Passivation remediation of weakly alkaline Cd-contaminated soils using combined treatments of biochar and sepiolite

Abstract

Background

Cadmium (Cd) pollution in agricultural soils has become a priority environmental concern globally. A reasonable application of passivators is critical to address the problem. In this study, we examined the remediation effects of rice husk biochar (rBC) and sepiolite (Sep) as single and combined (rBC + Sep) treatments on Cd pollution in a weakly alkaline soil using three maize cultivars (Liyu 16, Zhengdan 958, and Sanbei 218) as test crops. We also explained the mechanisms involved in the remediation effects.

Results

The pseudo-second-order kinetic equation and Langmuir model could well describe the adsorption process of rBC + Sep for Cd²⁺. Compared with the control treatment (CK), soil available Cd concentration decreased by 29.51–36.34% under rBC + Sep treatment (p< 0.05) and the Cd concentrations in maize grains of Liyu 16, Zhengdan 958, and Sanbei 218 decreased by 38.08–47.85%, 37.25–45.61%, and 33.96–46.15%, respectively (p< 0.05). Following passivation treatment, soil available Cd concentration decreased and gradually changed from the exchangeable and carbonate binding forms to the Fe/Mn oxide and residual forms. The bioconcentration factors of Liyu 16 (0.05–0.09) and Sanbei 218 (0.05–0.09) were lower than those of Zhengdan 958 (0.07–0.13). In addition, rBC +Sep treatment increased soil pH and soil electrical conductivity, but the differences were not significant (p> 0.05).

Conclusions

The application of 0.2% rBC + 0.5% Sep composite passivation material to weakly alkaline Cd-contaminated soil can effectively reduce the Cd concentration of soil and maize.

Introduction

Soil cadmium (Cd) pollution has become an important problem affecting human health and environmental safety in China and many other patrs of the world (Fan et al. 2020). According to the National Soil Pollution Status Survey Bulletin, in 2014, 7.0% of soils in China exceeded the standard for Cd pollution (Zhao et al. 2015). Statistics indicate that approximately 10 million ha of land in China is contaminated by heavy metals and over 12 million tons of grains are contaminated by heavy metals each year (Zhou et al. 2014; Sun et al. 2017). Cd is a toxic heavy metal that is silver-white in color and has a long half-life, making it difficult to reduce in the environment and easy to migrate and transform. The entry of Cd into the food chain poses a threat to human health (Fan et al. 2020). Therefore, the remediation of Cd pollution in soil environments is urgent.

In recent years, in-situ passivation remediation has emerged as a promising technique for treating contaminated soils because of its high efficiency and low cost. Commonly used passivators include clay minerals and biochar. Biochar is a carbon-rich material produced from biomass under high-temperature and oxygen-limited conditions. The porous structure and oxygen-rich functional groups on the surface of biochar enable the adsorption and passivation of heavy metals through mechanisms such as surface complexation, ion exchange, electrostatic adsorption, and coprecipitation (Ahmad et al. 2014; Xiao et al. 2018; Cheng et al. 2020). Additionally, biochar can be produced from a wide range of raw materials, providing a waste-to-resource solution and enhancing crop yield. Clay minerals possess large pore volumes and specific surface areas and are effective in passivating heavy metal-contaminated soils through adsorption, ion exchange, and co-precipitation, as well as in reducing Cd absorption in crops (Covelo et al. 2007; Song et al. 2011). Both biochar and clay minerals possess excellent adsorption capabilities owing to their unique surface structures, making them highly desirable passivators for the study of soil heavy metal pollution. Zhang et al. (2018) found that a montmorillonite and wheat straw biochar composite was effective in adsorbing norfloxacin from solution through electrostatic interactions, hydrogen bonding, and pore filling. Xie (2020) investigated the adsorption of Pb²⁺ in weakly alkaline soil using a montmorillonite and rice husk biochar (rBC) composite and found that the primary mechanism was chemisorption, with physical adsorption also playing a role. As different passivators exhibit varying degrees of adsorption capacity for heavy metals, the selection of an appropriate passivator has become a key factor in the success of in-situ passivation and remediation.

Sepiolite (Sep) is a natural clay mineral that possesses a unique fibrous porous structure and large surface area, which enables it to adsorb both organic and inorganic ions with good potential for heavy metal adsorption (Abad-Valle et al. 2016). As the pH increases, surface complexation and isomorphous substitution coexist, enhancing its ability to adsorb and passivation heavy metals (Basta and McGowen 2004; Liang et al. 2014). However, studies have shown that when Sep is applied as a single passivation agent, its effect on Cd passivation decreases over time (Zhang et al. 2023).

This study aimed to examine the combined use of rBC and Sep for the prevention and control of Cd pollution in soil. To achieve this, we studied the passivation effects and mechanisms of different compound ratios of rBC and Sep on Cd in a weakly alkaline agricultural soil using three maize cultivars (Liyu 16, Zhengdan 958, and Sanbei 218) as test crops. The results of this study can be used to inform environmental management for Cd and other heavy metal pollution.

Materials and methods

Experimental site

The experimental site was located in a Cd-polluted farmland in Nansunzhuang, Dongli District, Tianjin City (39°09′N, 117°31′E), China, which belongs to the Haihe River Basin and has a temperate semi-humid continental monsoon climate. The average annual precipitation is 643.8 mm, the frost-free period is 237 d, and the average annual temperature is 13.5 ºC. The soil is classified as fluvo-aquic soil with a pH of 8.33, electric conductivity (EC) of 256.3 µS·cm⁻¹, and soil dissolved organic carbon (DOC) of 93.96 mg·kg⁻¹. The soil Cd, Pb, As, and Cr concentrations are 1.30, 29.79, 17.41, and 36.67 mg·kg⁻¹, respectively.

Material preparation and characterization

The rice husk was air-dried, sieved to remove gravel and debris, and ground using a high-speed universal grinder. The biochar was produced by pyrolyzing the rice husk at 600 ºC for 2 h under anaerobic conditions using a continuous carbonization machine. After the pyrolysis, the biochar was cooled to room temperature and stored in bags. The basic chemical properties of the passivation agents are listed in Table 1. The surface morphologies and structural characteristics of the samples were analyzed using a scanning electron microscope (SEM; Hitachi SU3500) equipped with an energy-dispersive spectrometer and a transmission electron microscope (EDS; JEOL JEM2010). The phase compositions and crystal structures of the samples were obtained using X-ray diffraction (XRD; Rigaku D/Max 2500) in the scanning range of 10–90°. Fourier transform infrared spectrophotometer (FTIR; Nicolet 380) was used to determine changes in functional groups, with a measured wavenumber range of 400–4000 cm⁻¹ and a resolution of 1 cm⁻¹.

Table 1 Basic chemical properties of amendments

Batch adsorption experiment

A total of 1 g of rBC, 1 g of Sep, and a mixture of 0.5 g of rBC and 0.5 g of Sep (rBC + Sep) were separately placed into 500 mL of Cd nitrate solution with an initial concentration of 100 mg·L⁻¹ of Cd²⁺. The solutions were stirred constantly for 24 h at 25 ºC and 300 rpm using a magnetic stirrer. Samples were taken at various time intervals (1, 3, 5, 10, 15, 20, 30, 40, 60, 120, 180, 240, 300, 360, 480, 600, 720, and 1440 min) and the Cd²⁺ concentration was measured using inductively coupled plasma mass spectrometry (ICP-MS; iCAP Q, USA) after passing through 0.45-μm filter.

For a series of concentrations of Cd²⁺ (30, 60, 90, 120, 150, and 200 mg·L⁻¹), 0.1 g of the adsorbent was added and the optimal adsorption time was determined by shaking the solution at 180 rpm and 25 ºC. The supernatant was filtered through a 0.45-μm filter and the Cd²⁺ concentration was determined using ICP-MS.

Field experiment

In May 2020, a field experiment was conducted to remediate the soil pollution. The experiment consisted of eight treatments, including a control treatment with no application of a remediation agent (CK) and treatments with the application of 0.1% and 0.2% rBC (A1, A2), 0.2% and 0.5% Sep (S1, S2), 0.1% rBC and 0.2% Sep (A1S1), 0.2% rBC and 0.2% Sep (A2S1), and 0.2% rBC and 0.5% Sep (A2S2). Each treatment was replicated three times in a single plot area of 20 m². Three cultivars of maize (Zea mays L.) were planted in each treatment: Zhengdan 958 (common cultivar), Liyu 16 (low-accumulation cultivar), and Sanbei 218 (low-accumulation cultivar), which were provided by the Tianjin Luhe Seed Co. The remediation agent was applied evenly to the topsoil (0–20 cm) of each plot using a tillage machine and mixed uniformly. After a 20-day equilibrium period, maize was planted and managed under normal production conditions, including regular weeding, insecticide application, and watering. Maize samples were harvested, dried, and sieved in September 2020. The Cd concentrations of the roots, stems, leaves, and grains of maize were measured. Soil samples were collected from the experimental plots, excluding residues such as gravel, plant roots, and stems. After grinding and air-drying, the samples were passed through 20-, 50-, and 100-mesh nylon sieves to determine the available Cd concentration and fraction of the soil.

Analysis and determination

Soil sample analysis

The total Cd concentration in the soil was determined using HNO₃-HF and the Cd concentration in the digestion solution was determined using ICP-MS (Ultimate 3000-i CAP QC). The available Cd in the soil was extracted using the chelating agent diethylenetriaminepentaacetic acid (DTPA) at a solid-to-liquid ratio of 1:5, and the Cd concentration in the supernatant was determined using ICP-MS after extraction. Quality assurance and control were performed using duplicate samples with three replicates, blanks, and standard reference materials (NIST 2586).

The speciation of Cd in the soil was determined using the Tessier sequential extraction method considering five forms: exchangeable, carbonate-bound, Fe/Mn oxide-bound, organic-bound, and residual. Exchangeable Cd was extracted using MgCl₂·6H₂O, with a solid-to-liquid ratio of 8:1; carbonate-bound Cd was extracted using NaOAc, with a solid-to-liquid ratio of 8:1; Fe/Mn oxides-bound Cd was extracted using NH₂OH·HCl, with a solid-to-liquid ratio of 20:1; organic-bound Cd was extracted using HNO₃ and H₂O₂, with a solid-to-liquid ratio of 8:1; and residual Cd was calculated by subtracting the sum of Cd concentration in the first fourth steps from the total soil Cd concentration.

Soil pH was measured using the glass electrode method using a pH meter (HJ 615-2011) (1:2.5 w/v). The sample EC was measured using a conductivity meter (Five Easy Plus; Mettler Toledo) (1:5 w/v). Soil dissolved organic carbon concentration was determined using the potassium dichromate oxidation method (Wang et al. 2018). Catalase activity was determined by potassium permanganate titration (Bandara et al. 2017). Urease activity was determined using indophenol blue colorimetry and that of alkaline phosphatasewas determined using disodium phenyl phosphate colorimetry (Meena et al. 2016).

Determination of Cd concentration in maize

The Cd concentrations of the roots, stems, leaves, and grains of maize were determined using the HNO₃ digestion method. Wheat flour (GBW08503c) was used for quality control and the Cd concentrations in the digested solution were measured using ICP-MS.

Data analysis

All treatments were independently repeated three times and correlation analysis and graphing were performed using SPSS 17.0 and Origin 2018, respectively.

The Cd²⁺ adsorption capacity (q) was calculated as follows:

$$ q\,=\,\frac{\left({\text{C}}_{0}-{\text{C}}_{\text{t}}\right)\times{\text{V}}}{\text{m}}$$

(1)

where m is the amount of adsorbent material (g); V is the total volume of the solution (L); and C₀ and C_t are the mass concentration of Cd²⁺ in the solution (mg·L⁻¹) before adsorption and at adsorption time t, respectively.

The pseudo-first-order kinetic Eq. (2) and pseudo-second-order kinetic Eq. (3) were used to fit the adsorption kinetics of Cd²⁺ on different adsorption materials.

$${{q}}_{{t}}\,=\,{{q}}_{{e}}{(1-}{{e}}^{-{{K}}_{1}{{t}}}{)}$$

(2)

$${{q}}_{{t}}\,=\,\frac{{{K}}_{2}{{{q}}}_{{e}}^{2}{{t}}}{\text{1} + {{K}}_{2}{{{q}}}_{{e}}{{t}}}$$

(3)

where K₁ is the rate constant (min⁻¹) of the pseudo-first-order kinetic model; t is the adsorption time (min); K₂ is the rate constant of pseudo-second-order kinetic model (g·mg⁻¹·min⁻¹); and q_e and q_t are the Cd²⁺ adsorption capacity (mg·g⁻¹) at adsorption equilibrium and t, respectively.

The Langmuir and Freundlich models were used for the non-linear fitting of the adsorption isotherm and were calculated using Eqs. (4) and (5), respectively:

$${{q}}_{{e}}\,=\,\frac{{{q}}_{{m}}{{{K}}}_{{L}}{{{C}}}_{{e}}}{{1} + {{K}}_{{L}}{{{C}}}_{{e}}}$$

(4)

$${{q}}_{{e}}\,=\,{{{K}}_{{F}}{{C}}}_{{e}}^{\frac{1}{{\text{n}}}}$$

(5)

where q_e is the equilibrium adsorption capacity (mg·g⁻¹), C_e is the equilibrium concentration (mg·L⁻¹), q_m is the saturated adsorption capacity (mg·g⁻¹), K_L (L·mg⁻¹) and K_F (mg·g⁻¹·L·mg⁻¹)^1/n are Langmuir adsorption constants and Freundlich adsorption constants, respectively, and n is a constant that indicates the adsorption strength and varies with the inhomogeneity of the material.

The accumulation and transport of Cd in maize grains were calculated using the bioconcentration factor (BCF) (Eq. (6)) and translocation factor (TF) (Eq. (7)). The translocation factor refers to the transport of heavy metals within the plants. The larger the bioconcentration factor, the faster the plant absorbs heavy metals, and the larger the transport coefficient, the stronger the ability of plants to transport heavy metals from roots to shoots (Wang et al. 2019).

$$\text{BCF} = \frac{{{C}}_{\text{grain}}}{{{C}}_{\text{soil}}}$$

(6)

$$\text{TF} = \frac{{{C}}_{\text{grain}}}{{{C}}_{\text{root}}}$$

(7)

where C_grain, C_root, and C_soil (mg·kg⁻¹) are the Cd concentrations of maize grains, roots, and soil, respectively.

Results and discussion

Performance of Cd adsorption by rBC, Sep, and rBC + Sep

The fitting curves of the adsorption kinetic models for Cd²⁺ by rBC, Sep, and rBC + Sep are shown in Fig. 1a. The adsorption capacity of the different adsorption materials for Cd²⁺ showed a rapidly increasing pattern (0–120 min) and reached adsorption equilibrium after 120 min. This could be attributed to the presence of abundant adsorption sites on the adsorbent during the initial stage. As the adsorption time increased, the gradual saturation of the available adsorption sites resulted in a decrease in the adsorption rate (Mane and Babu 2013). As shown in Table 2, the pseudo-second-order kinetic fitting R² values of the three adsorbents were 0.92, 0.93, and 0.97, respectively, which were higher than those of the pseudo-first-order kinetics. The pseudo-second-order kinetic model better fitted the adsorption process of Cd²⁺ by rBC, Sep, and rBC + Sep, indicating that the adsorption process of Cd²⁺ was mainly chemical.

Fig. 1

Adsorption kinetics a and adsorption isotherm b of Cd²⁺ on rBC, Sep, and rBC + Sep

Table 2 Adsorption kinetic parameters of Cd²⁺ on rBC, Sep, and rBC + Sep

Langmuir and Freundlich models were used to fit the adsorption isotherm data of Cd²⁺ onto the adsorbent (Fig. 1b), and the fitting parameters are shown in Table 3. The results revealed that the Langmuir isotherm model provided a better fit for the adsorption of Cd²⁺ (R²> 0.97). This suggests that the adsorption of Cd²⁺ by rBC and Sep primarily involved a uniform adsorption process in a monolayer, which is consistent with the findings of Foo and Hameed (2010). The theoretical maximum adsorption amount of rBC, Sep, rBC + Sep for Cd²⁺ predicted by the Langmuir model was in the order of rBC + Sep (43.53 mg·g⁻¹) > rBC (27.88 mg·g⁻¹) > Sep (17.77 mg·g⁻¹), which agreed with the results of the adsorption kinetics experiment.

Table 3 Isotherm parameters for Cd²⁺adsorption on rBC, Sep, and rBC + Sep

Adsorption mechanisms of rBC, Sep, and rBC + Sep on Cd

The surface structure and elemental concentration of the biochar were characterized using SEM–EDS. Figure 2a–c show the rBC and Sep materials before Cd²⁺ adsorption and Fig. 2d–f show the rBC and Sep materials after Cd²⁺ adsorption. rBC before adsorption shows a layered and overlapping shape, with a loose texture and many wrinkles, whereas Sep exhibits a smooth, straight fiber structure. After adsorption, Cd²⁺ appeared uniformly on the surface of the adsorbent, indicating that Cd²⁺ was successfully adsorbed. From a structural perspective, the successful adsorption of Cd²⁺ may be due to the abundant pores and internal cavity structures on the surface of rBC as well as the large smooth surface of Sep. The EDS spectrum analysis revealed that Sep contains a large amount of metal elements and it is tentatively speculated that Cd²⁺ may undergo a chemical reaction with certain metal elements in Sep to form precipitates (Zhang et al. 2019).

Fig. 2

SEM–EDS image of rBC, Sep, and rBC + Sep before a–c and after d–f Cd²⁺ adsorption

The FTIR spectra of rBC, Sep, and rBC + Sep is shown in Fig. 3. A peak at 3427 cm⁻¹ was observed in rBC, Sep, and rBC + Sep, which corresponded to the stretching vibration of -OH and adsorption (Chang et al. 2019; Li et al. 2015). The two weak absorption peaks observed at approximately 2916 and 2848 cm⁻¹ were attributed to the stretching vibrations of -CH₂ and -CH₃, respectively. The absorption peak at 1432 cm⁻¹ was assigned to the stretching vibration of -CH₂. The vibration peak of rBC + Sep shifted to around 1594 cm⁻¹, corresponding to the C = C vibration peak. The characteristic peak at 1018 cm⁻¹ was assigned to the stretching vibration of -C-O. Compared to rBC, Sep exhibited a sharper and larger absorption peak at 1201 cm⁻¹, corresponding to the stretching vibration of the C-O bond. The absorption peaks at 875 and 667 cm⁻¹ indicated the presence of single and polycyclic compounds. The peak at 464 cm⁻¹ was attributed to the bending vibration of the Si-O-Si bond.

Fig. 3

The FTIR spectra of rBC, Spe and rBC + Spe before and after Cd²⁺ adsorption

Upon the adsorption of Cd²⁺, the characteristic peaks representing -OH in the biochar decreased from approximately 3427 to 3388 cm⁻¹, indicating a possible reaction between Cd²⁺ and -OH on the biochar surface, leading to the corresponding carbonate or hydroxide precipitation (Xu et al. 2013). Furthermore, an enhancement of the broad peak at 1018 cm⁻¹ was observed upon the adsorption of Cd²⁺ following the Sep treatment, suggesting the potential involvement of Si-O-Si groups on the surface of the Sep in the adsorption of Cd²⁺.

The XRD analysis of rBC, Sep, and their composite materials revealed a broad diffraction peak of rBC at approximately 2θ= 25°, which is a characteristic peak of graphite structure (Fig. 4). The peaks of rBC-Cd at 2θ= 23.5°, 30.3°, 36.4°, 43.8°, 49.9°, and 58.2° correspond to crystal planes (012), (104), (110), (202), (116), and (122) of CdCO₃, respectively, suggesting that precipitation is the adsorption mechanism of Cd²⁺ on rBC (Fig. 4a). The XRD pattern of Sep showed diffraction peaks at 2θ= 23.0°, 29.4°, 47.5°, and 48.5°, corresponding to crystal planes (012), (104), (018), and (116) of CaCO₃, respectively. Compared with that of Sep, the XRD pattern of Sep-Cd showed a new peak at 2θ= 31.7° (Fig. 4b), which was attributed to the crystal plane (-111) of Cd(OH)₂, indicating that the precipitation of O–H groups and Cd²⁺ may be one of the mechanisms through which Sep adsorbs Cd²⁺ (Xu et al. 2009). After the adsorption of Cd²⁺ by the composite material of rBC and Sep, the XRD pattern of the composite material showed the characteristic peaks of CdCO₃ and Cd(OH)₂. Specifically, in the rBC + Sep-Cd material, there were peaks at 2θ= 23.5°, 36.4°, 47.5°, and 48.5°, corresponding to crystal planes (012), (110), (018), and (116) of CdCO₃, respectively, and a peak at 2θ= 31.7°, corresponding to the crystal plane (-111) of Cd(OH)₂. This indicated that carbonate precipitation occurred during the adsorption of Cd²⁺ by rBC + Sep (Fig. 4c).

Fig. 4

The XRD patterns of rBC (a), Sep (b) and rBC + Sep (c) before and after Cd²⁺ adsorption

Effects of rBC, Sep, and rBC + Sep on available Cd concentration and Cd fractions in soil

As shown in Fig. 5, the application of rBC, Sep, and rBC + Sep resulted in a decrease in the soil available Cd concentration with an increase in the amount of adsorbent applied compared to the CK treatment. Therefore, passivating agents play a crucial role in reducing the soil Cd concentration. In the single material adsorption, the available Cd concentration of soil treated with A2 and S2 decreased by 25.88% and 34.54%, respectively, compared with CK. In the combined treatment, the available Cd concentration in the soil treated with A2S2 was the lowest, showing a 38.24% reduction compared with CK (p< 0.05).

Fig. 5

Soil available Cd concentration under different treatments. Different letters show significant differences (p< 0.05)

For the distribution of heavy metal forms, exchangeable and carbonate-bound states are considered to be biologically available, whereas residual states are considered to be the most stable and least biologically available form (Lock et al. 2003). In soil without the application of passivating agents, the predominant forms of Cd were the Fe/Mn oxide-bound state (35.24%) and the exchangeable state (28.30%), with the residual state accounting for only 6.45% and possessing some potential biological availability (Fig. 6), consistent with the findings of Zhang et al. (2020). The application of passivating agents decreased the proportion of exchangeable Cd in the soil to varying degrees, with the decreasing range increasing with the amount of passivating agent applied, ranging from 24.33 to 34.73%. Compared to CK, the exchangeable Cd decreased by 26.60% and 28.97%, respectively, when rBC was applied alone. This is because rBC has a porous carbon framework structure with clear and distinct pore profiles that can more effectively adsorb and retain heavy metals in the soil, thereby reducing their biological availability (Beesley et al. 2011). Compared to the application of rBC alone, S2 achieved the largest decrease in exchangeable Cd at 34.73%. Thawornchaisit and Polprasert (2009) found that different passivating agents had different effects on soil remediation efficiency, which is consistent with the results of this study. Furthermore, the proportion of carbonate-bound Cd decreased to varying degrees. The proportion of carbonate-bound Cd accounted for 25.44% without passivation treatment but decreased by 19.82% to 23.53% following passivation treatment. The reduction was smaller when rBC and Sep were applied separately, decreasing by 0.02–0.06 units. When 0.2% rBC and 0.5% Sep were applied simultaneously, the proportion of carbonate-bound Cd decreased to 19.96%. After passivation treatment, the available Cd concentration in the soil decreased, gradually transforming from exchangeable and carbonate-bound forms to Fe/Mn oxides and residual forms, resulting in reduced mobility and bioavailability (Seshadri et al. 2017). This finding was consistent with the results of the present study. In this study, the Fe/Mn oxide-bound Cd increased to varying degrees (3.89% to 10.71%) following passivation treatment. The proportion of Fe/Mn oxide-bound Cd in the CK treatment was 35.24%, and the differences under different application rates were not significant (p> 0.05). The passivation treatment effectively increased the proportion of residual Cd. The passivation treatment effectively increased the proportion of residual Cd by 115.45–294.20%. Uchimiya et al. (2010) also showed that the input of biochar promoted the transformation of exchangeable Cd into residual forms, possibly because residual heavy metals are generally composed of primary and secondary silicates, sulfides, and other stable secondary minerals. In addition, biochar input increases soil pH and available silicon concentration, leading to the formation of new, structurally stable silicate precipitates with soil heavy metal ions and silicate ions, thereby increasing the proportion of residual forms (Neumann and zur Nieden 2001).

Fig. 6

Soil Cd fractions under different treatments. Exchangeable (F1), Carbonate-bound (F2), Fe/Mn oxides-bound (F3), Organic (F4), Residual (F5)

Effects of rBC, Sep, and rBC + Sep on Cd concentration in maize

Figure 7 depicts the impact of rBC and Sep treatments on Cd accumulation in various parts of the three maize cultivars. The Cd accumulation pattern was in the order of roots > leaves > stems > grains. Following the application of the passivation agents to the soil, a significant reduction in Cd concentration was observed in all parts of the maize plants. The range of Cd reduction in the roots of Liyu 16 varied from 9.66 to 32.77%. The decrease in Liyu 16 was the largest, although it was not statistically different from that in the two other cultivars (p> 0.05). However, a significant reduction in Cd concentration was observed in the roots of Zhengdan 958 and Sanbei 218 (p< 0.05). Under 0.2% rBC + 0.5% Sep, the maximum reductions in Cd concentration were 40.04% and 46.07% in the roots of Zhengdan 958 and Sanbei 218, respectively. A significant reduction in Cd content was also observed in the stems of all three maize cultivars (p< 0.05), with the reduction increasing with increasing passivation concentration. The reduction in Cd concentration in the stems of Liyu 16, Zhengdan 958, and Sanbei 218 ranged from 10.27–40.47%, 3.12–36.53%, and 6.60–53.60%, respectively. In the leaves of Zhengdan 958, A2S2 treatment decreased Cd concentration by 16.05%, which was an insignificant reduction in Cd concentration (p> 0.05) compared with CK treatment. However, a significant reduction in Cd concentration was observed in the leaves of Liyu 16 and Sanbei 218 (p< 0.05), with maximum reductions of 32.62% and 42.10%, respectively.

Fig. 7

Effects of different treatments on Cd concentration of different parts of maize for different cultivars. (a): Liyu 16, (b): Zhengdan 958, (c): Sanbei 218. Error bars represent ± SE (n = 3). Different letters show significant differences (p< 0.05)

In this experiment, the accumulation of Cd in grains of the three cultivars of maize in the CK exceeded the Cd limit of ≤ 0.1 mg·kg⁻¹ set by the China's National Food Safety Standard – Maximum Levels for Contaminants in Foods (GB2762-2022). Following the application of passivation agents, a significant reduction in Cd concentration was observed in the grains of all three maize cultivars (p< 0.05), with reductions of 13.80% to 31.40% observed in the grains of Zhengdan 958. Under A2S2 treatment, the Cd concentration in the grains of Liyu 16, Zhengdan 958 and Sanbei 218 decreased by 47.85%, 45.61% and 46.15%, respectively, indicating that the application of 0.2% rBC + 0.5% Sep passivation material had the greatest inhibitory effect on Cd in the grains of Liyu 16, which may be attributed to the different genotypes and tolerance mechanisms of the same plant species. Further studies are needed to elucidate the underlying mechanism (Bhargava et al. 2012). The Cd concentration in the grain of the low-accumulating cultivar Liyu 16 was lower than 0.1 mg·kg⁻¹ (except for S1 treatment) after treatment, which met the standard. However, under A1, S1, S2, and A1S1 treatments of the common cultivar Zhengdan 958 and the S1 treatment of the low-accumulating cultivar Sanbei 218, the Cd concentration in the grain was above 0.1 mg·kg⁻¹. Among them, the most effective treatment for reducing Cd concentration in grains was A2S2 treatment, which showed reductions of 47.85%, 45.61%, and 46.15% compared to CK. The combined treatment of rBC + Sep effectively reduced the Cd concentration in grains, which met the requirements of the above standards.

Effects of rBC, Sep, and rBC + Sep on transport ability of Cd in maize

We evaluated the ability of plants to accumulate and transport the Cd using bioconcentration and translocation factors (Cui et al. 2017). The Cd bioconcentration factors of the two low-accumulating maize cultivars, Liyu 16 and Sanbei 218, were significantly lower than that of the common cultivar Zhengdan 958 (Fig. 8a). This finding was consistent with that of Wu et al. (2005), who found differences in Cd absorption, transport, and accumulation among different wheat genotypes. This phenomenon may be attributed to the genetic regulation of Cd accumulation in crops, which is closely associated with the ability of different cellular organelles in the roots, stems, and leaves to fix, store, and tolerate Cd, as well as the influence of genes involved in Cd regulation (Yue et al. 2018; Feng et al. 2020). Moreover, after applying the passivator, the Cd bioconcentration factor in the grains of the three maize cultivars significantly decreased (p< 0.05). In Liyu 16, the bioconcentration factor of Cd in grains decreased by 2.89–20.37% compared to the CK. The Cd bioconcentration factor in Zhengdan 958 ranged from 0.07 to 0.13, with a mean of 0.09. In Sanbei 218, under the A2S2 treatment, the Cd bioconcentration factor decreased significantly by 17.71% compared with CK (p< 0.05). Variance in heavy metal uptake and adaptation capabilities among different maize cultivars may be attributed to various factors, including differential gene expressions within the crop and disparities in the application of passivating agents (Feng et al. 2020; Guo et al. 2018).

Fig. 8

The bioconcentration (a) and translocation factors (b) of Cd for different treatments. Different letters show significant differences (p< 0.05)

A higher translocation factor indicates a stronger ability of the heavy metal to transfer from the roots to the aboveground parts (Sun et al. 2008). Zhi et al. (2014) suggested that the criteria for screening low-accumulating crop cultivars must satisfy both the bioconcentration and translocation factors of the cultivar of less than 1. In this study, the Cd translocation factor of the three maize cultivars from roots to the grains was less than 1 (Fig. 8b), indicating that Cd mainly accumulated in the roots of maize. In Liyu 16, the Cd translocation factor from roots to grains ranged from 0.24 to 0.35, with a mean of 0.29. In Zhengdan 958, the Cd translocation factor ranged from 0.19 to 0.35, with a mean of 0.30. In Sanbei 218, the Cd translocation factor ranged from 0.21 to 0.27, with a mean of 0.25. These findings indicated that the Cd transferred from the soil to the grains was very low and most of the Cd was intercepted by the roots, stems, and leaves. Due to the decrease of Cd concentration in different parts of maize under the treatment of adding passivating agent, it is feasible to use passivating agent to repair Cd-contaminated farmlands.

Influence of soil chemical properties and enzyme activities under different treatments

As shown in Table4, the application of rBC + Sep resulted in a moderate increase in soil pH compared to CK but the differences were insignificant (p> 0.05). The magnitude of the increase generally increased with the application of passivating agents, with a pH increasing by 0.17 under A2S2 treatment. Compared with CK, except for S1 treatment, soil electrical conductivity under each treatment increased to varying degrees but the differences were insignificant (p> 0.05). When the application rates of rBC and Sep were 0.2% and 0.2%, the electrical conductivity increased by 25.10%. Except for under A2, soil dissolved organic carbon concentration showed no significant difference compared to CK (except for S1, p> 0.05), with increases of 6.91–14.28%.

Table 4 Effects of different treatments on basic properties of soil

Soil enzyme activity, as an organic component of the soil, is involved in the transformation of substances and energy in the soil, thereby affecting soil metabolism (Jimenez et al. 2002). Soil enzymes are sensitive to heavy metals (Zhang et al. 2015) and their activity is affected to varying degrees when exposed to heavy metal stress. In this study, after the application of different rates of rBC and Sep, the activities of urease, alkaline phosphatase, and catalase in the soil increased to varying degrees compared to CK. Urease activity exhibited an initial increase and subsequent decrease when rBC and Sep were applied in combination. When the Sep application rate was 0.2% and the rBC application rate was 0.2%, the activity of soil urease was higher than that at an application rate of 0.1% for Sep. However, when the Sep application rate increased to 0.5%, the activity of urease decreased. Under A2S1, the activity of soil urease increased significantly by 46.86% (p< 0.05) compared to CK (Fig. 9a). Under A2S1 and A2S2, the activity of soil alkaline phosphatase increased significantly by 3.99 and 3.68 units, respectively (p< 0.05) (Fig. 9b). Application of various concentrations of passivators induced an increase in catalase activity by 6.16–11.20%. Notably, the highest catalase activity was achieved when a rate of 0.5% Sep was applied, exhibiting a significant difference compared to that of CK (p< 0.05) (Fig. 9c).

Fig. 9

Effects of different treatments on soil urease (a), alkaline phosphatase (b) and catalase (c) activities. Different letters show significant differences (p< 0.05)

Conclusion

Pseudo-second-order kinetic and Langmuir models could well explain the adsorption process of Cd by rBC, Sep, and rBC + Sep, indicating that the adsorption of Cd²⁺ was mainly chemical. The combined application of rBC and Sep significantly reduced the concentration of DTPA-Cd in the soil, which tended to change from a more active exchangeable and carbonate-bound state to a more stable residual and Fe/Mn oxidation-bound state. In addition, Cd concentrations in roots, stems, leaves and grains of Liyu 16, Sanbei 218 and Zhengdan 958 were decreased under different passivator treatments. The application of single and composite materials can increase soil enzyme activity, as well as soil pH and electrical conductivity, but the differences for the latter two were not significant. Therefore, it is recommended to apply 0.2% rBC + 0.5% Sep combined material in weakly alkaline Cd-contaminated soil.