Superior simultaneous immobilization of cadmium and arsenic by ferrihydrite-modified biochar: performance and mechanisms

The contrasting geochemical behaviors of cationic cadmium (Cd) and anionic arsenic (As) in co-contaminated soil pose a significant challenges for their simultaneous immobilization. To address this issue, three materials were developed: pristine biochar (BC) derived from rice straw, pure ferrihydrite (FH), and ferrihydrite-engineered biochar (FB) produced through iron-oxide impregnation. These materials were applied to contaminated soil at 1.0%-3.0% (w/w) to evaluate their effects on soil properties and metalloid stabilization. The results showed that the FB treatment significantly improved soil pH, cation exchange capacity (CEC), electrical conductivity (EC), and organic matter (OM) content. Compared to control, the FB3% treatment produced the strongest effects after 120 days, increasing CEC and OM by 19.31% and 17.97%, respectively. Soil pH and EC also increased by 11.4% and 7.5% relative to the control. Moreover, FB3% treatment significantly reduced bioavailable Cd and As by 51.58% and 72.94%, outperforming BC and FH. The sequential extraction results indicated that FB promoted the transformation of Cd and As from exchangeable to residual fractions. After 120 days, exchangeable Cd and As decreased by 55.53% and 45.50% relative to the control. Material characterization revealed that ferrihydrite loading increased the Fe content, surface area of the biochar, and functional groups (-OH and C=O) of biochar. Overall, FB exhibited superior immobilization performance through electrostatic attraction, surface complexation, and co-precipitation, demonstrating its strong potential as an efficient amendment for remediating Cd/As co-contaminated soils.

1 Introduction

Cadmium (Cd) and arsenic (As) are widespread contaminants that pose significant toxicity risks to soil biota, plants, animals, and humans (Cui et al., 2005; Guo et al., 2010). Rice, a staple food for more than half of the global population and primarily cultivated in Asia, is particularly susceptible to elevated Cd and As levels in paddy soils. Such contamination can result in excessive accumulation in rice plants, threatening food security and negatively affecting grain yield and quality (Khan et al., 2008). Specifically, Cd predominantly occurs in cationic forms, whereas As generally exists as anions. This contrast creates significant challenges for a single remediation agent to immobilize both elements simultaneously (Vaňková et al., 2021). Moreover, their co-existence in soils can lead to complex interactions, including competition for sorption sites and changes in solubility, which may reduce remediation efficiency (Zhao et al., 2019). Conventional physical and chemical remediation techniques, including soil replacement, leaching, chemical immobilization, have shown clear remediation effects. However, their application is limited by high costs, risks of secondary pollution, potential damage to soil structure, raising concerns about their long-term sustainability (Basta et al., 2001; He et al., 2019). Consequently, the development of safe, cost-effective, and environmentally friendly stabilization materials and strategies for heavy metal-contaminated soils has become a critical research focus in the field of soil environmental science.

Biochar has gained increasing attention for environmental remediation because of its diverse feedstock sources, highly porous structure, abundant functional groups, and strong heavy metal adsorption capacity (Huang et al., 2025; Zhou et al., 2025). Biochar has been shown to effectively reduce the mobility and bioavailability of heavy metals in soils through ion exchange, surface complexation, precipitation, and pH modulation mechanisms (Kookana et al., 2011; Ahmad et al., 2014). However, pristine biochar exhibits limited adsorption capacity and selectivity toward anionic heavy metals such as As (Sharma et al., 2024).

To address this limitation, functionalizing biochar with metal oxides has emerged as an effective strategy. Ferrihydrite, a precursor of crystalline iron oxides commonly occurring as nanoparticulate colloids in natural environments (Wang and Wang, 2019), possesses a high specific surface area and strong affinity for both Cd and As. Its ability to form inner-sphere complexes facilitates the immobilization of heavy metals through arsenate adsorption, inner-sphere complexation, co-precipitation or incorporation, and isomorphous substitution. (Hafeez et al., 2022). When immobilized onto biochar, ferrihydrite can form stable complexes and precipitates with Cd²⁺ and AsO₄^3-, substantially enhancing adsorption and immobilization capacity (He et al., 2018; Guo et al., 2021). Previous studies have demonstrated that iron-modified biochar exhibits improves soil pH regulating, enhancing cation exchange capacity (CEC), stabilizing organic matter, and reducing the bioavailability of heavy metal, particularly in acidic soils and co-contaminated environments (Ibrahim et al., 2022; Qu et al., 2022). Ferrihydrite modification enriches surface functional groups such as hydroxyl and carboxyl groups and enhances heavy metal immobilization through complexation, precipitation, and redox reactions. (Hafeez et al., 2022). In addition, ferrihydrite modification also alters the pore structure of biochar, thereby improving its ability to immobilize contaminants (Lima and Rodrigues, 2025). Studies indicate that ferrihydrite-modified biochar reduces particle aggregation, increases the availability of functional groups, and intensifies complexation and redox reactions, thereby enhancing Cd and As stabilization in soil systems (Guo et al., 2021). Similarly, co-application of ferrihydrite and biochar increased the specific surface area and pore connectivity of the composite material, thereby enhancing its adsorption capacity for Cd(II) and As(V) (Lalmalsawmdawngliani et al., 2024). Despite these advances, the long-term stabilization pathways and speciation dynamics governing the performance of iron-modified biochar in Cd–As co-contaminated soils remain insufficiently understood. Therefore, this study provides a mechanistic and time-resolved evaluation of Cd and As stabilization in co-contaminated soils by simultaneously tracking their geochemical speciation under BC, FH, and FB during a 120-day incubation. Unlike previous work that primarily focuses on short-term adsorption or single-metal systems, our study integrates FTIR/SEM characterization with soil chemical and speciation analyses to elucidate how ferrihydrite loading, enhanced functional groups, and Fe-oxide coatings drive element-specific immobilization pathways. We further distinguish the mechanisms responsible for Cd and As stabilization—including pH buffering, Fe-oxyhydroxide complexation, co-precipitation, and fraction transformation—thereby providing new insights into the synergistic remediation potential of ferrihydrite-engineered biochar in Cd–As co-contaminated soils.

This study addresses these knowledge gaps through a time-resolved, mechanistic evaluation of Cd and As stabilization under biochar (BC), ferrihydrite (FH), and ferrihydrite-modified biochar (FB) during a 120-day soil incubation. Our work simultaneously integrates: 1) Long-term speciation tracking of both Cd and As using sequential extraction; 2) In-depth material characterization (FTIR, SEM) to link structural evolution with stabilization behavior; 3) Comparative assessment of BC, FH, and FB to isolate the added value of ferrihydrite loading; 4) Mechanistic differentiation of stabilization pathways including Fe-oxyhydroxide complexation, pH buffering, co-precipitation, and redox-driven transformations. These innovations allow us to directly connect ferrihydrite functionalization with shifts in contaminant speciation, quantified soil property changes, and evolving surface chemistry. Such integration provides new insights into the long-term remediation potential and mechanistic robustness of ferrihydrite-engineered biochar in Cd–As co-contaminated agricultural soils.

2 Materials and methods

2.1 Preparation and modification of biochar

2.1.1 Biochar preparation

Rice straw was air-dried, ground, and sieved through a 5 mm sieve. Biochar was produced by slow pyrolysis of rice straw at 550 °C for 2 h with a heating rate of 5 °C min^-1, under limited oxygen supply in a muffle furnace. After cooling to room temperature, the biochar was washed with ultrapure water and filtered through a 100-mesh sieve. The resulting biochar was collected and referred to as BC.

2.1.2 Biochar modification

The synthetic procedure of ferrihydrite was used as described by (Engel et al., 2021). The biochar-loaded ferrihydrite material was synthesized using ferrihydrite and biochar at a critical mass ratio of 2:1. 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(NO₃)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 4,500 rpm, washed twice with ultrapure water and dialyzed for 3 days in darkness to remove excess Na⁺ and NO₃⁻. 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. To prepare the ferrihydrite-engineered biochar (FB), biochar was first subjected to KOH activation by mixing it with 1 mol L^-1 KOH solution and stirring at 60 °C for 72 h. After activation, the biochar was separated by centrifugation (5,000 rpm), thoroughly washed with ultrapure water, frozen at −20 °C overnight, and then freeze-dried. The ferrihydrite impregnation was performed by dispersing the activated biochar in ultrapure water and gradually adding ferrihydrite at a mass ratio of 2:1 (ferrihydrite:biochar) under vigorous stirring. The suspension was maintained at 60 °C for 12 h to ensure uniform adsorption and surface deposition of ferrihydrite nanoparticles onto the biochar matrix. The resulting composite was separated by centrifugation, washed repeatedly with ultrapure water to remove unbound ferrihydrite, and dried at 60 °C to obtain the final ferrihydrite-engineered biochar (FB).

2.2 Characterization of biochar

The pH of BC and FB were analyzed in 1: 2.5 (w/v) ratio of solid and ultrapure water using a pH meter (PHSJ-3F, Inesa, China) (Zhu et al., 2017). 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 4,000 to 400 cm^-1.

2.3 Soil incubation experiment

2.3.1 Experimental design

The contaminated soil used for the incubation experiment was obtained from the top layer (0–20 cm) of a paddy field contaminated with both Cd and As in Jinhua City, Zhejiang Province, China (29°17′N, 120°01′E). The dried soil subsamples were homogenized and sieved (<2 mm). The sample properties were determined, and the results are as follows: pH, 6.67; EC, 82.45 uS cm⁻¹; total Cd, As: 0.89, 64.14 mg kg⁻¹, respectively; organic matter, 20.11 mg kg⁻¹. The soil incubation 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 and FB 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. The mixture was stirred to be well homogenized and then incubated at 25 °C ± 1 °C for 4 months. Throughout the whole experimental period, the soil humidity was kept at 60% water holding capacity by adding purified water reached at a constant weight once a day (Wu et al., 2016). During the 4-month incubation period, the sampled intervals were set as 7, 15, 30, 60 and 120 days of each treatment for analysis of the samples’ physical and chemical properties.

2.3.2 Metal(liod)s availability and redistribution

The concentrations of cadmium (Cd) and arsenic (As) in the soil solution were measured by inductively coupled plasma mass spectrometry (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).

2.4 Data analysis

Data were expressed as the standard deviation of the average value. Different letters were verified using the ANOVA tests by least-significant difference (LSD) (p < 0.05). Data processing was operated using SPSS (Version 19.0) software. The data shown in the tables and figures were analyzed using Excel 2010 and Origin 9.0.

3 Results

3.1 Characteristics of amendments

The physicochemical properties of the two biochar samples are shown in Table 1. BC exhibited a higher pH (11.48) than FB (10.86). FB displayed increased Fe content and surface area relative to BC, whereas BC had higher ash and oxygen content. SEM imaging (Figure 1) revealed that both biochars retained uniformly arranged tubular bundle structures, reflecting the morphology of the original biomass. SEM-EDS analysis confirmed a substantial increase in iron content from 4.41% to 12.96%, indicating successful ferrihydrite loading onto the biochar surface. FTIR spectra (Figure 2) showed characteristic peaks for C=O conjugated ketones (1,351 cm^-1), phenolic -OH (3,065 cm^-1), and ester C=O (780 cm^-1), with peak intensities of oxygen-containing functional groups enhanced after ferrihydrite modification. After incubation, both BC and FB exhibited noticeable changes in their surface spectral features as reflected in the FTIR results. For BC, the–OH band near 3400 cm^-1 and the C=O/C–O related absorptions within 1,700–1,000 cm^-1 showed moderate decreases in transmittance. For FB, however, the attenuation of these characteristic bands was more pronounced, indicating a greater degree of functional group transformation during the incubation process. These post-reaction spectral changes suggest that the two materials underwent different extents of surface modification, with FB participating more actively in ligand exchange, surface complexation, and metal-binding reactions relative to unmodified BC.

Table 1

Table 1. Physico-chemical characterization of pristine biochar (BC) and ferrihydrite-loaded biochar (FB).

Figure 1

Figure 1. SEM-EDS of pristine biochar (a,b) and ferrihydrite-loaded biochar (c,d).

Figure 2

Figure 2. (a) FTIR spectra of pristine biochar (BC) and ferrihydrite-loaded biochar (FB) before and after the reaction. (b) SEM-mapping images showing the distribution of Cd and As elements on the BC and FB.

3.2 Soil chemical properties

3.2.1 Soil pH and CEC

As shown in Figures 3a–c, BC, FH and FB addition significantly enhanced soil pH over 120 days, with stronger effects observed at higher application rates and longer incubation time (p < 0.05). FB treatments, particularly at 3%, elicited earlier and more persistent pH elevation, significantly increased by 5.8% on day 7 and maintained a 9.48% rise by day 120 relative to the control (p < 0.05). In contrast, BC-induced pH increases were noticeable mainly after day 30, with BC3% enhancing pH by 5.9% and 4.5% on days 30 and 60, respectively, relative to the control (p < 0.05).

Figure 3

Figure 3. Effects of pristine biochar (BC), ferrihydrite (FH) and ferrihydrite-loaded biochar (FB) on soil pH (a–c) and cation exchange capacity (d–f). Data points and error bars represent the mean ± standard deviation (n = 3).

Similarly, BC, FH and FB addition significantly improved soil CEC, with FB showing more rapid and pronounced effects (p < 0.05) (Figures 3d–f). On day 7, compared to the control FB2% increased CEC by 72.6%, while BC2% led to a 44.54% increase (p < 0.05). CEC enhancements persisted over time, with FB3% showing a 20.38% increase on day 30, and both BC2% and FB2% maintaining significant improvements by day 60. Even the lowest doses (BC1% and FB1%) sustained significant CEC increases by day 120 relative to the control (p < 0.05).

3.2.2 Soil EC and OM

The application of BC, FH and FB significantly increased soil EC over the incubation period (Figures 4a–c). On day 7, BC3% and FB3% treatments significantly enhanced EC by 191.38% and 36.11%, respectively, compared with the control (p < 0.05). By day 30, a significant increases were observed in the BC3% and FB1% treatments (p < 0.05), with EC rising by 143.56% and 31.68%, respectively relative to the control. On day 60, BC3% and FB3% treatments further increased EC by 137.2% and 15.5%, respectively, with EC reaching a maximum by day 120; BC3% maintained a significant 134.97% increase compared to the control (p < 0.05).

Figure 4

Figure 4. Effects of pristine biochar (BC), ferrihydrite (FH) and ferrihydrite-loaded biochar (FB) on soil electrical conductivity (a–c) and organic matter content (d–f) of soil. Data points and error bars represent the mean ± standard deviation (n = 3).

As shown in Figures 4d–f, the application of BC, FH and FB also significantly improved soil OM content arcoss the incubation period. On day 7, BC3% and FB3% treatments significantly increased soil OM by 100% and 21.58%, respectively, as conpared to the control (p < 0.05). By day 30, soil OM content in BC3% and FB3% treatments increased by 154.77% and 67.34%, respectively. On day 60, BC3% and FB3% treatments resulted in a increases of 102.09% and 31.01% of soil OM, respectively. By day 120, BC3% and FB3% treatments remained higher soil OM levels, with increased by 88.10% and 17.97%, respectively (p < 0.05), as compared with the control.

3.3 Bioavailability and fractionation of Cd and As in soil

3.3.1 Available Cd and As in soil

As presented in Figure 5a, BC, FH and FB treatments effectively decreased the concentration of available Cd in the soil ver the incubation period, with FB consistently outperforming BC and FH. Notably, FB3% treatment reduced available Cd by 39.01% as early as day 7, whereas BC showed no significant effect at this stage. From day 30 onward, both amendments exhibited increasing efficacy, with the highest reductions observed on day 120, BC3% and FB2% treatments significantly decreased available Cd by 49.78% and 51.58%, respectively (p < 0.05). Similarly, available As concentrations were significantly reduced by all treatments (Figure 5b). FB3% achieved a 58.1% reduction by day 7. On day 30, BC2% and FB2% reduced As availability by 16.7% and 46.2%, respectively (p < 0.05). By day 60, BC2% and FB3% decreased As by 21.2% and 9.4%, respectively. At day 120, BC3% and FB1% treatments led to 67.1% and 72.9% reductions in available As relative to the control (p < 0.05).

Figure 5

Figure 5. Effect of pristine biochar (BC), ferrihydrite (FH) and ferrihydrite-loaded biochar (FB) on soil-available concentration of Cd (a) and As (b). Data points and error bars represent the mean ± standard deviation (n = 3).

3.3.2 The fraction of Cd and As in soil

As shown in Figures 6a,b, the application of BC, FH and FB significantly altered the soil Cd speciation. On day 7, BC3% and FB3% treatments significantly reduced exchangeable Cd fraction by 29.03% and 16.5%, respectively (p < 0.05), while increasing the residual Cd fraction by 36.54% and 26.54%, respectively, compared to the control. By day 120, exchangeable Cd fraction concentration was further reduced by 27.34% and 55.53% under BC3% and FB3% treatments, respectively, accompanied by significant increases in the residual Cd fraction by 39.89% and 38.79% (p < 0.05). Similarly, BC3% treatment significantly decreased the exchangeable As fraction by 8.9% (p < 0.05), whereas FB3% treatment significantly increased the residual As fraction by 26.83% compared with the control on day 7 as depicted in Figures 6c,d. At day 120, FB1%, FB2%, and FB3% treatments significantly reduced exchangeable As fraction by 39.4%, 41.21%, and 45.5%, respectively, with FB3% treatment significantly enhanced the residual As fraction by 18.58% (p < 0.05).

Figure 6

Figure 6. The fractionation of Cd and As in soils following the applications of pristine biochar (BC), ferrihydrite (FH) and ferrihydrite-loaded biochar (FB). (a) Cd fractionation at day 7 (D7), (b) Cd fractionation at day 120 (D120), (c) As fractionation at D7, and (d) As fractionation at D120. Data points and error bars represent mean ± S.D. (n = 3). Different letters indicate a significant difference (p < 0.05).

4 Discussion

4.1 Effect of ferrihydrite modification on biochar properties

The physicochemical characteristics of biochar were markedly enhanced by hydration-based ferrihydrite modification. Incorporation of iron oxides increased the density of adsorption sites while preserving the intrinsic pore structure of the biochar, as evidenced by SEM analysis. The elevated Fe content in ferrihydrite-modified biochar (FB) confirmed successful loading and the formation of additional metallic active sites, which are conducive to heavy metal complexation and precipitation. A slight decrease in pH from 11.48 to 10.86 was observed in FB, likely resulting from hydrolysis of Fe³⁺ releasing H⁺ and partial occupation of alkaline functional groups by iron oxides (Muzaffar et al., 2024). FTIR spectra revealed increased intensities of oxygen-containing functional groups, including phenolic hydroxyl (-OH) and carbonyl (C=O) moieties, indicating enhanced surface chemical reactivity (Zhou et al., 2014). These structural modifications, combined with the introduction of redox-active Fe species, are expected to synergistically improve the adsorption and immobilization capacity of FB for heavy metals, particularly under acidic and co-contaminated soil conditions (Yuan et al., 2017). Overall, the formation of Fe-containing surface layers on the biochar matrix not only increases its metal-binding affinity but also provides additional sites for complexation, precipitation, and redox-mediated immobilization of Cd and As, highlighting the superior potential of FB for remediation of contaminated paddy soils. The more pronounced attenuation of–OH and C–O functional group signals in FB compared with BC indicates that ferrihydrite loading substantially increased the participation of these oxygen-containing moieties in metal-binding reactions. This spectral depletion suggests that Fe-oxyhydroxide domains introduced on the surface of FB provided a greater number of reactive coordination sites capable of forming inner-sphere complexes with Cd and As (Zhou et al., 2014). These FTIR-based observations align well with the fractionation results, which showed larger decreases in exchangeable Cd and As and greater enrichment in the residual fractions under FB relative to BC. The enhanced consumption of surface functional groups further supports the occurrence of ligand exchange, surface complexation, and co-precipitation processes facilitated by ferrihydrite modification, providing spectroscopic evidence that FB strengthened the immobilization pathways compared with pristine biochar (Muzaffar et al., 2024).

4.2 Effects of biochar on soil chemical properties

4.2.1 Soil pH

The results of this study indicated that both biochar, ferrihydrite and ferrihydrite-modified biochar effectively mitigated soil pH in varying degrees, with the most pronounced effects observed in the high application rates (3%) (Figures 3a–c). The ability of biochar to regulate soil pH is primarily attributed to its abundant alkaline functional groups, such as carboxyl and phenolic hydroxyl groups, as well as its relatively high CEC (Singh Yadav et al., 2023; Sarfraz et al., 2024). Additionally, biochar can increase soil pH by adsorbing hydrogen ions (H⁺) and releasing basic cations including K⁺, Ca²⁺, and Mg²⁺ into the soil solution (Kong et al., 2014; Shi et al., 2019). Compared to unmodified biochar, FB exhibited superior and more stable performance in maintaining elevated soil pH. This improved buffering capacity is likely associated with the iron oxide coating formed on the biochar surface, which enhances the material’s ability to neutralize H⁺ ions (Hao et al., 2018). Previous studies have confirmed that iron oxides such as goethite and hematite can adsorb protons via surface complexation mechanisms, thereby contributing to soil pH stabilization (Adegoke et al., 2014; Villalobos et al., 2025). Moreover, the iron components in ferrihydrite-modified biochar may facilitate the stabilization of soil organic matter and reduce the release of acidic substances, thus slowing the decline in soil pH (Wang and Wang, 2019; Jing et al., 2021). These findings indicate that ferrihydrite-modified biochar possesses greater potential for long-term pH regulation and buffering in contaminated soils (Hao et al., 2018). The pure ferrihydrite (FH) treatment also demonstrated a significant ability to regulate soil pH, although its effect was less pronounced compared to FB. This suggests that while ferrihydrite alone can enhance soil pH, the presence of biochar in the composite (FB) may offer a synergistic effect, enhancing the pH buffering capacity through both ferrihydrite’s neutralizing ability and biochar’s high cation exchange capacity and alkaline functional groups. The stronger attenuation of–OH and C–O bands in the FTIR spectra of FB indicates more intensive proton-binding and ligand exchange reactions on its surface, consistent with its higher capacity to neutralize soil acidity. This suggests that ferrihydrite loading enhanced the reactivity of surface functional groups involved in pH buffering.

4.2.2 Soil CEC

The observed increase in CEC following biochar application is closely related to its porous structure and high density of surface functional groups (Tesfaye et al., 2021). The negatively charged functional groups, particularly carboxyl and phenolic hydroxyl groups present on biochar surfaces, provide abundant sites for cation adsorption (e.g., K⁺, Ca²⁺, Mg²⁺), thereby significantly enhancing soil CEC (Wan et al., 2023). Furthermore, biochar may promote the accumulation of soil organic matter through physical adsorption and chemical complexation processes, which further contribute to improvements in CEC (Carter et al., 2013; Xia et al., 2022). Although the application of ferrihydrite (FH) alone also contributed to a moderate increase in soil CEC, its effect was notably weaker than that of FB. This suggests that the synergistic interaction between biochar’s porous carbon matrix and ferrihydrite’s active iron oxide surfaces plays a crucial role in amplifying cation adsorption capacity beyond what FH alone can achieve. Notably, FB exhibited an even greater capacity for enhancing soil CEC, particularly after 30 days of incubation. This enhanced performance can be attributed to the surface loading of iron oxides. Iron oxides, such as goethite and hematite, are known to form stable surface complexes that strengthen the soil’s ability to adsorb cations and improve overall nutrient retention (Yan et al., 2022). The more pronounced decrease in–OH and C=O/C–O band intensity observed for FB after incubation reflects stronger interactions between its surface functional groups and exchangeable cations, supporting the greater CEC enhancement relative to BC. These spectral changes corroborate the increased availability of coordination sites generated by ferrihydrite modification. Furthermore, the incorporation of ferrihydrite may increase the number of active functional sites on the biochar surface, thereby further enhancing CEC (Tao et al., 2020; Lalmalsawmdawngliani et al., 2024).

4.2.3 Soil EC

The application of pristine biochar (BC3%) resulted in a significantly higher enhancement of soil EC than ferrihydrite-modified biochar (FB3%), as shown in Figures 4a,c. The presence of iron oxides on the biochar surface, which promote the release and migration of soluble cations (Nguyen et al., 2018). Previous research has indicated that iron oxides, such as goethite and hematite, can interact with soil anions (e.g., phosphate, sulfate), altering the distribution and mobility of soluble salts and thereby influencing EC dynamics (Geelhoed et al., 1997). Although ferrihydrite (FH) alone slightly increased soil EC, its effect was markedly lower than that of FB, indicating that the composite’s enhanced performance arises from the synergistic interaction between ferrihydrite and the biochar matrix. The carbon framework of biochar likely improves ion diffusion and retention around ferrihydrite particles, thereby amplifying the overall electrolyte balance compared with FH alone. In addition, the microporous structure and abundant active sites of FB may further facilitate cation exchange and the efficient release of soluble nutrients, contributing to a more stable and sustained elevation of EC over time (Mallet et al., 2013). FTIR spectral weakening of oxygen-containing functional groups in FB suggests enhanced ligand exchange and surface reactions that may contribute to the gradual release or redistribution of ionic species, helping explain the more stable EC trends observed over time. This indicates that ferrihydrite-modified surfaces sustain electrochemical activity during incubation (Geelhoed et al., 1997). The enhancement of soil EC following biochar application is likely associated with its mineralization process and the release of soluble components. Previous studies have demonstrated that biochars derived from different feedstocks and subjected to varying pyrolysis temperatures exhibit distinct contents of soluble inorganic salts. In particular, high-temperature biochars are often enriched in carbonates and alkaline oxides, which promote the release of basic cations and subsequently increase soil EC levels (Uchimiya et al., 2011; Zhang et al., 2022). Moreover, the high specific surface area and well-developed porous structure of biochar facilitate nutrient adsorption and gradual release, thereby contributing to a sustained upward trend in EC (Hui, 2021).

4.2.4 Soil OM

Biochar is inherently rich in recalcitrant carbon components, and its porous structure offers physical protection to OM, reducing its decomposition rate (Ginebra et al., 2022). Biochar provides abundant surface area for the adsorption of dissolved organic matter and facilitates the formation of stable organo-mineral complexes through interactions with soil minerals, thereby prolonging OM residence time in soil (Li et al., 2020). The results indicate that high-dose biochar application can effectively enhance the stability and accumulation of soil organic carbon pools. Notably, FB demonstrated superior efficacy in promoting OM accumulation compared to unmodified biochar. Previous studies have revealed that iron oxides, such as goethite and hematite, can form strong, stable complexes with OM, reducing its microbial degradation and enhancing long-term stability (Yang et al., 2022). The reduction in intensities of FB’s C–O and C=O absorption bands after incubation implies stronger interactions with soil organic compounds, potentially via complexation or hydrogen bonding, which supports its greater contribution to OM stabilization. These spectral shifts further indicate that FB promotes the formation of more stable organo–mineral associations (Yang et al., 2022). Furthermore, the mineralization rate of FB is relatively low, which may suppress microbial degradation of organic matter and facilitate its long-term retention in soil (Hafeez et al., 2022). Although the FH treatment alone contributed to a modest increase in soil organic matter, its effect was notably weaker than that of FB. This suggests that the biochar matrix in FB enhances the stabilization of OM by providing additional sorption sites and improving the physical protection of organic residues that ferrihydrite alone cannot achieve (Gao et al., 2025).

4.3 Effects of biochar on the Cd and As availability and fractionation in soil

4.3.1 Availability and fraction of soil Cd

The addition of biochar has altered the fraction of Cd(II) in the soil. Specifically, the F1 fraction of Cd is readily bioavailable, whereas residual F4 Cd is highly stable (Xu et al., 2020). Both BC and FB effectively transformed Cd from exchangeable to residual fractions, reducing environmental risk (Ahmad et al., 2022). Biochar and its modified derivatives have been widely recognized for their ability to reduce the mobility and bioavailability of Cd in soils through multiple mechanisms, including adsorption, surface complexation, and precipitation processes (Shen et al., 2018). The immobilization of Cd by BC is primarily attributed to its high specific surface area and well-developed porous structure, which provide abundant binding sites for electrostatic adsorption (Gao et al., 2019). The surface of biochar is rich in oxygen-containing functional groups (e.g., carboxyl and hydroxyl groups), which can complex with Cd(II) to form stable coordination compounds. Additionally, mineral components such as carbonates and phosphates in biochar can react with Cd(II) to form insoluble precipitates, such as CdCO₃ or Cd₃(PO₄)₂, further enhancing Cd immobilization (Dai et al., 2021). Biochar application often leads to an increase in soil pH, which reduces Cd solubility and facilitates its transformation into less bioavailable forms (Jing et al., 2021).

However, the Cd stabilization capacity of unmodified BC remains limited relative to FB, likely due to the pH sensitivity and variable surface charge characteristics of BC, which can influence its Cd sorption efficiency (Ou et al., 2023). The iron oxides loaded on the surface of ferrihydrite-modified biochar serve as active sites for cation exchange, electrostatic adsorption, surface complexation, and co-precipitation reactions (Wei et al., 2023). Under redox-active soil conditions, ferrihydrite promotes the formation of stable Fe³⁺–Cd²⁺ co-precipitates, such as Cd–Fe hydroxide complexes, effectively reducing Cd mobility (Cheng et al., 2024). The more pronounced weakening of–OH and C–O functional group signals in FB after incubation demonstrates greater participation of these groups in Cd complexation, consistent with the observed reduction in exchangeable Cd fractions. The FTIR evidence supports the mechanism that ferrihydrite loading enhances inner-sphere complexation and co-precipitation with Cd (Dai et al., 2021). Furthermore, FB enhances the stabilization of soil organic matter, which facilitates the formation of stable organo-metallic complexes between Cd and humic or fulvic acids, further decreasing Cd bioavailability (Boostani et al., 2024; Wei et al., 2024). Iron oxides can also facilitate the formation of Fe–Cd hydroxide complexes, thereby reducing Cd mobility and promoting its transformation into non-extractable fractions (Dai et al., 2021). The FTIR spectroscopic data revealed a discernible intensification of oxygen-bearing functional groups, notably the phenolic hydroxyl (-OH) and carbonyl (C=O) moieties, following the incorporation of ferrihydrite. This enhanced signal profile serves as compeling spectroscopic evidence for the increased density of surface coordination sites. In addition to the FTIR evidence, the XPS analysis further confirms that both BC and FB participate in direct Cd surface complexation processes. The C 1s and O 1s spectra revealed significant shifts and peak-area reductions in C-O and O-H functional groups after Cd adsorption, demonstrating that these groups actively donated electron pairs to Cd²⁺ through inner-sphere complexation (Supplementary Figure. S1). This provides direct proof that ligand exchange occurred between surface -OH/-COOH groups and Cd, forming stable Cd-O coordination bonds. For FB, the XPS changes were more pronounced: the O 1s spectra showed substantial increases in metal–oxygen (Cd-O) peaks, indicating strong coordination of Cd with Fe-O-H groups derived from ferrihydrite. This confirms the formation of Fe-O-Cd inner-sphere complexes, consistent with Fe-mediated co-precipitation processes. The stronger binding of Cd on FB is also evidenced by the decreased intensity of oxygen-containing groups (C–O, O–H), suggesting their enhanced participation in complexation reactions.

Consequently, the primary mechanism by which FB drives the conversion of highly labile Cd fractions into recalcitrant forms is the surface-controlled sequestration, facilitated by the formation of robust inner-sphere complexes between Cd and these augmented functional groups. (Liao et al., 2024). Ferrihydrite-modified biochar can alter the microscale redox potential in the rhizosphere, favoring the transformation of labile Cd into more stable carbonate- or oxide-bound fractions (Wang et al., 2024). The structural integration of iron oxides within the carbon matrix may limit Cd adsorption under fluctuating pH and redox conditions, thereby enhancing its long-term immobilization in soil systems (Su et al., 2024). Moreover, FB can regulate the soil redox environment, further decreasing the solubility of Cd²⁺ and accelerating its transformation into more stable mineral-bound forms (Liao et al., 2024). In contrast, the ferrihydrite (FH) treatment alone exhibited a limited ability to immobilize Cd, primarily due to its relatively low specific surface area and absence of carbon-based sorption domains. Although FH facilitated partial Cd transformation into less mobile fractions through Fe–Cd co-precipitation, its stabilization efficiency remained inferior to that of FB, highlighting the synergistic role of the biochar matrix in enhancing surface complexation and long-term Cd sequestration (Su et al., 2024). Although FB exhibited strong Cd immobilization performance during the incubation period, its long-term stability under dynamic soil conditions warrants further consideration. Over extended timescales, Fe(III) reduction or the partial dissolution of ferrihydrite under anaerobic or acidic environments may lead to the remobilization of Cd previously incorporated into Fe–Cd hydroxide phases. Moreover, the gradual oxidation–reduction cycling typical of paddy soils could alter surface functional groups and weaken inner-sphere complexation, thereby increasing the risk of secondary Cd release (Su et al., 2024).

4.3.2 Availability and fraction of soil As

The results of this study demonstrated that both biochar and ferrihydrite-modified biochar significantly influenced As speciation in soil by promoting its transformation from highly mobile exchangeable forms to more stable residual fractions, thereby reducing its environmental risk. The porous structure and large specific surface area of BC provide numerous active sites for As adsorption, primarily through electrostatic interactions (Vithanage et al., 2017). In addition, oxygen-containing surface functional groups, such as carboxyl and hydroxyl groups, can form stable complexes with As species, contributing to its immobilization (He et al., 2019). Biochar may also influence As redox transformations in soil; for instance, under anaerobic conditions, BC can promote the reduction of As(V) to the more soluble and mobile As(III), thereby affecting its environmental behavior. Furthermore, carbonate and metal oxide components within BC can participate in precipitation reactions with As, forming insoluble mineral phases such as calcium arsenate (Ca₃(AsO₄)₂), which enhance As retention within the soil matrix (Ibrahim et al., 2022). The results of this study demonstrated that high-dose BC application can significantly improve As stabilization in soil (Wu et al., 2017).

Although the FH treatment also contributed to As stabilization by providing reactive Fe–OH sites for surface complexation, its immobilization efficiency was notably lower than that of FB. In contrast, FB exhibited significantly greater As stabilization effects throughout the experiment. The superior As immobilization performance of FB is primarily due to the surface-loaded iron oxides, which exhibit a strong affinity for As and facilitate its immobilization via electrostatic attraction, surface complexation, and co-precipitation (Liao et al., 2024). Iron oxides readily form stable Fe–As hydroxide complexes, effectively reducing As solubility and mobility in the soil environment (Kumpiene et al., 2021). The substantial attenuation of–OH bands in FB indicates intensive ligand exchange between Fe-oxyhydroxide surfaces and arsenate species, aligning with the marked decline in exchangeable As. This FTIR response confirms that ferrihydrite-modified surfaces provide abundant reactive sites for As(V) inner-sphere complexation (Ibrahim et al., 2022). Complementing the FTIR findings, XPS analysis provided direct molecular-level evidence of strong interactions between As and surface functional groups: after adsorption, the O 1s spectra displayed significant increases in metal-oxygen signals characteristic of As-O coordination, while concurrent shifts in O-H and C-O peaks demonstrated ligand exchange reactions wherein surface hydroxyl and carboxyl groups were replaced by arsenate species, leading to strong chemisorptive binding. FB exhibited the most pronounced spectral changes among all treatments, indicating a higher density of reactive sites capable of forming stable bidentate binuclear and mononuclear As complexes, which underscores its elevated immobilization capacity. Furthermore, reductions in C-O and O-H peak intensities in both C 1s and O 1s spectra confirm the participation of oxygen-containing functional groups on the carbon matrix in As adsorption, supporting an integrated multi-mechanistic stabilization process. Collectively, these spectroscopic results demonstrate that the superior performance of FB arises from its enhanced surface reactivity and its ability to drive ligand exchange, inner-sphere complexation, and stable metal-oxygen coordination, thereby providing a robust and durable pathway for As immobilization in contaminated soils.

In addition, Number studies have shown that ferrihydrite-modified biochar may alter soil microbial communities, promoting the growth of iron-oxidizing or arsenic-transforming bacteria that indirectly enhance As stabilization (Sang et al., 2024). In addition, the porous structure and high specific surface area of the modified biochar can facilitate the physical entrapment of As-bearing nanoparticles, reducing their bioavailability and mobility in soil (Mohammad Eisa et al., 2020). Additionally, FB improves the stability of soil organic matter, promoting the formation of stable organo-metallic complexes between As and humic or fulvic acids, further enhancing As immobilization (Yang et al., 2015). Furthermore, FB may modulate the soil redox environment, facilitating the oxidation of As(III) to As(V), which exhibits a higher affinity for iron minerals, thereby enhancing the long-term stabilization of As within the soil matrix (Zhang et al., 2023). Moreover, ferrihydrite-modified biochar facilitates the mineralization of As, driving its transformation into less bioavailable, more stable fractions, thereby reducing the potential environmental risks associated with As contamination (Gemeinhardt et al., 2006). The FB composite exhibited a significant intensification of hydroxyl (-OH) stretching vibrations, a signature associated with both the Biochar matrix and the Fe-O-H functionalities of the loaded ferrihydrite. These O-H groups serve as the dominant surface coordination ligands, facilitating the strong, specific chemisorption of As(V) oxyanions (Teng et al., 2025). Specifically, the enhanced Fe-O-H density provides abundant sites for the formation of binuclear or mononuclear inner-sphere surface complexes, which is the thermodynamically favored pathway for As(V)immobilization on iron oxides. Crucially, the formation of stable inner-sphere surface complexes and the mineralization of As driven by FB create a highly chemically stable state that significantly minimizes the potential for secondary release of As under fluctuating environmental conditions, thereby assuring long-term stabilization within the soil matrix (Gemeinhardt et al., 2006). Furthermore, the persistent modulation of the soil redox environment and the protection of soil organic matter suggest that the superior immobilization efficacy of FB is robust and offers a sustainable strategy for mitigating the ecotoxicological risks associated with As contamination over extended periods.

5 Conclusion

The application of BC, FH, and FB markedly improved soil physicochemical properties, with the strongest effects observed at the 3% application rate. Among the amendments, FB achieved the highest remediation efficiency, reducing bioavailable Cd and As by 51.58% and 72.94% after 120 days and promoting their conversion from exchangeable to residual fractions. This enhanced performance is primarily attributed to the ferrihydrite-loaded surface, which provides abundant reactive functional groups and Fe oxyhydroxide sites that facilitate electrostatic adsorption, inner-sphere complexation, and co-precipitation. SEM analyses corroborated the formation of Fe-rich surface coatings and increased chemical reactivity, confirming the strong affinity of FB for heavy metal immobilization. The FTIR spectral changes, particularly the stronger post-reaction attenuation of–OH and C–O bands in FB, provide direct evidence of its enhanced surface reactivity and support the mechanistic inference that ferrihydrite loading intensifies ligand exchange and complexation pathways for Cd and As stabilization. These findings indicate that ferrihydrite-modified biochar is an effective short-term amendment for reducing Cd and As mobility in co-contaminated soils, with application rate being a key determinant of stabilization efficiency. However, the study is limited by the absence of post-incubation characterization and Fe quantification in soil, which restricts deeper interpretation of stabilization pathways. In future work, we plan to optimize the material granularity and develop appropriate particle-separation procedures to enable reliable post-reaction characterization and strengthen mechanistic analysis. Moreover, field-scale evaluation under fluctuating redox conditions is required to verify the long-term performance and practical applicability of FB in contaminated agricultural soils.

Data availability statement

The raw data supporting the conclusions of this article will be made available by the authors, without undue reservation.

Author contributions

XT: Data curation, Writing – original draft. WY: Data curation, Supervision, Writing – review and editing. HC: Validation, Writing – review and editing. XW: Formal Analysis, Resources, Writing – review and editing. JT: Visualization, Writing – review and editing. XY: Data curation, Formal Analysis, Writing – review and editing. ZL: Resources, Visualization, Writing – review and editing. DL: Funding acquisition, Project administration, Writing – review and editing. WX: Funding acquisition, Project administration, Writing – review and editing.

Funding

The author(s) declared that financial support was received for this work and/or its publication. This work was supported by the “Pioneer” and “Leading Goose” R&D Program of Zhejiang Province, China (No. 2023C02005), Department of Agriculture and Rural Affairs of Zhejiang Province (No. 2025SNJF018), National Natural Science Foundation of China (No. 42407016), Research and Development Foundation of Zhejiang A&F University (No. 2020LFR052), Zhejiang Provincial Department of Education of General Scientific Research Projects (No. Y202456091), and 2025 Zhejiang Provincial Innovation and Entrepreneurship Training Program for College Students (No. S202510341117).

Conflict of interest

The author(s) declared that this work was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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The author(s) declared that generative AI was not used in the creation of this manuscript.

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Supplementary material

The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fenvs.2025.1717178/full#supplementary-material

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