Removal of Cr(VI) from aqueous solutions by nZVI-loaded sludge-derived biochar: performance and mechanism

The Highland Railway is a major strategic deployment for Chinese development, a great project for long-term development, and with a hundred-year plan. In regional terms, it is necessary to further promote the economic development of the southwest region of China as well as the Tibetan region. But with the construction of the plateau railway requires a large number of metal materials (e.g. rails, steel bars, sound insulation panels, etc.), and the process of producing these metal materials will inevitably produce a large amount of heavy metal wastewater, which will cause serious environmental pollution if not treated in time. Chromium (Cr), one of the essential materials for alloy materials, generates a large amount of Cr wastewater during the manufacturing process of alloy materials and causes serious pollution to the environment (Altun et al. 2021; Chen et al. 2021). Chromium (Cr) is extensively used as an important raw material for production in industries such as metallurgy, electroplating, and alloy manufacturing (Chen et al. 2015). Cr in water mainly exists in the form of Cr(VI) and Cr(III), of which Cr(VI) is soluble, migratory, toxic, and hazardous to humans (Dong et al. 2017; Fei et al. 2021). At present, the treatment methods for Cr(VI)-containing wastewater include precipitation, photocatalysis, adsorption, and ion exchange (Hou et al. 2020; Jia et al. 2021; Jiang et al. 2021). The adsorption methods are of low operating cost, simple operation and high efficiency, and thus have received widespread attention (Pi et al. 2021).

Nanometer zero-valent iron (nZVI) hold abundant pore structure, strong reducibility, and high surface activity, and has a good removal effect on heavy metals, which has been widely used in environmental pollution (Qian et al. 2019; Qiu et al. 2020). However, the small particle size of nZVI, its high surface energy, and its magnetic properties, and hence its tendency to agglomerate, led to a significant reduction in specific surface area and reduction capacity, limiting its application in water treatment (Qiu et al. 2020; Zhao et al. 2020). To address this problem, researchers had attempted to use different materials as nZVI carriers to reduce particle agglomeration (Qiu et al. 2020; Tang et al. 2021). Biochar has a well-developed pore structure and a large number of functional groups on the surface and has good adsorption properties for heavy metals (Zhang et al. 2019; Zou et al. 2021). Biochar loaded with nZVI can effectively reduce the agglomeration effect of nZVI, and improve the reactivity of nZVI (Yi et al. 2020). Previous studies have shown that nZVI-loaded biochar can effectively improve the Cr removal from aqueous solutions and reduce the amount of chromium in aqueous solutions (Qiu et al. 2020; Zhou et al. 2021). Zhou et al. (2020) showed that biochar loaded with nZVI was able to double the Cr(VI) adsorption capacity compared to raw biochar. In addition, a number of studies have shown that the specific surface area of virgin biochar is small. For this reason, ball milling was used to increase the specific surface area of the biochar to provide more sites for nZVI loading (Zou et al. 2021). Previous studies had shown that ball milling not only promoted interfacial effects between materials, but also improved the removal of contaminants from composites (Tang et al. 2021; Zou et al. 2021).

In this study, the nZVI-loaded sludge-derived biochar materials (nZVI@SDBC) were prepared and investigated their Cr(VI) adsorption performance from aqueous solutions. The purposes of the work were to: (1) investigate the effects of Cr(VI) removal under different environmental conditions; (2) reveal the mechanism of Cr(VI) removal by nZVI@SDBC; (3) evaluate the nZVI@SDBC for the Cr(VI) removal from actual electroplating wastewater; (4) analyze the cost per m³ of effluent removed by nZVI@SDBC. This work analyses the feasibility of nZVI@SDBC for the adsorption of Cr(VI) generated during the construction of highland railways from multiple perspectives (removal performance, mechanism, actual wastewater removal, and cost).

Potassium dichromate (VI) (K₂Cr₂O₇), sodium borohydride (NaBH₄), ferrous sulphate hexahydrate (FeSO₄·7H₂O), and other reagents were purchased from Sinopharm Chemical Reagent Co., Ltd (AR, Shanghai, China). Different mass concentrations of Cr(VI) storage solution was prepared using K₂Cr₂O₇. 0.1415 g K₂Cr₂O₇ was dissolved in 10 mL of deionised water and the solution was transferred to a 100 mL volumetric flask. 0.1 mL HNO₃ (1 mol/L) was added to the volumetric flask and the solution was volumetised. The solution configured above was a 50 mg/L Cr(VI).

Pore structure parameters were measured by specific surface area analyzer and analyzed by BET method (BET, Tristar II Plus 2.02, USA). Scanning electron microscopy (SEM, JSM-7500F, JEOL, Japan) and transmission electron microscope (TEM, JEM-2100F, JEOL, Japan) were used to observe the microscopic morphology of the samples. The material composition of the samples were surveyed by an X-ray diffraction analyzer (XRD, D8X, Bruker, Germany) and the raw data were processed by jade 6.5 software. X-ray photoelectron spectrometer (XPS, Escalab 250Xi, Thermo Fisher Scientific, USA) was used to analyze the elemental composition and valence changes of the samples, and Avantage software was used to perform peak splitting analysis. The functional groups in nZVI@SDBC(2:1) were investigated using Fourier transform infrared spectroscopy (FTIR, Nicolet-460, Thermo Fisher, USA). nZVI@SDBC(2:1) was dispersed in 0.01 mol/L NaCl (1:1000, w/v) ultrasonically for 1 h, and the solution pH was adjusted to 2.0–10.0. The zero potential of different pH solutions were measured by a zeta potential analyzer (Zetasizer, Malvern, UK) and the pH_zpc is calculated according to the zeta potential-pH curve.

Preparation of SDBC (Qiu et al. 2020): 10 g of dried powdered sludge was weighed and placed in a crucible, covered tightly with a crucible lid and the crucible was placed in a muffle furnace at 500 °C for 2 h. The black solid material was obtained as sludge-derived biochar. The sludge-derived biochar was treated by ball milling (JC-QM, Qingdao Juchuang Environmental Protection Group Co., Ltd, China) and the sieved ball milling sludge-derived biochar was collected after passing through an 80 mesh sieve, noted as SDBC.

Preparation of nZVI (Qiu et al. 2020): To a conical flask containing 70 mL of distilled water and 30 mL of ethanol, 7.447 g FeSO₄·7H₂O was added and stirred with a stirrer (SN-JJ-1, Shanghai Shangyi Instruments & Equipment Co., Ltd, China) at a constant speed (150 r/min) for 20 min until completely dissolved. Under an N₂ inert environment, 1.125 g of NaBH₄ was added and stirred for 30 min. After the reaction was complete, the precipitate was washed with distilled water and anhydrous ethanol two times in a turn. The obtained precipitate was dried in a vacuum desiccator at 60 °C for 8 h. After drying is complete, the obtained precipitate was recorded as nZVI and collected and stored in a sealed container.

Preparation of nZVI@SDBC (Qiu et al. 2020): 1.0 g of SBDC was dissolved in 100 mL of distilled water. Different masses (4.965 g, 9.930 g, 14.894 g, and 19.860 g) of FeSO₄·7H₂O were weighed according to the different mass ratios of Fe to SDBC (1:1, 2:1, 3:1, and 4:1). The different weights of FeSO₄·7H₂O were then dissolved in a solution containing SDBC. Different ratios of nZVI@SDBC were then prepared according to the method used for the synthesis of nZVI. The samples were labeled as nZVI@SDBC(1:1), nZVI @SDBC(2:1), nZVI@SDBC(3:1), and nZVI@SDBC(4:1) depending on the ratio.

The effect of Fe:SDBC mass ratio was investigated. 50 mL Cr(VI) solutions (50 mg/L) and 20 mg adsorbent were mixed. The pH of the solution (about 2.5) was adjusted to 3.0 by 0.1 mol/L HCl and NaOH, and the adsorption temperature was 25°C. Then, the mixed solution was reacted at a shaker of 150 r/min. After the reaction was completed, the Cr_total concentration of the solution was admeasured by an inductively coupled plasma optical emission spectrometer (ICP-OES, FL1008M018, Cary, USA)

The initial pH of the solution was set at 2, 3, 4, 5, 6, 7, 8, 9, and 10. The solution pH was measured by means of a pH meter (PHSJ-4A, Shanghai INESA Scientific Instrument Co., Ltd, Shanghai, China). The initial Cr(VI) concentration was 50 mg/L and the adsorption temperature was 25 °C. 50 mL Cr(VI) solution and 20 mg absorbent were added to a 150 mL flask. The flask was reacted in a shaker at 150 r/min for 300 min and the Cr(VI) concentration in the solution was measured after the reaction.

The effect of dosage (0.1, 0.2, 0.3, 0.4, 0.5, 1.0 and 2.0 g/L) on the Cr(VI) removal was investigated. The adsorption condition was pH 3.0, Cr(VI) concentration 50 mg/L and adsorption temperature 25 °C. 50 mL Cr(VI) solution and 20 mg absorbent were mixed, and then the solution was reacted in a thermostatic oscillator for 300 min.

The effect of coexisting cations (Ca²⁺, Na⁺, Mg²⁺, Pb²⁺, Fe³⁺ and Ni²⁺) and anions (Cl⁻, SiO²⁻₃, NO⁻₃, CO²⁻₃, SO²⁻₄ and PO³⁻₄) on the Cr(VI) removal was investigated under pH 3.0, Cr(VI) concentration 50 mg/L and 25 °C. 20 mg absorbent was added into 50 mL Cr(VI)-containing solution with relevant ion concentrations (0–50 mg/L) (supplementary materials). The solution was reacted for 300 min.

500 mL Cr(VI) solution (50 mg/L) and 200 mg absorbent were added to a 1,000 mL flask, and the initial pH was adjusted to 3.0. The mixture was reacted in the shaker at different temperatures (15–35 °C) to research the effect of reaction time (5–300 min) on the Cr(VI) removal. The experimental data was analyzed by the pseudo-first-order kinetic model and pseudo-second-order kinetic model (supplementary materials). The thermodynamic equations were used to calculate the thermodynamic parameters (supplementary materials). In addition, chi-square tests were used to test the fit of the kinetic models (Foo & Hameed 2010).

Under the condition of pH 3.0, 50 mL Cr(VI) solution of different concentrations (50–150 mg/L) and 20 mg absorbent were added to 150 mL flasks, and the reaction was carried out in thermostatic oscillator at 25 °C for 300 min. After the reaction was completed, the remaining Cr(VI) concentration in the solution was measured. The date of adsorption results was analyzed by the Langmuir isotherm model and modified Langmuir isotherm model, Freundlich isotherm model, and Temkin isotherm model (supplementary materials).

The mixture solution was filtered and nZVI@SDBC(2:1) was collected after reaction with Cr(VI). The adsorbed nZVI@SDBC(2:1) was added to 0.1 mol/L NaOH solution and shaken for 2 h. After desorption was completed, filtered the solution and collected the desorbed nZVI@SDBC(2:1). The desorbed nZVI@SDBC(2:1) was washed with distilled water three times and dried at 60 °C for 12 h. The above experiments were repeated five times and the Cr(VI) removal efficiency was calculated after each experiment.

The actual electroplating wastewater was taken from the Baohe Swan Electroplating Plant (Chengdu, China). The actual wastewater parameters are shown in the supplementary material (Table 1). 50 mL solutions and 20 mg nZVI@SDBC(2:1) were mixed. And then, the reaction was carried out in a thermostatic oscillator at 25 °C for 720 min.

As the mass ratio increased, the Cr(VI) removal efficiency by the nZVI@SDBC was the first to increase and then decrease. At a mass ratio of 2:1, the largest removal efficiency (98.35%) of Cr(VI) by nZVI@SBDC(2:1) was achieved. The reason for this was presumably that the loading nZVI can enhance the removal efficiency of SDBC (Qiu et al. 2020). Therefore, nZVI@SDBC(2:1) was chosen for the follow-up experiments.

The wavenumber at 3,420 cm⁻¹ in SDBC was associated with O-H stretching vibrations (Chen et al. 2015). The characteristic peak at 1,620 cm⁻¹ corresponded to the vibration of the C = O deformation conjugate (Qiu et al. 2020). The characteristic peak at a wavenumber of 1,430 cm⁻¹ was considered to be the -COOH group (Yi et al. 2020). However, the vibrational peak at 1,070 cm⁻¹ was thought to be due to C-O and Si-O groups (Qian et al. 2019; Qiu et al. 2020). After loading Fe⁰, the wavenumbers at 3,420 cm⁻¹, 1,630 cm⁻¹, and 1,410 cm⁻¹ showed a significant change, indicating that the O-containing groups (R-OH and R-COOH) in nVZI@SDBC(2:1) may chelate with the Fe ion (Jia et al. 2021). Furthermore, the peak at 565 cm⁻¹ corresponded to the Fe-O bond, which also indicated that nZVI was successfully loaded on SDBC (Qiu et al. 2020; Yi et al. 2020). Compared to nVZI@SDBC(2:1), there was no significant change in the FTIR spectrum after Cr(VI) removal. In contrast, the characteristic peaks of C = O and R-COOH were slightly shifted to 1,630 cm⁻¹ and 1,410 cm⁻¹, respectively. The intensity of the Fe-O at 526 cm⁻¹ increased, indicating that more Fe₂O₃ was produced after the reaction with Cr(VI) (Yi et al. 2020). The peak at 3,410 cm⁻¹ was also enhanced, and this phenomenon may be caused by the formation of Fe₂O₃. This result suggested that the R-OH and R-COOH in nVZI@SDBC(2:1) underwent complexation with Cr(VI) (Jia et al. 2021; Tang et al. 2021).

To further analyze the Fe species, XRD (Figure 3(f)) was used to analyze the Fe species in SDBC and nZVI@SDBC(2:1). 2θ = 21.28°, 26.82°, 35.16°, 36.94°, 42.76°, and 46.08° in the XRD pattern of SDBC corresponded to SiO₂, while 2θ = 29.74° and 39.74° corresponded to CaCO₃ (Chen et al. 2015; Qiu et al. 2020). After loading with nZVI, the peaks associated with SiO₂ and CaCO₃ in nZVI@SDBC(2:1) weakened or disappeared. Most notably, the presence of Fe⁰ was evidenced by the detection of a characteristic peak of 44.68° in the XRD pattern of nZVI@SDBC (2:1) (Qian et al. 2019; Qiu et al. 2020; Yi et al. 2020). The diffraction peak at 2θ = 35.48° was the Fe₂O₃ peak (Yi et al. 2020; Zhou et al. 2021). After Cr(VI) removal, the Fe⁰ peak disappeared. This could be the involvement of Fe0 in the removal of Cr(VI). The diffraction peak at 2θ = 35.50° was Cr₂FeO₄, presumably due to the reduction of Cr(VI) to Cr(III) by Fe⁰ and the formation of Cr(III)-Fe(II) precipitates (Qiu et al. 2020). After adsorption, the diffraction peak of Fe₂O₃ was enhanced, which may have allowed the reaction of Fe⁰ with Cr(VI) to form Fe₂O₃ (Yi et al. 2020).

The residual concentration of Cr_total, Cr(III), and Cr(VI) in the solution were measured separately (Figure 4(b)). The residual concentration of Cr_total and Cr(VI) increased with increasing solution initial pH. The residual concentration of Cr(III) first increased and then decreased, indicating that nZVI@SDBC(2:1) reduced some of the Cr(VI) to Cr(III) (Qian et al. 2019). The species distribution of Cr(VI) and Cr(III) in aqueous solution was modeled by Visual MINTEQ 3.1 software. At pH < 6.0, Cr(VI) was mainly present as the anions Cr₂O₇²⁻ (more than 90%) and CrO₄²⁻ (less than 10%) (Fig. S2a). At solution pH < 5.0, Cr(III) was mainly present as Cr³⁺ and Cr(OH)²⁺ (Fig. S2b). At solution pH less than the zero potential point (pH_PZC), the nZVI@SDBC(2:1) surface was protonated and positively charged (Qiu et al. 2020). As a result of the electrostatic effect, it will in turn enhance the Cr(VI) removal by nZVI@SDBC(2:1) and also promote the adsorption-reduction reaction of Cr(VI) to form Cr(III). Thus, electrostatic action contributed significantly to the removal of Cr(VI) under acidic conditions (Yi et al. 2020). However, the generated Cr(III) reacted with Fe(II) in solution by precipitation thereby forming a Cr₂FeO₄ precipitate (Qiu et al. 2020). When the solution pH was above the pH_PZC, the adsorbent surface was deprotonated and negatively charged. The adsorbent and Cr(VI) repealed each other and reduced the opportunity for oxidation-reduction, thus leading to a decrease in adsorption capacity (Yi et al. 2020). Additionally, Cr(III) generated by Fe⁰/Fe(II) reduction can form insoluble Cr(OH)₃ precipitates when the initial pH was above 5.0 (Equation (3)) (Yi et al. 2020). It had also been reported that under weakly acidic and basic conditions Cr(III) was able to form metal hydroxides (Cr_xFe_1−x(OH)₃/Cr_xFe_1−XOOH) with Fe³⁺ (Equations (4) and (5)) (Dong et al. 2017; Qiu et al. 2020). During the reaction, H⁺ was generated thus causing a drop in pH, which was also identical to the experimental data (Figure 3(a)) (Qiu et al. 2020; Tang et al. 2021).

To further understand, the adsorption-reduction of Cr(VI) by nZVI@SDBC(2:1). Fe_total, Fe²⁺, and Fe³⁺ were determined in the solution after reaction equilibration (Figure 4(c)). At lower pH, the leaching concentration of Fe was very high, which also provided more exposure of Cr(VI) to Fe0/Fe(II). Furthermore, the levels of Fe_total and Fe³⁺ were essentially similar, while the levels of Fe²⁺ were very small. One reason for this was that Fe(II) was oxidized by dissolved oxygen, but under acidic conditions, this process was very slow (Qiu et al. 2020). In addition, dissolved oxygen was limited in the closed vial, and Fe(II) was very stable. This also accelerated the reduction of Cr(VI) and the formation of Cr₂FeO₄ with Cr(III) in an aqueous solution, which was probably the reason why the amount of Fe(II) in the solution was very low (Yi et al. 2020). At the same time, the reduction process consumed a large amount of H⁺ in the solution, increasing the solution pH (Qiu et al. 2020). However, under weakly acidic or alkaline conditions, the nZVI may be oxidized, thus reducing the amount of Fe species available for Cr(VI) reduction. Therefore, this led to a reduction in the adsorption capacity of Cr(VI).

From an economic point of view, it was necessary to study the effect of adsorbent dosage on Cr(VI) removal. The Cr(VI) removal efficiency by nZVI@SDBC(2:1) gradually increased from 40.33% to 98.66% when the dosage was increased from 0.1 g/L to 0.4 g/L (Figure 4(d)). The reason for this was that with increasing dosage of nZVI@SDBC(2:1), the number of active sites increased rapidly, resulting in a rapid increase in Cr(VI) removal efficiency (Yi et al. 2020). After the dosage was greater than 0.4 g/L, the removal efficiency remained more or less constant. However, as the dosage continues to increase and the Cr(VI) content of the solution was limited, the reduction of Cr(VI) removal efficiency slowed down and stabilized. Yet, the Cr(VI) adsorption capacity of nZVI@SDBC(2:1) decreased with increasing dosage, from 201.65 mg/g to 24.84 mg/g. The Cr(VI) adsorption capacity decreased as the dosage increased, due to the decrease in Cr(VI) adsorption per unit mass of nZVI@SDBC(2:1). To cause unnecessary waste, a dosage of 0.4 g/L was chosen as the optimum dosage.

Ca²⁺ and Mg²⁺ are the main sources of hardness in natural waters, with Ca²⁺ being the most abundant element in most freshwaters. In contrast, Mg²⁺ is second only to Ca²⁺ and Na⁺ in natural waters. Pb²⁺, Fe³⁺, and Ni²⁺ are heavy metal elements essential in the manufacture of alloying materials. To further research the ability of the adsorbent to remove Cr(VI), the effect of coexisting cations on the Cr(VI) removal will be investigated (Figure 4(e)). The presence of Ca²⁺, Na⁺, and Mg²⁺ had negligible effect on Cr(VI) removal by nZVI@SDBC(2:1). However, the concentration of Pb²⁺, Fe³⁺, and Ni²⁺ in the solution was increased from 0 to 50 mg/L, and the removal efficiency of Cr(VI) by nZVI@SDBC(2:1) decreased to 53.78%, 83.89%, and 60.34%, respectively. It was presumed that coexists ions (Pb²⁺, Fe³⁺, and Ni²⁺) competed with Cr(VI) for adsorption sites (Wang et al. 2013). In addition, the competition strengthened as the concentration of coexisting ions increased, resulting in a significant inhibition of Cr(VI) removal.

In actual water bodies, carbonate ion (CO²⁻₃) is the main substance that makes up the alkalinity of the water column and is found at high levels in natural water bodies. Sulfate ion (SO²⁻₄), silicate ion (SiO²⁻₃), nitrate ion (NO⁻₃), phosphate ion (PO³⁻₄), and chloride ion (Cl⁻) are common anions in water bodies (Lv et al. 2013). Therefore, the effect of these anions on the Cr(VI) adsorption will be investigated (Figure 4(f)). The Cl⁻, SiO²⁻₃, NO⁻₃, and CO²⁻₃ had essentially no effect on the Cr(VI) removal. There was a significant inhibition of Cr(VI) removal by nZVI@SDBC(2:1) by SO²⁻₄ and PO³⁻₄, with the removal efficiency decreasing to 70.64% and 61.34% respectively. SO²⁻₄ and PO³⁻₄ had similar chemical structures to CrO²⁻₄ (one S, P, or Cr element combined with four O elements), and this had a strong inhibitory effect on the Cr(VI) removal (Zou et al. 2021).

To study the adsorption process and rate-controlling steps of the absorbent on Cr(VI), the pseudo-first-order kinetic model and pseudo-second-order kinetic model were chosen to analyze the experimental data. The fitted results and fitted parameters were exhibited in Figure 5 and Table 2, respectively. The linear correlation coefficient of the pseudo-second-order kinetic model was greater than that of the pseudo-first-order kinetic model (Dehghani et al. 2018). This result indicated that the adsorption process of Cr(VI) by nZVI@SDBC(2:1) was chemisorption (Jia et al. 2021; Zou et al. 2021). The chi-square test demonstrated that the χ²-value of the pseudo-second-order kinetic model was smaller than that of the pseudo-first-order kinetic model, which also confirmed that the pseudo-second-order kinetic model can better describe the adsorption process (Foo & Hameed 2010). Notably, the q_e fitted by the pseudo-first-order kinetic model differed significantly from the actual adsorption capacity, while the q_e fitted by the pseudo-second-order kinetic model was comparable to the actual value, indicating that the pseudo-second-order kinetic model can better describe the adsorption process (Zou et al. 2021). Besides, the adsorption equilibrium time was 150 min at adsorption temperatures of 15 °C and 20 °C. Whereas, at 25 °C, 30 °C, and 35 °C, the adsorption equilibrium time was 60 min.

The thermodynamic parameters were listed in Table 3. The enthalpy (ΔH₀ > 0) illustrated that the adsorption process belonged to endothermic reaction. Worthy of note, the negative ΔG₀ indicated the reaction was spontaneous (Dehghani et al. 2020; Yi et al. 2020). The gradual decrease of ΔG₀ with warming suggested that high temperature favored Cr(VI) adsorption. This phenomenon illustrated that the adsorption reaction of Cr(VI) was spontaneous and endothermic (Yi et al. 2020). The positive value of ΔS⁰ illustrated that the high temperature increased the degree of freedom at the interface between the solid and liquid phases of the adsorption system (Dehghani et al. 2018).

The adsorption isotherm is extremely important in determining the mode of action of the adsorbent on the adsorbate. The Langmiur and modified Langmiur (Figure 6(b)), Freundich (Figure 6(c)) and Temkin (Figure 6(d)) adsorption isotherm models were chosen to analyse the experimental data. The correlation coefficients of the modified Langmuir model (R² = 0.999) were both higher than those of the Freundlich model (R² = 0.784) and the Temkin model (R² = 0.800). The results showed that the maximum adsorption capacity (178.96 mg/g) obtained from the modified Langmuir equation simulations and the actual adsorption capacity (179.78 mg/g) were comparable (Table 4). This illustrated that the modified Langmuir model was more suitable to interpret the Cr(VI) adsorption process by nZVI@SDBC(2:1) (Zhai et al. 2021). The results also demonstrated that the modified model was able to better analyse the Cr(VI) adsorption process by nZVI@SDBC(2:1). The Cr(VI) adsorption on the adsorbent surface was homogeneous monolayer adsorption and there was no interaction between the adsorption sites (Qian et al. 2019; Zhao et al. 2020; Tang et al. 2021). The separation factor R_L = 0.003–0.009 demonstrated that the adsorption process was favorable (Dehghani et al. 2018). Compared with other adsorbents (Table S1), the q_max of nZVI@SDBC(2:1) for Cr(VI) removal in this study was better than most of the adsorbents, and nZVI@SDBC(2:1) can achieve faster adsorption equilibrium. This indicates that nZVI@SDBC(2:1) could better apply the removal of Cr(VI) from water-soluble.

In Figure 7(b), the peaks at 576.77 eV and 586.55 eV were Cr(III) peaks, and the peaks at 578.26 eV and 588.15 eV were Cr(VI) peaks (Qiu et al. 2020). The content of Cr(III) and Cr(VI) was 52.01% and 47.99%, respectively. This result further suggested that Cr(VI) was reduced by nZVI@SDBC(2:1) to Cr(III) (Yi et al. 2020). The main peaks of the pre-reaction O 1s were Fe-O (529.90 eV), C-O (531.07 eV), and C = O (532.15 eV) (Figure 7(c)) (Zhou et al. 2021). The increase in the peak area of Fe-O and decrease in the peak area of C = O after adsorption suggested that biochar can act as an electron transfer medium and participate in the reaction by gaining and losing electrons through certain functional groups on the surface (Qiu et al. 2020; Tang et al. 2021; Zou et al. 2021). Notably, the binding energies of C-O (531.27 eV) and C = O (532.23 eV) were increased after the reaction. This result may be attributed to the complexation of Cr(VI) with -OH or -COOH in nZVI@SDBC(2:1) (Pi et al. 2021). The weak peak in the Fe 2p spectrum of nZVI@SDBC(2:1) before adsorption (Figure 7(d)) attributable to Fe⁰ (706.92 eV) indicated the successful synthesis of nZVI (Qiu et al. 2020; Yi et al. 2020). Binding energies of 723.11 eV and 710.76 eV correspond to Fe(II) (Zhou et al. 2020; Zou et al. 2021). Binding energies of 725.26 eV and 712.78 eV correspond to Fe(III) (Tang et al. 2021). The disappearance of the Fe⁰ peak after adsorption, the decrease in the relative amount of Fe(II), and the increase in the relative amount of Fe(III) suggested that nZVI and Fe(II) were involved in the removal of Cr(VI) (Qiu et al. 2020; Yi et al. 2020). The XPS analysis illustrated that the -OH and -COOH in nZVI@SDBC(2:1) were related to the Cr(VI) adsorption, along with the reduction of Cr(VI) by nZVI (Jia et al. 2021; Pi et al. 2021; Zou et al. 2021).

From the above results, we can deduce the Cr(VI) removal mechanism on nZVI@SDBC(2:1).

• (1)

  Part of the Cr(VI) (Cr₂O²⁻₇/HCrO⁻₄) was removed directly by adsorption on nZVI@SDBC(2:1) through electrostatic interaction and complexation of O-containing functional groups (Qiu et al. 2020; Zou et al. 2021).

• (2)
  Part of nZVI was oxidised to Fe²⁺ (Equation (6)), and then Fe⁰/Fe²⁺ reduced Cr(VI) to Cr(III)(Cr³⁺/Cr(OH)²⁺) (Equations (7)–(10)) (Qiu et al. 2020; Yi et al. 2020; Zou et al. 2021). Finally, Cr(III) (Cr³⁺/Cr(OH)²⁺) combined with Fe²⁺ to become Cr₂FeO₄ precipitate (Equations (11) and (12)) (Tang et al. 2021; Zou et al. 2021). The above was the possible mechanism of Cr(VI) removal by nZVI@SDBC(2:1).

  

  (6)

  

  (7)

  

  (8)

  

  (9)

  

  (10)

  

  (11)

  

  (12)

When dealing with actual wastewater, an assessment of the production costs of the material is essential (Jaafarzadeh et al. 2018). The cost of electricity for industrial use in Chengdu (Sichuan, China) is 0.0916 KW/h. During the preparation of the SDBC, a tube resistance furnace (Power 4 KWh) was used for 2 h pyrolysis. In addition, the prices of FeSO₄·7H₂O and NaBH₄ were 49.1313 USD/t and 588.8130 USD/kg respectively. However, the cost per m³ of wastewater treated is 0.1936 USD/m³. At an optimum dosage of 0.4 g/L, 400 kg of nZVI@SDBC(2:1) is required per 1 m³ of wastewater treated. Additionally, nZVI@SDBC(2:1) had good reproducibility in the five-cycle experiment. The total cost of the process for treating 1 m³ of wastewater was 325.7162 USD/m³. Based on the cost calculation, nZVI@SDBC(2:1) is feasible for the treatment of Cr(VI)-containing wastewater.

The nZVI@SDBC material was successfully prepared by liquid-phase reduction using SDBC as a carrier. The nZVI loading on SDBC effectively avoided nZVI agglomeration. The Cr(VI) maximum adsorption on nZVI-SDBC(2:1) was 178.05 mg/g at pH 3.0, dosage 0.4 g/L and temperature 25 °C. The Cr(VI) adsorption on nZVI-SDBC(2:1) conform to the modified Langmuir isotherm model and the pseudo-second-order kinetic model, indicating that the adsorption process was homogeneous chemisorption. The Cr(VI) removal mechanism by nZVI@SDBC(2:1) was a complexation of O-containing groups, the formation of Cr(III)-Fe(II) precipitation, electrostatic interaction, and reduction. The Cr(VI) removal by nZVI@SDBC(2:1) was maintained at over 90% after five replicate experiments. nZVI@SDBC is capable of removing most of the Cr(VI) from real electroplating wastewater. The cost of using nZVI@SDBC(2:1) to remove 1 m³ of actual wastewater is approximately 325.7162 USD/m³. The results show that the nZVI@SDBC material has a good effect on the removal of Cr(VI)-containing wastewater generated during the preparation of metal materials in the construction of highland railways, as well as solves the problem of potential pollution generated during the construction of highland railways.

This work was supported by National Key R&D Program (2018YFC1505404) and Sichuan Provincial Science and Technology Department Science and Technology Plan Project (No.2021YJ0033).

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.