all detected compounds, except for three compounds (clarithromycin, propranolol, and linuron). The wastewater pH (6–8) did not significantly affect removal efficiency significantly, and kinetic tests demonstrated rapid adsorption. The potential for biochar reuse was confirmed through three consecutive batch adsorption cycles. A comparative study between biochar and powdered activated carbon (PAC) revealed some differences in efficiency, primarily attributed to the larger surface area of PAC. π - π interactions, hydrogen bonding, and pore-filling were proposed as possible adsorption mechanisms based on the adsorption efficiency and biochar characterization. Keywords: biochar; agricultural waste; raspberry canes; adsorption; pharmaceuticals; pesticides; contaminants of emerging concern; municipal wastewater 1. Introduction Pharmaceutically active compounds (PhACs), personal care products, polar pesticides, polyand perfluoroalkyl substances (PFAS), and industrial chemicals are among the many examples of the wide range of mostly unregulated substances that are commonly found in various environmental compartments, which are referred to as contaminants of emerging concern, CECs [ 1 , 2 ]. The presence of CECs in environmental resources remains a significant challenge due to the excessive production and consumption of these chemicals. Many substances, recognized as CECs, are present in the environment, with concentrations ranging from ng/L to µ g/L; in some cases, emissions can be as high as 100 mg/L [ 3 ]. Emissions of CECs into environmental resources include various industrial capacities, households, wastewater treatment plants (WWTPs), hospital effluents, direct Molecules 2025,30, 4803 https://doi.org/10.3390/molecules30244803 Molecules 2025,30, 4803 2 of 28 disposal of unused or expired drugs, manufacturing, landfill leachates, livestock activities, aquaculture, and soil fertilization with sewage sludge and/or livestock waste [ 4 – 9 ]. Among these sources, WWTP effluent discharge is considered the main route for CEC distribution into environmental resources [ 7 ]. WWTPs have been identified as the main sources of CECs in the environment because they are not designed to efficiently remove organic micropollutants [ 3 , 8 – 10 ]. Accordingly, many CECs survive the conventional wastewater treatment processes [ 11 ]. According to Serbian regulations [ 12 ], urban WWTP effluents discharged into surface waters are generally characterized by suspended solids, dissolved organic matter, and nutrients, particularly in recipients sensitive to eutrophication. Similar requirements were applied in the European Union until recently. In 2025, the European Parliament adopted a revised version of the Urban Wastewater Treatment Directive [ 13 ], which also proposes the removal of CECs by introducing a fourth step (quaternary treatment) in urban wastewater treatment. WWTPs are expected to improve effluent quality, reducing the micropollutant contamination of surface waters and/or enabling water reuse. Various chemical, physical, and biological treatment techniques have been evaluated for their efficacy in removing these pollutants, but the most commonly investigated are advanced oxidation processes, membrane filtration, and adsorption [14,15]. Since the adoption of the Water Protection Act in 2016 in Switzerland and certain parts of Germany, ozonation and adsorption onto have been employed as hybrid or standalone processes to remove micropollutants in wastewater treatment plants [ 15 – 17 ]. The literature provides data on the application of powdered activated carbon (PAC) in biological effluent treatment, but also its direct addition to aeration basins with activated sludge, where the biological degradation of organic matter and adsorption simultaneously occur. The main advantage of this method is that the addition of the adsorbent can be easily implemented into existing aerobic-activated sludge processes, making it more cost-effective compared to a separate treatment stage involving PAC. However, comparative studies of PAC dosing in secondary treatment effluent versus dosing in aeration basins have shown that the latter approach is less efficient and requires almost double the amount of PAC [ 18 ]. These approaches have reduced the concentration of a wide range of micropollutants by more than 80% [16]. Adsorption on AC efficiently removes various pollutants from water, including CECs. Still, it is expensive due to high AC production costs and possible nonrenewable resource origins [ 9 , 14 ]. However, this technique is technically feasible and simple [ 16 ]. Therefore, low-cost materials, including biochar, have been investigated as substitutes for AC. Biochar is a carbon-rich material produced from biomass at high temperatures and in a poor oxygen atmosphere in the process known as pyrolysis [ 19 ]. Agricultural, forestry, and industrial waste, as well as municipal solid waste, manure, and sewage sludge, can all be used for its production. During pyrolysis, thermal decomposition occurs within a temperature range of 250–900 ◦ C. Depending on the temperature and residence time, pyrolysis can be categorized as slow, fast, or flash pyrolysis [ 19 – 22 ]. The physicochemical properties of biochar depend on the type of feedstock, the type of pyrolysis, and operating conditions (temperature, heating rate, and residence time) [ 21 , 22 ]. Generally, biochar is characterized by a porous structure, a relatively large specific surface area, and numerous functional groups. Biochars obtained by pyrolysis typically exhibit lower porosity and specific surface area; therefore, they are often subjected to further activation processes to enhance these properties [ 20 ]. One way to improve the properties of biochar is the physical activation in a CO 2 atmosphere at high temperatures. It was found that CO 2 -activated biochar had up to five times higher specific surface area and pore volume compared to before activation, where the properties were highly dependent on the temperature of the activation process [ 23 ]. These properties make biochar similar to activated carbon, making it Molecules 2025,30, 4803 3 of 28 a suitable adsorbent for water and wastewater treatment. The main advantage of biochar is its cost-effective production [ 19 ]. Consequently, the potential application of biochar for the removal of various pollutants from water, including CECs, has been widely investigated. Biochar derived from raspberry canes is a promising but under-explored option for sustainable wastewater treatment. Raspberry is widely cultivated in Western Serbia. Consequently, raspberry canes are generated as waste in large quantities after harvest and are usually burned in fields [ 24 ]. According to the data from the Statistical Office of the Republic of Serbia, the areas under raspberry cultivation in 2021, 2022, and 2023 were approximately 20,000 hectares, with a total production of 110,000 tons and yields ranging between 5.2 and 5.9 tons per hectare. Some estimates of the waste biomass generated during raspberry pruning indicate a significant annual output of 1.5–2 tons per hectare [ 25 ], which amounts to about 30,000 tons. Under the concept of a circular economy, the possible application of biochar from raspberry waste addresses waste management issues and transforms agricultural waste into valuable resources. According to the existing literature, the production, characterization, and application of biochar from this type of biomass have not been reported. The main aim of this study was to investigate the properties of biochar obtained from this kind of agricultural waste biomass and its possible application as an adsorbent for CEC removal from municipal wastewater. To address all these challenges, this study examines the effectiveness of biochar produced from waste raspberry canes for removing contaminants of emerging concern (CECs) from real municipal wastewater. The main objectives were to evaluate its adsorption efficiency toward a broad set of pharmaceuticals and pesticides, compare its performance with commercial powdered activated carbon (PAC), assess the influence of pH and initial pollutant concentration, and determine its reusability over several cycles. Raspberry canes were selected due to their availability as low-value agricultural residue and their suitability for conversion into porous carbon materials. The adsorption capacity and efficiency of the process depend on many factors, such as the physicochemical properties of the adsorbent and adsorbate, the concentration of the adsorbate, other substances present in the water, and operational conditions. In the case of highly complex and variable matrices, such as wastewater, the efficiency of an adsorbent can only be determined through tests on real wastewater samples. Therefore, all the experiments were conducted on real municipal WWTP effluent. By analyzing biochar application on real wastewater samples, this study provides insights into the simultaneous and complex influence of multiple factors on the adsorption process and the removal rate of CECs by raspberry cane-based biochar. 2. Results and Discussion To identify and quantify CECs in municipal wastewater, sampling was conducted on both the influent and the effluent. It is well known that the quality of wastewater, particularly the influent of the WWTP, varies over time; thus, composite samples provide more reliable data regarding the types of CECs that can be expected in the wastewater. The quality of the effluent also fluctuates over time. However, for further effluent treatment using biochar after the biological process, grab samples are more important as they provide insights into concentrations of individual compounds and their (potentially extreme) variations. Table S1 (Supplementary Materials) summarizes the target compounds investigated in water samples, while Table 1lists the compounds that were detected, along with their respective concentrations. Molecules 2025,30, 4803 4 of 28 Table 1. CECs found in municipal WWTP effluent. CEC * (ng/L) Campaign Influent I II III IV V Effluent Effluent Effluent Effluent Effluent Atenolol 0.50 19.4 458 15.3 39.2 54.1 Bezafibrate <MLOQ <MLOQ <MLOQ <MLOQ 275 <MLOQ Carbamazepine 161 319 177 726 507 413 Clarithromycin 82.3 75.3 218 115 60.9 46.9 Famotidine 6.23 27.5 <MLOQ <MLOQ 11.8 18.1 Furosemide 294 <MLOQ 955 <MLOQ <MLOQ <MLOQ HCTZ (hydrochlorthiazide) 574 816 1127 992 803 1459 Losartan 389 <MLOQ 382 <MLOQ <MLOQ <MLOQ Propranolol <MLOQ 78.1 1767 93.4 61.0 83.5 Ranitidine <MLOQ <MLOQ <MLOQ <MLOQ 54.6 <MLOQ Salbutamol <MLOQ 37.5 <MLOQ 36.1 49.7 47.5 Sotalol 178 310 321 265 589 481 Acetamiprid <MLOQ 35.6 <MLOQ 35.3 54.1 188 Diazinon <MLOQ 32.9 <MLOQ 45.9 54.6 66.5 Linuron <MLOQ 68.7 <MLOQ 58.8 51.5 55.2 Methamidophos <MLOQ 119 <MLOQ 146 <MLOQ 91.7 Tebuconazole <MLOQ 49.2 <MLOQ 58.5 98.2 <MLOQ MLOQ—method limit of quantification. * Standard pharmaceutical nomenclature was used for PhACs. According to the literature data, 70% of PhACs in wastewater originate from households, 20% from livestock farming, 5% from hospital wastewater, and the remaining 5% from other sources, with certain variations observed during different seasons [ 26 ] and depending on the population socioeconomic composition and the wide range of consumer products used daily [9]. As shown in Tables 1and S1, some compounds were not detected in any of the analyzed effluent samples, and some appear sporadically (famotidine, furosemide, losartan, salbutamol, acetamiprid, diazinon, linuron, methamidophos, tebuconazole). Certain compounds (atenolol, carbamazepine, clarithromycin, sotalol), however, were present in all five effluent samples, either at similar or varying concentrations, indicating their nearly constant or variable emission. Literature data suggest that the presence of most compounds listed in Table 1is characteristic of municipal wastewater [9,26]. 2.1. Adsorption of CECs by Biochar The adsorption properties of activated biochar (BC) for removing CECs from WWTP effluent were investigated through batch adsorption to assess the effects of wastewater pH, initial CEC concentration, biochar dose, and time on the adsorption process. Additionally, the feasibility of using biochar through three cycles of multi-step batch adsorption was evaluated, and its efficiency was compared with that of commercial activated carbon. Preliminary tests were conducted to identify the most efficient biochar (either pristine or modified) for additional treatment of WWTP effluent. The results demonstrated that activated biochar is more efficient (Table S2). Thus, all further experiments were focused on activated biochar as an adsorbent. The Regulation on Limit Values of Emissions of Pollutants in Water and Deadlines for Their Achievement [ 12 ] allows a pH range of 6 to 9 for wastewater discharge into surface waters. Effluent samples from WWTP during the sampling period were within the permitted range (8.0–8.4). Therefore, the effect of other pH values within the range on the adsorption efficiency of CECs by biochar was also examined. The results are presented in Table 2. Molecules 2025,30, 4803 5 of 28 Table 2. Removal efficiencies of CECs by biochar at different effluent pH values (contact time 24 h, dose 1 g/L). CEC Campaign I Removal (%) * C0(ng/L) pH 6 pH 7 pH 8 Original pH Atenolol 19.4 100 100 100 100 Carbamazepine 319 100 100 100 100 Clarithromycin 75.3 33.9 30.3 36.2 36.5 Famotidine 27.5 100 100 100 100 HCTZ 816 100 100 100 100 Propranolol 78.1 35.1 37.3 47 38.8 Salbutamol 37.5 100 100 100 100 Sotalol 310 100 100 100 100 Acetamiprid 35.6 100 100 100 100 Diazinon 32.9 100 100 100 100 Linuron 68.7 37.1 34.2 36.2 35.8 Methamidophos 119 100 100 100 100 Tebuconazole 49.2 100 100 100 100 * An efficiency of 100% was achieved when the residual concentration of the compound of interest after adsorption was below the quantification limit of the applied method (MLOQ) (Table S1). The obtained activated biochar shows high efficiency for the simultaneous removal of PhACs and pesticides from effluent at different pH values. Within the examined pH range, the minor differences observed were minor and, from a wastewater treatment perspective, not operationally significant. Real effluents vary over time in many parameters (e.g., DOC, ionic strength, competing solutes), and pH is only one of several factors that can influence adsorption. Therefore, small pH-related variations are unlikely to affect overall treatment performance. These results suggest the potential for the direct application of the investigated biochar for the treatment of the WWTP effluent, regardless of pH variations. Literature provides various data on the impact of solution pH on the adsorption of PhACs and pesticides. For example, Fu et al. [ 27 ] reported enhanced adsorption capacity of atenolol on corncob-based biochars with increased solution pH from 2 to 9. As the pH increased above 9, the adsorption capacity decreased, indicating that the optimum pH for atenolol removal is 8 and 9, depending on the biochar modification. Naghdi et al. [ 28 ] reported an enhancement of carbamazepine adsorption on pine wood nanobiochar as pH increased from 3 to 6. However, authors showed that a further increase in pH from 6 to 8 did not affect the adsorption process, which is consistent with the results presented in Table 2. Similar results were reported by Chu et al. [ 29 ]. Namely, the adsorption of carbamazepine onto non-modified and bleached biochar from pine sawdust did not vary in the pH range of 4–8. Within the solution pH range 3–9, carbonized coconut shell-based biochars modified with H 3 PO 4 and NaOH showed the highest removal of diazinon at pH 7 (over 85%). At the same time, only pyrolyzed material was more efficient in diazinon removal at pH 3, reaching 92% [ 30 ]. Moussavi et al. [ 31 ] also reported the best diazinon removal at neutral pH on NH 4 Cl-induced activated carbon. The adsorption capacity of tebuconazole on waste-activated sludge-based biochar dropped slightly as the solution pH increased from 3 to 10, from 12.4 to 11.4 mg/g [ 32 ]. On the other hand, pH did not affect the adsorption of linuron across the same range [ 32 ]. Similarly, Moraes et al. [ 33 ] reported a lower efficiency of linuron adsorption on commercially available activated carbon, and ZnCl 2 and HNO 3 spent coffee grounds-based activated carbons (~45% for all the adsorbents) over a wide pH range. However, the reported data were obtained in synthetic solutions, so different results can be expected in much more complex matrices such as real wastewater. The adsorption capacity of the biochar and the CEC removal depend on the concentration and type of pollutants and the coexistence of various contaminants [ 34 ], which Molecules 2025,30, 4803 6 of 28 is a common case in wastewater. Dissolved organic matter can negatively affect CEC adsorption due to competition for active sites and the potential to block microand mesopores, given their molecular weights range from several hundred daltons (Da) to over 100 kDa. Compounds with lower molecular weights may have a greater impact on CEC adsorption [ 35 ]. The content of organic matter varies in both wastewater and treated water, depending on the type of treatment applied. According to the average annual values (Section 3.3.1 in Section 3Materials and Methods), the content of organic matter (expressed as COD and BOD 5 ) in the effluent from the biological treatment at the WWTP falls within a very narrow range, suggesting that significant deviations in the efficiency of CEC removal by biochar shown in Table 3are not expected, regardless of change in CEC concentrations in a reasonable range. Table 3. Removal efficiencies of CECs in WWTP effluent by biochar (contact time 24 h, dose 1 g/L). CEC Campaign II C0, (ng/L) Removal (%) * Campaign III C0, (ng/L) Removal (%) * Atenolol 458 100 15.3 100 Carbamazepine 177 100 726 100 Clarithromycin 218 43.7 115 48.9 Furosemide 955 100 <MLOQ / HCTZ 1127 96.6 992 100 Losartan 382 100 <MLOQ / Propranolol 1767 42.3 93.4 52.7 Salbutamol <MLOQ / 36.1 100 Sotalol 321 97.7 265 100 Acetamiprid <MLOQ / 35.3 100 Diazinon <MLOQ / 45.9 100 Linuron <MLOQ / 58.8 26.9 * An efficiency of 100% was achieved when the residual concentration of the compound of interest after adsorption was below the quantification limit of the applied method (MLOQ) (Table S1). Numerous emerging contaminants in wastewater differ significantly in their physicochemical properties (including molecular structure, hydrophobicity, aromaticity, and polarity). As a result, their sorption onto biochar surfaces may arise from synergistic effects or competition among multiple compounds [ 34 ]. The quality of municipal wastewater, in terms of the type and concentration of these compounds, varies depending on the size of the settlement, lifestyle, population habits, and the time of year. To examine the efficiency and capacity of biochar at higher concentrations of CECs, the effluent was further spiked with the quantified compounds (Table 4). Furthermore, other selected compounds that had not been previously detected in the effluent (known to be quantifiable in WWTP effluents according to the previous studies) were spiked into the effluent. Spiked CEC concentrations of the effluent were chosen based on literature data, i.e., concentrations characteristic of wastewater in the Western Balkans region during the period from 2008 to 2023 [ 9 ]. The results are presented in Table 4. According to the data presented in Tables 3and 4, biochar is highly effective in removing CECs across a wide range of environmentally relevant concentrations. Only bezafibrate and diltiazem out of all compounds were removed with an efficiency of less than 91% at higher concentrations. This is in agreement with the reported results of Moraes et al. [ 33 ]. Namely, the lower efficiency of linuron adsorption on commercially available activated carbon and spent coffee grounds-based activated carbons (~45% for all the adsorbents) was noted. The removal efficiency of raspberry biochar did not decrease even when the effluent was spiked with additional compounds due to a sufficiently large Molecules 2025,30, 4803 7 of 28 surface area and active sites. Moreover, the removal of propranolol is much higher at higher concentrations. Table 4. Removal efficiencies of CECs in original and spiked effluent (contact time 24 h, dose 1 g/L). CEC Campaign IV Spiked Effluent C0(ng/L) Concentration Range (ng/L) Removal * (%) Atenolol 39.2 1605–7589 99.5–100 Bezafibrate 275 850–3533 72.6–91.5 Famotidine 11.8 927–4282 100 HCTZ 803 1544–5705 100 Carbamazepine 507 1467–4566 100 Clarithromycin 60.9 927–5787 96.2–100 Propranolol 61 893–5841 91.6–100 Ranitidine 54.6 947–3998 93.8–100 Salbutamol 49.7 1358–6928 97.9–100 Sotalol 589 2052–7552 100 Acetamiprid 54.1 745–3557 93–100 Diazinon 54.6 1499–4638 94.7–100 Methamidophos <MLOQ 158–208 100 Linuron 51.5 796–3094 100 Tebuconazol 98.2 1751–4386 100 Acetaminophen <MLOQ 1333–4831 100 Diltiazem <MLOQ 172–761 80–97 Dimethoate <MLOQ 542–2165 100 Ethopropohos <MLOQ 1606–5518 100 Furosemide <MLOQ 1294–3996 100 Imazalil <MLOQ 389–1080 100 Carbofuran <MLOQ 205–784 100 Losartan <MLOQ 579–2402 93.3–100 Methidation <MLOQ 506–2305 100 Omethoate <MLOQ 114–306 100 Propioconazole <MLOQ 2334–5831 100 * An efficiency of 100% was achieved when the residual concentration of the compound of interest after adsorption was below the quantification limit of the applied method (MLOQ) (Table S1). An adsorption study of carbamazepine on spent brewery grains-based biochars showed that the adsorption capacity was significantly lower in a synthetic solution compared to real wastewater. It was assumed that some inorganic and organic matter present in wastewater can affect adsorption by competition for active sites or block access to the pores on the adsorbents [ 36 ]. Diniz et al. [ 37 ] also reported a decrease in the adsorption capacity of different CECs in wastewater compared to synthetic solutions up to 50% despite higher doses of commercially available activated carbon. These results imply that even if the adsorption is affected by a complex matrix, removal of various CECs is possible by Molecules 2025,30, 4803 8 of 28 adsorption. However, tests in real wastewater are only valid for predicting the performance of adsorbents during their actual application. The dose of biochar is a crucial parameter influencing the quality of the treated water and the overall cost of the adsorption process. The optimal dose depends on the type of adsorbent, wastewater composition, and the required pollutant removal. Usually, it is determined by laboratory testing or based on previous experience, and it is confirmed during the initial implementation phase at the industrial scale. At high concentration levels of pollutants, increasing the dose of biochar can significantly enhance the removal due to the larger overall specific surface area and the greater number of active sites. Nevertheless, above a certain dose, biochar may form agglomerates, resulting in a reduced total surface area and thus decreased efficiency. This phenomenon is more likely to occur with biochars produced from materials with a high lignin content [ 34 ]. Moreover, the higher the lignin content, the higher the yield and fixed carbon in biochar that may assist the adsorption of organic contaminants [ 38 ]. The lignin content in lignocellulosic materials is typically up to 35% [ 39 ]. In raspberry canes, it is relatively high, exceeding 30%, depending on the particle size [ 24 ], but not as high as in coconut shells (around 50%) [ 40 ]. The efficiency of biochar at different doses was investigated, and the results are presented in Figure 1. 0 20 40 60 80 100 Removal (%) 1 g/l 0.5 g/l 0.2 g/l Figure 1. Removal efficiencies of CECs depending on the applied biochar dose. Results showed that the lowest investigated dose (0.2 g/L) may be sufficient for the efficient removal of CECs from WWTP effluent and that the removal efficiency does not decrease with a decrease in biochar dose from 1 to 0.2 g/L. The doses applied in this study are somewhat higher than typical doses of commercial activated carbon used in water and wastewater treatment. However, they are in the range of biochar doses reported in the literature for organic pollutants removal [ 41 – 44 ]. Naghdi et al. [ 28 ] reported an increased removal efficiency of carbamazepine from 53% to 87% with the increase in pinewood-based biochar dose of 0.2–1 g/L. Higher doses did not show any improvement. Biochar produced from pineapple leaf non-fibrous material showed an acetamiprid removal efficiency of 84.3% at 5 g of adsorbent per liter of solution and negligible removal increase at 7 g/L, suggesting biochar saturation [ 45 ]. Yang et al. [ 32 ] reported that the efficient removal of tebuconazole (87.8%) and linuron (95%) from a single-component solution was achieved with 1.25 g/L of waste-activated sludge-based biochar modified with FeCl3. Molecules 2025,30, 4803 9 of 28 Adsorption is a time-dependent process, and for performance optimization, it is crucial to determine the time required for the process to reach equilibrium. For the practical application of an adsorbent, its adsorption capacity is equally important as the reasonable period necessary for the process to end. Adsorption kinetics of PhACs and pesticides are shown in Figure 2. 0 5 10 15 20 25 30 0 200 400 600 800 1000 q (ng/g) t (min) Atenolol Carbamazepine Clarithromycin HCTZ Propranolol Salbutamol Sotalol (a) 0 5 10 15 20 25 30 0 40 80 120 160 q (ng/g) t (min) Acetamiprid Diazinon Linuron Methamidophos Tebuconazole (b) Figure 2. Adsorption kinetics of: (a) PhACs; (b) Pesticides on biochar from WWTP effluent. Adsorption of all CECs found in WWTP effluent onto biochar was very rapid. All detected compounds reached equilibrium within the first 5 min of contact. Literature reports much longer contact times between biochar and wastewater. Coimbra et al. [ 46 ] showed that the adsorption of selected pharmaceuticals onto pyrolyzed pulp mill sludge reached equilibrium within 200 min from a solution and secondary municipal WWTP effluent. Silva et al. [ 47 ] conducted an adsorption study of selected pharmaceuticals on laboratory-produced activated carbon and commercial activated carbon from a solution and biological treatment effluent of municipal wastewater. Adsorption kinetics experiments showed that it took 60–240 min for all the compounds to reach equilibrium on both adsorbents in the mixture and wastewater. Fast adsorption greatly improves the economic efficiency of the process, allowing the use of smaller reactors and lower doses of adsorbent. Short contact times, ranging from 18 to 30 min [ 48 , 49 ] to 0.7 to 3 h [ 50 ], are enough for PAC applications in wastewater treatment. When PAC is used for wastewater treatment, in some cases, the contact time or the concentration of pollutants could be insufficient to reach the adsorption capacity, resulting in an unsaturated adsorbent. For more effective exploitation, batch adsorption can be conducted as a multistage process with a countercurrent flow of untreated water and PAC. Namely, after the initial use, the spent PAC is separated from the treated water and reintroduced to the new amount of fresh, untreated water. This can significantly reduce the overall demand for PAC for wastewater treatment. As previously mentioned, a relatively short contact time is typically sufficient when PAC is used in a batch mode. Given the very short contact time required to complete the adsorption of detected CECs from the effluent onto biochar (Figure 2), a contact time of 45 min was chosen for the multistage adsorption tests. Three cycles were conducted, and the results are presented in Figure 3. Almost all the compounds found in the effluent, under the applied conditions, are removed with maximum efficiency across all three stages of the multi-step adsorption, which is attributed to the sufficiently large total adsorption capacity of the biochar (Table 4). However, the removal efficiencies of clarithromycin, propranolol, and linuron were lower but within the characteristic range for this adsorbent/adsorbate system. To provide stronger evidence of adsorption stability, the adsorption capacities (q, ng/g) were Molecules 2025,30, 4803 16 of 28 with the biochar FTIR spectrum. The same peaks are visible around 3400 cm −1 , in the area of 2800–3000 cm −1 , around 1630 cm −1 , and 1640 cm −1 . The peak visible at 1384 cm −1 is assigned to C=C groups of the aromatic rings [ 60 ], and 1110 cm −1 to –OH stretching in COH group [61]. 4000 3500 3000 2500 2000 1500 1000 500 40 50 60 70 80 90 100 Transmittance (%) Wavenumber (cm -1 ) Biochar PAC Figure 7. FTIR absorption spectra of biochar and PAC. To know surface chemistry, the pH of dispersion of biochar was determined since it is an important characteristic affecting adsorption and depends on the solution pH and surface functional groups. Namely, higher pyrolysis temperatures produce alkaline biochar due to the loss of acidic groups (thus, a larger amount of total basic groups) and higher content of alkali and alkaline earth metals in ash [ 36 , 62 ]. The raspberry cane-based biochar is alkaline with a pH value of 10.8. 2.4. Adsorption Mechanism of CECs Considering all the factors affecting the adsorption process and the diverse properties of PhACs and pesticides (molecular structure, hydrophobicity, aromaticity, and polarity), it is clear that the mechanism of adsorption is a complex process that includes different interactions such as pore-filling, hydrogen bonding, π - π electron donor-acceptor interactions, hydrophobic, and electrostatic interactions, etc. [ 34 , 38 ]. Understanding these interactions in complex matrices would help manage wastewater treatment processes and develop efficient methods for their removal. The molecular size of the examined CECs (Table S4) and the pore structure of the biochar and PAC represent very important parameters affecting the adsorption capacity since molecules with smaller dimensions can easily penetrate the adsorbent pores. According to the results of textural characterization, both adsorbents are predominantly microporous materials with similar values of average pore diameter and share of micropore volume. A significantly higher specific surface area obtained for PAC is related to its higher total pore volume. Implying that all the tested molecules of CECs are sufficiently small to access the pores (e.g., the maximum size of carbamazepine is 1.17 nm [ 63 ]), the higher adsorption capacity of PAC for some CECs, compared to biochar, could be attributed to its higher specific surface area and volume of micropores. However, since the initial concentrations of all identified CECs in the real wastewater sample are quite low, the previously mentioned difference is likely to be the consequence of other possible adsorption mechanism scenarios. Hydrophobic interaction is another possible adsorption mechanism occurring due to the tendency of non-polar groups of CECs to aggregate in aqueous solution and therefore, minimize their contact with water molecules. The CECs identified in the real wastewater sample have different hydrophobicity, with logK ow values for PhACs from − 0.07 for Molecules 2025,30, 4803 17 of 28 HCTZ to 3.48 for propranolol, and pesticides from 0.80 for acetamiprid to 3.81 for diazinon (Table S4). In general, compounds with a higher value of logK ow tend to adsorb better on the surface of carbon-based materials due to their more hydrophobic nature. However, adsorption test results showed that both biochar and PAC are very effective adsorbents for highly polar CECs, as well as for the more hydrophobic ones, possibly due to the contribution of the different densities of polar moieties on their surface. The elemental composition of the CO 2 -activated biochar (higher C content and lower O/C ratio) suggests a moderately hydrophobic carbon framework, which can still contribute to the adsorption of neutral or non-polar species. Considering the obtained removal efficiencies of CECs by both adsorbents, a direct correlation between them and the appropriate logK ow values could not be determined, suggesting that these interactions are not the ones controlling the adsorption process. During the adsorption process, electrostatic interactions may occur between ionizable functional groups of CECs and ionizable functional groups present on the surface of biochar and PAC. The raspberry cane-based biochar is alkaline with a pH value of 10.8, while the PAC is slightly acidic with a pH value of 5.02. The pH of the solution is a very important parameter determining the surface charge of sorbents and the degree of dissociation of the PhAC and pesticides [ 34 ]. The pH value of the examined real wastewater sample was around 8. Under these conditions, the number of protonated surface groups on the alkaline biochar is expected to be relatively low, while the more acidic PAC surface generally carries a higher proportion of protonated functionalities at this pH. According to the literature data (Table S4), the majority of the tested PhACs have pK a values above the pH of the solution, suggesting that their molecules mainly exist in neutral form in the wastewater sample [ 64 ]. For example, the pK a of atenolol is 9.6. A significant part of atenolol is positively charged at low pH, but as pH increases above 4, molecules become more neutral [ 27 ]. Thus, at the pH of WWTP effluent (6-8), electrostatic interactions between biochar and atenolol are insignificant. The same conclusions regarding electrostatic interactions between raspberry biochar and carbamazepine, clarithromycin, propranolol, ranitidine, salbutamol, sotalol, and linuron could be drawn. However, almost all the compounds were removed successfully, suggesting that some other mechanisms involved in adsorption are dominant. The lower adsorption efficiency of clarithromycin can be explained by repulsive forces, but also by the fact that it is a larger molecule compared to other components, which facilitates diffusion into pores. The comparative study of biochar and PAC efficiency showed better clarithromycin removal by PAC, probably due to a larger surface area and average pore diameter obtained for the measuring range (1–100 nm). The same assumption regarding electrostatic interactions can be made in the case of pesticides, diazinon and tebuconazole, with pK a values below the pH of the solution, 2.6 (Table S4) and 5.03 [ 65 ], respectively. Both compounds, possibly carrying a negative charge, were completely removed by biochar and PAC, implying that the obtained adsorption capacities are not caused by electrostatic interactions. In addition, Moussavi et al. [ 31 ] reported that NaCl as a background electrolyte and ammonia slightly improved the diazinon adsorption, probably due to the compression of the diazinon electrical double layer, and the formation of a complex diazinon-ammonia that adsorbed on activated carbon, respectively [ 31 ]. Since the adsorption of diazinon in this study was conducted in real wastewater, the effect of these pollutants could contribute to the removal efficiency. The most significant differences in the adsorption capacity of the examined adsorbents were obtained in the case of acetamiprid and linuron. The pK a of acetamiprid is 0.7 (Table S4), which is far below the pH of the wastewater sample, indicating the possible existence of its negatively charged molecules. However, considering the chemical structure, acetamiprid molecules do not contain functional groups that can be deprotonated, referring to their probable neutral Molecules 2025,30, 4803 18 of 28 form under the given experimental conditions and insignificant contribution of electrostatic interactions in the adsorption mechanism. Linuron, with a pK a value of 12.3 (Table S4), has nitrogen atoms in the urea group as potential protonatable sites on the molecule. Since the higher degree of protonation would require more acidic conditions (pH < 7), at pH 8, linuron is more likely to exist in its neutral, deprotonated state. In the case that a small proportion of protonated species exists, it could be adsorbed by the negatively charged PAC and electrostatically repulsed by the positively charged surface of biochar. Two types of hydrogen bonding interactions, as another possible adsorption mechanism, can occur between the examined molecules of CECs and the surface of biochar and PAC. The FTIR spectra confirmed the presence of hydroxyl groups at the surface of the sorbents which can act as H-donors, while the oxygen and nitrogen atoms of the PhACs and pesticides can act as H-acceptors, resulting in dipole–dipole hydrogen bonding [ 66 ]. Considering the chemical structure of all CECs, it can be suggested that this type of interaction is significant during the adsorption process since all the molecules have oxygenand/or nitrogen-containing species that could be involved in dipole–dipole formation (carbonyl, hydroxyl, ester, amino, imino, nitrile groups, etc.). Another type of H-bonding, Yoshida type, occurring between the hydroxyl groups present at the surface of the sorbents and aromatic rings of the adsorbate [ 67 ] could considerably enhance the adsorption affinity of both adsorbents. Given the abundance of aromatic structures in tested CECs (benzene rings, as well as nitrogenand sulfur-containing fused heterocycles), it is plausible that these interactions play an important role in their adsorption. The only exception is related to the clarithromycin molecule. Due to the absence of its aromatic character, its removal efficiency is decreased on both adsorbents. For example, a predominantly neutral carbamazepine can form hydrogen bonds between carbamazepine –NH 2 and oxygen functional groups present at biochar surfaces, such as –OH and C=O [ 28 ]. More efficient adsorption of carbamazepine could be expected at higher pH due to lower concentrations of H + ions when its hydrogen donor groups can interact with hydrogen bonding acceptors or P donors in pine wood nano biochar [ 28 ]. FTIR spectrum of raspberry biochar confirmed the presence of relevant functional groups, so the high carbamazepine removal efficiencies obtained in this study in the pH range of 6–8 could be explained by the same mechanism. Atenolol also has –NH 2 and –OH groups; thus, can adsorb onto biochar surface by the interaction of dipole moment with –OH groups on biochar detected by FTIR [ 68 ]. The aromatic structure of both sorbents and almost all tested CECs (except clarithromycin) might indicate that π - π interactions tend to exert a dominant influence, occurring between the aromatic rings of CECs and the aromatic structure of graphene layers of the biochar and PAC. This is supported by the FTIR evidence of C=C structure. Besides these interactions, nπ interactions can also take place between surface oxygen groups in sorbents ( − COOH and –OH), acting as electron donors, and aromatic rings of CECs molecules acting as electron acceptors [ 69 ]. The investigated pH range (6–8) showed no influence on HCTZ adsorption on raspberry biochar, and it was reduced below the quantification limit of the analytical method. Since HCTZ was predominantly neutral under these conditions, the suggested adsorption mechanisms are π - π interactions and hydrogen bonding. Similar findings were reported by Diniz et al. [ 37 ]. However, the share of this type of interaction in the overall mechanism for each tested pollutant depends on the amount of surface oxygen-containing groups present on the surface of the biochar and PAC. Based on the previous discussion, it can be suggested that the prevailing microporous structure of the biochar and PAC provides abundant adsorption sites, but, like electrostatic and hydrophobic interactions, is not the determining reason for the high adsorption capacity. Hydrogen bonding, as well as π - π and nπ interactions, appear to be the controlling Molecules 2025,30, 4803 19 of 28 mechanisms providing strong sorption affinity of biochar and PAC for the tested PhACs and pesticides (Figure 8). Figure 8. Proposed mechanisms of adsorption. 3. Materials and Methods 3.1. Synthesis and Activation of Biochar Raspberry canes were air-dried, ground, and sieved through a vibrating screen to isolate the particle size fraction of 1.03–2.00 mm, which was used as the feedstock for biochar production. Biochar was synthesized via slow pyrolysis, defined as a thermochemical decomposition of lignocellulosic biomass at moderate heating rates ( typically <20 ◦C/min ) and relatively long residence times under an inert atmosphere [ 70 , 71 ]. Biochar was synthesized in a horizontal fixed-bed quartz reactor following the temperature program described below. A mass of 4 g of the pre-processed raspberry cane material was placed into a quartz sample holder. To ensure an oxygen-free atmosphere, the pyrolysis reactor was flushed with nitrogen for 10 min before heating. The system was then heated to 700 ◦ C under a continuous N 2 flow (110 mL/min) at a heating rate of 10 ◦ C/min, including a holding period of 15 min at 350 ◦ C. The residence time at 700 ◦ C was 2 h. After the pyrolysis step, the reactor was cooled to room temperature under an N 2 flow (110 mL/min). The obtained biochar was subsequently subjected to physical activation under a CO 2 flow (110 mL/min). The sample was heated at 10 ◦ C/min up to 800 ◦ C and held for 1 h (without an additional holding period at intermediate temperatures). Following activation, the reactor was cooled to room temperature under N 2 flow (110 mL/min), after which the biochar sample was removed and weighed. To verify that no background mass loss originated from the reactor system itself, a blank run was performed under identical heating and gas-flow conditions using an empty quartz sample holder. No measurable mass change was observed. Thus, all recorded mass losses during pyrolysis and activation were attributed exclusively to biomass decomposition. A mass balance check was carried out by measuring the mass of Molecules 2025,30, 4803 20 of 28 the biomass before pyrolysis (m 0 ) and the mass of the resulting biochar after each step (m). These values were used to calculate the biochar yield according to Equation (1). Y(%) = m/m0·100 (1) 3.2. Adsorbent Characterization We assessed the morphology and elemental composition of the biochar by using a High-Resolution Scanning Electron Microscope (HRSEM), Thermo Fisher Scientific Apreo 2C (Waltham, MA, USA). Fourier Transform Infra-Red Spectroscopy (FTIR) analysis was performed on a BRUKER Vertex 70 IR spectrometer (wavenumber range of 400–4000 cm −1 with a resolution of 2 cm −1 ) (BRUKER, Billerica, MA, USA) on a solid sample prepared by the KBr method. The Brunauer–Emmett–Teller specific surface analyses were conducted using an Anton Paar NOVAtouch LX2 instrument (Graz, Austria) after degassing materials for 6 h at 110 ◦ C in a vacuum. Brunauer–Emmett–Teller (BET) specific surface area of materials was calculated from the adsorption branch, in the relative pressure range from 0.05 to 0.3 P/P 0 . Barrett, Joyner, and Halenda (BJH) calculations of pore diameter were performed on the desorption branch of the isotherm. Bichar pH was measured after mixing 0.5 g of biochar in 50 mL of deionized water for 1 h and allowing it to stand for another 30 min [72]. 3.3. Adsorption Experiments 3.3.1. Wastewater We performed all adsorption experiments on a real wastewater sample, i.e., effluent of biological treatment from an urban WWTP. The wastewater treatment process includes preliminary (mechanical treatment: coarse and fine screening, grit removal), primary sedimentation, biological treatment (removal of biodegradable organic matter and nutrients), and subsequent reduction of total phosphorus and suspended solids using a membrane process (with the addition of coagulants) before discharge into the receiving body. Table 8presents the average annual values of general quality parameters of treated WWTP effluent at the discharge point into the receiving water body. The quality of the effluent meets the requirements of the Serbian Regulation on Emission Limit Values for Pollutants in Water and Deadlines for Their Achievement [12] for the discharge of treated municipal wastewater into the receiving body. However, it is to be expected that soon, Serbia will have to adopt EU regulations as an EU candidate, including the revised Urban Wastewater Directive. Wastewater sampling was conducted on the influent as a 24-h composite sample for CEC analysis. We collected effluent samples as grab samples (50 L per campaign) of the secondary treatment effluent in the period March-June 2024. Before each set of experiments, wastewater sampling was conducted (campaigns I–V) due to variations in the wastewater composition and the fact that certain CECs go through biotic and abiotic degradation processes over time. Consequently, the number of quantified compounds and their concentrations varied across different investigations. For this reason, the initial concentrations of CECs in the effluent were determined for each experimental set, and removal efficiency was calculated based on these initial concentrations. Molecules 2025,30, 4803 21 of 28 Table 8. Quality parameters of treated effluent from the WWTP biological treatment process (period 2020–2023). Annual Average ELV [12] (Discharge into Receiving Body) Parameter Value Removal (%) Emission Limit Values Requested Removal (%) Suspended solids (mg/L) 5–10.1 97–98 35 (mg/L) 90 pH 6.8–8.1 - 6–9 - COD (mg O2/L) 27–36 96 125 (mg/L) 75 BOD5(mg O2/L) 3–5.2 98–99 25 (mg/L) 70–90 Total nitrogen (mg/L) 5.7–7.2 89–92 10 mg/L 70–80 Total phosphorus (mg/L) 0.51–1 93–95 1 mg/L 80 COD—Chemical Oxygen Demand. BOD5—Biological Oxygen Demand. 3.3.2. Batch Adsorption We assessed the adsorption performance of the biochar in the batch mode with two replicates. All the results are presented as mean values. For the kinetic study, 1 g of biochar was agitated with 1 L of a real wastewater sample for 5, 10, 20, 30, 60, 90, 180, 360 min, 24 h, and 48 h on a laboratory shaker (IKA KS 260 basic). The mixture was then filtered with a glass-fiber filter using 0.45 µ m pore-size membranes, and the filtrates were further analyzed. For the adsorption isotherm study, we spiked a real wastewater sample with pre-selected CECs (15 PhACs, 12 pesticides in current use) to obtain diverse initial concentrations of the micropollutants. The selection of CECs to be included in the pre-selected list was done taking into consideration the CECs identified during the previously undertaken studies [ 7 , 9 , 73 , 74 ], which involved the analysis of surface and wastewater samples from the region of interest. The influence of the pH levels (6, 7, 8, and the original pH of the wastewater sample) and the adsorbent dose (1.0, 0.5, and 0.2 g/L) was also examined. The pH of the wastewater samples was adjusted using HCl and NaOH. The adsorption capacity was calculated as follows: q (ng/L) = (C0−C)/m, (2) where C 0 (ng/L) is the initial concentration of the organic compound, C (ng/L) is the concentration of the same compound after adsorption, and m (g) is the mass of biochar used per liter of effluent. Were determined the CEC concentrations by high-performance liquid chromatography coupled to triple quadrupole mass spectrometry (UHPLC–MS/MS). For compounds that were not detected or detected in a concentration lower than the method’s limit of quantification (MLOQ) after adsorption, for the calculation of adsorption capacity, we expressed residual concentrations as MLOQ/2 (Table S1). We also investigated the possibility of multi-stage batch adsorption. The biochar was mixed with a real wastewater sample (1 g/L, original pH) for 45 min. After filtration, the biochar was transferred to a clean and dry Erlenmeyer flask and used again under the same conditions as a fresh aliquot of the wastewater sample. Three consecutive adsorption cycles were performed. Repeatability for the components’ adsorption was evaluated by conducting three tests under identical conditions (including the same sample of wastewater) with a relative standard deviation (RSD) of 15, 8 and 15% for clarithromycin, propranolol, and linuron, respectively. For all other components, the final concentrations after adsorption tests were lower than MLOQ. Molecules 2025,30, 4803 22 of 28 3.3.3. Sample Preparation for UHPLC–MS/MS We prepared the water samples according to the method described by Gros et al. [ 75 ], Petrovi´c et al. [ 73 ], and Raki´c et al. [ 76 ]. Selected CECs were extracted from water samples by solid-phase extraction (SPE) (Oasis HLB cartridges, 6 mL, 200 mg; Waters Corp., Milford, MA, USA) using a Baker vacuum system (J.T. Baker, Deventer, The Netherlands). Briefly, the cartridges were conditioned with 5 mL of methanol (MeOH, J.T. Backer, Deventer, The Netherlands), followed by 5 mL of HPLC water (18.2 M Ω· cm, Milli-Q purification system, Millipore, Molsheim, France). Water samples (i.e., 100 mL effluent) were pre-concentrated on Oasis HLB cartridges. Afterward, the cartridges were washed with 5 mL of HPLC water and dried under vacuum for 3 h to remove excess water. The retained compounds were eluted using methanol (2 × 4 mL). The extracts were fortified with 20 µ L of a mixture of internal standards (Table S1) to a final concentration of 40 ng/mL. The extracts were evaporated to dryness at 30 ◦ C under a gentle nitrogen stream and reconstituted in 1 mL of the first gradient of the mobile phase. Before evaporation, internal standards were added to compensate for possible matrix effects (ion suppression or enhancement) and potential losses during evaporation, but not for losses incurred during the extraction process. 3.3.4. Instrumental Analysis Instrumental analysis of selected CECs in the prepared extracts was performed using high-performance liquid chromatography coupled to triple quadrupole mass spectrometry (UHPLC–MS/MS, Thermo Fisher Scientific, Waltham, MA, USA), based on the method previously described by Raki´c et al. [ 74 ]. The separation of the analyzed CECs was performed using a Rapid Resolution HD ZORBAX Eclipse Plus C18 column (100 × 2.1 mm, i.d. 1.8 µ m, Agilent Technologies, Santa Clara, CA, USA) with a flow rate of 400 µ L/min. The column temperature was maintained at 30 ◦ C, while the sample injection volume was 10 µL . The mobile phase consisted of eluent A, containing water/formic acid (Sigma-Aldrich, St. Louis, MO, USA) (99.9:0.1, v/v) and 5 mM ammonium acetate (Sigma-Aldrich, St. Louis, MO, USA), and eluent B, which consisted only of MeOH. The gradient program was set as follows: 5% of eluent B, increasing to 95% in 4.5 min, rising to 100% in the following 0.1 min, and held for 1.9 min and then returned to the initial condition. The total run time was 13 min. For identification and quantification of CECs we used a triple quadrupole mass spectrometer (MS/MS) TSQ Vantage (Thermo Fisher Scientific, Waltham, MA, USA) equipped with a heated-electrospray ionization probe (HESI-II, Thermo Scientific, Waltham, MA, USA). The parameters of the ion source were set as follows: spray voltage—3.4 kV, vaporizer temperature, 250 ◦ C; sheath gas pressure, 40 arbitrary units; auxiliary gas pressure, 10 arbitrary units; capillary temperature, 270 ◦ C. The target compounds were analyzed in selected reaction monitoring (SRM) mode. All data were processed using Xcalibur 2.1.0 software (Thermo Fisher Scientific, Waltham, MA, USA). 4. Conclusions Within this study, we tested the adsorption characteristics of biochar obtained through the slow pyrolysis of dry raspberry stalks for the removal of CECs from the effluent of the biological treatment of WWTP. The results demonstrated that adsorption occurs in a very short time (around 5 min), enabling the efficient removal of most detected compounds with minimal contact time, which in turn allows for the use of smaller treatment reactors. The smallest tested dose of biochar, 0.2 g/L, is sufficient to achieve high removal efficiencies, and no significant variation in performance was observed across the common pH range for treated municipal wastewater, eliminating the need for pH adjustment and reducing operational costs. Additionally, biochar retained its adsorption performance across three consecutive batch cycles, confirming its potential for reuse. A comparative study showed Molecules 2025,30, 4803 23 of 28 that raspberry cane-based biochar performance is similar to commercial powdered activated carbon (PAC) in removing a broad spectrum of PhACs and pesticides from complex wastewater matrices. To the best of our knowledge, this is the first study to investigate biochar produced from raspberry canes for the removal of CECs from real wastewater. What also stands out in this work is that the tests were carried out using real wastewater samples, which are far more complex than synthetic solutions typically used in laboratory studies. Our main objective was to assess the practical performance of biochar in real wastewater and to determine whether its removal efficiency meets the requirements of the new EU Urban Wastewater Treatment Directive (2024) [ 13 ], which requires 80% removal of selected micropollutants. These results give confidence that the biochar is applicable in real-world conditions. Furthermore, the study considers the valorization of regional agricultural waste within a circular economy concept, as well as the challenges of waste management, and contributes to the development of cost-effective and environmentally friendly wastewater treatment strategies. Altogether, the obtained results confirm that raspberry cane-based biochar is an effective, low-cost, and sustainable alternative for removing CEC from wastewater, offering both environmental and operational benefits to modern wastewater treatment. However, future research should focus on continuous-flow column studies and breakthrough curve analyses, which are necessary to evaluate performance under realistic hydraulic conditions typical of industrial wastewater treatment, as well as scalability for pilot systems. We should also examine challenges related to regeneration methods and the disposal of spent material. A techno-economic assessment, combined with life-cycle analysis, would help clarify the competitiveness of raspberry-cane biochar relative to powdered activated carbon (PAC) and other commercial adsorbents. Supplementary Materials: The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/molecules30244803/s1, Table S1: Validation parameters: instrumental linearity, the method limits of quantification (MLOQ), recoveries, and precision (%RSD); Table S2: Removal efficiencies of nonactivated and activated biochar (1 g/L, 24 h); Table S3: CEC removal efficiencies by biochar and PAC (dose 20 mg/L, 45 min); Table S4. Properties of CECs. References [66,77–80] are cited in the Supplementary Materials. Author Contributions: Conceptualization, D.L., V.V. and N.Ð.-M.; methodology, D.L., V.V., J.Ž. and I.A.; software, I.A. and J.Ž.; validation, I.A. and J.Ž.; formal analysis, I.A., J.Ž. and D.L.; investigation, D.L., V.V., S.P. and M.P.; resources, N.Ð.-M.; data curation, D.L., I.A. and J.Ž.; writing—original draft preparation, D.L., V.V. and S.P.; writing—review and editing, D.L., V.V., J.Ž., I.A., S.P. and M.P.; visualization, D.L.; supervision, N.Ð.-M.; project administration, N.Ð.-M.; funding acquisition, N.Ð.-M. All authors have read and agreed to the published version of the manuscript. Funding: Funded by the European Union. Views and opinions expressed are, however, those of the author(s) only and do not necessarily reflect those of the European Union or EU executive agency. Neither the European Union nor the granting authority can be held responsible for them. This study is conducted under the project TwiNSol-CECs, which has received funding from the Horizon Europe programme under grant agreement no.101059867. Institutional Review Board Statement: Not applicable. Informed Consent Statement: Not applicable. Data Availability Statement: Data is contained within the article or Supplementary Materials. Conflicts of Interest: The authors declare no conflicts of interest. The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript; or in the decision to publish the results. Molecules 2025,30, 4803 24 of 28 Abbreviations The following abbreviations are used in this manuscript: WWTPs Wastewater treatment plants CECs Contaminants of emerging concern PAC Powdered activated carbon PhACs Pharmaceutically active compounds PFAS polyand perfluoroalkyl substances AC Activated carbon MLOQ Method limit of quantification HCTZ Hydrochlorothiazide DOC Dissolved organic carbon SBET BET specific surface area PV Pore volume HRSEM High-resolution scanning electron microscopy EDX X-ray energy dispersion FTIR Fourier transform infra-ed spectroscopy BET Brunauer–Emmett–Teller specific surface CODbBiodegradable Chemical Oxygen Demand BOD5Biological Oxygen Demand UHPLC–MS/MS high-performance liquid chromatography coupled to triple quadrupole mass spectrometry References 1. Bilal, M.; Adeel, M.; Rasheed, T.; Zhao, Y.; Iqbal, H.M.N. Emerging contaminants of high concern and their enzyme-assisted biodegradation—A review. Environ. Int. 2019,124, 336–353. [CrossRef] 2. Gkotsis, G.; Nika, M.-C.; Nikolopoulou, V.; Alygizakis, N.; Bizani, E.; Aalizadeh, R.; Badry, A.; Chadwick, E.; Cincinelli, A.; Claßen, D.; et al. Assessment of contaminants of emerging concern in European apex predators and their prey by LC-QToF MS wide-scope target analysis. Environ. Int. 2022,170, 107623. [CrossRef] [PubMed] 3. Yadav, D.; Rangabhashiyam, S.; Verma, P.; Singh, P.; Devi, P.; Kumar, P.; Hussain, C.M.; Gaurav, G.K.; Kumar, S. Environmental and health impacts of contaminants of emerging concerns: Recent treatment challenges and approaches. Chemosphere 2021,272, 129492. [CrossRef] 4. Gros, M.; Rodríguez-Mozaz, S.; Barceló, D. Rapid analysis of multiclass antibiotic residues and some of their metabolites in hospital, urban wastewater, and river water by ultra-high-performance liquid chromatography coupled to quadrupole linear ion trap tandem mass spectrometry. J. Chromatogr. A 2013,1292, 173–188. [CrossRef] 5. Gaw, S.; Thomas, K.V.; Hutchinson, T.H. Sources, impacts and trends of pharmaceuticals in the marine and coastal environment. Philos. Trans. R. Soc. B Biol. Sci. 2014,369, 20130572. [CrossRef] 6. Chen, M.; Cooper, V.I.; Deng, J.; Amatya, P.L.; Ambrus, D.; Dong, S.; Stalker, N.; Nadeau-Bonilla, C.; Patel, J. Occurrence of pharmaceuticals in Calgary’s wastewater and related surface water. Water Environ. Res. 2015,87, 414–424. [CrossRef] [PubMed] 7. ˇ Celi´c, M.; Gros, M.; Farré, M.; Barceló, D.; Petrovi´c, M. Pharmaceuticals as chemical markers of wastewater contamination in the vulnerable area of the Ebro Delta (Spain). Sci. Total Environ. 2019,652, 952–963. [CrossRef] 8. Vasi´c, V.; Kuki´c, D.; Š´ciban, M.; Ðuriši´c-Mladenovi´c, N.; Veli´c, N.; Pajin, B.; Crespo, J.; Farre, M.; Šereš, Z. Lignocellulose-based biosorbents for the removal of contaminants of emerging concern (CECs) from water: A review. Water 2023,15, 1853. [CrossRef] 9. Ðuriši´c-Mladenovi´c, N.; Živanˇcev, J.; Anti´c, I.; Raki´c, D.; Buljovˇci´c, M.; Pajin, B.; Llorca, M.; Farre, M. Occurrence of contaminants of emerging concern in different water samples from the lower part of the Danube River Middle Basin—A review. Environ. Pollut. 2024,363, 125128. [CrossRef] 10. Llamas-Dios, M.I.; Vadillo, I.; Jiménez-Gavilán, P.; Candela, L.; Corada-Fernández, C. Assessment of a wide array of contaminants of emerging concern in a Mediterranean water basin (Guadalhorce river, Spain): Motivations for an improvement of water management and pollutants surveillance. Sci. Total Environ. 2021,788, 147822. [CrossRef] 11. Munro, K.; Martins, C.P.; Loewenthal, M.; Comber, S.; Cowan, D.A.; Pereira, L.; Barron, L.P. Evaluation of combined sewer overflow impacts on short-term pharmaceutical and illicit drug occurrence in a heavily urbanised tidal river catchment (London, UK). Sci. Total Environ. 2019,657, 1099–1111. [CrossRef] Molecules 2025,30, 4803 25 of 28 12. ELV—Regulation on Limit Values for Emissions of Pollutants into Waters and Deadlines for Achieving Them. “Official Gazette of RS”, No. 67/2011, 48/2012, 1/2016. Available online: https://www.paragraf.rs/propisi/uredba-granicnim-vrednostima-emisijezagadjujucih-materija-u-vode.html (accessed on 16 March 2025). 13. European Parliament and Council. Directive (EU) 2024/3019 of 27 November 2024 on the treatment of urban wastewater (recast of Directive 91/271/EEC). Off. J. Eur. Union 2024. Available online: https://eur-lex.europa.eu/legalcontent/EN/TXT/PDF/?uri=OJ:L_202403019 (accessed on 16 March 2025). 14. Rizzo, L.; Malato, S.; Antakyali, D.; Beretsou, V.G.; Ðoli´c, M.B.; Gernjak, W.; Heath, E.; Ivancev-Tumbas, I.; Karaolia, P.; Ribeiro, A.R.L.; et al. Consolidated vs new advanced treatment methods for the removal of contaminants of emerging concern from urban wastewater. Sci. Total Environ. 2019,655, 986–1008. [CrossRef] [PubMed] 15. Pistocchi, A.; Alygizakis, N.A.; Brack, W.; Boxall, A.; Cousins, I.T.; Drewes, J.E.; Finckh, S.; Gallé, T.; Launay, M.A.; McLachlan, M.S.; et al. European scale assessment of the potential of ozonation and activated carbon treatment to reduce micropollutant emissions with wastewater. Sci. Total Environ. 2022,848, 157124. [CrossRef] 16. Eggen, R.I.L.; Hollender, J.; Joss, A.; Schärer, M.; Stamm, C. Reducing the discharge of micropollutants in the aquatic environment: The benefits of upgrading wastewater treatment plants. Environ. Sci. Technol. 2014,48, 7683–7689. [CrossRef] 17. Hagemann, N.; Schmidt, H.-P.; Kägi, R.; Böhler, M.; Sigmund, G.; Maccagnan, A.; McArdell, C.S.; Bucheli, T.D. Wood-based activated biochar to eliminate organic micropollutants from biologically treated wastewater. Sci. Total Environ. 2020,730, 138417. [CrossRef] 18. Qian, J.; Riede, P.; Abbt-Braun, G.; Parniske, J.; Metzger, S.; Morck, T. Removal of organic micropollutants from municipal wastewater by powdered activated carbon—Activated sludge treatment. J. Water Process Eng. 2022,50, 103246. [CrossRef] 19. Dai, Y.; Zhang, N.; Xing, C.; Cui, Q.; Sun, Q.; Zhang, X. The adsorption, regeneration and engineering applications of biochar for removal of organic pollutants: A review. Chemosphere 2019,223, 12–27. [CrossRef] [PubMed] 20. Yaashikaa, P.R.; Senthil Kumar, P.; Varjani, S.; Saravanan, A. A critical review on the biochar production techniques, characterization, stability and applications for circular bioeconomy. Biotechnol. Rep. 2020,28, e00570. [CrossRef] 21. Ambaye, T.G.; Vaccari, M.; van Hullebusch, E.D.; Amrane, A.; Rtimi, S. Mechanisms and adsorption capacities of biochar for the removal of organic and inorganic pollutants from industrial wastewater. Int. J. Environ. Sci. Technol. 2021,18, 3273–3294. [CrossRef] 22. Adamovi´c, A.; Petronijevi´c, M.; Pani´c, S.; Cvetkovi´c, D.; Anti´c, I.; Petrovi´c, Z.; Ðuriši´c-Mladenovi´c, N. Biochar and hydrochar as adsorbents for the removal of contaminants of emerging concern from wastewater. Adv. Technol. 2023,12, 57–74. [CrossRef] 23. Chen, S.; Chen, T.; Huang, Q.; Zhan, M.; Li, X.; Yan, J. Activation of persulfate by CO 2 -activated biochar for improved phenolic pollutant degradation: Performance and mechanism. Chem. Eng. J. 2020,380, 122519. [CrossRef] 24. Kuki´c, D.; Ivanovska, A.; Vasi´c, V.; La ¯ darevi´c, J.; Kosti´c, M.; Š´ciban, M. The overlooked potential of raspberry canes: From waste to an efficient low-cost biosorbent for Cr(VI) ions. Biomass Convers. Biorefinery 2024,14, 4605–4619. [CrossRef] 25. eKapija. Thousands of Kilowatts Remain in Orchards—Production of Pellets and Briquettes from Nuts Increases the Price of Raspberries by 7%. Available online: https://www.ekapija.com/where-to-invest/2152201/hiljade-kilovata-ostaje-u-vocnjacimaproizvodnja-peleta-i-briketa-od-orezina-uvecava (accessed on 7 November 2025). 26. Das, S.; Ray, N.M.; Wan, J.; Khan, A.; Chakraborty, T.; Ray, M.B. Micropollutants in wastewater: Fate and removal processes. In Physico-Chemical Wastewater Treatment and Resource Recovery; Farooq, R., Ahmad, Z., Eds.; IntechOpen: London, UK, 2017. [CrossRef] 27. Fu, C.; Zhang, H.; Xia, M.; Lei, W.; Wang, F. The single/co-adsorption characteristics and microscopic adsorption mechanism of biochar-montmorillonite composite adsorbent for pharmaceutical emerging organic contaminant atenolol and lead ions. Ecotoxicol. Environ. Saf. 2020,187, 109763. [CrossRef] 28. Naghdi, M.; Taheran, M.; Pulicharla, R.; Rouissi, T.; Brar, S.K.; Verma, M.; Surampalli, R.Y. Pine-wood derived nanobiochar for removal of carbamazepine from aqueous media: Adsorption behavior and influential parameters. Arab. J. Chem. 2019,12, 5292–5301. [CrossRef] 29. Chu, G.; Zhao, J.; Liu, Y.; Lang, D.; Wu, M.; Pan, B.; Steinberg, C.E.W. The relative importance of different carbon structures in biochars to carbamazepine and bisphenol A sorption. J. Hazard. Mater. 2019,373, 106–114. [CrossRef] [PubMed] 30. Baharum, N.A.; Nasir, H.M.; Ishak, M.Y.; Isa, N.M.; Hassan, M.A.; Aris, A.Z. Highly efficient removal of diazinon pesticide from aqueous solutions by using coconut shell-modified biochar. Arab. J. Chem. 2020,13, 6106–6121. [CrossRef] 31. Moussavi, G.; Hosseini, H.; Alahabadi, A. The investigation of diazinon pesticide removal from contaminated water by adsorption onto NH4Cl-induced activated carbon. Chem. Eng. J. 2013,214, 172–179. [CrossRef] 32. Yang, J.; Sun, H.; Liu, Y.; Wang, X.; Valizadeh, K. The sorption of Tebuconazole and Linuron from an aqueous environment with a modified sludge-based biochar: Effect, mechanisms, and its persistent free radicals study. J. Chem. 2021,2021, 2912054. [CrossRef] 33. Moraes, L.E.Z.d.; Marcoti, F.A.O.; Lucio, M.A.N.; Rocha, B.C.d.S.; Rocha, L.B.; Romero, A.L.; Bona, E.; Peron, A.P.; Junior, O.V. Analysis and simulation of adsorption efficiency of herbicides Diuron and Linuron on activated carbon from spent coffee beans. Processes 2024,12, 1952. [CrossRef]