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Article

Stabilization of Soil Co-Contaminated with Mercury and Arsenic by Different Types of Biochar

by
Yang Wei
1,2,3,4,*,
Risheng Li
1,2,3,4,
Nan Lu
2,3,4 and
Baoqiang Zhang
1,2,3,4
1
Shaanxi Land Engineering Construction Group Co., Ltd., Xi’an Jiaotong University Technology Innovation Center for Land Engineering and Human Settlements, Xi’an 710075, China
2
Land Engineering Technology Transformation Center, Shaanxi Provincial Land Engineering Construction Group Co., Ltd., Xi’an 710075, China
3
Institute of Land Engineering and Technology, Shaanxi Provincial Land Engineering Construction Group Co., Ltd., Xi’an 710075, China
4
Key Laboratory of Degraded and Unused Land Consolidation Engineering, Ministry of Natural Resources, Xi’an 710075, China
*
Author to whom correspondence should be addressed.
Sustainability 2022, 14(20), 13637; https://doi.org/10.3390/su142013637
Submission received: 24 August 2022 / Revised: 17 October 2022 / Accepted: 19 October 2022 / Published: 21 October 2022
Browse Figures
Versions Notes

Abstract

:
Mercury (Hg) and arsenic (As) are toxic and harmful heavy metals, with exceedance rates of 1.6% and 2.7%, respectively, in soils across China. Compared to soils contaminated with Hg or As alone, co-contaminated soils pose complex environmental risks and are difficult to remediate. Biochar is widely used as a soil amendment to adsorb and immobilize pollutants such as heavy metals. However, only a few studies have explored the efficiency of biochars produced from different crop straws to reduce the bioavailability of heavy metals in co-contaminated soils, and the effects on soil biological properties are often overlooked. The aim of this study was to investigate changes to the physicochemical properties, enzyme activities, and heavy metal bioavailability of an industrial soil co-contaminated with Hg and As upon the addition of different biochars from reed, cassava, and rice straws (REB, CAB, and RIB, respectively). The soil was amended with 1% biochar and planted with spinach in pots for 30 days. RIB was more effective than REB and CAB in increasing the soil pH, organic matter content, and cation exchange capacity. RIB and CAB exhibited similar positive effects on the soil dehydrogenase, catalase, invertase, and urease activities, which were higher than those of REB. The exchangeable fraction of both metals decreased upon biochar addition, and the residual fraction showed the opposite trend. All biochar amendments reduced the bioconcentration factors of heavy metals (especially Hg) in plants and decreased the metal bioavailability in soil. RIB is the optimal amendment for the stabilization of soil co-contaminated with Hg and As.

1. Introduction

In the wake of rapid economic development, heavy metal contamination caused by irrational exploitation, inappropriate environmental management, and improper legal regulations has become a global environmental issue. For example, 1108–3784 t of mercury (Hg) and 28,400–94,000 t of arsenic (As) are released into global soil every year owing to anthropogenic activity. Many heavy metal(loid)s, including Hg and As, are characterized by high toxicity, low biodegradability, and long environmental persistence [1]. They can accumulate in crop plants and then enter the human body through the food chain. All forms of Hg are toxic, with certain carcinogenicity, which can poison the nervous, reproductive immune, cardiovascular, and cerebrovascular systems [2]. In addition, long-term As exposure can lead to cancer of the skin or visceral organs [3], and oral administration of 0.1 g As2O3 can be fatal [4]. Compared to soils contaminated with Hg or As alone, co-contaminated soils pose complex environmental risks [3]. Therefore, the remediation of soil co-contaminated with Hg and As is of practical significance from a human health perspective.
According to the national soil pollution survey in China, the exceedance rates of Cd, Ni, As, Cu, Hg, Pb, Cr, and Zn are 7.0%, 4.8%, 2.7%, 2.1%, 1.6%, 1.5%, 1.1%, and 0.9%, respectively [5,6]. Although Cd is the most important heavy metal contaminant, soil pollution by As and Hg is also a non-negligible problem, especially in industrial wasteland. Wastewater, waste residue, and waste gas discharged from industrial production are the major sources of soil Hg and As [7]. In addition, the extensive use of Hg- and As-containing pesticides in agricultural production causes soil pollution [8]. Similar to Cd, Hg often exists as cations (e.g., Hg2+) in the soil, whereas As mostly occurs as oxyanions (e.g., H2AsO4− and HAsO32−) [6]. Given the different behaviors of cations and oxyanions in the complex environmental matrix, it is difficult to remediate soils co-contaminated with Hg (or Cd) and As.
Numerous studies have been devoted to the removal or immobilization of heavy metals in soil using organic and inorganic additives such as biochar (e.g., [9]). As a major agricultural country, China contributes 15.0% of the global straw output [10]. However, the straw utilization rate is only 80%, and incineration or disposal can easily lead to resource waste and environmental pollution [11]. The transformation of straw into biochar through pyrolysis is a sustainable technology for waste utilization and carbon (C) sequestration [12]. Compared with traditional physical and chemical techniques for soil remediation, biochar addition is pollution-free and cost-effective.
Biochar can be used as a soil amendment to adsorb and immobilize heavy metals via different mechanisms [13]. First, biochar is characterized by high electronegativity and cation exchange capacity (CEC), with abundant functional groups (–OH, –COOH, –C=O–, and C=N) on the surface. It can interact directly with heavy metals via electrostatic attraction, ion exchange, complexation, and precipitation [14]. Second, biochar is rich in C and has a well-developed pore structure. It can indirectly affect heavy metal speciation by altering soil physicochemical properties, which, in turn, reduces the bioavailability of heavy metals [15].
The adsorption performance of biochars prepared from different feedstocks varies, due to differences in their elemental composition, functional groups, specific surface area, surface properties, and microscopic morphology [16]. Some researchers investigated the efficiency of biochars produced from cotton, corn, and rape straws for Hg adsorption and immobilization [17,18,19], whereas others applied biochars derived from wheat, peanut, and sugarcane straws to remove As from soil and water [20,21,22]. However, no previous studies have explored the use of biochars obtained from different crop straws to treat soil co-contaminated with Hg and As. The adsorption capacity and mechanisms of biochars on Hg–As in co-contaminated soil may be different from those on other heavy metals, such as Cd–Pb or Cd–Zn.
Biochar amendment may improve the soil biological properties [23] in addition to reducing the bioavailability of toxic heavy metals. In particular, the activity of soil enzymes has attracted attention for the remediation of heavy metal-contaminated soils [24]. Soil enzymes secreted by microorganisms play a critical role in nutrient cycling, redox reactions, and organic matter (OM) decomposition. Therefore, the soil enzyme activity reflects the intensity of biochemical reactions in soil and can be used to evaluate the quality of soils contaminated with heavy metals [24]. Biochar amendment can directly and indirectly affect the soil enzyme activity [25]. However, the relationships between specific physicochemical properties, enzyme activities, and metal bioavailability in biochar-amended soil contaminated with heavy metals have not been systematically studied.
Reed is a wetland plant with a wide range of sources and high levels of cellulose and hemicellulose. Reed straw-derived biochar is characterized by a loose internal structure, large specific surface area, and abundant active groups such as hydroxyl and carboxyl groups [26]. Thus, it may have a high ability to adsorb and stabilize heavy metals in soil. In addition, cassava and rice are common food crops, and their straw accounts for a large proportion of tropical agricultural waste [9]. Therefore, using reed, cassava, and rice straws to prepare biochar for the stabilization of heavy metal-contaminated soil can realize the comprehensive utilization of agricultural waste.
In the present study, we selected reed, cassava, and rice straw as feedstocks to prepare different types of biochars. A pot experiment was carried out to investigate biochar-induced changes in the physicochemical and biological properties as well as the heavy metal bioavailability of soil co-contaminated with Hg and As. We also discussed the feasibility of using different biochars to stabilize heavy metal-co-contaminated soil. The present study can provide new insights into the efficiency of biochar for the green remediation of heavy metal-contaminated soil and may represent a theoretical reference for the comprehensive utilization of agricultural waste.

2. Materials and Methods

2.1. Materials

Soil co-contaminated with Hg and As was collected within a chemical plant that produced fumaric acid and its affiliated chemical products. There were coal yards, oil tank plants, and garbage dumps around the chemical plant, leading to substantial wastewater discharge and soil pollution. At the sampling site, weeds were removed from the surface, and soil was collected at the depth of 0–50-cm using a shovel. The soil sample was transported to the laboratory and air-dried. After removing debris, the sample was ground and passed through a 2 mm sieve. The soil was classified as silty loam, and its basic physicochemical properties are summarized in Table 1.
Reed and rice straw samples were collected in Fuping County, Weinan City (109°11.78′ E, 34°42.12′ N) in Shaanxi Province, and the cassava straw sample was collected in Wuming County, Nanning City (107°58.26′ E, 23°19.27′ N) in Guangxi Province, China. Different types of biochars were produced by pyrolysis [27]. Briefly, the straw samples were washed with ultrapure water to remove surface dust and dirt, followed by oven-drying at 60 °C for >2 days to remove moisture. The dry straw samples were milled to <1 mm pieces. Each straw sample was loaded into a covered crucible and placed in a muffle furnace, then heated to 550 °C at a heating rate of 5 °C min−1 for 4 h. Afterward, the crucible was placed into a glass desiccator and cooled to room temperature. The biochar sample was then collected and ground to pass through a 0.149 mm sieve. The biochars obtained from reed, cassava, and rice straw samples are referred to as REB, CAB, and RIB, respectively.

2.2. Experimental Design

A pot experiment (Table 2) was conducted in a greenhouse from 13 May 2021 to 11 June 2021. Before the experiment, equal amounts of soil samples were mixed with 1% REB, CAB, or RIB and filled into pots (20 cm diameter and 20 cm depth, 3.0 kg soil per pot). Basal fertilizer (N; 75 mg pot−1) was applied before sowing. Spinach seeds (Chunqiu Seed Industry Co., Ltd., Shouguang, China) were sown into the pots (40 seeds per pot) and covered with ~1 cm of soil. A blank control group without biochar amendment was included, and three replicates were used for each treatment group. After sowing, the pots were arranged in a randomized block design, with their positions changed regularly to maintain ventilation and light conditions. Water was added every other day to control the soil moisture at ~25%. Spinach plants were harvested 30 days after planting in order to determine their total Hg and As contents. Then, potted soil samples were collected to determine their physicochemical properties, enzyme activities, and heavy metal speciation.

2.3. Physicochemical Analysis

Air-dried soil subsamples were used for physicochemical analyses. In brief, mechanical composition tests were conducted using an MS2000 laser particle size analyzer (Malvern Instruments Ltd., Malvern, UK). pH and electrical conductivity (EC) measurements were made in 1:2.5 soil/deionized water slurries (w/w) using an S220-K acidometer and an S230-K conductivity meter (Mettler Toledo Instruments Ltd., Zurich, Switzerland), respectively. Total organic carbon (TOC) contents were determined by potassium dichromate oxidation with external heating and then converted to OM content using a conversion factor of 1.724 [24]. Calcium carbonate (CaCO3) content was determined by a titrimetric method [28]. CEC values were determined using the ammonium acetate method [29].
Biochar yields were calculated as the mass ratio of obtained biochar to crop straw used for pyrolysis. pH and EC measurements were conducted in 1:10 biochar/deionized water (w/w) slurries that had been agitated for 90 min at 20 °C. Sulfur (S) content was analyzed using a Flash EA 1112 elemental analyzer (Thermo Fisher Scientific, Cleveland, OH, USA). Silica (SiO2) content was determined by alkali fusion [30]. Scanning electron microscopy (SEM) images and pore sizes were obtained using an FEI Q45 scanning electron microscope (FEI Corp., Hillsboro, OR, USA). Specific surface areas were measured using a NOVA4200E analyzer (Quantachrome Instruments, Boynton Beach, FL, USA). Zeta potentials were measured using a Zeta potentiometer (DelsaMax Pro; Beckman Coulter, Brea, CA, USA).

2.4. Heavy Metal Analysis

An atomic fluorescence spectrometer (AFS-9760, Haiguang Instruments Corp., Beijing, China) was used to determine the total Hg and As contents in soil and plant samples according to the GB/T 22105.1-2008–GB/T 22105.2-2008 and GB 5009.17-2014–GB 5009.11-2014 Chinese standards, respectively.
The bioconcentration factors (BCFs) of Hg and As from soil to plants were calculated using Equation (1) [31]:
BCF = C p t C s o
where Cpt (mg kg−1) is the Hg or As content in harvested plants, and Cso (mg kg−1) is the initial Hg or As content in soil samples.
The Tessier sequential extraction procedure was used to separate soil Hg and As into exchangeable (F1), carbonate-bound (F2), iron (Fe)/magnesium (Mn) oxide-bound (F3), organic-bound (F4), and residual (F5) fractions [32]. A higher proportion of F1 corresponds to more active and bioavailable heavy metals, whereas a higher F5 fraction denotes a lower bioavailability of heavy metals in the soil [33].

2.5. Enzyme Activity Assay

Fresh soil subsamples were used to measure enzyme activities. In brief, dehydrogenase activities were measured by the triphenyl tetrazolium chloride assay and expressed in μg day−1 of triphenyl formamidine produced per gram of soil [34]. Catalase activities were measured by potassium permanganate titration and expressed in μmol min−1 of potassium permanganate (0.1 mol L−1) consumed per gram of soil [35]. Invertase activities were assayed by 3,5-dinitrosalicylic acid colorimetry and expressed as in mg day−1 of glucose produced per gram of soil [36]. Urease activities were assayed by sodium phenate–sodium hypochlorite colorimetry and expressed in mg day−1 of NH3-N produced per gram of soil [36].

2.6. Statistical Analysis

Statistical analyses were performed using the SPSS 18.0 software (SPSS Inc., Chicago, IL, USA). One-way analysis of variance (ANOVA) was used to determine differences in sample means across different treatment groups.

3. Results and Discussion

3.1. Comparison of Biochars from Different Feedstocks

Table 3 summarizes the main properties of the three biochars, which showed significant variations with the feedstock type. Overall, the biochar samples exhibited considerable differences in terms of yield, pH, EC, pore size, specific surface area, S content, and SiO2 content. The yield of RIB was remarkably higher than those of CAB and REB. All three biochars were alkaline, with a mean pH between 10.11 and 10.81. Similar pH values (10.76 and 10.13, respectively) were reported for biochars from canola and wheat straws (at 550 °C) [37,38]. Biochars with alkaline pH could facilitate the treatment of heavy metal-contaminated soils [39]. The EC measures the content of dissolved salts and is a key indicator of biochar quality for use as an amendment of heavy metal-contaminated soil [40]. In the present study, the EC of CAB was similar to that of RIB, whereas REB exhibited a considerably lower value. Previous studies have shown that sulfur and silicon can change the speciation of heavy metals and reduce their bioavailability to plants, thereby realizing the passivation and immobilization of heavy metals [41,42]. Biochar composition analysis revealed that RIB had the highest S and SiO2 contents, while lower S and SiO2 contents were observed in REB and CAB, respectively.
The SEM images showed that REB, CAB, and RIB had porous structures (Figure 1). In particular, REB and CAB contained more macropores, while RIB had smaller pore sizes with a more regular and compact pore arrangement. The mean pore size of the different biochars followed the order RIB < CAB < REB, while the specific surface area decreased in the order RIB > CAB > REB. The heavy metal adsorption capacity of biochar is strongly affected by its specific surface area and porosity; in particular, higher specific surface area and porosity values correspond to a stronger adsorption capacity for heavy metals [43]. Therefore, the present results indicate that, compared with REB and CAB, RIB had a greater ability to adsorb and immobilize heavy metals in soil.

3.2. Effects of Biochar Amendments on Soil Physicochemical Properties

All of the soil physicochemical properties analyzed in this study were significantly improved upon the addition of different biochars (Figure 2). Compared with the unamended control, the highest soil pH was observed for the RIB treatment, while the REB and CAB treatments led to smaller and similar increases in soil pH (Figure 2a). Other studies also reported soil pH increases following the application of wheat straw and rice husk biochars [44,45], although a decrease in soil pH was observed in some cases [46]. The soil EC showed the largest increase upon CAB and RIB addition (Figure 2b), consistent with a previous study by Igalavithana et al. using pinecone biochar [47]. Compared with the unamended control, each biochar amendment significantly increased the soil OM, TOC, and CEC values, with the RIB treatment showing the strongest effect (Figure 2c–e). Many studies also found that biochars produced from rice straw, wheat straw, and empty palm fruit bunches had a positive effect on soil OM [24], TOC [44,48], and CEC [49], respectively.
The improvements in soil physicochemical properties following biochar application may be a direct effect of the biochar components or result from a combination of different physicochemical factors. For example, the negatively charged functional groups, including phenolic, hydroxyl, and carboxyl groups on the surface of biochar, can bind to the H+ ions in the soil, causing an increase in its pH [50]. Biochar can also increase the soil NO3−–N content, which in turn leads to a pH increase [24]. NO3−–N may react with H+ to suppress soil acidification. The marked EC increases observed for the three biochar treatments might be due to the higher EC of the biochars used in the experiment compared with that of the original soil. In addition, Tang et al. reported that biochar treatment can increase the available phosphorus (P) and potassium (K) contents in soil, leading to an increase in EC [24].
Biochar is a C-rich material with an organic C content of up to 90%, depending on the feedstock [51]. Biochar can increase the soil TOC content by adsorbing small organic molecules and then promoting their polymerization through surface catalysis [52]. Biochar can also efficiently increase the soil CEC, and a positive relationship between soil CEC and organic C content has been identified [53]. Liang et al. reported that organic C increases the soil CEC through a higher surface area with more cation adsorption sites, a higher charge density per unit surface area, or a combination of the two factors [54].

3.3. Changes in Soil Enzyme Activities upon Biochar Addition

The soil enzyme activities serve as biological indicators for evaluating the quality of soil contaminated by heavy metals [20]. Dehydrogenase and catalase are oxidoreductase enzymes that directly alter ionic valence states and contribute to the detoxification of heavy metals [55]. Invertase and urease are involved in the transformation of soil nutrients such as C and N [55]. Compared with the unamended control, each biochar amendment enhanced the activities of dehydrogenase, catalase, invertase, and urease enzymes in soil samples (Figure 3). Overall, RIB exhibited the greatest positive effect on soil enzyme activities (with increases of 29.2–89.3%), while REB displayed the least positive effect (13.9–27.7%).
Consistent with our results, many studies have concluded that biochar application can enhance the activity of soil enzymes such as dehydrogenase [44,56], urease, and invertase [57]. The enhancement of soil enzyme activities by biochar amendment may be attributed to the stimulation of soil microorganisms that secrete enzymes. Gomez et al. showed that biochar addition increased microbial abundance and altered the community composition of a sandy loam soil [58]. When biochar was added to contaminated soil, the resistance of microbial groups to cadmium (Cd) and lead (Pb) tended to increase [27]. Biochar can promote the growth of microorganisms in soil by improving its quality, thus enhancing the soil enzyme activity [59]. Another factor that could influence the soil enzyme activity is that heavy metal stress levels in the soil may change following biochar addition [60].
Interestingly, in some cases, biochar may inhibit the activity of soil enzymes such as dehydrogenase, catalase [24], invertase, and urease [25]. The enzymatic activities mainly depend on soil properties including OM content, pH value, and mineral nutrients [61]. Different biochar feedstocks, production methods, and amendment doses may result in distinct changes in soil enzyme activity [25,62].

3.4. Bioavailability of Heavy Metals in Biochar-Amended Soil

Soil heavy metals exist in different forms, with varied migration ability and biological toxicity [63]. Following the sequential extraction procedure, F5 (residual fraction) is generally considered to be an unavailable form, because it is fixed in the soil matrix and cannot be absorbed or utilized by plants. Conversely, F1 (exchangeable fraction) is considered a bioavailable form, whereas the other three fractions (F2–F4) are relatively difficult for plant uptake or utilization [64]. In the present study, Hg and As showed different distributions among the five fractions in soil samples with and without biochar amendments (Figure 4 and Figure 5). In the unamended control, F4-Hg was the predominant Hg fraction (63.0%), followed by F1-Hg (19.0%), whereas F5-Hg, F2-Hg, and F3-Hg only accounted for small proportions of the total Hg (4.4–7.0%; Figure 4a). In contrast, soil As mainly comprised F5-As (40.3%) and F1-As (33.2%), with small proportions of F4-As, F2-As, and F3-As (4.1–13.2%; Figure 4b). The different distributions of Hg and As in the five fractions suggest a different bioavailability of these heavy metals in the original soil [65].
In all treatments, F4-Hg was the highest among the five fractions, despite an increasing trend in the contribution of F5-Hg in soil with biochar amendments (Figure 4a). A similar trend was observed for F5-As, which was the highest fraction in biochar-amended soil. For both heavy metals, the F1 fraction decreased significantly in all three biochar-treated soils compared with that in the unamended control. This indicated that REB, CAB, and RIB could effectively reduce the bioavailability of heavy metals in co-contaminated soil by facilitating the conversion of Hg and As from exchangeable to residual form. Among the biochar-amended soils, the largest decrease in F1-Hg and F1-As fractions was observed for the RIB treatment, along with the largest increase in F5-Hg and F5-As fractions. The increase in F5 suggests heavy metal passivation. F2 and F3 fractions of Hg and As showed no marked differences among all treatments. Compared with the unamended control, the F4 fraction moderately increased in all biochar-amended soils, but there was no significant difference among all the treatments. In summary, the results indicate that RIB exhibited the greatest effect on the stabilization and remediation of soil co-contaminated with Hg and As.
After biochar addition, some of the exchangeable Hg and As fractions in soil were converted into organic-bound and residual fractions with lower bioavailability. The effects of different biochars on the bioavailability of soil Hg and As observed in the present study are supported by many previous reports. For example, Gamboa et al. found that, in a contaminated soil treated with biochar, the immobile Hg fraction increased by 76% compared with the control [66]. In addition, Gu et al. showed that the bioavailable fraction of As in soil exhibited a marked decrease of up to 18.0% after biochar treatment [67]. Yu et al. also indicated that the concentrations of bioavailable As species decreased upon biochar addition to contaminated soils [68].
The reduction of Hg and As bioavailability may be partly attributed to the direct effect of biochar on soil heavy metals. The three biochar materials had negative charges on their surface, with large specific surface area, many functional groups, and a high pH. All these factors can facilitate the immobilization of Hg cations through electrostatic interactions and chelation with functional groups on the biochar surface [57]. Gamboa et al. also indicated that highly porous biochars with abundant polar functional groups on the surface, along with high pH, EC, CEC, and ash percentage, can favor the adsorption and stabilization of Hg [66]. On the other hand, the adsorption of As oxyanions by biochar may be mainly related to its huge specific surface area and well-developed pore structure. After examining the microscopic morphology of the present systems, an obvious microporous structure could be observed in all three biochar materials after carbonization (Figure 1). Overall, REB exhibited few isolated pores with relatively large size and the smallest specific surface area. However, RIB had a hierarchical pore structure with the smallest mean pore size and largest specific surface area, which may explain its strong ability to adsorb and immobilize both Hg and As in the co-contaminated soil.
Considering the elemental composition, all three biochars contained S and SiO2 (Table 3). Hg in soil can adsorb to the surface of biochar and react with S groups (e.g., sulfate ester and sulfate radicals) loaded on biochar to form stable residual Hg-sulfide [42]. Compared with REB and CAB, RIB had a higher S content, which resulted in stronger immobilization of Hg. In addition, the water-soluble Si in Si-rich straw biochar plays a role in the adsorption and immobilization of heavy metals. Si can promote the conversion of exchangeable heavy metals with higher activity to the residue state [69]. Compared with reed and cassava, rice prefers Si, and its straw is rich in phytolith (SiO2·H2O). Consequently, RIB had the highest SiO2 content among the three biochars (Table 3), which could enhance the passivation of heavy metals.
Furthermore, biochar-induced changes in soil properties (e.g., CEC and enzymatic activity) may indirectly reduce the bioavailability of soil heavy metals. Mohamed et al. indicated that the soil CEC is positively correlated with the residual fraction of heavy metals [70]. Liu et al. found a negative correlation between soil enzyme activity and the bioavailable content of heavy metals [71]. In this study, biochar amendments resulted in increased CEC and enhanced dehydrogenase, catalase, invertase, and urease activities in the soil. In turn, these changes effectively increased the F5-Hg and F5-As fractions, thereby reducing the bioavailability of heavy metals. In the present study, RIB resulted in the highest soil CEC and enzymatic activity, posing a strong effect on the immobilization of heavy metals.
Soil OM acts as an adsorbent for heavy metals, because it contains different functional groups (–OH and –COOH) that can easily bind metal ions to form strong complexes with low bioavailability [55,72]. In biochar-amended soil samples, the increase of OM content could benefit plant growth and development, enhance plant stress resistance, and reduce plant uptake of available heavy metals. Since RIB contributed the most to soil OM content, this amendment could promote the transformation of heavy metals into the unusable residual form most prominently.
Generally, soil pH is considered one of the most important factors controlling heavy metal bioavailability [57,73], because it controls the adsorption–desorption, dissolution–precipitation, and other processes of soil minerals. Biochar is an alkaline substance, and its application increases the soil pH, thereby affecting the adsorption and desorption of soil minerals. In turn, the increased pH promotes the transformation of active heavy metal species into stable residuals and thus stabilizes heavy metals such as Hg. Among the three biochars produced in the present study, RIB displayed the strongest effect on increasing the soil pH, which contributed to the effective reduction of Hg bioavailability. Unlike Hg, As is relatively stable in acid soil, and lowering the pH in an alkaline soil environment is conducive to As stabilization, while the increase of pH can activate soil As to a certain extent [6]. Soil As exists mainly in the form of oxyanions, such as H2AsO4− and HAsO32−. When pH increases, the OH− concentration in the soil increases correspondingly. As OH− competes with H2AsO4− and HAsO32− for the adsorption sites on biochar surface, the adsorption and stabilization of As in the soil may be decreased. However, in this study, the strong adsorption capability of biochars covered their mobilizing effect on As bioavailability, and significantly decreased the exchangeable As concentration (Figure 5). Meanwhile, both effects of the increase in soil pH and adsorption of biochars significantly enhanced the stabilization of Hg2+ and reduced the bioavailability of Hg in soil. Moreover, the test soil is derived from loess parent material with high CaCO3 content. The application of biochar could increase the contact area between Ca and As in the soil, promoting their reaction and co-precipitation, and consequently reducing the bioavailability of As.

3.5. Bioconcentration of Heavy Metals in Plants after Soil Stabilization with Biochar

The BCF indicates the ability of plants to accumulate heavy metals [74]. This factor can directly measure the difficulty in the migration and utilization of soil heavy metals, and indirectly reflect the efficiency of biochar for the stabilization of heavy metal-contaminated soils [75]. Irrespective of biochar amendments, the BCF of Hg (0.337) was markedly higher than that of As in spinach (0.095; Figure 6), indicating a greater risk of Hg migration in the soil and subsequent accumulation in plants.
For both Hg and As, the BCFs tended to decrease in the biochar-amended soil samples compared with the unamended control. These results suggest that the addition of biochars could enhance Hg and As adsorption by soil and effectively reduce heavy metal uptake by plant tissues. Yu et al. also reported that biochar addition could inhibit the accumulation of Hg in cabbage plants by preventing its migration [76]. Similar results were observed by Zama et al. showed that both silicon-modified and unmodified biochars were effective in minimizing the BCFs of As in spinach [77]. Among the different treatments, the BCFs of Hg and As in plants were the lowest in the RIB treatment and the highest in the REB treatment. These results indicate that the plant availability of Hg and As in soil decreased after amendment, and RIB had the best remediation effect on Hg and As, which was consistent with the results of the heavy metal speciation analysis.

4. Conclusions

The addition of different biochars from crop straw waste changed the physicochemical and biological properties of an industrial soil co-contaminated with Hg and As. All biochar amendments considerably improved the soil properties tested and led to enhanced soil enzyme activities. Moreover, the biochar amendments decreased the bioavailability of Hg and As, enhanced their adsorption on soil, and effectively reduced heavy metal uptake by plant tissues. The reduction of Hg and As bioavailability was attributed to the direct effect of biochar on soil heavy metals and the indirect effect of biochar-induced changes in soil properties. Compared with the biochars produced from reed and cassava straws, rice straw-derived biochar was the optimal amendment for the stabilization of co-contaminated soil, because it had the best pore and surface structures and exhibited the greatest positive effect on soil properties. Overall, the present results provide guidance for the use of biochars in the remediation of soil co-contaminated with Hg–As and possibly Cd–As. However, further studies are needed in order to determine the optimal dose of rice straw-derived biochar and the most suitable environmental conditions for its field application.

Author Contributions

Conceptualization, Y.W. and R.L.; methodology, Y.W.; validation, Y.W. and N.L.; formal analysis, Y.W. and B.Z.; writing—original draft preparation, Y.W.; writing—review and editing, Y.W. and N.L. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Technology Innovation Center for Land Engineering and Human Settlements, Shaanxi Land Engineering Construction Group Co., Ltd. and Xi’an Jiaotong University, China, grant number 2021WHZ0094.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

All data are presented in this study and thus contained within the article. There are no other available data in any publicly accessible repository.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. SEM images of biochar samples prepared by pyrolysis of different feedstocks. (a) REB, reed straw biochar; (b) CAB, cassava straw biochar; (c) RIB, rice straw biochar.
Figure 1. SEM images of biochar samples prepared by pyrolysis of different feedstocks. (a) REB, reed straw biochar; (b) CAB, cassava straw biochar; (c) RIB, rice straw biochar.
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Figure 2. Effects of biochar amendments on soil physicochemical properties: (a) pH, (b) EC, (c) OM, (d) TOC, (e) CEC. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
Figure 2. Effects of biochar amendments on soil physicochemical properties: (a) pH, (b) EC, (c) OM, (d) TOC, (e) CEC. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
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Figure 3. Effect of biochar amendments on soil enzyme activities: (a) dehydrogenase, (b) catalase, (c) invertase, and (d) urease. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
Figure 3. Effect of biochar amendments on soil enzyme activities: (a) dehydrogenase, (b) catalase, (c) invertase, and (d) urease. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
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Figure 4. Distribution of (a) Hg and (b) As heavy metals among five fractions in soil samples with different biochar amendments: F1, exchangeable fraction; F2, carbonate fraction; F3, Fe/Mn oxide-bound fraction; F4, organic-bound fraction; F5, residual fraction. REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
Figure 4. Distribution of (a) Hg and (b) As heavy metals among five fractions in soil samples with different biochar amendments: F1, exchangeable fraction; F2, carbonate fraction; F3, Fe/Mn oxide-bound fraction; F4, organic-bound fraction; F5, residual fraction. REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar.
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Figure 5. Effects of biochar amendments on different fractions of heavy metals: (a) exchangeable Hg (F1-Hg), (b) residual Hg (F5-Hg), (c) exchangeable As (F1-As), and (d) residual As (F5-As). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05).
Figure 5. Effects of biochar amendments on different fractions of heavy metals: (a) exchangeable Hg (F1-Hg), (b) residual Hg (F5-Hg), (c) exchangeable As (F1-As), and (d) residual As (F5-As). REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05).
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Figure 6. Bioconcentration factors (BCFs) of heavy metals in spinach plants after biochar remediation. REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05).
Figure 6. Bioconcentration factors (BCFs) of heavy metals in spinach plants after biochar remediation. REB, reed straw biochar; CAB, cassava straw biochar; RIB, rice straw biochar. Error bars represent standard deviations of the means. Different letters above bars indicate significant differences between mean values of the treatment groups (p < 0.05).
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Table
Table 1. Physicochemical properties of the soil sample used in experiments.
Table 1. Physicochemical properties of the soil sample used in experiments.
Particle Size Distribution (%)TexturepHOrganic Matter
(g kg−1)		CaCO3
(%)		Total Hg
(mg kg−1)		Total As
(mg kg−1)
Clay
(0–2 μm)		Silt
(2–50 μm)		Sand
(50–2000 μm)
11.2971.1617.55Silty loam8.968.738.958.74106
Table
Table 2. Experimental treatments used in the study.
Table 2. Experimental treatments used in the study.
TreatmentBiochar TypeBiochar Dose (g pot−1)Plant
1Reed straw biochar (REB)30 (1%)Spinach
2Cassava straw biochar (CAB)
3Rice straw biochar (RIB)
4Control0
Table
Table 3. Main properties of biochars from reed straw (REB), cassava straw (CAB), and rice straw (RIB).
Table 3. Main properties of biochars from reed straw (REB), cassava straw (CAB), and rice straw (RIB).
PropertyBiochar
REBCABRIB
Yield (%)29.24 ± 2.57 c32.20 ± 1.35 b45.71 ± 0.61 a
pH10.42 ± 0.13 b10.11 ± 0.09 c10.81 ± 0.24 a
Electrical conductivity (μS cm−1)689.21 ± 1.36 b1028.05 ± 2.73 a1005.36 ± 2.59 a
S (%)0.32 ± 0.02 c0.51 ± 0.14 b0.69 ± 0.25 a
SiO2 (%)4.02 ± 0.91 b3.58 ± 0.24 b11.08 ± 0.37 a
Average pore size (μm)10.36 ± 1.24 a8.25 ± 1.69 b3.09 ± 0.83 c
Specific surface area (m2 g−1)10.61 ± 0.37 c18.45 ± 0.27 b23.05 ± 1.36 a
Zeta potential (mV)−12.69 ± 0.21 c−14.18 ± 0.44 b−15.96 ± 0.18 a
Data are shown as mean ± standard deviation. Different letters in the same column indicate significant differences between mean values corresponding to different treatment groups (p < 0.05).
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Wei, Y.; Li, R.; Lu, N.; Zhang, B. Stabilization of Soil Co-Contaminated with Mercury and Arsenic by Different Types of Biochar. Sustainability 2022, 14, 13637. https://doi.org/10.3390/su142013637

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Wei Y, Li R, Lu N, Zhang B. Stabilization of Soil Co-Contaminated with Mercury and Arsenic by Different Types of Biochar. Sustainability. 2022; 14(20):13637. https://doi.org/10.3390/su142013637

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Wei, Yang, Risheng Li, Nan Lu, and Baoqiang Zhang. 2022. "Stabilization of Soil Co-Contaminated with Mercury and Arsenic by Different Types of Biochar" Sustainability 14, no. 20: 13637. https://doi.org/10.3390/su142013637

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