Reduction and immobilization of Cd(II) and As(III) using sulfur-ferrihydrite-biochar as an amendment in water and soil: Investigation of the Mechanism of Remediation

Reduction and immobilization of Cd(II) and As(III) using sulfur-ferrihydrite-biochar as an amendment in water and soil: Investigation of the Mechanism of Remediation | Research Square window.SnipcartSettings = { analytics: { enabled: false } }; (function() { var accessVector = localStorage.getItem('access_vector') || ''; window.dataLayer = window.dataLayer || []; if (accessVector) { window.dataLayer.push({ user: { profile: { profileInfo: { snid: accessVector } } } }); } })(); (function(w,d,s,l,i){w[l]=w[l]||[];w[l].push({'gtm.start':new Date().getTime(),event:'gtm.js'});var f=d.getElementsByTagName(s)[0],j=d.createElement(s),dl=l!='dataLayer'?'&l='+l:'';j.async=true;j.src='https://www.googletagmanager.com/gtm.js?id='+i+dl;f.parentNode.insertBefore(j,f);})(window,document,'script','dataLayer','GTM-K279D39R'); Browse Preprints In Review Journals COVID-19 Preprints AJE Video Bytes Research Tools Research Promotion AJE Professional Editing AJE Rubriq About Preprint Platform In Review Editorial Policies Our Team Advisory Board Help Center Sign In Submit a Preprint Cite Share Download PDF Research Article Reduction and immobilization of Cd(II) and As(III) using sulfur-ferrihydrite-biochar as an amendment in water and soil: Investigation of the Mechanism of Remediation Xuqiao Wu, Xiaowen Teng, Dong Huang, Ijlal Ahmad, Hanbo Chen, and 7 more This is a preprint; it has not been peer reviewed by a journal. https://doi.org/ 10.21203/rs.3.rs-6195683/v1 This work is licensed under a CC BY 4.0 License Status: Posted Version 1 posted You are reading this latest preprint version Abstract The transformation behaviors of arsenic (As) and cadmium (Cd) in contaminated soils are generally complex process due to their distinct chemical and physical characteristics, which poses challenge for remediation. This study proposes an efficient strategy for the simultaneous immobilization of Cd and As using sulfur-ferrihydrite-modified biochar (SFB) as an organic amendment. A series of experiments, including batch and pot experiments, was conducted under controlled conditions. The results showed that the maximum sorption capacities of Cd and As by SFB were 76.69 mg kg -1 and 8.28 mg kg -1 , respectively, which were significantly higher than those of biochar (BC), ferrihydrite (FH) and ferrihydrite-biochar (FB). This higher sorption capacity is attributed to synergistic interactions between biochar and ferrihydrite. The sorption process of Cd and As by SFB follows the Langmuir isothermal sorption model and the pseudo-second-order kinetic model, indicating a combination of physical sorption and chemisorption mechanisms. The removal mechanisms for As primarily involve precipitation, oxidation and complexation, while those for Cd mainly include ion exchange, complexation, precipitation, and electrostatic sorption. Application of SFB reduced the bioavailable forms of Cd and As in the soil, shifting their chemical forms toward more stable residual states and enhancing immobilization. Overall, the SFB is a novel and effective adsorbent by immobilizing Cd and As in agricultural soils, promoting safer crops production in contaminated field. Modified biochar Cadmium Arsenic Sorption mechanisms Immobilization Soil remediation Figures Figure 1 Figure 2 Figure 3 Figure 4 Figure 5 Figure 6 Figure 7 Figure 8 1. Introduction Hazardous toxic metal(liod)s, such as cadmium (Cd) and arsenic (As), primarily originate from anthropogenic activities and the weathering of soil-forming rocks and mineral posing serious threats to plants and human health (Lyu et al., 2022 ). Cd and As are recognized highly toxic and mobile metals that are taken up by the plants in higher concentration when presents in soil and if these plants are consumed by human or animals posing serious health hazards. Chronic exposure to these metalloids can cause irreversible damage to the kidneys, bones, and nervous system (Adnan et al., 2022 ). In China, over 2.0×10 5 km 2 cultivated land is contaminated by Cd and As, particularly in the southern regions (Sun et al., 2024 , 33; Gong et al, 2020 ). However, the remediation of those contaminated soils with Cd and As remains highly challenging due to their complex chemical behavior and mobile properties in soil (Qiao et al., 2018 ; Zhou et al., 2022 ). Specifically, Cd exists in the cationic form, whereas As generally occurs in anionic forms, which resulting in the difficult for a single remediation agent to simultaneously immobilize both Cd and As in soil (Shen et al., 2020 ; Vankova et al., 2021 ; Zhou et al., 2022 ). Moreover, their co-existence in the environment may lead to complex interactions, such as competition for sorption and solubility changes that may reduce their remediation efficiency (Huang et al., 2022 ). Thus, it is essential to seek an effective remediation strategy to simultaneously immobilize Cd and As in contaminated soil, while elucidating the mechanisms underlying their sorption and stabilization. In situ stabilization to make it non available to plants is one of the most widely used and proven technique for remediating contaminated soils. Amendments such as ferrihydrite and sulfur (S) materials have shown great potential in reducing the mobility and bioavailability of heavy metal(liod) (Qu et al., 2022a ; Liu et al., 2015 ; Palansooriya et al., 2020 ; Zeng et al., 2024 ). Moreover, these amendments are generally effective in immobilizing cationic metals such as Cd by forming stable precipitates or minerals, while they fail to stabilize As in their co-contaminated soil due to its anionic nature (Lee et al., 2022 ). Biochar a product of the anaerobic pyrolysis of agricultural, industrial, and household waste (Zou et al., 2022 ), is widely employed one of the organic amendments in soil remediation due to its high porosity, abundant functional groups and strong ion exchange capacity (Yang et al., 2021 ). These properties make biochar highly effective immobilizing agent for cationic pollutants, such as Cd, through a mechanisms including cation exchange, electron donation, and chelation. However, the negative charge of biochar surface limits its capacity to adsorb anionic pollutants such as As, due to electrostatic repulsion (Tan et al., 2017 ; Zhi et al., 2014 ). Therefore, chemical modification of biochar through integration with ferrihydrite is an innovative strategy to enhance its performance in soils contaminated with both Cd and As. Ferrihydrite, a precursor of crystalline iron oxides is found in natural environments as nanoparticle colloids (Wang et al., 2020 ). It has high surface area with strong binding capacity, and ability to form inner-sphere complexes that make ferrihydrite a powerful immobilizer of both cations and anions including Cd and As (Cui et al., 2022 ; Fu et al., 2021 ; Hong et al., 2022 ; Xiu et al., 2018 ). Moreover, ferrihydrite exhibits a strong tendency for anions complexation via mechanisms, such as formation of arsenate oxyanions, inner-sphere sorption, co-precipitation/incorporation, and isomorphous substitution (Hu et al., 2018 ; Hu et al., 2020 ). Preliminary studies also have indicated that ferrihydrite had effectively stabilized Cd and As in the solid phase, thereby reduced their mobility and bioavailability (Fan et al., 2021 ; Jeong et al.,2017; Ouyang et al.,2021). However, the tendency of anhydrous ferrous iron in ferrihydrite significantly reduces the efficiency of metal(liod) sorption and partitioning (Tian et al., 2018 ). Several research studies ave suggested that integration of ferrihydrite with biochar can mitigate agglomeration, enhancing the availability of active binding sites within biochar porous structure and improving the immobilization of Cd and As (Bai et al., 2022 ; Du et al., 2021 ). Sulfur modifications offer another promising avenue for enhancing immobilization and reducing bioavailability of heavy metal(liod)s in soils. Sulfur plays a vital role in plant responses to heavy metal(liod)s stress by contributing to thiol-rich peptides, such as phytochelatins, that bind heavy metal(liod)s and mitigate their toxicity. Iron-sulfur modified biochar could serve as an effective strategy for improving the remediation efficiency of soils co-contaminated with As and Cd (Wu et al., 2019 ). The large surface area, high reactivity, and amphoteric properties of iron oxides make them highly suitable for the chemical immobilization of various soil pollutants, as iron-based materials effectively remove heavy metal(liod)s via ion exchange and precipitation (Wang et al., 2018 ). Furthermore, the addition of sulfur can enhance the reduction and precipitation of heavy metal(liod)s through the provision of reducing species (S²⁻) (Yang et al., 2019 ). Rajendran et al. ( 2019 ) reported that sulfur enrichment onto iron-based biochar has effectively reduced Cd content in soil and lowered Cd accumulation in rice tissues. Yin et al. ( 2019 ) found that sulfur-doped Fe 3 O 4 nanoparticles exhibited excellent sorption performance for As(V) because sulfur doping activates Fe atoms, creating sorption sites. Qiao et al. (2020) reported that sulfide-modified α-FeOOH increased the sorption capacity of As (V) from 153.8 to 384.6 mg g − 1 compared to pristine α-FeOOH. Overall, iron-sulfur-modified biochar shows significant potential for enhancing the simultaneous remediation of Cd and As in contaminated soils. However, the mechanisms underlying its dual immobilization capabilities require further investigation. Therefore, in this study, an iron-sulfur-modified biochar was prepared to investigate the efficiency of its immobilization potential and mechanisms of stabilization of Cd(II) and As(III) in soil through batch sorption experiment. The specific objectives of this study were: (1) to evaluate the sorption behavior and mechanisms of the sulfur-ferrihydrite-biochar amendment for immobilization of Cd(II) and As(III); (2) To elucidate the binding mechanism of Cd(II) and As(III) on sulfur-ferrihydrite-biochar composites. This study provided an in-depth understanding of mechanism into the interfacial interaction between Cd(II), As(III), and sulfur-ferrihydrite-biochar composites, particularly in co-contaminated systems. The findings will help in advancing remediation strategies and improving the geochemical reaction understanding of Cd(II) and As(III) in natural environments. 2. Materials and methods 2.1 Materials Rice straw was collected from a farm located in Taizhou, Zhejiang Province, China, and was used as feedstock to produce pristine biochar. Fe(NO 3 ) 3 ·9H 2 O, SC 2 and NaOH (all AR grade) were selected to modify the biochar NaAsO 2 and CdCl 2 (both AR grade) were used to prepare the stock solutions for the batch sorption experiment. All reagents were purchased from Reagent Co. Ltd (China), and all the solutions were prepared by using ultrapure water. 2.2 Preparation and modification of biochar. 2.2.1 Biochar preparation Rice straw was air-dried and filtered, sieved through a 5 mm sieve. Biochar was produced by slow pyrolysis of rice straw at 550℃ for 2 h with a heating rate of 5℃ min − 1 , under limited oxygen supply in a muffle furnace. After cooling to room temperature, the biochar was washed with the ultrapure water and filtered through a 100-mesh sieve. The sieved biochar was collected and was referred to as BC. 2.2.2 Biochar modification The synthetic procedure of ferrihydrite was used as described by Du et al. ( 2021 ). A 2-line ferrihydrite (FH) was prepared by adding 1.5 mol L − 1 NaOH drop wise to a 0.1 mol L − 1 Fe(NO 3 ) 3 solution until the pH reached 7.0, and was allowed to stand for 1 h, afterward the supernatant was decanted. The precipitate was separated by centrifugation at 4500 rpm, washed twice with ultrapure water and dialyzed for three days in darkness to remove excess Na + and NO 3− . The residue was dried at 60°C in a thermostatic drying oven for 8 h, After cooling at room temperature, the residues was passed through a 100-mesh sieve. The resulting ferrihydrite was collected and referred to as FH. Biochar-loaded ferrihydrite was prepared at a 2: 1 mass ratio of ferrihydrite to biochar. The biochar was added to 1 mol L − 1 KOH solution and stirred at 60°C for 72 h. The mixture was then kept for 24 h and then centrifuged at 5000 rpm. The precipitate was washed thoroughly with ultrapure water and frozen overnight at -20°C, and put into freeze-dried to complete dryness. The resulting material was designated as FB. The FB was grounded and passed through a 0.15 mm sieve. Sulfur addition was performed by mixing 100 mL of 0.40 mol L − 1 NaOH solution with carbon disulfide (CS 2 ) at a ratio of 1: 1.5. The mixture was stirred for 4 h at 25°C and sonicated for 1 hour to obtain a sulfur modified solution. The solution was further stirred for 16 h at 40°C using a magnetic stirrer, then cooled to room temperature and filtered through 0.45 µm filters. A 10 g of the prepared biochar was added to 100 mL sulfur modified mixture solution and stirred for 8 h at 45°C with a magnetic stirrer. The mixture was dried, filtered and passed through a 100-mesh sieve, and designated the product as SFB. 2.3 Batch sorption experiments A series of batch experiments were conducted at 25℃ to investigate the sorption characteristic of Cd(II), and As(III) by adding BC, FH, FB and SFB. All solutions used in batch tests contained 0.1 mol L − 1 NaOH and 0.1 mol L − 1 HCl as ionic strength adjuster, and the experiments were performed in triplicates. The solution concentrations of Cd(II) and As(III) before and after sorption were determined through inductively coupled plasma optical emission spectrometry (ICPOES, 720 ES, Agilent, USA). The adsorbents were all used at the rate of 1.5 g L − 1 . Each set of experiments had three replicates. The oscillation condition was set at 180 rpm, 25℃, for 24 h. The optimum pH for sorption was determined in the dual pollutant system (5mg L − 1 As(III) solution and 40mg L − 1 Cd(II) solution) with pH settings of 2, 3, 4, 5, 6, 7, 8 and 9. For the amendments dosage experiments, solutions were extracted at different dosages (0.2, 0.5, 1, 1.5 and 2g L − 1 ) in a solution containing 40 mg L − 1 Cd(II) and 5 mg L − 1 As(III). For kinetic experiment, solutions were extracted at different intervals (5, 10, 30, 60, 120, 240, 360, 540, 720, 960, 1200 and 1440 min) in the solution containing 40 mg L − 1 Cd(II) and 5 mg L − 1 As(III). The sorption isotherms were studied for Cd(II) and As(III) mixed solutions at concentrations of 3.2–80 mg L − 1 and 0.4–10 mg L − 1 . For thermodynamics, the sorption of Cd(II) and As(III) at 35℃, 25℃ and 15℃ was also investigated. All samples were filtered through 0.45 µm microporous membranes, acidified, and determined concentrations on an inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7000, Japan). 2.3.2 Modeling for metal(liod)s sorption (1) To investigate the change of binding rate between Cd(II) and As(III) and adsorbents with time, kinetic investigations were performed. The uptake results were then fitted by pseudo-first (Formula 1) and second-order (Formula 2), and intra-particle diffusion models for describing the kinetic course of Cd(II) and As(III)adsorption by adsorbents quantitatively (Bai et al., 2022 ): \(\:{Q}_{t}={Q}_{e}(1-{e}^{-k1t})\) (Formula 1) where, Q t : sorption capacity at time t (mg g − 1 ); Q e : Equilibrium sorption capacity (mg g − 1 ); t: Time (min); k 1 : Pseudo-first-order rate constant. \(\:\frac{t}{{q}_{t}}=\frac{1}{{k}_{2}{q}_{e}^{2}}+\frac{t}{{q}_{e}}\) (Formula 2) where, Q t : sorption capacity at time t (mg g − 1 ); Q e : Equilibrium sorption capacity (mg g − 1 ); t: Time (min); k 2 : Pseudo-second-order rate constant. (2) Isotherm rate were then performed at different temperatures by further elucidating the Cd(II) and As(III) uptake behaviors with adsorbents. Two recognized isotherm models, including the Langmuir sorption isotherm (Formula 3) and Freundlich parameters (Formula 4) equation is given by Bai et al. ( 2022 ): \(\:{Q}_{e}=\frac{{Q}_{m}{bC}_{e}}{1+{bC}_{e}}\) (Formula 3) where, Q e : Equilibrium sorption capacity (mg g -1 ); Q m : Maximum sorption capacity (mg g -1 ); b: Langmuir constant related to the affinity of binding sites (L mg -1 ); C e : Equilibrium concentration of the sorption in the solution (mg L -1 ). \(\:{Q}_{e}={k}_{f}{C}_{e}^{1/n}\) (Formula 4) where, Q e : Equilibrium sorption capacity (mg g -1 ); C e : Equilibrium concentration of the sorption in the solution (mg L -1 ); K f : Freundlich constant indicative of sorption capacity; n: Empirical constant indicative of sorption intensity. 2.3.3 Characterization of amendments before and after metal(liod)s sorption The physicochemical properties of BC, FH, FB and SFB were analyzed in 1: 2.5 (w/v) ratio of solid and ultrapure water. The pH measured using the methods of Zhu et al. ( 2017 ), and using a pH meter (PHSJ-3F, Inesa, China). The surface morphologies and element distribution were determined by using a scanning electronic microscope equipped with energy dispersive spectrometer (SEM-EDS, Gemini 300, ZEISS, Germany). The specific surface areas were calculated by the Brunauer–Emmett–Teller (BET) method. A Fourier transform infrared spectroscope (FTIR, Nicolet 5700, Thermo Fisher Scientific, USA) was used to identify the surface functional groups between different materials in the range of 4000 to 400 cm − 1 . An X-ray photoelectron spectrometer (XPS, K-Alpha+, Thermo Fisher Scientific, USA) with an Al Kα source at 12 kV and 72 W was used to characterize the chemical state of elements and surface structures of ABF before and after the sorption. 2.4 Pot experiment 2.4.1 Experimental design The pot experiment was conducted in a greenhouse at Zhejiang A&F University, China. Plastic pots (diameter 10 cm; height 12 cm) were filled with 1.5 kg of the Cd and As contaminated soil. The BC, FH, FB and SFB were added into the contaminated soil at application rates of 1.0%, 2.0%, 3.0% (dry weight, w/w), and soil without amendments as the control. The experimental treatments were replicated thrice. 2.4.2. Metal(liod)s availability and redistribution Two distinct methodologies were employed to determine the Cd-As concentrations in the soil solution using hydride generation-atomic fluorescence spectrometry (HG-AFS) and inductively coupled plasma-mass spectrometer (ICP-MS). The geochemical fractions of Cd-As were sequentially extracted using the BCR sequential extraction procedure to extract the acid soluble fraction (F1), reducible fraction (F2), oxidizable fraction (F3), and residual fraction (F4) (Ure et al., 1993 ). 2.5 Data quality control and statistical analysis All batch sorption experiments were conducted in triplicate, ensuring a relative standard deviation of triplicate analysis with probability level less than 5%. The plastic ware and glassware used in the experiment and analysis were soaked in 3% nitric acid for 24 h and rinsed with deionized water. The statistical analyses were performed using SPSS (version 26.0, SPSS Inc., Chicago, IL, USA) statistical package. The experimental data were expressed as mean ± standard error (n = 3). The significant differences ( p < 0.05) were evaluated using the analysis of variance (ANOVA) technique. The results were visualized using Origin 2021 (OriginLab Corporation, Northampton, MA, USA) for data graphing. The XPS data were analyzed using the Thermo Avantage program (Thermo Fisher Scientific). 3. Results and discussion 3.1 Characterization of the amendments The morphology and physio-chemical composition of adsorbents are shown in Fig. 1 . The SEM images revealed that BC exhibited visible pores on its plain surface. In contrast, FB displayed a rough surface cover with attachments, and the partial pores were apparently blocked. The SEM-EDS showed a significant increase in iron content ranging from 4.41–12.96% was recorded, confirming that ferrihydrite was successfully loaded onto the biochar surface. The surface of SFB exhibited greater roughness and a more granular structure compared to pristine biochar. The EDS analysis detected a sulfur peak, with sulfur content ranging from 0 to 3.36% on the surface of SFB, confirming successful sulfur loading onto the biochar surface. This might be attributable to the process of modification wherein the sulfur group increased the viscosity between biochar particles and iron particles, resulting in a denser surface coverage. The results of EDS analysis further showed that sulfur and iron were completely saturated on the surface of biochar (SFB) (Zeng et al., 2023 ). 3.2 The sorbent for Cd(II) and As(Ⅲ) sorption 3.2.1. Effect of initial solution pH The sorption performance of adsorbents on heavy metal(liod)s is affected by the pH of the solution, the removal efficiencies of BC, FH, FB and SFB were compared for the removal of Cd(II) and As(III) were evaluated across a pH range of 3.0–9.0. Figure 2 A and 2 B shows the sorption data under different pH range. In the binary solute (Cd-As coexistence) system, the sorption of Cd(II) increased linearly in the range of pH 3.0–5.0, whereas most of the adsorbents showed a slow increasing followed by a slight decreasing trend when the solution pH raised beyond 5, and eventually reached the equilibrium point of sorption at pH value of 7.0. The electrostatic repulsion of high concentration of H + under low pH conditions may lead to the low removal of Cd, that mainly exists as Cd(OH) + ions in aqueous solution and precipitates as Cd(OH) 2 at pH > 8 (Reddy et al., 2014). Therefore, the advantage of Cd-generated precipitation under high pH conditions is the main reason for the high removal rate. The solution pH is an important factor affecting the sorption effect of adsorbent on As, and the existence of As ions in the solution morphology, the charge on the surface of the adsorbent and the degree of dissociation of hydroxyl groups are closely related to it (Smedley et al., 2002). The sorption amount of BC for As did not change significantly, while the rest of the adsorbents showed an increasing trend at pH range of 3–8 and the sorption amount decreased sharply when the pH exceed 8. The effect of pH also showed an increasing trend for As, indicating the existence of a certain co-precipitation between As and Cd. Due to the very limited sorption sites on the adsorbent surface, As usually exists in the form of arsenate ions (H 2 SO 4 − , HASO 4 2− and ASO 4 3− ) in the reaction system. These results showed that the competition between H + and arsenate for the sorption sites became more and more obvious with the increase of OH − content. Therefore, the sorption of As by FH gradually decreased with the increase of pH. FH and As formed adsorptive and specialized sorption through legend exchange reactions (Deng et al., 2018). H 2 AsO 4 − is the main form when the pH is 2–6. In the normal water environment within the pH range between 6 to 9, As is mainly exist in the form of HAsO 4 2− (Chen et al., 2011). At pH > 8, As is mainly in the form of the anion AsO 4 3− . With the increase in pH value, the surface of FH, FB and SFB gradually changed from net positive charge to negative charge and the main form of As changed from HAsO 4 2− to AsO 4 3− , the electrostatic repulsion between FH, FB, SFB and As increased that led to the decrease in its sorption capacity. Therefore, the sorption rates of FH, FB and SFB on As increased and then decreased with the increase of pH value. The removal efficiencies of the amendments for Cd(II) and As(III) increases with the increase of the initial pH, whereby SFB had a significantly higher sorption and removal capacity for Cd and As. 3.2.2. Effect of the sorbent dosage level The effects of adsorbents on the sorption of Cd(II) and As(III) under different dosage conditions are shown in Fig. 2 C and 2 D. The sorption of Cd(II) by BC increased slowly although the increase was not significant with the increase of dosage indicating that BC had no significant effect on the sorption of Cd(II). When the dosage of FH, FB and SFB reached to 1.0 g L − 1 , the sorption of Cd(II) showed a slow upward increase indicating that the sorption of Cd(II) on biochar tended to be dynamically balanced that may be due to the higher pores numbers on the surface of biochar that were completely occupied by Cd(II). The sorption of As(III) by the adsorbent increased sharply at dosage from 0.2 to 1.0 g L − 1 . When the dosage level beyond 1.0 g L − 1 , the sorption capacity of As(III) increased slowly, indicating that the sorption capacity was close to saturation at 1.0 g L − 1 . In addition, the sorption of As by SFB increased sharply when the dosage level increased from 0.2 to 1.5 g L − 1 , and tended to be stable when the dosage level beyond 1.5 g L − 1 , indicating that sorption level reached to saturation. Overall, the dosage of 1.5 g L − 1 is critical limit of SFB when used as a adsorbent for Cd and As. 3.3 Sorption kinetics As shown in Fig. 3 A, Cd sorption increased with contact time and BC reached to equilibrium after 4 h, while the remaining three adsorbents reached equilibrium after 8 h. Similarly, the sorption equilibrium time of four adsorbents for Fig. 3 B (As) was 4 h, that was faster than that of Cd. Ahmad et al. ( 2018 ) reported that sorption can be divided into three phases: an initial fast phase due to physical sorption (P1); a slow rising phase due to chemical sorption (P2); and a final phase of reaching of sorption to equilibrium (P3). The kinetic process of sorption of Cd by adsorbent has the largest proportion of P2 for chemo sorption. This may be due to the presence of As that leads to the combination of Cd and As to form complexes making P2 larger whereas the sorption kinetic process for As, P2 accounted for the major part and was chemo-sorption. The quasi-primary and quasi-secondary kinetic models fitted well with the sorption process, and the parameters of the kinetic equations for the sorption of Cd and As were shown in Fig. 4 , Table. The correlation coefficient (R 2 ) of the secondary kinetic model of the four adsorbents was higher than that of the primary kinetic model, indicating that the sorption of As and Cd mainly occurred through chemical and physical sorption. 3.4 Sorption isotherms In order to further explore the sorption and immobilization mechanism of Cd(II) and As(III) in the composite solution, isothermal sorption experiments were carried out at 25°C, and the results are shown in Fig. 4 A and 4 B. It can be seen that with the increase of the initial concentration of the composite solution, the sorption level of the three adsorbents on Cd(II) and As(III) gradually increased, and finally reached saturation. It is assumed that with the increase of the initial concentration of Cd(II) and As(III) in the solution, the probability of contact between the heavy metal(liod)s and the biochar increased, which is conducive to the adsorption of Cd(II) and As(III) to the active sites on the surface of the biochar material, however, due to the limited number of active sites on the surface of the adsorbent, the process of sorption will be saturated when the initial concentration of the heavy metal(liod)s solution increases to a certain limit. Figure 4 shows the effect of different concentrations on Cd and As adsorption by the sorbent, and the sorption data of Cd and As fitted with both Freundlich and Langmuir models as shown in Fig. 4 A and B. Sorption of Cd by BC, FH, FB and SFB, the Langmuir model fitted better than Freundlich model with R 2 values of 0.91–0.96 and 0.84–0.92, respectively. This indicates that the Cd sorption process after modification of the biochar surface was monolayer sorption instead of multi-layer sorption (Chen et al., 2020 ; Li et al., 2017 ). Moreover, the maximum sorption capacity of SFB (Q m = 76.69 mg kg − 1 ) was higher than that of BC (Q m = 41.36 mg kg − 1 ), and the Cd sorption capacity of SFB was higher than that of BC in the Cd concentration range of 0–80 mg kg − 1 . This result clearly indicates that the Cd sorption capacity of BC was significantly improved after ferrihydrite and sulfur modification. From the correlation coefficients (R 2 ), the sorption of As(III) by FH, FB, and SFB at 298.15 K is more in line with the Freundlich model, suggesting that the sorption behavior of As(III) in these two materials is controlled by multilayered sorption on the inhomogeneous surface (Fu et al., 2021 ). In addition, n > 1 was observed to be favorable for sorption at all temperatures. Comparing to BC, FH and FB, the SFB showed higher sorption capacity for As(III) and Cd(II). The sorption of Cd(II) and As(III) was significantly enhanced by the co-doping of ferrihydrite and S. 3.5 Sorption thermodynamics In order to further investigate the energy changes during the sorption process, sorption thermodynamic studies of Cd(II) and As(III) were carried out at different temperature (i.e., 15°C, 25°C, and 35°C). Figure 5 shows the results of sorption of Cd(II) and As(III) by adsorbent under different temperature conditions. With the increase of temperature, the sorption effect of adsorbent on Cd(II) and As(III) shown a non significant increase (Lin et al., 2017 ), indicating that the temperature did not more effect on the sorption of Cd(II) and As(III). 3.6 Cd(II) and As(III) removal mechanisms 3.6.1. BET analysis BET analysis shown that the specific surface areas of BC, FH, FB and SFB are 64.20, 294.91, 175.35 and 185.16 m 2 g − 1 , respectively (Table. 1). The reason may be that the specific surface of FH is larger than other amendments. The complexation of ferrihydrite on the tunnels or pores of biochar may also lead to a decrease of the specific surface area of FB compared with that of pure FH, as reported by previous studies (Hong et al., 2022 ; Li et al., 2022 ). The N 2 sorption and desorption isotherms of BC, FB and SFB are type IV isotherms morphologically, indicating that BC, FB and SFB mainly includes mesopores or macropores. In contrast, those of FH are indicative of the presence of almost micropores (Huang et al., 2022 ). The isotherm of FB and SFB is of type IV, and includes an H4 hysteresis loop. This result indicates that mesopores or macropores may also be present, despite that the material is dominated by micropores, which also confirms the combination of ferrihydrite and biochar (Li et al., 2020 ). After modification, the S content of pristine biochar significantly increased to 3.36%. Similar results also have been reported by Wu et al. ( 2019 ). which showed that the sulfur content of sulfur modified biochar is approximately 3.34%. After ferrihydrite modification, the surface area and the pore volume from 64.20 m 2 g − 1 and 0.0696 cm 3 g − 1 for BC to 175.35 m 2 g − 1 and 0.1263 cm 3 g − 1 for FB, to 185.16 m 2 g − 1 and 0.1305 cm 3 g − 1 for SFB, increasing by 173.13%-188.41% and 81.47%-87.50%. Thus, after ferrihydrite and sulfur modification biochar significantly increase the surface area and the pore volume. According to Table 1, after ferrihydrite and sulfur modification, the content of S and Fe of biochar increased, indicating that ferrihydrite and sulfur are completely loaded. The C and O content of biochar increased, resulting in an increased polarity with an increase in oxygen-containing functional groups (Ahmad et al., 2014 ; Ennis et al., 2012 ). The increase of O and C content indicates that C on the surface of biochar is oxidized, and oxygen-containing functional groups such as carboxyl group, hydroxyl group and carbonyl group may be formed on the surface (Chen et al., 2020 ). The pH of BC was 11.48, and that of FH was 4.36. The pH of biochar modified with ferrihydrite and sulfur hydrate decreased significantly. Table. 1. Physico-Chemical Characterization of adsorbents Parameter BC FH FB SFB Yield (% dry wt.) 50.12 —— —— —— ash (% dry wt.) 36.61 —— —— —— pH 11.48 4.36 10.86 10.58 C (%) 10.58 0.00 61.43 36.72 O (%) 39.05 34.27 15.37 32.23 S (%) 0.00 0.00 0.00 3.36 Fe (%) 4.41 65.73 12.96 7.98 Surface area(m 2 ·g −1 ) 64.20 294.91 175.35 185.16 Pore volume(cm 3 ·g −1 ) 0.0696 0.1804 0.1263 0.1305 Pore diameter (nm) 4.84 2.46 3.05 2.89 3.6.2. FTIR analyses The functional groups of biochar have significant influence on its sorption capacity. The FTIR spectra of BC before and after modification and after absorption of Cd and As ion were observed and the results are show in Fig. 6 . Two major transmittance peaks at 1590.30 cm − 1 and 1020.15 cm − 1 were detected in the BC sorption system, which may be due to the stretching vibrations of the C = O and C-O-C functional groups on the biochar, respectively (Novais et al., 2018 ; Shi et al., 2019 ; Wang et al., 2018 ; Zhao et al., 2018 ). The intensification of the -OH groups was due to the introduction of oxygen-containing functional groups after the alkali modification, which was accompanied by a shift of the C-O-C stretching vibration from 1590.30 to 1611.19 (Zeng et al., 2023 ). In FH, the transmittance peak at 3363.96 cm − 1 and 576.12 cm − 1 is attributed to Fe-OH and Fe-O, respectively, while the peak at 1623.88 cm − 1 showing the hydration changes on ferrihydrite surfaces (Pawar et al., 2018 ; Zeng et al., 2024 ). The FTIR results of adsorbents before and after sorption of Cd(II) and As(III) in mixed Cd-As system implied the potential binding mechanisms for the Cd(II) and As(III) in the adsorbents. As shown in Fig. 6 , several peaks of BC, FH, FB and SFB shift after sorption, and new peak also appeared. Comparatively, the absorbed Cd-As, reinforcing or decreasing hydrogen bonding and provoking the enhancement or weakening of -OH; the same phenomenon is reported by Lyu et al. ( 2022 ). The peaks of -OH (3363.96 cm − 1 ) was enhanced and shifted to 3371.64 cm − 1 after the sorption of metal(liod)s, suggesting that the oxygen sites and hydroxyl group are the main complexation sites during the process (Jiang et al., 2023 ). The show of the vibration of C = C, and C = O at 1632.19 cm − 1 implied that the biochar or ferrihydrite contributed to the stabilization of Cd(II) (Zeng et al., 2024 ). The vibration of Fe-O bond at 576.12 cm − 1 were shifted to 570.15 cm − 1 , that further indicated the complexation of Cd(II) and As(III) with iron oxides of FH, FB and SFB after the sorption process (Zhang et al., 2020 ). 3.6.3. XPS analyses The elemental composition and oxidation state of the sorbent surface area was determined using XPS spectroscopy. Figure 7 shows the survey spectra of the sorbents before and after Cd and As sorption. Interestingly, the new peaks at 1326, 1362, 412, 405 eV from As/Cd-laden the sorbents were due to electron from the As 2p and Cd 3d levels, respectively, indicating that As and Cd were captured by the sorbent. The homogeneous distribution of Cd and As of the sorbent that sorbents has abundant functional groups, and the distribution mapping of Cd and As revealed metal(iold)s successfully adsorbed. The high-resolution XPS spectra of each element before and after the sorbent sorption are shown in Fig. 7 . The C 1s spectrum of the sorbent can be resolved into three chemical bonds, C-C/C = C i.e., C-OH, C-OOH, with peaks at 284, 285 and 288 eV, respectively (Fig. 7 A and 7 E). After sorption, the C-OH groups showed a slight decrease in intensity and there was a corresponding decreas in C-OOH, C-C/C = C, which may be due to the oxygen-containing functional groups bound to Cd and As, enhancing the molar ratio of the carbon substrate (Lyu et al.,2022). For the mixed Cd-As systems, the relative intensities of the carbon peaks changed little with respect to the case of untreated BC, suggesting that the metal(iold)s binding in the systems mostly occurs in the ferrihydrite region of the composite, casing little change in the BC. The attribution of O 1s were as follows (Fig. 7 B and 7 F): -OH bond at 530 eV, C-O/C = O bond at 532 eV, lattice oxygen inside ferrihydrite, Fe-OH (surface Fe and -OH bond) at 531 eV (Hong et al., 2022 ; Jacukowicz et al., 2020; Lin et al., 2017 ). XPS spectrum of O 1s in SFB, where the peak at 535.33 eV occurred corresponds to Fe-O-H (Alchouron et al., 2020 ). In the Cd-As mixture system, the relative strength of Fe-OH of FH, FB and SFB decreased, suggesting the formation of a ferrihydrite-Cd-As ternary complex, which would promote the removal of Cd and As. This may be due to the interaction of Cd with Fe-OH/-OH and ligand exchange of As with Fe-OH (Xu et al., 2019 ; Zhao et al., 2019 ). The Fe 2p spectra before and after sorption of Cd(II) and As(Ⅲ) on the adsorbent are shown in Fig. 7 C and 7 G. It can be seen that Fe is mainly loaded on the surface of biochar in the form of Fe(II) and Fe(III). Before and after sorption, the occurrence of displacement indicates the involvement of ferrihydrite in the uptake process. The S 2p of XPS spectra are shown in Fig. 7 D and 7 H. The content of sulfur on BC increased after modification. The three peaks of SFB in XPS were located at 164.16, 165.49 and 168.79 eV, which represent C-S, C = S and S(Ⅳ/Ⅱ)-O, respectively (Cato et al., 2018 ; Zhang et al., 2019 ). Park et al. ( 2019 ) reported that the sulfur existence in S-modified wood biochar are mainly in the form of oxidized and thiophene sulfur, which is similar to the XPS results for sulfur. After Cd and As sorption, S(Ⅳ/Ⅱ)-O of SFB decreased from 39.64–23.29%, which indicated that sulfite reacts with Cd(II) and As(Ⅲ) and is consumed during the sorption process. During decomposition of organic matter, oxygen is consumed and sulphate is reduced to sulfite, which may have dissolved Cd oxide to form metal(liod) ions, which react with hydrogen sulfide to form metal(iold) sulfides (Chen et al., 2020 ). Depending on the acid-base reaction, positively charged metal(liod) species, such as Cd(II) or Cd (OH) + , may interact with sulfur groups (Qiao et al., 2023 ). Kubier et al. ( 2019 ) also suggested that sulphate and bisulfide anions are some of the most stable Cd complexes ligands, and final complex formation is related to ligand concentrations. 3.7 Effects of adsorbents on the Cd and As availability and fractionation in soil The concentration of dissolved Cd(II) and As(Ⅲ) significantly decreased under SFB application and changed with the time (Fig. 8 ). The SFB addition resulted in a decline in the concentration of the CaCl 2 -extractable Cd in soil (Fig. 8 A). Compared to the control, significant ( p < 0.05) decreases were observed after the addition of different radio of SFB. Addition of SFB significantly decreased Cd from 0.15 mg kg − 1 in the control to 0.06 mg kg − 1 -0.07 mg kg − 1 in the difference SFB treatment, with a decreasing varied from 32.59–41.26%. The lowest content of the CaCl 2 -extractable Cd was observed in the 3%SFB. The NaH 2 PO 4 -extractable As concentration in soil decreased (Fig. 8 B). Compared to the control, significant ( p < 0.05) decreases were observed after the addition of different radio of SFB. Addition of SFB significantly decreased As from 2.40 mg kg − 1 in the control to 0.86 mg kg − 1 -1.20 mg kg − 1 in the difference SFB treatment, with a decreasing from 50.06–64.06%. The lowest content of the CaCl 2 -extractable Cd was observed in the 3%SFB. The addition of biochar has also altered the distribution fraction of Cd(II) and As(Ⅲ) in the soil (Fig. 8 ). Specifically, the F1 fraction of Cd-As is easily absorbed by plants; however, the F4 fraction of Cd-As is stable and not readily available to plants (Xu et al., 2020 ; Yang et al., 2023 a). The F1, F2 and F3 fraction of Cd decreased in the range of 17.50, -26.66%, 12.86–25.84%, and 37.29–53.60%, respectively; whereas the F4 fraction of Cd increased by 53.65to -87.89% in the SFB treatments, compared to Control (Fig. 8 C). Under 1.0%, 2.0%, 3.0% SFB application, the percentage of F1, F2 and F3 fraction of As decreased by 33.47–38.77%, 36.08–45.90% and 7.53–28.87%, respectively, while F4 fraction of As increased by 33.55 to 41.86%, compared to Control (Fig. 8 D). These results showed that the absorbents transferred Cd(II) and As(Ⅲ) from relatively labile fractions to less toxic and more stable fractions, thereby reduced the bioavailability of Cd in soil, particularly in the 3% SFB that was highly effective compared to other treatments. These results dictate that the SFB was more effective in immobilizing Cd(II) and As(Ⅲ) compared to control. SFB mainly decreased the F1, F2 and F3 fractions of soil Cd(II) and As(Ⅲ) but increased the more stable fraction of Cd(II) and As(Ⅲ) (F4), facilitating the transformation of Cd(II) and As(Ⅲ) chemical speciation and changing Cd(II) and As(Ⅲ) bioavailability (Sun et al., 2024 ). 4. Conclusion Sulfur-ferrihydrite-biochar composites were developed for the simultaneous removal of Cd and As. The SFB exhibited a significant increase in surface area, pore structure, and functional groups, which enhanced its capacity for Cd and As sorption. The composite material demonstrated strong sorption properties across a wide pH range in both binary systems. The primary removal mechanisms include electrostatic attraction, ion exchange, precipitation and formation of various complexes. In Cd-As mixture systems, Cd and As primarily form tri/quaternary complexes with S/FH, facilitating the removal of both metals(iod)s. The presence of As can promote the sorption of Cd on the sulfur-ferrihydrite-biochar composite material. The enhanced sorption of As may be due to several possible mechanisms, including the change of surface charge of the composite material due to the presence of As and the increase of the dispersion of biochar on ferrihydrite. The introduction of SFB led to reductions in Cd and As concentrations by 34.98% and 78.27%, respectively, in water and soil, indicating its effectiveness for remediating Cd and As co-contaminated soil. The study also showed that the sorption kinetics and potential binding mechanisms of heavy metal(liod)s, particularly in systems where cations and anions coexist. The sulfur-ferrihydrite-biochar composite can reduce environmental risks associated with Cd and As contamination by immobilizing them in the solid phase, and considered as a promising material for soil and water remediation. Considering the prevalence of precipitation and transformation processes of ferrihydrite and S in natural environments, special attention needs to be paid to the interactions between hydrated Fe, S, heavy metal(liod)s and environmental factors. Declarations Funding This work was supported by the Natural Science Foundation of China [grant number 42407016, 32271532], Research and Development Fund of Zhejiang A & F University [grant number 2020LFR052], and Young Scientists Fund of the National Natural Science Foundation of China [grant number 42407016]. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Data availability Data will be made available on request. CRediT authorship contribution statement Xuqiao Wu: Conceptualization, Formal analysis, Writing – original draft. Xiaowen Teng: Data curation, Writing – original draft, Software. Dong Huang: Formal analysis, Visualization. Ijlal Ahmad: Writing – review & editing. Hanbo Chen: Supervision. Yaqian Li: Software. Dubin Dong: Methodology. Yanxin Tang: Validation, Visualization. Yini Wang: Investigation. Song Li: Supervision. 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Three-dimensional porous graphene oxide-maize amylopectin composites with controllable pore-sizes and good adsorption-desorption properties: Facile fabrication and reutilization, and the adsorption mechanism. Ecotoxicol. Environ. Saf., 176 , 11-19. Zhi, M., Liu, S., Hong, Z., Wu, N. (2014). Electrospun activated carbon nanofibers for supercapacitor electrodes. RSC Adv., 4(82) , 43619-43623. Zhou, S, Liu, Z., Sun, G., Zhang, Q., Cao, M., Tu, S., Xiong, S. (2022). Simultaneous reduction in cadmium and arsenic accumulation in rice ( Oryza sativa L. ) by iron/iron-manganese modified sepiolite. Sci. Total Environ., 810 , 152189. Zhu, F., Hou, J., Xue, S., Wu, C., Wang, Q., Hartley, W. (2017). Vermicompost and gypsum amendments improve aggregate formation in bauxite residue. Land Degrad. Dev., 28(7) , 2109-2120. Zou, R., Qian, M., Wang, C., Mateo, W., Wang, Y., Dai, L., Lin, X., Zhao, Y., Huo, E., Wang, L., Zhang, X., Kong, X., Ruan, R., Lei, H. (2022). Biochar: From by-products of agro-industrial lignocellulosic waste to tailored carbon-based catalysts for biomass thermochemical conversions. Chem. Eng. J., 441 , 135972. Additional Declarations No competing interests reported. Cite Share Download PDF Status: Posted Version 1 posted You are reading this latest preprint version Research Square lets you share your work early, gain feedback from the community, and start making changes to your manuscript prior to peer review in a journal. As a division of Research Square Company, we’re committed to making research communication faster, fairer, and more useful. We do this by developing innovative software and high quality services for the global research community. Our growing team is made up of researchers and industry professionals working together to solve the most critical problems facing scientific publishing. Also discoverable on Platform About Our Team In Review Editorial Policies Advisory Board Help Center Resources Author Services Accessibility API Access RSS feed Manage Cookie Preferences © Research Square 2026 | ISSN 2693-5015 (online) Privacy Policy Terms of Service Do Not Sell My Personal Information {"props":{"pageProps":{"initialData":{"identity":"rs-6195683","acceptedTermsAndConditions":true,"allowDirectSubmit":true,"archivedVersions":[],"articleType":"Research Article","associatedPublications":[],"authors":[{"id":428003585,"identity":"0bf85f90-20d0-4a63-be60-f00df8921770","order_by":0,"name":"Xuqiao Wu","email":"","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":false,"prefix":"","firstName":"Xuqiao","middleName":"","lastName":"Wu","suffix":""},{"id":428003586,"identity":"db46089b-36be-4af8-93d8-81da8f76ab2c","order_by":1,"name":"Xiaowen Teng","email":"","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":false,"prefix":"","firstName":"Xiaowen","middleName":"","lastName":"Teng","suffix":""},{"id":428003587,"identity":"886ed713-6420-45dd-9a9c-bd7951cf1e41","order_by":2,"name":"Dong Huang","email":"","orcid":"","institution":"Pujiang County Ecological Civilization Promotion Center","correspondingAuthor":false,"prefix":"","firstName":"Dong","middleName":"","lastName":"Huang","suffix":""},{"id":428003588,"identity":"80d5b35e-3d66-4b9f-a826-54695813b6d2","order_by":3,"name":"Ijlal Ahmad","email":"","orcid":"","institution":"The University of Agriculture","correspondingAuthor":false,"prefix":"","firstName":"Ijlal","middleName":"","lastName":"Ahmad","suffix":""},{"id":428003591,"identity":"653526e9-c2d0-4560-88d5-c6d67c4bf845","order_by":4,"name":"Hanbo Chen","email":"","orcid":"","institution":"Zhejiang University of Science \u0026 Technology","correspondingAuthor":false,"prefix":"","firstName":"Hanbo","middleName":"","lastName":"Chen","suffix":""},{"id":428003592,"identity":"d1331d86-72da-4074-9e1e-6a7482ed617d","order_by":5,"name":"Yaqian Li","email":"","orcid":"","institution":"People's Government of Yanguan Town","correspondingAuthor":false,"prefix":"","firstName":"Yaqian","middleName":"","lastName":"Li","suffix":""},{"id":428003593,"identity":"48258faa-e952-4621-b15e-0944b8fb90d7","order_by":6,"name":"Dubin Dong","email":"","orcid":"","institution":"Central South University of Forestry and Technology","correspondingAuthor":false,"prefix":"","firstName":"Dubin","middleName":"","lastName":"Dong","suffix":""},{"id":428003594,"identity":"c21b0ec8-1398-410a-87e6-10fce773f15b","order_by":7,"name":"Yanxin Tang","email":"","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":false,"prefix":"","firstName":"Yanxin","middleName":"","lastName":"Tang","suffix":""},{"id":428003595,"identity":"fb6982f0-0934-46b9-a1bf-42809e6ac9a8","order_by":8,"name":"Yini Wang","email":"","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":false,"prefix":"","firstName":"Yini","middleName":"","lastName":"Wang","suffix":""},{"id":428003596,"identity":"b8908b1b-b9d3-446e-8d94-1bfaab10e3ed","order_by":9,"name":"Song Li","email":"","orcid":"","institution":"Jiangsu Vocational College of Agriculture and Forestry","correspondingAuthor":false,"prefix":"","firstName":"Song","middleName":"","lastName":"Li","suffix":""},{"id":428003597,"identity":"c83abe0b-37eb-415c-bbaf-ab02f4eff712","order_by":10,"name":"Dan Liu","email":"","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":false,"prefix":"","firstName":"Dan","middleName":"","lastName":"Liu","suffix":""},{"id":428003598,"identity":"31e7aa85-a585-4e93-b0e6-885b740e1220","order_by":11,"name":"Weijie Xu","email":"data:image/png;base64,iVBORw0KGgoAAAANSUhEUgAAAZAAAAAyAQMAAABI0h/eAAAABlBMVEX///8AAABVwtN+AAAACXBIWXMAAA7EAAAOxAGVKw4bAAAA30lEQVRIiWNgGAWjYLCCBB4JBn4g/QGIGRuI1iLZwMA4g3gtIGBwgFgtBjdyDB88kLGQMz5/+GEzD4ON7IYDzM8e4NMiOSPH2ADoMGOzA8cMgVrSjDccYDM3wKeFXyJ3mwRQS+K2gw3mj3kYDiduOMDDJoFPC5tE7vYfQC31m5vZPwJt+U9YC8gWUIglGLDxgBx2gLAWyZ73n0EOM5xxhqewcY5BsvHMw2xmeLUYHE9L/Pizp06ev//4xoY3FXayfcebn+HVAgaMPXATgJiZoHoQ+EGUqlEwCkbBKBipAACxUkWKZMrGgwAAAABJRU5ErkJggg==","orcid":"","institution":"Zhejiang A\u0026F University","correspondingAuthor":true,"prefix":"","firstName":"Weijie","middleName":"","lastName":"Xu","suffix":""}],"badges":[],"createdAt":"2025-03-10 12:53:10","currentVersionCode":1,"declarations":"","doi":"10.21203/rs.3.rs-6195683/v1","doiUrl":"https://doi.org/10.21203/rs.3.rs-6195683/v1","draftVersion":[],"editorialEvents":[],"editorialNote":"","failedWorkflow":false,"files":[{"id":78644143,"identity":"55b7f073-0d79-4578-a064-f89419877f92","added_by":"auto","created_at":"2025-03-17 07:22:44","extension":"png","order_by":1,"title":"Figure 1","display":"","copyAsset":false,"role":"figure","size":643011,"visible":true,"origin":"","legend":"\u003cp\u003eSEM-EDS and SEM-mapping of adsorbents. Treatments: biochar (BC), Ferrihydrite (FH), Biochar loaded with ferrihydrite (FB) and sulfur modified biochar loaded with ferrihydrite (SFB). Data points represent mean ± S.D. (n = 3).\u003c/p\u003e","description":"","filename":"1.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/7df5a4c9b53f9445b7280806.png"},{"id":78643880,"identity":"05cc90fb-7fbe-4c5a-a2a4-7f93be32c857","added_by":"auto","created_at":"2025-03-17 07:14:44","extension":"png","order_by":2,"title":"Figure 2","display":"","copyAsset":false,"role":"figure","size":137662,"visible":true,"origin":"","legend":"\u003cp\u003eEffect of pH (A; B) and addition of adsorbents dosage (C; D) on Cd (II) and As(III) sorption to the sorbent. Data points represent mean ± S.D (n = 3).\u003c/p\u003e","description":"","filename":"2.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/540984def571267ba1243f1d.png"},{"id":78643885,"identity":"67ffd06d-6907-4a17-b516-3902fa7cdca4","added_by":"auto","created_at":"2025-03-17 07:14:44","extension":"png","order_by":3,"title":"Figure 3","display":"","copyAsset":false,"role":"figure","size":136688,"visible":true,"origin":"","legend":"\u003cp\u003eThe effect of time for the sorption of Cd(II) and As(III) and the sorption kinetic modeling using pseudo first-order and pseudo second-order. Data points represent mean ± S.D (n = 3). (BC: biochar; FH: Ferrihydrite, FB: Biochar loaded with ferrihydrite; SFB: sulfur modified biochar loaded with ferrihydrite).\u003c/p\u003e","description":"","filename":"3.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/35d629620f19f9327cd24591.png"},{"id":78643878,"identity":"bca21128-bcfa-4a9a-b877-ce727e6e6ed0","added_by":"auto","created_at":"2025-03-17 07:14:44","extension":"png","order_by":4,"title":"Figure 4","display":"","copyAsset":false,"role":"figure","size":185976,"visible":true,"origin":"","legend":"\u003cp\u003eEffect of initial concentration on Cd(II) and As(III) sorption to the adsorbents and Isothermal parameters of Cd(II) and As(III) sorption on the adsorbents. Data points represent mean ± S.D (n = 3). (BC: biochar; FH: Ferrihydrite, FB: Biochar loaded with ferrihydrite; SFB: sulfur modified biochar loaded with ferrihydrite).\u003c/p\u003e","description":"","filename":"4.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/2eb40068e68698ac5acca1eb.png"},{"id":78646003,"identity":"fc8c9598-ca64-45d0-9b94-3d799396c33e","added_by":"auto","created_at":"2025-03-17 07:38:47","extension":"png","order_by":5,"title":"Figure 5","display":"","copyAsset":false,"role":"figure","size":47167,"visible":true,"origin":"","legend":"\u003cp\u003esorption of Cd(II) and As(III) by the sorbent at different temperatures. Cd(II) in the composite solution by the sorbent at 288.15 K (A), 298.15 K (B) and 308.15K (C); As(III) in the composite solution by the sorbent at 288.15 K (D), 298.15 K (E) and 308.15K (F).Data points represent mean ± S.D. (n = 3).\u003c/p\u003e","description":"","filename":"5.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/530b67204fa0e0f8ba9dafe4.png"},{"id":78644144,"identity":"42cd0f90-5e83-45c9-9457-08fcaa99e97b","added_by":"auto","created_at":"2025-03-17 07:22:44","extension":"png","order_by":6,"title":"Figure 6","display":"","copyAsset":false,"role":"figure","size":110886,"visible":true,"origin":"","legend":"\u003cp\u003eFTIR spectra of adsorbents before and after Cd(II) and As(III) sorption. Data points represent mean ± S.D. (n = 3).(BC: biochar; FH: Ferrihydrite, FB: Biochar loaded with ferrihydrite; SFB: sulfur modified biochar loaded with ferrihydrite).\u003c/p\u003e","description":"","filename":"6.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/6f0a5e64e59ab8a7cc1ddbfd.png"},{"id":78644859,"identity":"a4f7e048-fcde-470b-92bb-0c36b70f281d","added_by":"auto","created_at":"2025-03-17 07:30:44","extension":"png","order_by":7,"title":"Figure 7","display":"","copyAsset":false,"role":"figure","size":729070,"visible":true,"origin":"","legend":"\u003cp\u003eXPS spectra of adsorbents before (A; B; C; D) and after (E; F; G; H) Cd(II) and As(III).. Data points represent mean ± S.D. (n = 3).(BC: biochar; FH: Ferrihydrite, FB: Biochar loaded with ferrihydrite; SFB: sulfur modified biochar loaded with ferrihydrite).\u003c/p\u003e","description":"","filename":"7.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/0f4df2b11e24be054df8e2b2.png"},{"id":78643896,"identity":"9d3e240b-6a17-482b-9a20-821932fff247","added_by":"auto","created_at":"2025-03-17 07:14:44","extension":"png","order_by":8,"title":"Figure 8","display":"","copyAsset":false,"role":"figure","size":75139,"visible":true,"origin":"","legend":"\u003cp\u003eThe fraction of Cd(II) (C) and As(III) (D) in soil after amended and Available Cd(II) (A) and As(III) (B) content of soil after amended by SFB. Data points and error bars represent mean ± S.D. (n = 3).Different letters indicate a significant difference (\u003cem\u003ep\u003c/em\u003e\u0026lt; 0.05).\u003c/p\u003e","description":"","filename":"8.png","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/4056c2398ca5943012c70bd1.png"},{"id":79036243,"identity":"84d67a07-9fe5-4297-8f88-58896d6d1110","added_by":"auto","created_at":"2025-03-23 08:31:34","extension":"pdf","order_by":0,"title":"","display":"","copyAsset":false,"role":"manuscript-pdf","size":3129959,"visible":true,"origin":"","legend":"","description":"","filename":"manuscript.pdf","url":"https://assets-eu.researchsquare.com/files/rs-6195683/v1/61e46dee-6566-40be-ba86-4b6ddc2b3e75.pdf"}],"financialInterests":"No competing interests reported.","formattedTitle":"\u003cp\u003eReduction and immobilization of Cd(II) and As(III) using sulfur-ferrihydrite-biochar as an amendment in water and soil: Investigation of the Mechanism of Remediation\u003c/p\u003e","fulltext":[{"header":"1. Introduction","content":"\u003cp\u003eHazardous toxic metal(liod)s, such as cadmium (Cd) and arsenic (As), primarily originate from anthropogenic activities and the weathering of soil-forming rocks and mineral posing serious threats to plants and human health (Lyu et al., \u003cspan citationid=\"CR31\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Cd and As are recognized highly toxic and mobile metals that are taken up by the plants in higher concentration when presents in soil and if these plants are consumed by human or animals posing serious health hazards. Chronic exposure to these metalloids can cause irreversible damage to the kidneys, bones, and nervous system (Adnan et al., \u003cspan citationid=\"CR1\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). In China, over 2.0\u0026times;10\u003csup\u003e5\u003c/sup\u003e km\u003csup\u003e2\u003c/sup\u003e cultivated land is contaminated by Cd and As, particularly in the southern regions (Sun et al., \u003cspan citationid=\"CR44\" class=\"CitationRef\"\u003e2024\u003c/span\u003e, 33; Gong et al, \u003cspan citationid=\"CR14\" class=\"CitationRef\"\u003e2020\u003c/span\u003e). However, the remediation of those contaminated soils with Cd and As remains highly challenging due to their complex chemical behavior and mobile properties in soil (Qiao et al., \u003cspan citationid=\"CR37\" class=\"CitationRef\"\u003e2018\u003c/span\u003e; Zhou et al., \u003cspan citationid=\"CR67\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Specifically, Cd exists in the cationic form, whereas As generally occurs in anionic forms, which resulting in the difficult for a single remediation agent to simultaneously immobilize both Cd and As in soil (Shen et al., \u003cspan citationid=\"CR42\" class=\"CitationRef\"\u003e2020\u003c/span\u003e; Vankova et al., \u003cspan citationid=\"CR48\" class=\"CitationRef\"\u003e2021\u003c/span\u003e; Zhou et al., \u003cspan citationid=\"CR67\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Moreover, their co-existence in the environment may lead to complex interactions, such as competition for sorption and solubility changes that may reduce their remediation efficiency (Huang et al., \u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Thus, it is essential to seek an effective remediation strategy to simultaneously immobilize Cd and As in contaminated soil, while elucidating the mechanisms underlying their sorption and stabilization.\u003c/p\u003e \u003cp\u003eIn situ stabilization to make it non available to plants is one of the most widely used and proven technique for remediating contaminated soils. Amendments such as ferrihydrite and sulfur (S) materials have shown great potential in reducing the mobility and bioavailability of heavy metal(liod) (Qu et al., \u003cspan citationid=\"CR39\" class=\"CitationRef\"\u003e2022a\u003c/span\u003e; Liu et al., \u003cspan citationid=\"CR29\" class=\"CitationRef\"\u003e2015\u003c/span\u003e; Palansooriya et al., \u003cspan citationid=\"CR34\" class=\"CitationRef\"\u003e2020\u003c/span\u003e; Zeng et al., \u003cspan citationid=\"CR60\" class=\"CitationRef\"\u003e2024\u003c/span\u003e). Moreover, these amendments are generally effective in immobilizing cationic metals such as Cd by forming stable precipitates or minerals, while they fail to stabilize As in their co-contaminated soil due to its anionic nature (Lee et al., \u003cspan citationid=\"CR23\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). Biochar a product of the anaerobic pyrolysis of agricultural, industrial, and household waste (Zou et al., \u003cspan citationid=\"CR69\" class=\"CitationRef\"\u003e2022\u003c/span\u003e), is widely employed one of the organic amendments in soil remediation due to its high porosity, abundant functional groups and strong ion exchange capacity (Yang et al., \u003cspan citationid=\"CR57\" class=\"CitationRef\"\u003e2021\u003c/span\u003e). These properties make biochar highly effective immobilizing agent for cationic pollutants, such as Cd, through a mechanisms including cation exchange, electron donation, and chelation. However, the negative charge of biochar surface limits its capacity to adsorb anionic pollutants such as As, due to electrostatic repulsion (Tan et al., \u003cspan citationid=\"CR45\" class=\"CitationRef\"\u003e2017\u003c/span\u003e; Zhi et al., \u003cspan citationid=\"CR66\" class=\"CitationRef\"\u003e2014\u003c/span\u003e). Therefore, chemical modification of biochar through integration with ferrihydrite is an innovative strategy to enhance its performance in soils contaminated with both Cd and As.\u003c/p\u003e \u003cp\u003eFerrihydrite, a precursor of crystalline iron oxides is found in natural environments as nanoparticle colloids (Wang et al., \u003cspan citationid=\"CR50\" class=\"CitationRef\"\u003e2020\u003c/span\u003e). It has high surface area with strong binding capacity, and ability to form inner-sphere complexes that make ferrihydrite a powerful immobilizer of both cations and anions including Cd and As (Cui et al., \u003cspan citationid=\"CR9\" class=\"CitationRef\"\u003e2022\u003c/span\u003e; Fu et al., \u003cspan citationid=\"CR13\" class=\"CitationRef\"\u003e2021\u003c/span\u003e; Hong et al., \u003cspan citationid=\"CR15\" class=\"CitationRef\"\u003e2022\u003c/span\u003e; Xiu et al., \u003cspan citationid=\"CR52\" class=\"CitationRef\"\u003e2018\u003c/span\u003e). Moreover, ferrihydrite exhibits a strong tendency for anions complexation via mechanisms, such as formation of arsenate oxyanions, inner-sphere sorption, co-precipitation/incorporation, and isomorphous substitution (Hu et al., \u003cspan citationid=\"CR17\" class=\"CitationRef\"\u003e2018\u003c/span\u003e; Hu et al., \u003cspan citationid=\"CR16\" class=\"CitationRef\"\u003e2020\u003c/span\u003e). Preliminary studies also have indicated that ferrihydrite had effectively stabilized Cd and As in the solid phase, thereby reduced their mobility and bioavailability (Fan et al., \u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2021\u003c/span\u003e; Jeong et al.,2017; Ouyang et al.,2021). However, the tendency of anhydrous ferrous iron in ferrihydrite significantly reduces the efficiency of metal(liod) sorption and partitioning (Tian et al., \u003cspan citationid=\"CR46\" class=\"CitationRef\"\u003e2018\u003c/span\u003e). Several research studies ave suggested that integration of ferrihydrite with biochar can mitigate agglomeration, enhancing the availability of active binding sites within biochar porous structure and improving the immobilization of Cd and As (Bai et al., \u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2022\u003c/span\u003e; Du et al., \u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2021\u003c/span\u003e).\u003c/p\u003e \u003cp\u003eSulfur modifications offer another promising avenue for enhancing immobilization and reducing bioavailability of heavy metal(liod)s in soils. Sulfur plays a vital role in plant responses to heavy metal(liod)s stress by contributing to thiol-rich peptides, such as phytochelatins, that bind heavy metal(liod)s and mitigate their toxicity. Iron-sulfur modified biochar could serve as an effective strategy for improving the remediation efficiency of soils co-contaminated with As and Cd (Wu et al., \u003cspan citationid=\"CR51\" class=\"CitationRef\"\u003e2019\u003c/span\u003e). The large surface area, high reactivity, and amphoteric properties of iron oxides make them highly suitable for the chemical immobilization of various soil pollutants, as iron-based materials effectively remove heavy metal(liod)s via ion exchange and precipitation (Wang et al., \u003cspan citationid=\"CR49\" class=\"CitationRef\"\u003e2018\u003c/span\u003e). Furthermore, the addition of sulfur can enhance the reduction and precipitation of heavy metal(liod)s through the provision of reducing species (S\u0026sup2;⁻) (Yang et al., \u003cspan citationid=\"CR55\" class=\"CitationRef\"\u003e2019\u003c/span\u003e). Rajendran et al. (\u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2019\u003c/span\u003e) reported that sulfur enrichment onto iron-based biochar has effectively reduced Cd content in soil and lowered Cd accumulation in rice tissues. Yin et al. (\u003cspan citationid=\"CR58\" class=\"CitationRef\"\u003e2019\u003c/span\u003e) found that sulfur-doped Fe\u003csub\u003e3\u003c/sub\u003eO\u003csub\u003e4\u003c/sub\u003e nanoparticles exhibited excellent sorption performance for As(V) because sulfur doping activates Fe atoms, creating sorption sites. Qiao et al. (2020) reported that sulfide-modified α-FeOOH increased the sorption capacity of As (V) from 153.8 to 384.6 mg g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e compared to pristine α-FeOOH. Overall, iron-sulfur-modified biochar shows significant potential for enhancing the simultaneous remediation of Cd and As in contaminated soils. However, the mechanisms underlying its dual immobilization capabilities require further investigation.\u003c/p\u003e \u003cp\u003eTherefore, in this study, an iron-sulfur-modified biochar was prepared to investigate the efficiency of its immobilization potential and mechanisms of stabilization of Cd(II) and As(III) in soil through batch sorption experiment. The specific objectives of this study were: (1) to evaluate the sorption behavior and mechanisms of the sulfur-ferrihydrite-biochar amendment for immobilization of Cd(II) and As(III); (2) To elucidate the binding mechanism of Cd(II) and As(III) on sulfur-ferrihydrite-biochar composites. This study provided an in-depth understanding of mechanism into the interfacial interaction between Cd(II), As(III), and sulfur-ferrihydrite-biochar composites, particularly in co-contaminated systems. The findings will help in advancing remediation strategies and improving the geochemical reaction understanding of Cd(II) and As(III) in natural environments.\u003c/p\u003e"},{"header":"2. Materials and methods","content":"\u003cdiv id=\"Sec3\" class=\"Section2\"\u003e \u003ch2\u003e2.1 Materials\u003c/h2\u003e \u003cp\u003eRice straw was collected from a farm located in Taizhou, Zhejiang Province, China, and was used as feedstock to produce pristine biochar. Fe(NO\u003csub\u003e3\u003c/sub\u003e)\u003csub\u003e3\u003c/sub\u003e\u0026middot;9H\u003csub\u003e2\u003c/sub\u003eO, SC\u003csub\u003e2\u003c/sub\u003e and NaOH (all AR grade) were selected to modify the biochar NaAsO\u003csub\u003e2\u003c/sub\u003e and CdCl\u003csub\u003e2\u003c/sub\u003e (both AR grade) were used to prepare the stock solutions for the batch sorption experiment. All reagents were purchased from Reagent Co. Ltd (China), and all the solutions were prepared by using ultrapure water.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec4\" class=\"Section2\"\u003e \u003ch2\u003e2.2 Preparation and modification of biochar.\u003c/h2\u003e \u003cdiv id=\"Sec5\" class=\"Section3\"\u003e \u003ch2\u003e2.2.1 Biochar preparation\u003c/h2\u003e \u003cp\u003eRice straw was air-dried and filtered, sieved through a 5 mm sieve. Biochar was produced by slow pyrolysis of rice straw at 550℃ for 2 h with a heating rate of 5℃ min\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, under limited oxygen supply in a muffle furnace. After cooling to room temperature, the biochar was washed with the ultrapure water and filtered through a 100-mesh sieve. The sieved biochar was collected and was referred to as BC.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec6\" class=\"Section3\"\u003e \u003ch2\u003e2.2.2 Biochar modification\u003c/h2\u003e \u003cp\u003eThe synthetic procedure of ferrihydrite was used as described by Du et al. (\u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2021\u003c/span\u003e). A 2-line ferrihydrite (FH) was prepared by adding 1.5 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e NaOH drop wise to a 0.1 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e Fe(NO\u003csub\u003e3\u003c/sub\u003e)\u003csub\u003e3\u003c/sub\u003e solution until the pH reached 7.0, and was allowed to stand for 1 h, afterward the supernatant was decanted. The precipitate was separated by centrifugation at 4500 rpm, washed twice with ultrapure water and dialyzed for three days in darkness to remove excess Na\u003csup\u003e+\u003c/sup\u003e and NO\u003csup\u003e3\u0026minus;\u003c/sup\u003e. The residue was dried at 60\u0026deg;C in a thermostatic drying oven for 8 h, After cooling at room temperature, the residues was passed through a 100-mesh sieve. The resulting ferrihydrite was collected and referred to as FH.\u003c/p\u003e \u003cp\u003eBiochar-loaded ferrihydrite was prepared at a 2: 1 mass ratio of ferrihydrite to biochar. The biochar was added to 1 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e KOH solution and stirred at 60\u0026deg;C for 72 h. The mixture was then kept for 24 h and then centrifuged at 5000 rpm. The precipitate was washed thoroughly with ultrapure water and frozen overnight at -20\u0026deg;C, and put into freeze-dried to complete dryness. The resulting material was designated as FB.\u003c/p\u003e \u003cp\u003eThe FB was grounded and passed through a 0.15 mm sieve. Sulfur addition was performed by mixing 100 mL of 0.40 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e NaOH solution with carbon disulfide (CS\u003csub\u003e2\u003c/sub\u003e) at a ratio of 1: 1.5. The mixture was stirred for 4 h at 25\u0026deg;C and sonicated for 1 hour to obtain a sulfur modified solution. The solution was further stirred for 16 h at 40\u0026deg;C using a magnetic stirrer, then cooled to room temperature and filtered through 0.45 \u0026micro;m filters. A 10 g of the prepared biochar was added to 100 mL sulfur modified mixture solution and stirred for 8 h at 45\u0026deg;C with a magnetic stirrer. The mixture was dried, filtered and passed through a 100-mesh sieve, and designated the product as SFB.\u003c/p\u003e \u003c/div\u003e \u003c/div\u003e \u003cdiv id=\"Sec7\" class=\"Section2\"\u003e \u003ch2\u003e2.3 Batch sorption experiments\u003c/h2\u003e \u003cp\u003eA series of batch experiments were conducted at 25℃ to investigate the sorption characteristic of Cd(II), and As(III) by adding BC, FH, FB and SFB. All solutions used in batch tests contained 0.1 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e NaOH and 0.1 mol L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e HCl as ionic strength adjuster, and the experiments were performed in triplicates. The solution concentrations of Cd(II) and As(III) before and after sorption were determined through inductively coupled plasma optical emission spectrometry (ICPOES, 720 ES, Agilent, USA).\u003c/p\u003e \u003cp\u003eThe adsorbents were all used at the rate of 1.5 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. Each set of experiments had three replicates. The oscillation condition was set at 180 rpm, 25℃, for 24 h. The optimum pH for sorption was determined in the dual pollutant system (5mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e As(III) solution and 40mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e Cd(II) solution) with pH settings of 2, 3, 4, 5, 6, 7, 8 and 9.\u003c/p\u003e \u003cp\u003eFor the amendments dosage experiments, solutions were extracted at different dosages (0.2, 0.5, 1, 1.5 and 2g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e) in a solution containing 40 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e Cd(II) and 5 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e As(III). For kinetic experiment, solutions were extracted at different intervals (5, 10, 30, 60, 120, 240, 360, 540, 720, 960, 1200 and 1440 min) in the solution containing 40 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e Cd(II) and 5 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e As(III). The sorption isotherms were studied for Cd(II) and As(III) mixed solutions at concentrations of 3.2\u0026ndash;80 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 0.4\u0026ndash;10 mg L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. For thermodynamics, the sorption of Cd(II) and As(III) at 35℃, 25℃ and 15℃ was also investigated. All samples were filtered through 0.45 \u0026micro;m microporous membranes, acidified, and determined concentrations on an inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7000, Japan).\u003c/p\u003e \u003cdiv id=\"Sec8\" class=\"Section3\"\u003e \u003ch2\u003e2.3.2 Modeling for metal(liod)s sorption\u003c/h2\u003e \u003cp\u003e(1) To investigate the change of binding rate between Cd(II) and As(III) and adsorbents with time, kinetic investigations were performed. The uptake results were then fitted by pseudo-first (Formula 1) and second-order (Formula 2), and intra-particle diffusion models for describing the kinetic course of Cd(II) and As(III)adsorption by adsorbents quantitatively (Bai et al., \u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2022\u003c/span\u003e):\u003c/p\u003e \u003cp\u003e \u003cspan class=\"InlineEquation\"\u003e \u003cspan class=\"mathinline\"\u003e\\(\\:{Q}_{t}={Q}_{e}(1-{e}^{-k1t})\\)\u003c/span\u003e \u003c/span\u003e (Formula 1)\u003c/p\u003e \u003cp\u003ewhere, Q\u003csub\u003et\u003c/sub\u003e: sorption capacity at time t (mg g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e); Q\u003csub\u003ee\u003c/sub\u003e: Equilibrium sorption capacity (mg g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e); t: Time (min); k\u003csub\u003e1\u003c/sub\u003e: Pseudo-first-order rate constant.\u003c/p\u003e \u003cp\u003e \u003cspan class=\"InlineEquation\"\u003e \u003cspan class=\"mathinline\"\u003e\\(\\:\\frac{t}{{q}_{t}}=\\frac{1}{{k}_{2}{q}_{e}^{2}}+\\frac{t}{{q}_{e}}\\)\u003c/span\u003e \u003c/span\u003e (Formula 2)\u003c/p\u003e \u003cp\u003ewhere, Q\u003csub\u003et\u003c/sub\u003e: sorption capacity at time t (mg g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e); Q\u003csub\u003ee\u003c/sub\u003e: Equilibrium sorption capacity (mg g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e); t: Time (min); k\u003csub\u003e2\u003c/sub\u003e: Pseudo-second-order rate constant.\u003c/p\u003e \u003cp\u003e(2) Isotherm rate were then performed at different temperatures by further elucidating the Cd(II) and As(III) uptake behaviors with adsorbents. Two recognized isotherm models, including the Langmuir sorption isotherm (Formula 3) and Freundlich parameters (Formula 4) equation is given by Bai et al. (\u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2022\u003c/span\u003e):\u003c/p\u003e \u003cp\u003e \u003cspan class=\"InlineEquation\"\u003e \u003cspan class=\"mathinline\"\u003e\\(\\:{Q}_{e}=\\frac{{Q}_{m}{bC}_{e}}{1+{bC}_{e}}\\)\u003c/span\u003e \u003c/span\u003e (Formula 3)\u003c/p\u003e \u003cp\u003ewhere, Q\u003csub\u003ee\u003c/sub\u003e: Equilibrium sorption capacity (mg g\u003csup\u003e-1\u003c/sup\u003e); Q\u003csub\u003em\u003c/sub\u003e: Maximum sorption capacity (mg g\u003csup\u003e-1\u003c/sup\u003e); b: Langmuir constant related to the affinity of binding sites (L mg\u003csup\u003e-1\u003c/sup\u003e); C\u003csub\u003ee\u003c/sub\u003e: Equilibrium concentration of the sorption in the solution (mg L\u003csup\u003e-1\u003c/sup\u003e).\u003c/p\u003e \u003cp\u003e \u003cspan class=\"InlineEquation\"\u003e \u003cspan class=\"mathinline\"\u003e\\(\\:{Q}_{e}={k}_{f}{C}_{e}^{1/n}\\)\u003c/span\u003e \u003c/span\u003e (Formula 4)\u003c/p\u003e \u003cp\u003ewhere, Q\u003csub\u003ee\u003c/sub\u003e: Equilibrium sorption capacity (mg g\u003csup\u003e-1\u003c/sup\u003e); C\u003csub\u003ee\u003c/sub\u003e: Equilibrium concentration of the sorption in the solution (mg L\u003csup\u003e-1\u003c/sup\u003e); K\u003csub\u003ef\u003c/sub\u003e: Freundlich constant indicative of sorption capacity; n: Empirical constant indicative of sorption intensity.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec9\" class=\"Section3\"\u003e \u003ch2\u003e2.3.3 Characterization of amendments before and after metal(liod)s sorption\u003c/h2\u003e \u003cp\u003eThe physicochemical properties of BC, FH, FB and SFB were analyzed in 1: 2.5 (w/v) ratio of solid and ultrapure water. The pH measured using the methods of Zhu et al. (\u003cspan citationid=\"CR68\" class=\"CitationRef\"\u003e2017\u003c/span\u003e), and using a pH meter (PHSJ-3F, Inesa, China). The surface morphologies and element distribution were determined by using a scanning electronic microscope equipped with energy dispersive spectrometer (SEM-EDS, Gemini 300, ZEISS, Germany). The specific surface areas were calculated by the Brunauer\u0026ndash;Emmett\u0026ndash;Teller (BET) method. A Fourier transform infrared spectroscope (FTIR, Nicolet 5700, Thermo Fisher Scientific, USA) was used to identify the surface functional groups between different materials in the range of 4000 to 400 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. An X-ray photoelectron spectrometer (XPS, K-Alpha+, Thermo Fisher Scientific, USA) with an Al Kα source at 12 kV and 72 W was used to characterize the chemical state of elements and surface structures of ABF before and after the sorption.\u003c/p\u003e \u003c/div\u003e \u003c/div\u003e \u003cdiv id=\"Sec10\" class=\"Section2\"\u003e \u003ch2\u003e2.4 Pot experiment\u003c/h2\u003e \u003cdiv id=\"Sec11\" class=\"Section3\"\u003e \u003ch2\u003e2.4.1 Experimental design\u003c/h2\u003e \u003cp\u003eThe pot experiment was conducted in a greenhouse at Zhejiang A\u0026amp;F University, China. Plastic pots (diameter 10 cm; height 12 cm) were filled with 1.5 kg of the Cd and As contaminated soil. The BC, FH, FB and SFB were added into the contaminated soil at application rates of 1.0%, 2.0%, 3.0% (dry weight, w/w), and soil without amendments as the control. The experimental treatments were replicated thrice.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec12\" class=\"Section3\"\u003e \u003ch2\u003e2.4.2. Metal(liod)s availability and redistribution\u003c/h2\u003e \u003cp\u003eTwo distinct methodologies were employed to determine the Cd-As concentrations in the soil solution using hydride generation-atomic fluorescence spectrometry (HG-AFS) and inductively coupled plasma-mass spectrometer (ICP-MS). The geochemical fractions of Cd-As were sequentially extracted using the BCR sequential extraction procedure to extract the acid soluble fraction (F1), reducible fraction (F2), oxidizable fraction (F3), and residual fraction (F4) (Ure et al., \u003cspan citationid=\"CR47\" class=\"CitationRef\"\u003e1993\u003c/span\u003e).\u003c/p\u003e \u003c/div\u003e \u003c/div\u003e \u003cdiv id=\"Sec13\" class=\"Section2\"\u003e \u003ch2\u003e2.5 Data quality control and statistical analysis\u003c/h2\u003e \u003cp\u003eAll batch sorption experiments were conducted in triplicate, ensuring a relative standard deviation of triplicate analysis with probability level less than 5%. The plastic ware and glassware used in the experiment and analysis were soaked in 3% nitric acid for 24 h and rinsed with deionized water. The statistical analyses were performed using SPSS (version 26.0, SPSS Inc., Chicago, IL, USA) statistical package. The experimental data were expressed as mean\u0026thinsp;\u0026plusmn;\u0026thinsp;standard error (n\u0026thinsp;=\u0026thinsp;3). The significant differences (\u003cem\u003ep\u003c/em\u003e\u0026thinsp;\u0026lt;\u0026thinsp;0.05) were evaluated using the analysis of variance (ANOVA) technique. The results were visualized using Origin 2021 (OriginLab Corporation, Northampton, MA, USA) for data graphing. The XPS data were analyzed using the Thermo Avantage program (Thermo Fisher Scientific).\u003c/p\u003e \u003c/div\u003e"},{"header":"3. Results and discussion","content":"\u003cdiv id=\"Sec15\" class=\"Section2\"\u003e\n \u003ch2\u003e3.1 Characterization of the amendments\u003c/h2\u003e\n \u003cp\u003eThe morphology and physio-chemical composition of adsorbents are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e1\u003c/span\u003e. The SEM images revealed that BC exhibited visible pores on its plain surface. In contrast, FB displayed a rough surface cover with attachments, and the partial pores were apparently blocked. The SEM-EDS showed a significant increase in iron content ranging from 4.41\u0026ndash;12.96% was recorded, confirming that ferrihydrite was successfully loaded onto the biochar surface. The surface of SFB exhibited greater roughness and a more granular structure compared to pristine biochar. The EDS analysis detected a sulfur peak, with sulfur content ranging from 0 to 3.36% on the surface of SFB, confirming successful sulfur loading onto the biochar surface. This might be attributable to the process of modification wherein the sulfur group increased the viscosity between biochar particles and iron particles, resulting in a denser surface coverage. The results of EDS analysis further showed that sulfur and iron were completely saturated on the surface of biochar (SFB) (Zeng et al., \u003cspan class=\"CitationRef\"\u003e2023\u003c/span\u003e).\u003c/p\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec16\" class=\"Section2\"\u003e\n \u003ch2\u003e3.2 The sorbent for Cd(II) and As(Ⅲ) sorption\u003c/h2\u003e\n \u003cdiv id=\"Sec17\" class=\"Section3\"\u003e\n \u003ch2\u003e3.2.1. Effect of initial solution pH\u003c/h2\u003e\n \u003cp\u003eThe sorption performance of adsorbents on heavy metal(liod)s is affected by the pH of the solution, the removal efficiencies of BC, FH, FB and SFB were compared for the removal of Cd(II) and As(III) were evaluated across a pH range of 3.0\u0026ndash;9.0. Figure\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e2\u003c/span\u003eA and \u003cspan class=\"InternalRef\"\u003e2\u003c/span\u003eB shows the sorption data under different pH range. In the binary solute (Cd-As coexistence) system, the sorption of Cd(II) increased linearly in the range of pH 3.0\u0026ndash;5.0, whereas most of the adsorbents showed a slow increasing followed by a slight decreasing trend when the solution pH raised beyond 5, and eventually reached the equilibrium point of sorption at pH value of 7.0. The electrostatic repulsion of high concentration of H\u003csup\u003e+\u003c/sup\u003e under low pH conditions may lead to the low removal of Cd, that mainly exists as Cd(OH)\u003csup\u003e+\u003c/sup\u003e ions in aqueous solution and precipitates as Cd(OH)\u003csub\u003e2\u003c/sub\u003e at pH\u0026thinsp;\u0026gt;\u0026thinsp;8 (Reddy et al., 2014). Therefore, the advantage of Cd-generated precipitation under high pH conditions is the main reason for the high removal rate.\u003c/p\u003e\n \u003cp\u003eThe solution pH is an important factor affecting the sorption effect of adsorbent on As, and the existence of As ions in the solution morphology, the charge on the surface of the adsorbent and the degree of dissociation of hydroxyl groups are closely related to it (Smedley et al., 2002). The sorption amount of BC for As did not change significantly, while the rest of the adsorbents showed an increasing trend at pH range of 3\u0026ndash;8 and the sorption amount decreased sharply when the pH exceed 8. The effect of pH also showed an increasing trend for As, indicating the existence of a certain co-precipitation between As and Cd. Due to the very limited sorption sites on the adsorbent surface, As usually exists in the form of arsenate ions (H\u003csub\u003e2\u003c/sub\u003eSO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e\u0026minus;\u003c/sup\u003e, HASO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e2\u0026minus;\u003c/sup\u003e and ASO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e3\u0026minus;\u003c/sup\u003e) in the reaction system. These results showed that the competition between H\u003csup\u003e+\u003c/sup\u003e and arsenate for the sorption sites became more and more obvious with the increase of OH\u003csup\u003e\u0026minus;\u003c/sup\u003e content. Therefore, the sorption of As by FH gradually decreased with the increase of pH. FH and As formed adsorptive and specialized sorption through legend exchange reactions (Deng et al., 2018). H\u003csub\u003e2\u003c/sub\u003eAsO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e\u0026minus;\u003c/sup\u003e is the main form when the pH is 2\u0026ndash;6. In the normal water environment within the pH range between 6 to 9, As is mainly exist in the form of HAsO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e2\u0026minus;\u003c/sup\u003e (Chen et al., 2011). At pH\u0026thinsp;\u0026gt;\u0026thinsp;8, As is mainly in the form of the anion AsO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e3\u0026minus;\u003c/sup\u003e. With the increase in pH value, the surface of FH, FB and SFB gradually changed from net positive charge to negative charge and the main form of As changed from HAsO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e2\u0026minus;\u003c/sup\u003e to AsO\u003csub\u003e4\u003c/sub\u003e\u003csup\u003e3\u0026minus;\u003c/sup\u003e, the electrostatic repulsion between FH, FB, SFB and As increased that led to the decrease in its sorption capacity. Therefore, the sorption rates of FH, FB and SFB on As increased and then decreased with the increase of pH value. The removal efficiencies of the amendments for Cd(II) and As(III) increases with the increase of the initial pH, whereby SFB had a significantly higher sorption and removal capacity for Cd and As.\u003c/p\u003e\n \u003c/div\u003e\n \u003cdiv id=\"Sec18\" class=\"Section3\"\u003e\n \u003ch2\u003e3.2.2. Effect of the sorbent dosage level\u003c/h2\u003e\n \u003cp\u003eThe effects of adsorbents on the sorption of Cd(II) and As(III) under different dosage conditions are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e2\u003c/span\u003eC and \u003cspan class=\"InternalRef\"\u003e2\u003c/span\u003eD. The sorption of Cd(II) by BC increased slowly although the increase was not significant with the increase of dosage indicating that BC had no significant effect on the sorption of Cd(II). When the dosage of FH, FB and SFB reached to 1.0 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, the sorption of Cd(II) showed a slow upward increase indicating that the sorption of Cd(II) on biochar tended to be dynamically balanced that may be due to the higher pores numbers on the surface of biochar that were completely occupied by Cd(II). The sorption of As(III) by the adsorbent increased sharply at dosage from 0.2 to 1.0 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. When the dosage level beyond 1.0 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, the sorption capacity of As(III) increased slowly, indicating that the sorption capacity was close to saturation at 1.0 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. In addition, the sorption of As by SFB increased sharply when the dosage level increased from 0.2 to 1.5 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, and tended to be stable when the dosage level beyond 1.5 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, indicating that sorption level reached to saturation. Overall, the dosage of 1.5 g L\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e is critical limit of SFB when used as a adsorbent for Cd and As.\u003c/p\u003e\n \u003c/div\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec19\" class=\"Section2\"\u003e\n \u003ch2\u003e3.3 Sorption kinetics\u003c/h2\u003e\n \u003cp\u003eAs shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e3\u003c/span\u003eA, Cd sorption increased with contact time and BC reached to equilibrium after 4 h, while the remaining three adsorbents reached equilibrium after 8 h. Similarly, the sorption equilibrium time of four adsorbents for Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e3\u003c/span\u003eB (As) was 4 h, that was faster than that of Cd. Ahmad et al. (\u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e) reported that sorption can be divided into three phases: an initial fast phase due to physical sorption (P1); a slow rising phase due to chemical sorption (P2); and a final phase of reaching of sorption to equilibrium (P3). The kinetic process of sorption of Cd by adsorbent has the largest proportion of P2 for chemo sorption. This may be due to the presence of As that leads to the combination of Cd and As to form complexes making P2 larger whereas the sorption kinetic process for As, P2 accounted for the major part and was chemo-sorption. The quasi-primary and quasi-secondary kinetic models fitted well with the sorption process, and the parameters of the kinetic equations for the sorption of Cd and As were shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e4\u003c/span\u003e, Table. The correlation coefficient (R\u003csup\u003e2\u003c/sup\u003e) of the secondary kinetic model of the four adsorbents was higher than that of the primary kinetic model, indicating that the sorption of As and Cd mainly occurred through chemical and physical sorption.\u003c/p\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec20\" class=\"Section2\"\u003e\n \u003ch2\u003e3.4 Sorption isotherms\u003c/h2\u003e\n \u003cp\u003eIn order to further explore the sorption and immobilization mechanism of Cd(II) and As(III) in the composite solution, isothermal sorption experiments were carried out at 25\u0026deg;C, and the results are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e4\u003c/span\u003eA and \u003cspan class=\"InternalRef\"\u003e4\u003c/span\u003eB. It can be seen that with the increase of the initial concentration of the composite solution, the sorption level of the three adsorbents on Cd(II) and As(III) gradually increased, and finally reached saturation. It is assumed that with the increase of the initial concentration of Cd(II) and As(III) in the solution, the probability of contact between the heavy metal(liod)s and the biochar increased, which is conducive to the adsorption of Cd(II) and As(III) to the active sites on the surface of the biochar material, however, due to the limited number of active sites on the surface of the adsorbent, the process of sorption will be saturated when the initial concentration of the heavy metal(liod)s solution increases to a certain limit. Figure\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e4\u003c/span\u003e shows the effect of different concentrations on Cd and As adsorption by the sorbent, and the sorption data of Cd and As fitted with both Freundlich and Langmuir models as shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e4\u003c/span\u003eA and B. Sorption of Cd by BC, FH, FB and SFB, the Langmuir model fitted better than Freundlich model with R\u003csup\u003e2\u003c/sup\u003e values of 0.91\u0026ndash;0.96 and 0.84\u0026ndash;0.92, respectively. This indicates that the Cd sorption process after modification of the biochar surface was monolayer sorption instead of multi-layer sorption (Chen et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e; Li et al., \u003cspan class=\"CitationRef\"\u003e2017\u003c/span\u003e). Moreover, the maximum sorption capacity of SFB (Q\u003csub\u003em\u003c/sub\u003e = 76.69 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e) was higher than that of BC (Q\u003csub\u003em\u003c/sub\u003e = 41.36 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e), and the Cd sorption capacity of SFB was higher than that of BC in the Cd concentration range of 0\u0026ndash;80 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e. This result clearly indicates that the Cd sorption capacity of BC was significantly improved after ferrihydrite and sulfur modification. From the correlation coefficients (R\u003csup\u003e2\u003c/sup\u003e), the sorption of As(III) by FH, FB, and SFB at 298.15 K is more in line with the Freundlich model, suggesting that the sorption behavior of As(III) in these two materials is controlled by multilayered sorption on the inhomogeneous surface (Fu et al., \u003cspan class=\"CitationRef\"\u003e2021\u003c/span\u003e). In addition, n\u0026thinsp;\u0026gt;\u0026thinsp;1 was observed to be favorable for sorption at all temperatures. Comparing to BC, FH and FB, the SFB showed higher sorption capacity for As(III) and Cd(II). The sorption of Cd(II) and As(III) was significantly enhanced by the co-doping of ferrihydrite and S.\u003c/p\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec21\" class=\"Section2\"\u003e\n \u003ch2\u003e3.5 Sorption thermodynamics\u003c/h2\u003e\n \u003cp\u003eIn order to further investigate the energy changes during the sorption process, sorption thermodynamic studies of Cd(II) and As(III) were carried out at different temperature (i.e., 15\u0026deg;C, 25\u0026deg;C, and 35\u0026deg;C). Figure\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e5\u003c/span\u003e shows the results of sorption of Cd(II) and As(III) by adsorbent under different temperature conditions. With the increase of temperature, the sorption effect of adsorbent on Cd(II) and As(III) shown a non significant increase (Lin et al., \u003cspan class=\"CitationRef\"\u003e2017\u003c/span\u003e), indicating that the temperature did not more effect on the sorption of Cd(II) and As(III).\u003c/p\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec22\" class=\"Section2\"\u003e\n \u003ch2\u003e3.6 Cd(II) and As(III) removal mechanisms\u003c/h2\u003e\n \u003cdiv id=\"Sec23\" class=\"Section3\"\u003e\n \u003ch2\u003e3.6.1. BET analysis\u003c/h2\u003e\n \u003cp\u003eBET analysis shown that the specific surface areas of BC, FH, FB and SFB are 64.20, 294.91, 175.35 and 185.16 m\u003csup\u003e2\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, respectively (Table. 1). The reason may be that the specific surface of FH is larger than other amendments. The complexation of ferrihydrite on the tunnels or pores of biochar may also lead to a decrease of the specific surface area of FB compared with that of pure FH, as reported by previous studies (Hong et al., \u003cspan class=\"CitationRef\"\u003e2022\u003c/span\u003e; Li et al., \u003cspan class=\"CitationRef\"\u003e2022\u003c/span\u003e). The N\u003csub\u003e2\u003c/sub\u003e sorption and desorption isotherms of BC, FB and SFB are type IV isotherms morphologically, indicating that BC, FB and SFB mainly includes mesopores or macropores. In contrast, those of FH are indicative of the presence of almost micropores (Huang et al., \u003cspan class=\"CitationRef\"\u003e2022\u003c/span\u003e). The isotherm of FB and SFB is of type IV, and includes an H4 hysteresis loop. This result indicates that mesopores or macropores may also be present, despite that the material is dominated by micropores, which also confirms the combination of ferrihydrite and biochar (Li et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e). After modification, the S content of pristine biochar significantly increased to 3.36%. Similar results also have been reported by Wu et al. (\u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e). which showed that the sulfur content of sulfur modified biochar is approximately 3.34%. After ferrihydrite modification, the surface area and the pore volume from 64.20 m\u003csup\u003e2\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 0.0696 cm\u003csup\u003e3\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e for BC to 175.35 m\u003csup\u003e2\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 0.1263 cm\u003csup\u003e3\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e for FB, to 185.16 m\u003csup\u003e2\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 0.1305 cm\u003csup\u003e3\u003c/sup\u003e g\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e for SFB, increasing by 173.13%-188.41% and 81.47%-87.50%. Thus, after ferrihydrite and sulfur modification biochar significantly increase the surface area and the pore volume.\u003c/p\u003e\n \u003cp\u003eAccording to Table 1, after ferrihydrite and sulfur modification, the content of S and Fe of biochar increased, indicating that ferrihydrite and sulfur are completely loaded. The C and O content of biochar increased, resulting in an increased polarity with an increase in oxygen-containing functional groups (Ahmad et al., \u003cspan class=\"CitationRef\"\u003e2014\u003c/span\u003e; Ennis et al., \u003cspan class=\"CitationRef\"\u003e2012\u003c/span\u003e). The increase of O and C content indicates that C on the surface of biochar is oxidized, and oxygen-containing functional groups such as carboxyl group, hydroxyl group and carbonyl group may be formed on the surface (Chen et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e). The pH of BC was 11.48, and that of FH was 4.36. The pH of biochar modified with ferrihydrite and sulfur hydrate decreased significantly.\u003c/p\u003e\n \u003cp\u003e\u003cstrong\u003eTable. 1.\u003c/strong\u003e Physico-Chemical Characterization of adsorbents\u003c/p\u003e\n \u003cdiv align=\"\"\u003e\n \u003ctable border=\"1\" cellspacing=\"0\" cellpadding=\"0\" align=\"\" width=\"599\"\u003e\n \u003ctbody\u003e\n \u003ctr\u003e\n \u003ctd valign=\"top\" style=\"width: 186px;\"\u003e\n \u003cp\u003eParameter\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd valign=\"top\" style=\"width: 103px;\"\u003e\n \u003cp\u003eBC\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd valign=\"top\" style=\"width: 103px;\"\u003e\n \u003cp\u003eFH\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd valign=\"top\" style=\"width: 103px;\"\u003e\n \u003cp\u003eFB\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd valign=\"top\" style=\"width: 103px;\"\u003e\n \u003cp\u003eSFB\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eYield (% dry wt.)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e50.12\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eash (% dry wt.)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e36.61\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e\u0026mdash;\u0026mdash;\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003epH\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e11.48\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e4.36\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e10.86\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e10.58\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eC (%)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e10.58\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.00\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e61.43\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e36.72\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eO (%)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e39.05\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e34.27\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e15.37\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e32.23\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eS (%)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.00\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.00\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.00\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e3.36\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eFe (%)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e4.41\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e65.73\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e12.96\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e7.98\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003eSurface area(m\u003csup\u003e2\u003c/sup\u003e\u0026middot;g\u003csup\u003e\u0026minus;1\u003c/sup\u003e)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e64.20\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e294.91\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e175.35\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e185.16\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003ePore volume(cm\u003csup\u003e3\u003c/sup\u003e\u0026middot;g\u003csup\u003e\u0026minus;1\u003c/sup\u003e)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.0696\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.1804\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.1263\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e0.1305\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003ctr\u003e\n \u003ctd style=\"width: 186px;\"\u003e\n \u003cp\u003ePore diameter (nm)\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e4.84\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e2.46\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e3.05\u003c/p\u003e\n \u003c/td\u003e\n \u003ctd style=\"width: 103px;\"\u003e\n \u003cp\u003e2.89\u003c/p\u003e\n \u003c/td\u003e\n \u003c/tr\u003e\n \u003c/tbody\u003e\n \u003c/table\u003e\n \u003c/div\u003e\n \u003c/div\u003e\n \u003cdiv id=\"Sec24\" class=\"Section3\"\u003e\n \u003ch2\u003e3.6.2. FTIR analyses\u003c/h2\u003e\n \u003cp\u003eThe functional groups of biochar have significant influence on its sorption capacity. The FTIR spectra of BC before and after modification and after absorption of Cd and As ion were observed and the results are show in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e6\u003c/span\u003e. Two major transmittance peaks at 1590.30 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 1020.15 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e were detected in the BC sorption system, which may be due to the stretching vibrations of the C\u0026thinsp;=\u0026thinsp;O and C-O-C functional groups on the biochar, respectively (Novais et al., \u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e; Shi et al., \u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e; Wang et al., \u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e; Zhao et al., \u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e). The intensification of the -OH groups was due to the introduction of oxygen-containing functional groups after the alkali modification, which was accompanied by a shift of the C-O-C stretching vibration from 1590.30 to 1611.19 (Zeng et al., \u003cspan class=\"CitationRef\"\u003e2023\u003c/span\u003e). In FH, the transmittance peak at 3363.96 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e and 576.12 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e is attributed to Fe-OH and Fe-O, respectively, while the peak at 1623.88 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e showing the hydration changes on ferrihydrite surfaces (Pawar et al., \u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e; Zeng et al., \u003cspan class=\"CitationRef\"\u003e2024\u003c/span\u003e). The FTIR results of adsorbents before and after sorption of Cd(II) and As(III) in mixed Cd-As system implied the potential binding mechanisms for the Cd(II) and As(III) in the adsorbents. As shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e6\u003c/span\u003e, several peaks of BC, FH, FB and SFB shift after sorption, and new peak also appeared. Comparatively, the absorbed Cd-As, reinforcing or decreasing hydrogen bonding and provoking the enhancement or weakening of -OH; the same phenomenon is reported by Lyu et al. (\u003cspan class=\"CitationRef\"\u003e2022\u003c/span\u003e). The peaks of -OH (3363.96 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e) was enhanced and shifted to 3371.64 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e after the sorption of metal(liod)s, suggesting that the oxygen sites and hydroxyl group are the main complexation sites during the process (Jiang et al., \u003cspan class=\"CitationRef\"\u003e2023\u003c/span\u003e). The show of the vibration of C\u0026thinsp;=\u0026thinsp;C, and C\u0026thinsp;=\u0026thinsp;O at 1632.19 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e implied that the biochar or ferrihydrite contributed to the stabilization of Cd(II) (Zeng et al., \u003cspan class=\"CitationRef\"\u003e2024\u003c/span\u003e). The vibration of Fe-O bond at 576.12 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e were shifted to 570.15 cm\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e, that further indicated the complexation of Cd(II) and As(III) with iron oxides of FH, FB and SFB after the sorption process (Zhang et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e).\u003c/p\u003e\n \u003c/div\u003e\n \u003cdiv id=\"Sec25\" class=\"Section3\"\u003e\n \u003ch2\u003e3.6.3. XPS analyses\u003c/h2\u003e\n \u003cp\u003eThe elemental composition and oxidation state of the sorbent surface area was determined using XPS spectroscopy. Figure\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003e shows the survey spectra of the sorbents before and after Cd and As sorption. Interestingly, the new peaks at 1326, 1362, 412, 405 eV from As/Cd-laden the sorbents were due to electron from the As 2p and Cd 3d levels, respectively, indicating that As and Cd were captured by the sorbent. The homogeneous distribution of Cd and As of the sorbent that sorbents has abundant functional groups, and the distribution mapping of Cd and As revealed metal(iold)s successfully adsorbed.\u003c/p\u003e\n \u003cp\u003eThe high-resolution XPS spectra of each element before and after the sorbent sorption are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003e. The C 1s spectrum of the sorbent can be resolved into three chemical bonds, C-C/C\u0026thinsp;=\u0026thinsp;C i.e., C-OH, C-OOH, with peaks at 284, 285 and 288 eV, respectively (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eA and \u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eE). After sorption, the C-OH groups showed a slight decrease in intensity and there was a corresponding decreas in C-OOH, C-C/C\u0026thinsp;=\u0026thinsp;C, which may be due to the oxygen-containing functional groups bound to Cd and As, enhancing the molar ratio of the carbon substrate (Lyu et al.,2022). For the mixed Cd-As systems, the relative intensities of the carbon peaks changed little with respect to the case of untreated BC, suggesting that the metal(iold)s binding in the systems mostly occurs in the ferrihydrite region of the composite, casing little change in the BC.\u003c/p\u003e\n \u003cp\u003eThe attribution of O 1s were as follows (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eB and \u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eF): -OH bond at 530 eV, C-O/C\u0026thinsp;=\u0026thinsp;O bond at 532 eV, lattice oxygen inside ferrihydrite, Fe-OH (surface Fe and -OH bond) at 531 eV (Hong et al., \u003cspan class=\"CitationRef\"\u003e2022\u003c/span\u003e; Jacukowicz et al., 2020; Lin et al., \u003cspan class=\"CitationRef\"\u003e2017\u003c/span\u003e). XPS spectrum of O 1s in SFB, where the peak at 535.33 eV occurred corresponds to Fe-O-H (Alchouron et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e). In the Cd-As mixture system, the relative strength of Fe-OH of FH, FB and SFB decreased, suggesting the formation of a ferrihydrite-Cd-As ternary complex, which would promote the removal of Cd and As. This may be due to the interaction of Cd with Fe-OH/-OH and ligand exchange of As with Fe-OH (Xu et al., \u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e; Zhao et al., \u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e).\u003c/p\u003e\n \u003cp\u003eThe Fe 2p spectra before and after sorption of Cd(II) and As(Ⅲ) on the adsorbent are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eC and \u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eG. It can be seen that Fe is mainly loaded on the surface of biochar in the form of Fe(II) and Fe(III). Before and after sorption, the occurrence of displacement indicates the involvement of ferrihydrite in the uptake process. The S 2p of XPS spectra are shown in Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eD and \u003cspan class=\"InternalRef\"\u003e7\u003c/span\u003eH. The content of sulfur on BC increased after modification. The three peaks of SFB in XPS were located at 164.16, 165.49 and 168.79 eV, which represent C-S, C\u0026thinsp;=\u0026thinsp;S and S(Ⅳ/Ⅱ)-O, respectively (Cato et al., \u003cspan class=\"CitationRef\"\u003e2018\u003c/span\u003e; Zhang et al., \u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e). Park et al. (\u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e) reported that the sulfur existence in S-modified wood biochar are mainly in the form of oxidized and thiophene sulfur, which is similar to the XPS results for sulfur. After Cd and As sorption, S(Ⅳ/Ⅱ)-O of SFB decreased from 39.64\u0026ndash;23.29%, which indicated that sulfite reacts with Cd(II) and As(Ⅲ) and is consumed during the sorption process. During decomposition of organic matter, oxygen is consumed and sulphate is reduced to sulfite, which may have dissolved Cd oxide to form metal(liod) ions, which react with hydrogen sulfide to form metal(iold) sulfides (Chen et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e). Depending on the acid-base reaction, positively charged metal(liod) species, such as Cd(II) or Cd (OH)\u003csup\u003e+\u003c/sup\u003e, may interact with sulfur groups (Qiao et al., \u003cspan class=\"CitationRef\"\u003e2023\u003c/span\u003e). Kubier et al. (\u003cspan class=\"CitationRef\"\u003e2019\u003c/span\u003e) also suggested that sulphate and bisulfide anions are some of the most stable Cd complexes ligands, and final complex formation is related to ligand concentrations.\u003c/p\u003e\n \u003c/div\u003e\n\u003c/div\u003e\n\u003cdiv id=\"Sec26\" class=\"Section2\"\u003e\n \u003ch2\u003e3.7 Effects of adsorbents on the Cd and As availability and fractionation in soil\u003c/h2\u003e\n \u003cp\u003eThe concentration of dissolved Cd(II) and As(Ⅲ) significantly decreased under SFB application and changed with the time (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003e). The SFB addition resulted in a decline in the concentration of the CaCl\u003csub\u003e2\u003c/sub\u003e-extractable Cd in soil (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003eA). Compared to the control, significant (\u003cem\u003ep\u003c/em\u003e\u0026thinsp;\u0026lt;\u0026thinsp;0.05) decreases were observed after the addition of different radio of SFB. Addition of SFB significantly decreased Cd from 0.15 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e in the control to 0.06 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e-0.07 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e in the difference SFB treatment, with a decreasing varied from 32.59\u0026ndash;41.26%. The lowest content of the CaCl\u003csub\u003e2\u003c/sub\u003e-extractable Cd was observed in the 3%SFB. The NaH\u003csub\u003e2\u003c/sub\u003ePO\u003csub\u003e4\u003c/sub\u003e-extractable As concentration in soil decreased (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003eB). Compared to the control, significant (\u003cem\u003ep\u003c/em\u003e\u0026thinsp;\u0026lt;\u0026thinsp;0.05) decreases were observed after the addition of different radio of SFB. Addition of SFB significantly decreased As from 2.40 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e in the control to 0.86 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e-1.20 mg kg\u003csup\u003e\u0026minus;\u0026thinsp;1\u003c/sup\u003e in the difference SFB treatment, with a decreasing from 50.06\u0026ndash;64.06%. The lowest content of the CaCl\u003csub\u003e2\u003c/sub\u003e-extractable Cd was observed in the 3%SFB.\u003c/p\u003e\n \u003cp\u003eThe addition of biochar has also altered the distribution fraction of Cd(II) and As(Ⅲ) in the soil (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003e). Specifically, the F1 fraction of Cd-As is easily absorbed by plants; however, the F4 fraction of Cd-As is stable and not readily available to plants (Xu et al., \u003cspan class=\"CitationRef\"\u003e2020\u003c/span\u003e; Yang et al., \u003cspan class=\"CitationRef\"\u003e2023\u003c/span\u003ea). The F1, F2 and F3 fraction of Cd decreased in the range of 17.50, -26.66%, 12.86\u0026ndash;25.84%, and 37.29\u0026ndash;53.60%, respectively; whereas the F4 fraction of Cd increased by 53.65to -87.89% in the SFB treatments, compared to Control (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003eC). Under 1.0%, 2.0%, 3.0% SFB application, the percentage of F1, F2 and F3 fraction of As decreased by 33.47\u0026ndash;38.77%, 36.08\u0026ndash;45.90% and 7.53\u0026ndash;28.87%, respectively, while F4 fraction of As increased by 33.55 to 41.86%, compared to Control (Fig.\u0026nbsp;\u003cspan class=\"InternalRef\"\u003e8\u003c/span\u003eD). These results showed that the absorbents transferred Cd(II) and As(Ⅲ) from relatively labile fractions to less toxic and more stable fractions, thereby reduced the bioavailability of Cd in soil, particularly in the 3% SFB that was highly effective compared to other treatments. These results dictate that the SFB was more effective in immobilizing Cd(II) and As(Ⅲ) compared to control. SFB mainly decreased the F1, F2 and F3 fractions of soil Cd(II) and As(Ⅲ) but increased the more stable fraction of Cd(II) and As(Ⅲ) (F4), facilitating the transformation of Cd(II) and As(Ⅲ) chemical speciation and changing Cd(II) and As(Ⅲ) bioavailability (Sun et al., \u003cspan class=\"CitationRef\"\u003e2024\u003c/span\u003e).\u003c/p\u003e\n\u003c/div\u003e"},{"header":"4. Conclusion","content":"\u003cp\u003eSulfur-ferrihydrite-biochar composites were developed for the simultaneous removal of Cd and As. The SFB exhibited a significant increase in surface area, pore structure, and functional groups, which enhanced its capacity for Cd and As sorption. The composite material demonstrated strong sorption properties across a wide pH range in both binary systems. The primary removal mechanisms include electrostatic attraction, ion exchange, precipitation and formation of various complexes. In Cd-As mixture systems, Cd and As primarily form tri/quaternary complexes with S/FH, facilitating the removal of both metals(iod)s. The presence of As can promote the sorption of Cd on the sulfur-ferrihydrite-biochar composite material. The enhanced sorption of As may be due to several possible mechanisms, including the change of surface charge of the composite material due to the presence of As and the increase of the dispersion of biochar on ferrihydrite. The introduction of SFB led to reductions in Cd and As concentrations by 34.98% and 78.27%, respectively, in water and soil, indicating its effectiveness for remediating Cd and As co-contaminated soil. The study also showed that the sorption kinetics and potential binding mechanisms of heavy metal(liod)s, particularly in systems where cations and anions coexist. The sulfur-ferrihydrite-biochar composite can reduce environmental risks associated with Cd and As contamination by immobilizing them in the solid phase, and considered as a promising material for soil and water remediation. Considering the prevalence of precipitation and transformation processes of ferrihydrite and S in natural environments, special attention needs to be paid to the interactions between hydrated Fe, S, heavy metal(liod)s and environmental factors.\u003c/p\u003e"},{"header":"Declarations","content":"\u003cp\u003e\u003cstrong\u003eFunding\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eThis work was supported by the Natural Science Foundation of China [grant number 42407016, 32271532], Research and Development Fund of Zhejiang A \u0026amp; F University [grant number 2020LFR052], and Young Scientists Fund of the National Natural Science Foundation of China [grant number 42407016].\u0026nbsp;\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eDeclaration of Competing Interest\u0026nbsp;\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eThe authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.\u0026nbsp;\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eData availability\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eData will be made available on request.\u0026nbsp;\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eCRediT authorship contribution statement\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eXuqiao Wu:\u003c/strong\u003e Conceptualization, Formal analysis, Writing \u0026ndash; original draft.\u003csup\u003e\u0026nbsp;\u003c/sup\u003e\u003cstrong\u003eXiaowen Teng:\u003c/strong\u003e Data curation, Writing \u0026ndash; original draft, Software.\u003csup\u003e\u0026nbsp;\u003c/sup\u003e\u003cstrong\u003eDong Huang:\u003c/strong\u003e Formal analysis, Visualization.\u003csup\u003e\u0026nbsp;\u003c/sup\u003e\u003cstrong\u003eIjlal Ahmad:\u003c/strong\u003e Writing \u0026ndash; review \u0026amp; editing.\u003csup\u003e\u0026nbsp;\u003c/sup\u003e\u003cstrong\u003eHanbo Chen:\u003c/strong\u003e Supervision. \u003cstrong\u003eYaqian Li:\u003c/strong\u003e Software.\u003csup\u003e\u0026nbsp;\u003c/sup\u003e\u003cstrong\u003eDubin Dong:\u003c/strong\u003e Methodology.\u003cstrong\u003e\u0026nbsp;Yanxin Tang:\u003c/strong\u003e Validation, Visualization. \u003cstrong\u003eYini Wang:\u003c/strong\u003e Investigation. \u003cstrong\u003eSong Li:\u003c/strong\u003e Supervision. \u003cstrong\u003eDan Liu:\u003c/strong\u003e Conceptualization, Funding acquisition, Project administration, Resources.\u0026nbsp;\u003cstrong\u003eWeijie Xu:\u003c/strong\u003e Conceptualization, Writing \u0026ndash; review \u0026amp; editing, Funding acquisition, Project administration, Resources.\u003c/p\u003e"},{"header":"References","content":"\u003col\u003e\n\u003cli\u003eAdnan, M., Xiao, B., Xiao, P., Zhao, P., Li, R., Bibi, S. 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J., 441\u003c/em\u003e, 135972.\u003c/li\u003e\n\u003c/ol\u003e"}],"fulltextSource":"","fullText":"","funders":[],"hasAdminPriorityOnWorkflow":false,"hasManuscriptDocX":true,"hasOptedInToPreprint":true,"hasPassedJournalQc":"","hasAnyPriority":false,"hideJournal":true,"highlight":"","institution":"","isAcceptedByJournal":false,"isAuthorSuppliedPdf":false,"isDeskRejected":"","isHiddenFromSearch":false,"isInQc":false,"isInWorkflow":false,"isPdf":false,"isPdfUpToDate":true,"isWithdrawnOrRetracted":false,"journal":{"display":true,"email":"[email protected]","identity":"researchsquare","isNatureJournal":false,"hasQc":true,"allowDirectSubmit":true,"externalIdentity":"","sideBox":"","snPcode":"","submissionUrl":"/submission","title":"Research Square","twitterHandle":"researchsquare","acdcEnabled":true,"dfaEnabled":false,"editorialSystem":"","reportingPortfolio":"","inReviewEnabled":false,"inReviewRevisionsEnabled":true},"keywords":"Modified biochar, Cadmium, Arsenic, Sorption mechanisms, Immobilization, Soil remediation","lastPublishedDoi":"10.21203/rs.3.rs-6195683/v1","lastPublishedDoiUrl":"https://doi.org/10.21203/rs.3.rs-6195683/v1","license":{"name":"CC BY 4.0","url":"https://creativecommons.org/licenses/by/4.0/"},"manuscriptAbstract":"\u003cp\u003eThe transformation behaviors of arsenic (As) and cadmium (Cd) in contaminated soils are generally complex process due to their distinct chemical and physical characteristics, which poses challenge for remediation. This study proposes an efficient strategy for the simultaneous immobilization of Cd and As using sulfur-ferrihydrite-modified biochar (SFB) as an organic amendment. A series of experiments, including batch and pot experiments, was conducted under controlled conditions. The results showed that the maximum sorption capacities of Cd and As by SFB were 76.69 mg kg\u003csup\u003e-1\u003c/sup\u003e and 8.28 mg kg\u003csup\u003e-1\u003c/sup\u003e, respectively, which were significantly higher than those of biochar (BC), ferrihydrite (FH) and ferrihydrite-biochar (FB). This higher sorption capacity is attributed to synergistic interactions between biochar and ferrihydrite. The sorption process of Cd and As by SFB follows the Langmuir isothermal sorption model and the pseudo-second-order kinetic model, indicating a combination of physical sorption and chemisorption mechanisms. The removal mechanisms for As primarily involve precipitation, oxidation and complexation, while those for Cd mainly include ion exchange, complexation, precipitation, and electrostatic sorption. Application of SFB reduced the bioavailable forms of Cd and As in the soil, shifting their chemical forms toward more stable residual states and enhancing immobilization. Overall, the SFB is a novel and effective adsorbent by immobilizing Cd and As in agricultural soils, promoting safer crops production in contaminated field.\u003c/p\u003e","manuscriptTitle":"Reduction and immobilization of Cd(II) and As(III) using sulfur-ferrihydrite-biochar as an amendment in water and soil: Investigation of the Mechanism of Remediation","msid":"","msnumber":"","nonDraftVersions":[{"code":1,"date":"2025-03-17 07:14:39","doi":"10.21203/rs.3.rs-6195683/v1","editorialEvents":[{"type":"communityComments","content":0}],"status":"published","journal":{"display":true,"email":"[email protected]","identity":"researchsquare","isNatureJournal":false,"hasQc":true,"allowDirectSubmit":true,"externalIdentity":"","sideBox":"","snPcode":"","submissionUrl":"/submission","title":"Research Square","twitterHandle":"researchsquare","acdcEnabled":true,"dfaEnabled":false,"editorialSystem":"","reportingPortfolio":"","inReviewEnabled":false,"inReviewRevisionsEnabled":true}}],"origin":"","ownerIdentity":"6c0e464e-14bd-4c5f-aae0-09cced7299b9","owner":[],"postedDate":"March 17th, 2025","published":true,"recentEditorialEvents":[],"rejectedJournal":[],"revision":"","amendment":"","status":"posted","subjectAreas":[],"tags":[],"updatedAt":"2025-06-20T16:44:11+00:00","versionOfRecord":[],"versionCreatedAt":"2025-03-17 07:14:39","video":"","vorDoi":"","vorDoiUrl":"","workflowStages":[]},"version":"v1","identity":"rs-6195683","journalConfig":"researchsquare"},"__N_SSP":true},"page":"/article/[identity]/[[...version]]","query":{"redirect":"/article/rs-6195683","identity":"rs-6195683","version":["v1"]},"buildId":"XKTyCvWXoU3ODBz1xrDgd","isFallback":false,"isExperimentalCompile":false,"dynamicIds":[84888],"gssp":true,"scriptLoader":[]}