1 Introduction

Soil potentially toxic metals pollution has gained great attention globally, not only pose a serious threat to the ecosystem but can also be absorbed by the humans through the food chain and cause toxic effects to the human body (Bray et al., 2018; Sun et al., 2021a, 2021b). According to the China National Environmental Quality Standard for soil, approximately 19.4% of the collected agricultural soil samples from land in China was contaminated, whereas the pollution ratio of cadmium (Cd) and lead (Pb) was 7.0% and 1.5%, which ranked first and fifth, respectively (Yang et al., 2021; Zhao et al., 2015). A considerable area of soils is contaminated by both Cd and Pb, which are often found in farmland soil around mining areas (Li et al., 2018; Xie et al., 2019). Rice (Oryza sativa L.) is a staple food in Asian countries and is widely cultivated in China. However, rice contaminated with Cd and Pb has been reported in recent years. Thus, the need for a sustainable method to remediate the Cd and Pb co-contaminated paddy soil is an urgent issue necessary to protect human health.

Biochar is produced by biomass materials at a relatively low temperature (< 900 °C) under anaerobic conditions and is a material rich in carbon (Lehmann et al., 2011; Liu et al., 2021). A considerable number of studies have shown that biochar could improve the physicochemical properties of soil, reduce greenhouse gas emissions, increase fertilizer use efficiency, and promote crop production (An et al., 2021; Azeem et al., 2021; Li et al., 2021; Nan et al., 2021). In addition, numerous studies have claimed that biochar could immobilize Cd and Pb in soils due to its liming effect, large surface area, negatively charged surface, and the presence of oxygen-containing functional groups (Bian et al., 2013; Wang et al., 2015; Zeng et al., 2017). While there was a difference in the adsorption mechanism of biochar between Cd(II) and Pb(II), precipitation with ash mechanisms dominated Pb(II) sorption on biochar, and coordination with π electron mechanisms dominated Cd(II) sorption (Zhang et al., 2017). A previous study revealed that biochar with a high concentration of available phosphorus could immobilize Pb through the formation of stable hydroxypyromorphite (Pb5(PO4)3(OH)) by using X-ray diffraction analysis (Cao et al., 2011). Therefore, with the immobilization of Cd and Pb in the soil, the concentrations of Cd and Pb in rice were effectively reduced by biochar application. Some studies have shown that biochar addition did not significantly affect the bioavailability of potentially toxic metals in the soil or even increase the uptake of metals by plants (Kim et al., 2015; Li et al., 2018). Some researchers have claimed that the limited immobilization effect of biochar on potentially toxic metals is due to the properties of biochar. Therefore, the modification of biochar is becoming increasingly prevalent among recent studies, such as Fe-modified biochar, phosphate-modified active biochar, and sulfur-modified biochar (Rajendran et al., 2019; Sun et al., 2021b). However, few studies have focused on the effect of soil properties on the immobilization of Cd and Pb in different soils.

Previous studies have showed that potentially toxic metals availability in paddy soil and the accumulation of them in rice was significantly different among different soil parent materials (Elbana et al., 2018, Elbana &Selim, 2019, Hamid et al., 2020). In general, the concentrations of many pollutants reveal the following sequence: Mollic Fluvisols ≥ Eutric Gleysol ≥ Eutric Fluvisols (Du Laing et al., 2009, Rinklebe et al., 2007). Shaheen, (2009) reported that labile Cd was negatively correlated only with CaCO3 content in the soils studied, while labile Pb was negatively correlated with cation exchange capacity (CEC), clay, organic matter, total free aluminum oxides, and amorphous iron and aluminum oxide contents. Furthermore, the different components of clay minerals in soils result in a diversity of sorption capacities of soil for Cd(II) and Pb(II), and the adsorbed amount varies from species to species in the following order: montmorillonite > illite > kaolinite (Gupta & Bhattacharyya, 2006; He et al., 2001; Otunola & Ololade, 2020). It was reported that the application of lime significantly reduced the exchangeable Cd and the abundance of microorganisms in acidic soil but it had limited effects in neutral soil (Shangguan et al., 2019). Thus, the immobilization effect of biochar on soil potentially toxic metals may be affected by soil type.

Rice grows in a flooded soil system. The environmental factors (such as redox conditions (Eh), pH, and dissolved organic carbon) vary with the increasing flooding time (Secretan et al., 2018; Shaheen et al., 2017; Yang et al., 2021). Under anaerobic condition, the potentially toxic metals associated with Fe/Mn oxides could be released due to the reduction-induced dissolution of oxides, while the potentially toxic metals can be immobilized by forming precipitates with sulfide (Antoniadis et al., 2017; Beiyuan et al., 2020, Du Laing et al., 2009). Thus, the dynamic changes in the distribution and speciation of Cd and Pb in soil are critical for their immobilization and bioavailability. Moreover, rice is widely cultivated in China, covering an area of 296,940 km2 in 2019, and distributed in more than 25 provinces and regions throughout China (National Bureau of Statistics of China, 2020). The soil types for rice plant growth are diverse and from different regions. Therefore, it is necessary to explore the effect of biochar on the immobilization of Cd and Pb in different soils. The goals of this study are (1) to compare the immobilization effects of biochar on Cd/Pb in alkaline and acidic soils and (2) to investigate the critical impact factor of soil properties on the immobilization of Cd and Pb in soil when biochar is applied.

2 Methods and Materials

2.1 Soil and Biochar

Three typical soil samples, acid soil (soil A), weakly acidic soil (soil B), and alkaline soil (soil C), were sampled from the surface layer of uncontaminated fields for incubation batch experiments and pot experiments. Soil A was collected from Zhangzhou City, Fujian Province, China (24°30′N, 117°35′E), soil B was collected from Fuzhou City, Fujian Province, China (26°05′N, 119°14′E), and soil C was collected from Changzhi city, Shangxi Province, China (36°14′N, 113°12′E). The soil samples were air-dried, ground to pass through a 2-mm sieve. The procedures described by Jackson et al., (1979) were used to determine the soil properties. In brief, cation-exchange capacity (CEC) was determined using 1 M ammonium acetate buffered at pH 7.0. Soil organic matter was determined using the K2CrO7 wet oxidation method. Soil particle-size distribution was determined using the pipette method after hydrogen peroxide treatments.

Properties of the soils were listed in Table 1. Then the soils were spiked with Cd(NO3)2 and Pb(NO3)2 solution to realize a total of 10 mg Cd kg−1 and 2000 mg Pb kg−1 of the three soil samples. The artificially contaminated soils were aged for 1 month under natural conditions prior to experiments. Then, the soil samples were air-dried, ground and passed through a 2-mm sieve. Biochar used in this study was produced by rice straw, and pyrolysis under anaerobic conditions at 500 °C for 2 h. After pyrolysis, the biochar was grounded and pass through a 2-mm sieve. The biochar had a surface area of 18.64 m2 g–1 and a pH (H2O) of 10.3.

Table 1 The properties of the tested soils
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2.2 Batch Experiments

The batch experiments were conducted according to the procedure described by Li et al., (2018). There were two treatments per incubation for each type of contaminated soil: one was the control (CK), and the other was amendment with 5% (w/w) biochar (BC). For the biochar treatment, 10 g of 5% biochar-amended soil was added to a 50-mL tube and mixed with 20 mL of deionized water. The control was the same as the biochar treatment except that the soil was not amended with biochar. The six treatments were performed in triplicate. All the tubes were incubated at room temperature for 1, 3, 5, 10, 20, and 30 days. After incubation, the redox potential (Eh) of the soil for each tube was measured by an oxidation–reduction potential meter (FJA-6; ChuanDi, China). Then, the soil suspension was filtered through a Millipore filter (0.45 μm), and the filtrated solution was stored in plastic bottles for subsequent analysis. The soil solution was separated into two portions. The first was used for pH and electrical conductivity (EC) analysis by an electronic pH meter (PHS-3C, Leici, China) and an electrical conductivity meter (DDS-307A, Leici, China), respectively. The other portion was acidified by adding 0.1 mL of 6 M HNO3 for Cd and Pb analysis by inductively coupled plasma–mass spectrometry (ICP–MS; Agilent 7500cx; USA).

After 30 days of incubation, the soil samples were collected from the tubes, air-dried and ground to pass through a 2-mm sieve. The available extractable Cd and Pb were obtained by mixing 5 g of air-dried soil with 25 mL of CaCl2 (0.1 M) solution followed by 2 h of shaking at 25 °C and immediate filtering. The concentrations of Cd and Pb in the extraction solution were measured by ICP–MS.

2.3 Pot Experiments

The rice plant pot experiment was conducted in plastic boxes (5.5 × 5.5 × 6.5 cm) filled with 100 g of the 3 types of contaminated soils. The biochar treatment consisted of the addition of 5% (w/w) biochar (BC), and the control consisted of the treatment without any biochar (CK). There were three replicates in each treatment and 18 pots in total, which were arranged in a randomized block design. The rice cultivar Oryza sativa L. spp. indica named “Jingliangyouhuazhan” was used in this study. The seeds were sterilized with 10% (v/v) H2O2 solution for 1 min and rinsed with deionized water, and then germinated in moist conditions for 3 days. The germinated rice seedlings were transferred to a beaker and grown for 12 days. Then, the homogeneous seedlings were selected and transplanted into pots (two seedlings per pot). Each pot was submerged in water and kept at a water level of approximately 1 cm above the soil surface during the culture period.

After one month of growth, samples of soil and rice tissues were collected from each pot. The plant samples were washed by the tap water and deionized water and then separated into shoots and roots. The extraction of iron plaques was conducted according to the procedure described by Li et al., (2016). Thereafter, the rice plant samples were weighted fresh. Iron plaques on the root surface were extracted by submerging the roots in dithionite-citrate-bicarbonate (DCB) solution at 25 °C for 60 min, and the extracts were transferred to volumetric flasks. The concentrations of iron (Fe) and manganese (Mn) in the extraction were analyzed by flame atomic absorption spectrometry (AAS, PinAAcle 900; Perkin Elmer, USA).

Rice shoot samples and roots without iron plaque were air dried at 60 °C until a constant weight. Next, the dried samples were crushed by a disintegrator and ground into powder. The samples were completely digested with HNO3 and HClO4 at a rate of 5:1(v/v) on a heating block. Until the volume of digested solution was reduced to 1 mL, the solution was diluted with deionized water and filtered through a Millipore filter (0.45 μm). The concentrations of Cd and Pb in the solution were measured by ICP–MS.

2.4 The Distribution of Clay Minerals

The soil samples used in this study were analyzed to determine the mineralogical composition according to the method described by Omdi et al., (2018). Mineralogical composition was determined by X-ray diffraction using a diffractometer (Rigaku Ulima IV) with monochromatic Cu Kα radiation (scan step size: 0.02°) operated at 40 kV and 40 mA from 5 to 65° 2 theta on powdered soil and on the < 2 μm fraction. For the clay sample three X-ray patterns were recorded: air-dried (N), ethylene–glycol solvated for 24 h (EG), and heated at 500 °C for 4 h (H). The detailed procedures for this method are described in the Supplementary Information.

2.5 Data Analyses

The data shown are the mean ± standard deviation (SD). The statistical analysis was performed using the SPSS 19.0 statistical package (IBM Corp., Armonk, NY, USA, 2010). One-way analysis of variance and Duncan’s multiple range test were used to determine the significant differences between the means of different soil types at a probability level of 0.05. The differences between the means of control and biochar treatment were tested by a paired t test, with the level of 0.05. The figures were made using the software SigmaPlot 12.5.

3 Results

3.1 Properties and Clay Fractions of the Soils

Soil pH is generally considered as an index of the acid or base intensity of a soil and is presented in the Table 1. The pH of the three typical soils ranged in the order soil A (4.23) < soil B (5.54) < soil C (8.16). In addition, the contents of CEC and organic matter in soil C were 1.79 and 7.29 times higher than those in soil A, respectively. The semi-quantitative analysis of the clay minerals in the three studied soils is shown in Fig. 1 and Table S1. The clay fractions of soils A and B are dominated by kaolinite (40–42%), varying amounts of chlorite (23–27%), and illite–smectite mixed layer (22–26%), and traces of illite (9–11%). However, the clay fraction of Soil C is characterized by the dominance of illite–smectite (52%) and illite (30%), and the proportions of chlorite and kaolinite represent only 8 and 10%, respectively.

Fig. 1
figure 1

X-ray diffraction pattern of clay minerals in each sample (a–c) and the clay type and proportion of samples (d). N, EG, and H represented the clay sample in three X-ray patterns: air-dried (N), ethylene–glycol solvated for 24 h (EG), and heated at 500 °C for 4 h (H)

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3.2 Dynamic Changes in the pH, Eh, and EC of the Soil Solutions

The dynamic change in the soil solution pH was investigated, and the results are shown in the Fig. 2A–C. There was a large difference in the soil solution pH between the three soil samples, and the order was shown to be soil A < soil B < soil C. Initially, the pH slightly increased and then fluctuated to reach a stable value during the incubation period. A more significant change in the soil solution pH was observed for soils A and B with biochar application. Compared with that of the control treatment, biochar application significantly increased the pH of the soil solution from 4.55 to 5.51 for soil A and from 7.14 to 7.73 for soil B at day 30 of incubation. There was only a slight increase in soil solution pH at days 10 and 20 of incubation.

Fig. 2
figure 2

The effects of biochar on the pH, Eh and EC of the soil solution in the 3 types of soil. Data are mean ± SE (n = 3), the values with a small asterisk in the same incubation time differed significantly at 0.05 level. CK represents the treatment without any biochar, BC represents the treatment with 5% (w/w) biochar addition

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The influence of biochar on the Eh of the soil is shown in Fig. 2D–F. The Eh was reduced sharply from one to 10 days of incubation and then increased slightly for all treatments. The biochar application significantly increased the Eh of the three soil samples from 10 to 30 days. This implied that biochar inhibited the reduction of soils under flooding conditions.

The EC value increased between days 1 and 10 and then decreased slightly between days 10 and 30 in all treatments (Fig. 2G–I). In the control treatment, the EC of the soil solutions at the first day of incubation was shown in the order soil C > soil B > soil A. Compared to the control, the biochar treatment increased the EC by 498.1–607.1 μs cm−1 of soil A, 513.3–721.3 μs cm−1 of soil B, and 378.7–590.3 μs cm−1 for soil C. This indicated that the addition of biochar transferred some cations or anion to the soil.

3.3 The Dynamic Changes in Cd and Pb Concentrations in the Soil Solution

The concentrations of Cd and Pb in the soil solution were determined and are shown in Fig. 3. In the control treatment, the Cd concentrations of the soil solution in the three soil samples were quite different, and were illustrated in the order soil A > soil B > soil C. The concentration of Cd in the soil solution decreased from 635.6 to 248.9 μg L−1 for soil A and 29.2 to 9.2 μg L−1 for soil B increasing incubation time. With biochar application, the concentration of Cd in soil B solution was decreased significantly by 18.3–69.5% during the whole incubation period compared to the control. When the biochar treatment was applied to soil A, the Cd concentration decreased slightly between days 1 and 5 and then increased to 342.9 μg L−1 at day 10, which was more than the control treatment, while a decreasing trend was shown between days 10 and 30. For soil C, the Cd concentration of the control treatment fluctuated from 1.5 to 3.0 μg L−1, which decreased significantly on the first, third, and twentieth day of incubation for those solutions treated with the biochar treatment.

Fig. 3
figure 3

The effects of biochar on the Cd and Pb concentration of soil solution in the 3 types of soil. Data are mean ± SE (n = 3), the values with a small asterisk in the same incubation time differed significantly at 0.05 level. CK represents the treatment without any biochar, BC represents the treatment with 5% (w/w) biochar addition

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The concentration of Pb in the control soils A and B solutions decreased gradually with the increasing incubation time, and the concentration of Pb in the soil solution on the first day was nearly 2.2 and 1.7 times higher, respectively, than that on the thirtieth day of incubation. The Pb concentration in the soil C solution increased until it reached a maximum level on the tenth day of incubation and then decreased between days 10 and 30. Under the biochar treatment, the Pb concentration of the three types of soil solutions decreased significantly during the incubation, with a decrease of 33.6–92.0% for soil A, 22.9–50.3% for soil B, and 11.5–66.5% for soil C.

3.4 The Effects of Biochar on the Soil Available Cd and Pb Concentrations

The soil pH and the available Cd and Pb concentrations after 30 days of incubation were measured and are shown in Fig. 4. The available Cd concentration in the control soil samples (without biochar amendment) were presented in the order of soil A > soil B > soil C, and this trend was maintained as biochar amended in the three soil samples, which decreased by 15.2, 44.3, and 22.0%, respectively, compared with their respective controls. The available Pb concentration of Soil A was approximately 9.8 and 8.8 times that of soil B and soil C, respectively. Biochar significantly reduced the soil available Pb concentration of soil A, soil B, and soil C by 57.2, 52.3, and 40.3%, respectively.

Fig. 4
figure 4

The soil available Cd and Pb concentration under different treatments. The values in the figures are presented as mean ± SE (n = 3), different letters on top of each bar indicate significant differences (P < 0.05) among the treatments according to Duncan’s multiple range test. CK represents the treatment without any biochar, BC represents the treatment with 5% (w/w) biochar addition

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3.5 The Effects of Biochar on the Iron Plaques of Rice Grown

The formation of iron plaque on the root surface was determined by measuring the concentrations of Fe and Mn in DCB extracts, and the results are shown in Fig. 5. There was a significant difference between iron plaque formations among the three types of soil. In the control treatment, the amount of DCB-Fe and DCB-Mn on rice roots grown in the three soil samples decreased in the following order: soil B > soil A > soil C and soil B > soil C > soil A, respectively. Biochar application significantly reduced the amount of DCB-Fe on rice roots in soils A and B but increased the amount of DCB-Fe on rice roots in soil C. Compared with that of the control, biochar decreased the amount of DCB-Mn on rice roots in soil A but had no significant effect on DCB-Mn in soils B and C.

Fig. 5
figure 5

The Fe and Mn concentration of iron plaque. The values in the figures are presented as mean ± SE (n = 3), different letters on top of each bar indicate significant differences (P < 0.05) among the treatments according to Duncan’s multiple range test. CK represents the treatment without any biochar, BC represents the treatment with 5% (w/w) biochar addition

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3.6 Biomass and Concentrations of Cd and Pb in Rice Tissues

The uptake of potentially toxic metals by plants can more appropriately express their bioavailability in soil. The concentrations of Cd and Pb in rice shoots are shown in Fig. 6. The Cd concentrations in the shoots grown in Soil A and Soil B was much higher than that in the shoots grown in soil C and was reduced by 14.9% and 65.2% as a result of the biochar treatment, respectively. Contrary to expectations, biochar increased the Cd concentration of rice shoots grown in soil C. The order of Pb concentration in rice shoot was soil A > soil B > soil C. Biochar application decreased Pb levels of rice roots grown in the three soil samples, meanwhile, led to 48.9% and 36.6% reduction in rice shoots in soils A and B conditions, respectively. However, the biochar treatment resulted in an increased Pb content in the rice shoots of Soil C compared to that of the control.

Fig. 6
figure 6

The concentration of Cd and Pb in rice shoot. The values in the figures are presented as mean ± SE (n = 3), different letters on top of each bar indicate significant differences (P < 0.05) among the treatments according to Duncan’s multiple range test. CK represents the treatment without any biochar, BC represents the treatment with 5% (w/w) biochar addition

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Moreover, the biomass of roots and shoots from the different treatments was measured in this study and is presented in Supplementary Fig. S1. There were significant differences in root and shoot biomass between soil types. Under the control treatment, the shoot biomass of rice in the three types of soil was ranked as follows: soil C > soil A > soil B. With biochar application, the shoot biomass of plants for soil A and soil B was significantly increased by 105.5 and 64.4%, respectively, compared to that of the control. However, biochar reduced the shoot biomass of rice grown in soil C by 49.6%.

4 Discussion

4.1 Soil Bioavailability of Cd and Pb

The bioavailable concentrations of Cd and Pb in the three soil samples were quite different. The CaCl2-extracted Cd concentration in soils A and B was higher than that in soil C, and the CaCl2-extracted Pb concentration in soil A was more than eight times higher than that in soils B and C. This was due to several mechanisms.

First, the solubility of the trace elements that can occur as free hydrated cations generally increases with decreasing pH. With a decrease in the pH of the soil solution, there is an increase in the activity of H+, Al3+, and their positively charged hydroxides in the solution. These cations will compete with the potentially toxic metals for negative sorption sites (Hooda, 2010). In this study, the pH of the three typical soils was ranged in the order soil A (4.23) < soil B (5.54) < soil C (8.16). The CaCl2-extracted Cd and Pb concentrations of the soils under the control treatment were significantly negatively related to soil pH (Table S2), and the result was similar to those of previous studies.

Second, the soil CEC is an important factor in trace element retention or release. It was reported that soil with a high CEC could improve the sorption of Cd(II) from soil solution via chemisorption or ligand exchange on clay surfaces (Hooda, 2010). The type of clay mineral is of great importance to the soil CEC. It was reported that montmorillonite is characterized by a very high CEC between 80 and 100 cmol+ kg−1, illite has an average CEC between 15 and 40 cmol+ kg−1, and kaolinite has a relatively low CEC between 3 and 15 cmol+ kg−1 (Weil & Brady, 2017). The adsorption capacities of the three clay minerals for Cd are found to be in the order of montmorillonite, illite, and kaolinite (He et al., 2001). In the present study, the clay fraction of Soil C is dominated by illite–smectite and illite, while the clay fraction of soils A and B is dominated by kaolinite (Fig. 1). Accordingly, the amount of CEC of soil C is much greater than that of soils A and B (Table 1). In addition, it was found that the CaCl2-extracted Cd concentration of the soils under the control treatment was significantly negatively related to CEC and the proportion of illite and illite–smectite mixed layers of clay minerals (Table S2). Thus, the CaCl2-extracted Cd of Soil C was lower than that of soils A and B (Fig. 4).

Moreover, the soil organic material is another key factor affecting the bioavailability of Cd and Pb in the soils. Because of the relatively insoluble complexes of humic acid with trace metals, especially in an acid medium, these complexes can be considered to be organic storage for trace metals in soil (Kabata-Pendias, 2011). The affinity of humic substances to accumulate trace cations differ. A stronger fixation of Pb than of Cd by organic materials was shown in soils rich in humus. It was reported that the amounts of Cd and Pb sorbed by humic acid (HA) at pH 5–6 were 9 and 120 (in mg kg−1 of HA), respectively (Kabata-Pendias, 2011). In this study, the amount of organic materials in Soil A was much lower than that in soils B and C; thus, the CaCl2-extracted Pb in soil A was much higher than that in soils B and C.

Previous studies have shown that biochar application decreases the bioavailable Cd and Pb in soil; these processes include complexation, cation exchange, precipitation, electrostatic interactions, and chemical reduction, which result in the transformation of soluble metal species into stable metal species (such as Pb-hydroxide and Pb-phosphate) (Ahmad et al., 2014; Azhar et al., 2019; Wagner et al., 2015). In this study, biochar application increased the soil pH (Fig. S2), and the CaCl2-extracted Cd and Pb concentrations in all three soils decreased with biochar application (Fig. 4). Moreover, the decreasing amplitudes of the CaCl2-extracted Cd and Pb concentrations by biochar were different between soil types. The decreasing amplitude of the available Cd concentration is significantly negatively related to soil pH, CEC, and the proportion of illite and illite–smectite mixed layers of clay minerals (Table S3). The decreasing amplitude of the available Pb concentration was significantly negatively related to soil pH and soil organic materials. This result demonstrated that the stronger the stability of potentially toxic metals in soils, the more difficult it is to be changed by biochar, and the smaller the decrease of biochar on the concentration of CaCl2-extracted Cd and Pb in soil.

4.2 The Accumulation of Cd and Pb in Rice

Biochar application reduced the Cd and Pb concentrations in rice shoots grown in soils A and B, but biochar had no significant effect on the Pb content of rice shoots planted in soil C and even increased the Cd content of rice stems (Fig. 6).

The uptake of potentially toxic metals from soils is correlated with their soil bioavailability. It was reported that biochar induced the transformation of the acid-soluble fraction of Cd and Pb to reducible and residual fractions and inhibited their uptake and accumulation in rice plants (Li et al., 2016, Sun et al., 2021b; Xu et al., 2018; Yang et al., 2016). In this study, biochar application reduced the concentrations of CaCl2-extractable Cd and Pb in soils A and B, and then decreased the accumulation of the two elements in rice shoots (Fig. 6).

The accumulation of trace elements in rice is related not only to their availability in soil but also to their transportation from soil to rice. In this study, the concentration of Cd in rice shoots was significantly positive related to the amount of iron plaque on rice roots (P < 0.05) (Table S4). Iron plaque is formed where the oxygen released from roots reacts with reduced soluble Fe2+ to form a smooth regular reddish precipitate or irregular plaque coating on root surfaces (Cheng et al., 2014). The formation of iron plaque is influenced not only by the morphological and physiological properties of the plant but also by various soil-derived factors, such as Fe and Mn concentrations (Frommer et al., 2011). The redox conditions of soil affect the amounts of Fe2+ in the soil solution. Under reduction conditions, Fe2+ is produced by the reductive dissolution of Fe minerals in the soil matrix (Chen et al., 2008; Yamaguchi et al., 2014). With increasing of flooding time, the oxygen in the soil was consumed by microorganisms, and the soil Eh decreased. Interestingly, the Eh value of the control soil decreased more than that of the soil treated with biochar (Fig. 2). As a result, the amount of DCB-Fe in the control treatment of Soil A and Soil B was higher than that in the biochar treatment (Fig. 5). Generally, iron plaque has a high adsorption capacity for potentially toxic metals due to its iron hydroxide functional groups (Huang et al., 2021), and it has also been reported that iron plaque fails to affect potentially toxic metals uptake by plants (Liu, 2008). In this study, iron plaque acted as a storage of Cd and Pb for rice; hence, the accumulation of Cd and Pb in rice shoot declined in accordance with the decreased of iron plaque formation.

However, in contrast to soils A and B, the amount of iron plaque on the rice root surface of soil C increased with biochar addition (Fig. 5). This might be because the pH of soil C (8.16) is much higher than that of soil A (4.23) and soil B (5.54). The solubility of Fe was strongly decreased as the pH increased (Batty et al., 2000; Hooda, 2010). It was reported that iron plaque formation on rice roots was higher in soil with low pH (4.6) than in a soil of neutral pH (6.6) (Siddique et al., 2022). The pH of the soil C solution treated with biochar was lower than that of the control treatment after 30 days of incubation (Fig. 2). Therefore, with the increase in iron plaque on roots, more metals could be stored. In addition, the Cd concentration in the soil C solution of the biochar treatment was slightly higher than that of the soil C solution of the control treatment. Thus, biochar application increased the Cd concentration in rice shoots grown in soil C.

5 Conclusion

The effect of biochar on the mobility of Cd and Pb in soil and their bioaccumulation in rice shoots was significantly affected by soil properties. Biochar application reduced the CaCl2-extracted Cd and Pb in the three soils, and the decreasing amplitude of the available Cd concentration was significantly negatively (P < 0.05) related to soil pH, CEC, and the proportion of illite and illite–smectite mixed layers of clay minerals. The decreasing amplitude of the available Pb concentration with biochar addition was significantly negatively related to soil pH and soil organic materials. The Cd concentration of rice shoots is not only influenced by the soil availability of Cd but also positively related to the amount of iron plaque. Biochar application decreased the formation of iron plaque on rice roots grown in soils A and B due to the biochar slowing down the redox potential. However, biochar increased the amount of iron plaque on rice roots grown in soil C, which might be attributed to the higher alkalinity of soil C. As a result, biochar reduced the accumulation of Cd and Pb in rice shoots and promoted the biomass of rice tissues grown in soils A and soil B (acid soils), but it had the opposite effects on soil C (an alkaline soil).