Remediation of chromium-contaminated soil in semi-arid areas by combined chemical reduction and stabilization

ABSTRACT

The indoor simulation remediation of chromium-contaminated soil in semi-arid areas with chemical reduction-stabilization method was conducted to study the effects of remediation agent types and dosage, water (mL)-soil (g) ratio and remediation time on reduction and stabilization efficiency of the chromium-contaminated soil. The results showed that FeSO4 was the most effective reduction agent for reduction of Cr(VI) in soil. The optimally combined chemical reduction-stabilization remediation conditions were 2.5 times theoretical reaction dosage FeSO4 and 8% of modified corn stover biochar (KBC), 50% of water-soil ratio, and 7 days remediation time. The total Cr and Cr(VI) leaching concentrations and Cr(VI) content in soil were reduced from 65.65 mg L-1, 61.98 mg L-1 and 1000.00 mg kg-1 without treatment to 0.114 mg L-1, 0.125 mg L-1 and 18.909 mg kg-1 after treatment, respectively. They were according with comprehensive wastewater discharge standards and construction land soil contamination risk control standards, respectively.

Graphical abstract

KEYWORDS:

• Semi-arid area
• chromium contaminated soil
• chemical reduction
• stabilization
• soil remediation

1. Introduction

Chromium (Cr) is a common heavy metal element in the natural environment and an important raw material in industrial processes. Cr and its compounds are widely used in electroplating, tanning, printing and dyeing, metallurgy, metal processing, pigment manufacturing, chemical pharmaceutical, and other industries [1–4]. However, the emission of Cr-containing waste liquid, gas, and sludge produced by these industries during the production process may lead to soil and groundwater contamination in the surrounding areas [5]. The study showed that the Cr content in Chinese arable soils ranged from 0.05 to 3353.60 mg kg⁻¹, with an average Cr content of 78.94 mg kg⁻¹, which was significantly higher than its background value (57.30 mg kg⁻¹), indicating that human activities have introduced it into the soil environment [6]. According to the National Soil Pollution Investigation Bulletin in 2014, of which chromium pollution in China exceeds the risk value in the national soil sampling points by 1.1% [7]. In the soil environment, Cr usually exists in two dominant states: trivalent chromium (Cr(III)) and hexavalent chromium (Cr(VI)). Cr(III) usually exists in the form of hydroxides and oxides or is adsorbed onto clay particles, soil organic matter, metal oxyhydroxides, and so on, hindering its migration into soil and groundwater [8]. On the contrary, Cr(VI) exists mainly as relatively soluble oxygen-containing anions such as HCrO-4, CrO2-4, and Cr₂O2-7. Compared with Cr(III), Cr(VI) has greater solubility, stronger activity, and is easy to migrate in soil, so it is more toxic [9]. Many studies [10] have reported that Cr(VI) is 100 times more toxic than Cr(III), and Cr(VI) is easily absorbed by the human body and accumulates in the body. Long-term exposure to Cr(VI) may cause cancer, and the U.S. Environmental Protection Agency identified Cr(VI) as a Class A carcinogen [11]. Therefore, high priority must be given to the remediation of chromium-contaminated soil.

At present, there are two main methods to treat chromium-contaminated soil: one is to remove chromium directly from the contaminated soil, and the other is to change the form of chromium present in the soil. Reduction of Cr(VI) with high toxicity and high migration capacity to Cr(III) with low toxicity and low migration capacity, and Cr(III) was precipitated to reduce the toxicity, bioavailability, and migration ability of Cr(VI) in the natural environment. Based on the above treatment methods, researchers have proposed a variety of remediation techniques for chromium-contaminated soil, such as chemical reduction technology, solidification/stabilization technology, chemical leaching technology, electrokinetic remediation technology, and bioremediation technology [12–16]. Among them, chemical reduction technology is the most widely used treatment method for chromium-contaminated soil at home and abroad due to its advantages of simple operation, obvious treatment effect, and low cost [17]. Yang et al. [18] provided a detailed classification of reducing agent types, and common reducing agents include iron-based reducing agents (e.g. zero-valent iron and ferrous salts), sulfur-based reducing agents (e.g. polysulfide compounds and thiosulfates, etc.), and organic materials (e.g. biochar and humic acid). Lu et al. [19] were able to achieve 70% and 87% stabilization ratios for industrial chromium slag using FeSO₄ and Na₂S as reducing agents, respectively. Li et al. [20] found experimentally that Fe(NH₄)₂(SO₄)₂ effectively remediated soil contaminated with Cr (VI) and, acted as a fertilizer by supplying nitrogen because it contains ammonium.

Chemical reduction technology is usually used in combination with stabilization technology to remediate chromium-contaminated soil by combining reducing agents and stabilizers [21–24], so as to achieve the purpose of green remediation such as reducing the dosage of reducing agents and reducing the influence of reducing agents on the soil. Biochar is produced from carbon-rich biomass by pyrolysis under oxygen-limited conditions [25]. Some studies have reported that due to its highly developed porosity, high degree of surface reactivity, and excellent adsorption capacity [26–28], biochar has been increasingly recognized as the great potential to remove heavy metals [29,30], especially for Cr(VI) with highly redox activity [31–33] in soil. Raw biochar has a limited adsorption capacity for a high concentration of Cr(VI), and modification of biochar is needed to improve the adsorption capacity. Among the various modification methods [34–37], alkali-modification biochar can increase the specific surface area, change pH, introduce more oxygen-containing functional groups, and can facilitate the removal of anionic Cr(VI) by increasing the positive charge on the biochar surface [38]. At present, the mechanism of long-term stability of chromium-contaminated soil after remediation is still under study. It is the key to effectively improve the stability of remediation soil and ensure the optimal chromium endowment speciation in soil. With the rapid development of agriculture and industry, chromium contamination in the soil of the semi-arid northwest areas to deepen. However, there have been no reports on chemical reduction agents combined with stabilizers for remediation of chromium-contaminated soil in semi-arid areas.

This study compared the effects of four reduction agents, FeSO₄, Fe(NH₄)₂SO₄, Na₂S and Na₂SO₃, on the leaching concentration of total Cr and Cr(VI) under the conditions of different dosages, water (mL)-soil (g) ratio and remediation time through indoor simulation experiment, which determined the best reduction agent type and reduction conditions. On this basis, three stabilizers, calcium magnesium phosphate (CMP), quicklime (CaO) and modified corn stover biochar (KBC), were combined to screen the stabilizer type and dosage, and select the combination of reduction agent and stabilizer that is economical, efficient, environmentally friendly and stable in long-term effect. And the stabilization effectiveness before and after soil remediation was compared, so as to evaluate the practicality and feasibility of the combined chemical reduction and stabilization technology, with a view to providing theoretical and management basis for the remediation of chromium-contaminated soil in semi-arid areas, and feasible technical solution for its remediation and treatment.

2. Materials and methods

2.1 Cr(VI) contaminated soil preparation

The indoor simulated restoration experiment was conducted in September 2021 at Lanzhou City in northwestern China. This area belongs to a semi-arid region with low precipitation (180–450 mm annually), high evapotranspiration (1500–1600 mm annually), and high temperatures in the summer (16–28°C). The raw soil used is gray calcium soil, excavated from unpolluted soil at 0–20 cm depth with a spade in the campus greenery of Northwest Minzu University at Lanzhou city, and sub-samples collected were mixed to obtain a composite sample in accordance with the standard method [39]. The soil sample was transferred to the laboratory in plastic bags, naturally air-dried and milled, and then passed through a 2 mm sieve to remove large stones and grass clippings. The chemical and physical properties of raw soil are listed in Table 1. The chromium-contaminated soil used in the experiment was simulated soil. The soil with Cr(Ⅵ) content of about 1000 mg kg⁻¹ was prepared with potassium dichromate (K₂Cr₂O₇) solution. After mixing well with a glass rod, the mouth of the beaker was sealed with plastic film. The soil was cured under natural conditions for 7 days, dried in an oven at 60°C, and then milled to pass through a 0.149 mm sieve until ready for use.

Table 1. Chemical and physical properties of raw soil.

Property	pH	Water content (%)	Conductivity (μs cm^−1)	CEC (cmol kg^−1)	OM (g kg^−1)	Cr(Ⅵ) (mg kg^−1)	Total Cr Leaching concentration (mg L^−1)	Cr (VI) Leaching concentration (mg L^−1)
Value	8.06	1.31	448	27.75	20.87	1.758	0.083	0.077

2.2 Preparation and modification of biochar

The corn straw collected from the suburbs of Lanzhou City was washed repeatedly with tap water, dried in an oven at 60°C, and then milled to pass through a 0.25 mm sieve until ready for use. The corn straw powder was compressed in a crucible, and then pyrolyzed in a muffle furnace at 500°C for 2 h under constant temperature and oxygen limitation. After cooling to room temperature, milled to pass through a 0.149 mm sieve to obtain the corn stover biochar (BC), which was sealed and stored until ready for use.

Biochar (50 g) was added to 1000 mL of 2 mol L⁻¹ KOH solution which was fully soaked at a ratio of 1:10. The suspension was dispersed with an ultrasonic cleaner for 30 min, and then oscillated at a constant temperature of 25°C for 24 h. The resulting product was washed with deionized water until the pH of the waste liquid reached 7.0, and then dried in an oven at 60°C. The sample was milled to pass through a 0.149 mm sieve to obtain KBC which was sealed and stored until ready for use.

2.3 Experiment design

2.3.1 Screening of optimal reducing agent and dosage

Chromium-contaminated soil (100.0 g) was added to a 500 mL plastic beaker. The reducing agents ferrous sulfate (FeSO₄), ammonium iron(II) sulfate (Fe(NH₄)₂SO₄), sodium sulfide (Na₂S), and sodium sulfite (Na₂SO₃) were dissolved in 40 mL deionized water at 1, 1.5, 2, 2.5, and 3 times theoretical reaction dosage for the chemical reaction with chromium-contaminated soil. The samples maintained the water (mL)-soil (g) ratio at 40%. After stirring well with a glass rod, sealed the beakers with plastic film and reduced at room temperature for 7 days. At the same time, the soil without adding any reducing agent was used for the blank test. After reducing, the soil dried in an oven at 60°C, and then milled to pass through a 0.149 mm sieve, and put into a plastic bottle. The total Cr and Cr(VI) leaching concentrations in soil were measured to determine the optimal reducing agent and dosage. Three parallel experiments were performed for each test condition.

2.3.2 Screening of optimal reduction conditions for chromium-contaminated soil

According to the results of Section 2.3.1, the optimal reduction agent and dosage were selected, and the single-factor experiment was performed. The stabilization effectiveness of total Cr and Cr(VI) in Chromium-contaminated soil under different water (mL)-soil (g) ratios (40%, 50%, 60%, 80%, and 100%) and remediation times (3, 7, 14, 21, 28, and 30 d) was examined to determine the optimal reduction conditions. The test methods are the same described in Section 2.3.1.

2.3.3 Chromium-contaminated soil stabilization experiment

Chromium-contaminated soil (100.0 g) was added to a 500 mL plastic beaker. The stabilizers CMP, CaO, and KBC were added to the beaker at 5%, 10%, 15%, 20%, 25%, and 30% of the soil mass, respectively, and the total Cr and Cr(VI) leaching concentrations in soil were measured. The test methods are the same described in Section 2.3.1.

2.3.4 Remediation experiment of chromium-contaminated soil by combined chemical reduction and stabilization

Chromium-contaminated soil (100.0 g) treated with the optimal reduction agent and dosage for 3 days was added to a 500 mL plastic beaker. The stabilizers CMP, CaO, and KBC were added to the beaker at 0.5%, 1%, 1.5%, 2%, 2.5%, 3%, 5%, and 8% of the soil mass, respectively. The samples maintained the water (mL)-soil (g) ratio at 50%. After stirring well with a glass rod, sealed the beakers with plastic film, reduced at room temperature for 4 days. The total Cr and Cr(VI) leaching concentrations in soil and Cr(VI) content in soil after remediation were measured. The test methods are the same described in Section 2.3.1.

2.3.5 Chemical speciation analysis

Chromium was extracted by MgCl₂ solution in exchangeable (EX); by HAc-NaAc solution in carbonate-bound (CB); by NH₂OH•HCl solution in Fe-Mn oxide-bound (OX); by HNO₃-H₂O₂-NH₄Ac solution in organic matter-bound (OM); and by HNO₃-HClO₄-HF solution in residual (RS) (Table 2). Chromium in soil obtained from Tessier’s sequential extraction procedure [40] was also determined by an ultraviolet-visible spectrophotometer (UV-Vis, P2PC, Meproda Instruments Corporation, China).

Table 2. Tessier’s sequential extraction procedures.

Procedure	Fraction	Extractant	Shaking time and temperature
1	EX	8 ml 1 mol L^−1 MgCl2 (pH = 7)	1 h, 25 ± 2°C
2	CB	8 ml 1 mol L^−1 NaAc-HAc (pH = 5)	8 h, 25 ± 2°C
3	OX	20 ml 0.04 mol L^−1 NH2OH·HCl in 25% (V/V) Hac	4 h, 96 ± 3°C
4	OM	(1) 3 ml 0.02 mol L^−1 HNO3 + 5 ml 30% H2O2 (pH = 2) (2) 5 ml 30% H2O2 (pH = 2) (3) 5 ml 3.2 mol L^−1 NH4Ac in 20%(V/V) HNO3	(1) 2 h, 85 ± 2°C (2) 3 h, 85 ± 2°C (3) 30 min, 25 ± 2°C
5	RS	20 ml 16 mol L^−1 HNO3 + 5 ml 12 mol L^−1 HClO4 + 10 ml 23 mol L^−1 HF	Digestion to clarity and transparency

The bioavailability coefficient was introduced in the paper to study the distribution and bioavailability of each speciation of chromium in soil. The bioavailability coefficient is the sum of the EX and OX fractions content divided by the amount of the total fraction.

2.3.6 Determination of effective speciation Cr(VI) in soil

DTPA and EDTA were used to extract and determine the effective speciation Cr(VI) in chromium-contaminated soil before and after the combined chemical reduction and stabilization remediation, to verify whether the biological effective speciation Cr(VI) content in chromium-contaminated soil before and after the remediation was decreased, and correlation and regression analysis were performed.

1. DTPA extraction method: Chromium-contaminated soil (25.0 g) treated with the optimal reduction agent and dosage for 3 days was added to 50 mL 0.005 mol L⁻¹ DTPA-0.01 mol L⁻¹ CaCl₂-0.1 mol L⁻¹ TEA (pH 7.30) solution.

2. EDTA extraction method: Chromium-contaminated soil (5.0 g) treated with the optimal reduction agent and dosage for 3 days was added to 50 mL 0.05 mol L⁻¹ EDTA (pH 9.0) solution.

The above-treated samples oscillated in a water bath oscillator at 25°C and 180 r min⁻¹ for 2 h, and then centrifuged at 3000 r min⁻¹ for 10 min in a centrifuge. The supernatant was filtered through qualitative filter paper to measure the effective speciation Cr(VI) content.

2.4 Experiment methods

The determination of pH, water content, conductivity, cation exchange capacity, and organic matter content of raw soil was based on the ‘Soil Agrochemical Analysis Method’ [41]. The Cr(VI) content in soil was determined by the alkaline digestion method (HJ 1082–2019, Chinese National Standards) [42]; and 1,5-diphenylcarbohydrazide spectrophotometric method (GB/T 15555.4–1995, Chinese National Standards). The chromium leaching solutions were collected using an extraction agent containing sulfuric acid and nitric acid (HJ/T 300–2007, Chinese National Standards) [43]. The pH of the extraction agent used was 3.20 ± 0.05. Total Cr and Cr(VI) leaching concentrations were measured spectrophotometrically using 1,5-diphenylcarbohydrazide as the chromogenic reagent (GB/T 15555.5–1995 and GB/T 15555.4–1995, Chinese National Standards) [44]. The comprehensive wastewater discharge standards and soil control standards are listed in Table 3. The stabilization efficiency (η) of pharmaceuticals for total Cr and Cr (VI) in soil was calculated using the equation

$\eta =\left(1-\frac{C_i}{C_0}\right)\times 100\mathrm{\%}$

Table 3. Comprehensive wastewater discharge standards and soil control standards.

Parameter	Unit	Standard limit value	Limit value source
Cr (VI) Leaching concentration	mg L^−1	≤0.5	Comprehensive wastewater discharge standards (GB 8978–1996, Chinese National Standards) class I standard value
Total Cr Leaching concentration	mg L^−1	≤1.5
Cr(Ⅵ) content	mg kg^−1	≤78	Construction land soil contamination risk control standards (GB 36600–2018, Chinese National Standards) class II land screening value

where ${ }_0$ and ${ }_i$ are the total Cr and Cr (VI) concentrations (in mg L⁻¹) leached from the unstabilized and stabilized soil, respectively.

3. Results and discussion

3.1 Stabilization effectiveness of reducing agents on total Cr and Cr(VI) in soil

The changes in total Cr and Cr(VI) leaching concentrations with dosage during the treatment of chromium-contaminated soil with different reducing agents are shown in Table 4. The total Cr and Cr(VI) leaching concentrations in soil decreased gradually with the increase of reducing agent dosage. Under the same dosage of reducing agents, the reduction effect of iron-based reducing agents (Fe(NH₄)₂(SO₄)₂, FeSO₄) is better than that of sulfur-based reducing agents (Na₂SO₃, Na₂S). This is because due to the reduction reaction of Cr(VI) occurring at the liquid–liquid interface, ferrous salts can reduce Cr(VI) to Cr(III). At the same time, Fe²⁺ in soil reacts with water to form hydrated iron oxide, which adsorbs Cr₂O2-7 in soil to form insoluble composite precipitates Fe_xCr_1-x(OH)₃ or Fe₂Cr_2-xO₃. Therefore, total Cr and Cr(VI) leaching concentrations decreased significantly under the dual effects of chemical reduction and colloid adsorption. Comparing FeSO₄ with Fe(NH₄)₂(SO₄)₂ at 3 times theoretical reaction dosage, the stabilization effect of FeSO₄ is better than Fe(NH₄)₂(SO₄)₂. This is because the stabilization efficiency of FeSO₄ is greater than that of Fe(NH₄)₂(SO₄)₂ for the same dosage from the point of view of electron transfer for the role of reducing agent, so FeSO₄ was chosen as the optimal reducing agent in this study. The stability of FeSO₄ as the most effective reducing agent lies in the fact that Cr(VI) is rapidly reduced to Cr(III) under the reduction of Fe²⁺, and while toxicity is reduced, the resulting Fe³⁺ and Cr(III) exists in the form of precipitation. Due to the co-precipitation of Fe(OH)₃ and Cr(OH)₃, the reoxidation risk of Cr(III) is reduced after reduction. The results of Yin et al. [45] found that FeSO₄ had the best stabilization effect on chromium-contaminated soil and effectively reduced the total Cr and Cr(VI) contents in the leachate. Total Cr and Cr(VI) leaching concentrations in soil met the requirements of the discharge standards when the FeSO₄ dosage was the theoretical reaction dosage (2.5 times and 3 times), while the stabilization efficiency improvement was less obvious when the FeSO₄ dosage was 3 times theoretical reaction dosage. The main reason for this phenomenon is that chromium in soil exists partly on the outer surface of soil particles and partly wrapped in the inner part [46]. Cr(VI) attached to the outer surface of particles is stabilized by the reduction of FeSO₄ to Cr(III), while Cr(VI) stored in the inner part of soil particles is difficult to react with FeSO₄ effectively, making it difficult to further improve the stabilization efficiency, and the excessive FeSO₄ will lead to soil acidification and secondary pollution. Therefore, 2.5 times theoretical reaction dosage of FeSO₄ was chosen as the optimal dosage in this study.

Table 4. Effect of reductant types and dosage on the leaching concentration of total Cr and Cr(VI) in soil.

Reductant	Theoretical dosage (times)	Cr (VI) leaching concentration (mg L^−1)	Stabilization rate (%)	Total Cr leaching concentration (mg L^−1)	Stabilization rate (%)
Na2SO3	1	53.423	13.81	61.636	6.11
	1.5	48.037	22.50	57.927	11.76
	2	45.364	26.81	51.418	21.68
	2.5	43.545	29.74	51.055	22.23
	3	27.363	55.85	29.982	54.33
Na2S	1	26.818	56.73	27.600	57.96
	1.5	20.455	67.00	21.418	67.38
	2	17.945	71.05	18.782	71.39
	2.5	15.509	74.98	16.636	74.66
	3	2.382	96.16	4.236	93.55
Fe(NH4)2SO4	1	0.640	98.97	1.578	97.60
	1.5	0.567	99.09	1.489	97.73
	2	0.496	99.20	1.439	97.81
	2.5	0.459	99.26	1.316	98.00
	3	0.434	99.30	1.303	98.01
FeSO4	1	0.886	98.57	1.014	98.46
	1.5	0.795	98.72	0.959	98.54
	2	0.530	99.14	0.759	98.84
	2.5	0.485	99.22	0.724	98.90
	3	0.205	99.67	0.595	99.09

3.2 Water (mL)--soil (g) ratio and remediation time on stabilization effectiveness of total Cr and Cr(VI) in soil with FeSO₄ as reducing agent

The water (mL)-soil (g) ratio is one of the most important factors affecting the reduction of heavy metals in soil. To ensure adequate reaction between the reducing agent and the heavy metals, it is usually necessary to add appropriate deionized water for treatment. From the screening experiment results of reducing agent and dosage, the effect of different water (mL)-soil (g) ratios on total Cr and Cr(VI) leaching concentrations in soil is shown in Figure 1(a). When the water (mL)-soil (g) ratio increased from 40% to 50%, 60%, 80%, and 100%, total Cr and Cr(VI) leaching concentrations in soil were 0.724 mg L⁻¹, 0.674 mg L⁻¹, 0.759 mg L⁻¹, 0.814 mg L⁻¹, 0.850 mg L⁻¹ and 0.485 mg L⁻¹,0.471 mg L⁻¹, 0.504 mg L⁻¹, 0.522 mg L⁻¹, 0.544 mg L⁻¹, respectively. Therefore, when the water (mL)-soil (g) ratio is 50%, the stabilization effectiveness was obvious; at water (mL)-soil (g) ratio less than or greater than 50%, the stabilization effectiveness was less optimal. The analysis shows that when the water (mL)-soil (g) ratio was low, the slurry was thicker, Fe²⁺ and Cr₂O2-7 could not react completely, and stabilization effectiveness was poor. With the increase of water (mL)-soil (g) ratio, the migration ability of Cr₂O2-7 in soil was enhanced, and the reducing agent dissociated more reducing ions, which accelerated the reaction between Fe²⁺ and Cr₂O2-7, and improved the stabilization efficiency of Cr₂O2-7. However, when water (mL)-soil (g) ratio from 50% to 100%, the reducing agent concentration was continuously diluted, and its stabilization ability of total Cr and Cr(VI) decreased gradually, resulting in increased total Cr and Cr(VI) leaching concentrations in soil with increased water (mL)-soil (g) ratio. Therefore, the 50% of water (mL)-soil (g) ratio was selected as the best according to the soil remediation target value.

Figure 1. Effect of water -soil ratio (a) and remediation time (b) on the leaching concentration of total Cr and Cr(VI) in soil with FeSO₄ as reducing agent.

Based on the above experimental results, the effect of different remediation times on total Cr and Cr(VI) leaching concentrations in soil is shown in Figure 1(b). With the extension of remediation time, total Cr and Cr(VI) leaching concentrations in soil decreased gradually. When the remediation time was 3 days, total Cr and Cr(VI) leaching concentrations in soil were 0.763 mg L⁻¹ and 0.591 mg L⁻¹, respectively, while Cr(VI) leaching concentration in soil exceeded the requirements of the discharge standards. When the remediation time was increased from 7 days to 30 days, total Cr and Cr(VI) leaching concentrations in soil decreased from 0.674 mg L⁻¹ and 0.471 mg L⁻¹ to 0.371 mg L⁻¹ and 0.214 mg L⁻¹, respectively, and the stabilization efficiency of total Cr and Cr(VI) only increased by 0.46% and 0.41%, respectively. The stabilization effectiveness was not obvious, and the remediation time of 7 days most of the total Cr and Cr(VI) in soil had been stabilized and met the discharge standards. Therefore, the remediation time of 7 days was selected as the best according to the soil remediation target value.

3.3 Stabilization effectiveness of stabilizers on total Cr and Cr(VI) in soil

The changes in total Cr and Cr(VI) leaching concentrations with dosage during the treatment of chromium-contaminated soil with different stabilizers are shown in Figure 2. The total Cr and Cr(Ⅵ) leaching concentrations in soil decreased gradually with the increase of stabilizer dosage, which the stabilization effectiveness of KBC on Cr₂O2-7 in soil was the best. The total Cr and Cr(VI) leaching concentrations of untreated contaminated soil were 65.65 mg L⁻¹ and 61.98 mg L⁻¹. For the CMP-stabilized soil, the total Cr and Cr(VI) leaching concentrations decreased from 54.315 mg L⁻¹ and 53.114 mg L⁻¹ to 32.147 mg L⁻¹ and 30.497 mg L⁻¹, respectively, when the dosage of CMP was increased from 5% to 30%. Their stabilization efficiency increased from 17.27% and 14.30% to 51.03% and 50.80%, respectively. This is because CMP belongs to phosphate stabilizers, whose main components include Ca₃(PO₄)₂, CaSiO₃, and MgSiO₃ [47]. After CMP was incorporated into chromium-contaminated soil, the slowly released Ca²⁺ can combine with Cr₂O2-7 to generate stable CaCr₂O₇ precipitate, which reduces the migration activity of Cr(VI). For the CaO-stabilized soil, the total Cr and Cr(VI) leaching concentrations decreased from 27.206 mg L⁻¹ and 27.771 mg L⁻¹ to 11.658 mg L⁻¹ and 10.839 mg L⁻¹, respectively, when the dosage of CaO was increased from 5% to 30%. Their stabilization efficiency increased from 58.56% and 55.19% to 82.24% and 82.51%, respectively. This may be due to the hydrolysis reaction of CaO to generate Ca(OH)₂, which raises the pH of the soil, and the free Ca²⁺ in solution reacts with Cr₂O2-7 to generate CaCr₂O₇ precipitate. CaCr₂O₇ can exist stably in alkaline and neutral environments. For the KBC-stabilized soil, the total Cr and Cr(VI) leaching concentrations decreased from 53.709 mg L⁻¹ and 52.909 mg L⁻¹ to 31.982 mg L⁻¹ and 32.727 mg L⁻¹, respectively, when the dosage of KBC was increased from 5% to 30%. Their stabilization efficiency increased from 18.19% and 14.64% to 51.28% and 47.20%, respectively. This is because KBC, with its excellent porous structure and abundant oxygen-containing functional groups [7], adsorbs Cr(VI) on the surface and reductively transforms it into Cr(III), after which Cr(III) is stabilized by KBC through mechanisms such as surface precipitation, surface coordination, and ion exchange.

Figure 2. Effect of CMP (a), CaO (b) and KBC (c) on the leaching concentration of total Cr and Cr(VI) in soil.

In conclusion, the stabilization experiments showed that the total Cr and Cr(VI) leaching concentrations after stabilization with CMP, CaO, and KBC were still higher than the requirements of the emission standards. Therefore, further treatment is needed to enhance the stability of chromium in soil and ensure environmental safety.

3.3 Stabilization effectiveness of combined chemical reduction and stabilization remediation on chromium-contaminated soil

The effects of FeSO₄ combined with different dosages of stabilizers on the total Cr and Cr(VI) leaching concentrations and Cr(VI) content in soil are shown in Figure 3 and Figure 4. From the perspective of stabilization effectiveness, compared with FeSO₄ alone, the stabilization effectiveness of FeSO₄ combined with three kinds of stabilizers showed different degrees of change, and the overall stabilization effectiveness was KBC, CaO, and CMP in order from strong to weak. The total Cr and Cr(VI) leaching concentrations and Cr(VI) content of KBC at each dosage met the requirements of the emission standard, and the 8% dosage of KBC had the best stabilization effectiveness in remediating chromium-contaminated soil. At this time, the total Cr and Cr(VI) leaching concentrations and Cr(VI) content in soil were 0.125 mg L⁻¹, 0.114 mg L⁻¹, and 18.909 mg/kg, respectively. Their stabilization efficiency was 99.81%, 99.82%, and 98.11%, respectively. This is because, on the one hand, oxygen-containing functional groups such as C = O, -OH, -COOH, and C-O on the surface of KBC are involved in the adsorption of Cr(VI) and occur in a specialized adsorption behavior [48]; on the other hand, the pH of soil decreases after FeSO₄ treatment, and Fe³⁺ is adsorbed on KBC surface under acidic conditions and further reduced to Fe²⁺. Fe²⁺ can directly reduce Cr(VI) to Cr(III) [31]. In the reduction process, Fe²⁺ is electronically oxidized to Fe³⁺. Fe³⁺ can be re-adsorbed to the surface of KBC and act as an electron shuttle again, thus improving the stability effectiveness. Cr(III) is easily electrostatically adsorbed to form Cr(OH)₃ precipitate by the negatively charged KBC on the surface. At the same time, Cr(III) can form a poorly soluble Fe_xCr_1−x(OH)₃ or Fe₂Cr_2−xO₃ composite precipitate with Fe²⁺ and Fe³⁺, thus improving the removal rate of Cr(VI). The stabilization effectiveness of 8% dosage of CaO was taken second place. At this time, the total Cr and Cr(VI) leaching concentrations and Cr(VI) content in soil were 0.210 mg L⁻¹, 0.183 mg L⁻¹, and 24.000 mg kg⁻¹, respectively. Their stabilization efficiency was 99.66%, 99.70%, and 97.60%, respectively. This is because CaO has a strong chromium adsorption capacity, raises the pH of the soil, ionizes under alkaline conditions to produce OH⁻, and forms more stable Cr(OH)₃ precipitate with amorphous Cr(III). At the same time, Cr(III) forms a poorly soluble Fe_xCr_1−x(OH)₃ or Fe₂Cr_2−xO₃ composite precipitate [49] under the joint action of Fe²⁺. Compared with KBC and CaO, FeSO₄ combined with CMP has poor stability effectiveness.

Figure 3. Effect of combined chemical reduction and stabilization remediation with CMP (a), CaO (b) and KBC (c) in concert with FeSO₄ on the leaching concentration of total Cr and Cr(VI) in soil.

Figure 4. Effect of combined chemical reduction and stabilization remediation with CMP, CaO and KBC in concert with FeSO₄ on Cr(VI) content in soil.

3.5 Chemical speciation analysis

The biotoxicity and stability of chromium in soil are more largely determined by the bound morphology of chromium. The morphology of chromium and its transformation are important for the study of the environmental effects of heavy metals, the treatment, and remediation of contaminated soils, and their stability. The effect of FeSO₄ combined with different stabilizers on the morphological transformation of chromium in soil was studied by Tessier continuous extraction method. As Figure 5 shows, chromium species in chromium-contaminated soil were exchangeable (EX, 55.7%), carbonates-bound (CB, 1.9%), Fe-Mn oxides-bound (OX, 14.5%), organic matter-bound (OM, 11.3%) and the residual (RS, 16.5%), and the bioavailability coefficient of Cr(VI) was 0. 5763. Compared with the chromium-contaminated soil, the EX and CB fraction were almost completely converted to OX, OM, and RS when FeSO₄ alone treats chromium-contaminated soil. Especially, both OM and RS fractions increased by 22.5%, possibly because chromium in soil was mainly in the form of Cr(OH)₃ or Cr(III)/Fe(III) oxides/hydroxides [49] after FeSO₄ treatment, which accounted for the reduced total Cr and Cr(VI) leaching concentrations in the above remediation tests.

Figure 5. Effect of combined remediation with reductor of FeSO₄ and stabilizers of CMP (a), CaO (b) and KBC (c) on chromium speciation in soil. (Unstable-with FeSO₄ only, CK-blank control without remediation).

When 0.5% to 8% of CMP was added as stabilizer, the EX fraction was slightly higher than that in FeSO₄ treatment. Accordingly, the total Cr and Cr(VI) leaching concentrations in soil were also higher, which was consistent with the results of previous studies. The OX and OM fraction gradually decreased with the increase of the dosage, while the RS fraction showed a gradually increasing trend. This was because, CMP was a silica-aluminate complex containing phosphate radical (PO3-4), and the aqueous solution was alkaline. CMP could slowly release PO3-4 when added to chromium-contaminated soil, and free Cr(III) could react with PO3-4 to form the chemically extremely stable CrPO₄ precipitate. As the solubility of CrPO₄ was lower than that of Cr(OH)₃, more chromium occurred in the RS fraction. When 0.5% to 8% of Cao and KBC were added as stabilizers, the EX and CB fractions were basically consistent with that of FeSO₄ treatment. The OX and OM fractions gradually decreased with the increase of the dosage, while the RS fraction showed a gradually increasing trend. Among the three stabilizers, KBC has the best stabilization effectiveness. This was because, after the reduction of Cr(VI) to Cr(III) by Fe²⁺, Cr(III) had a similar ionic radius to Fe³⁺ and was easily adsorbed on the surface of Fe(OH)₃, while Cr(III) in the presence of KBC undergoes ion exchange and complexation reactions with the oxygen-containing functional groups in KBC, resulting in the morphology transformation of chromium to a more stable form. When the dosage of KBC was 8%, the RS fraction reached 90.4%, which was 2.3 times higher than that of chromium-contaminated soil and 5.5 times higher than that of FeSO₄ when treated alone, and the bioavailability coefficient of Cr(VI) was reduced to 0.0029, which was a remarkable stabilization effectiveness.

In summary, it was determined that the decreasing trend of total Cr and Cr(VI) leaching concentrations and Cr(VI) content in soil were significantly better than that of CaO and CMP when KBC was used as stabilizer, and the RS fraction was most superior. The optimally combined chemical reduction and stabilization remediation condition for stabilizing chromium in soil were 2.5 times the theoretical reaction amount FeSO₄ and 8% KBC.

3.6 Changes in effective speciation Cr(VI) in soil

Two chelating agents (DTPA and EDTA) were used to extract the effective speciation Cr(VI) in chromium-contaminated soil. The effect of different dosages of stabilizers combined with FeSO₄ on the effective speciation Cr(VI) content in chromium-contaminated soil was explored, as shown in Figure 6. With the increase of CMP, CaO, and KBC, the effective state Cr(VI) contents in soil gradually decreased, and KBC had the best stabilization effect. The effective speciation Cr(VI) content of DTPA extraction was significantly lower than that of EDTA, which may be due to the different types of chelating agents with different effects on the chelation of Cr(VI) in chromium-contaminated soil. Many studies had shown that EDTA, HCI, Mehlich3, DTPA, NH₄OAc, CaCl_2, and other extractants extracted the largest amount of soil heavy metals was EDTA [50], which was consistent with the results of this experimental study. The effective speciation Cr(VI) contents extracted by DTPA and EDTA were 0.098 mg kg⁻¹ and 1.663 mg kg⁻¹, respectively, when the dosage of KBC reached 8%.

Figure 6. Effect of combined chemical reduction and stabilization remediation on the effective speciation Cr(VI) content in soil for different types of extraction methods with EDTA (a) and DTPA (b).

The correlations between the effective speciation Cr(VI) content and Cr(VI) leaching concentration in soil for different types of extraction methods are shown in Table 5 and Figure 7. Significantly positive correlation (p < 0.01) was found between the effective speciation Cr(VI) content and Cr(VI) leaching concentration in soil. The correlation coefficients (R) of EDTA extraction method were 0.967 for CMP, 0.877 for CaO, and 0.798 for KBC, respectively; those of DTPA extraction method were 0.914 for CMP, 0.955 for CaO, and 0.811 for KBC, respectively. The fitted equation changes mainly conformed to the linear equation. This indicated that when the biologically effective speciation Cr(VI) content in soil decreases, the Cr(VI) concentration in the soil leachate was also affected and tends to decrease.

Figure 7. Stepwise regression analysis between Cr(VI) leaching concentration and EDTA -Cr(VI) content in soil with different remediation agency of CMP (a), CaO (b) and KBC (c) and DTPA-Cr(VI) content with CMP (d), CaO (e) and KBC (f).

Table 5. Correlation coefficients (R) between Cr(VI) leaching concentration and effective speciation Cr(VI) content in soil for different types of extraction methods with EDTA and DTPA (n = 9).

Stabilizer	Leaching concentration	Speciation	R
CMP	Cr(VI) leaching concentration	EDTA-Cr(VI)	0.967**
		DTPA-Cr(VI)	0.914**
CaO	Cr(VI) leaching concentration	EDTA-Cr(VI)	0.877**
		DTPA-Cr(VI)	0.955**
KBC	Cr(VI) leaching concentration	EDTA-Cr(VI)	0.798**
		DTPA-Cr(VI)	0.811**

Note: * Significant at level p < 0.05, ** significant at level p < 0.01.

4. Conclusions

In this study, the idea of combined treatment of chemical reduction and stabilization was adopted, and the optimal experiment conditions of the combined remediation method were determined as follows: FeSO₄ was used as the reducing agent, the dosage was 2.5 times theoretical reaction dosage, water (mL)-soil (g) ratio was 50%, and remediation time was 7 days. KBC was used as stabilizer, and its dosage was 10%. The bioavailability coefficient of Cr(VI) reduced from 0.5763 to 0.0029 under the above conditions to remediate the chromium-contaminated soil. The total Cr and Cr(VI) leaching concentrations and Cr(VI) content of the remediated soil were 0.114 mg L⁻¹, 0.125 mg L⁻¹, and 18.909 mg kg⁻¹, respectively. They were according with comprehensive wastewater discharge standards and construction land soil contamination risk control standards. The effective speciation Cr(VI) contents extracted by DTPA and EDTA were 0.098 mg kg⁻¹ and 1.663 mg kg⁻¹, respectively, which were significantly and positively correlated with the Cr(VI) leaching concentration. The results of chemical speciation analysis showed that the highly reactive EX and CB fractions decreased significantly when FeSO₄ treatment alone. After the reduction and stabilization of FeSO₄ combined with stabilizer, the OX and OM fractions gradually decreased with the increase of the dosage, while the RS fraction showed an increasing trend. These results indicated that chemical reduction and stabilization could effectively improve the stability of Cr(VI) in soil and reduce the toxicity of Cr(VI). These results were highly favorable and provide useful insights and directions for future efforts in chemical reduction and stabilization of chromium-contaminated soil.

Disclosure statement

No potential conflict of interest was reported by the author(s).

Additional information

Funding

This research was supported by the Fundamental Research Funds for the Central Universities (Grant number: 31920190014), the National Natural Science Foundation of China (Grant number: 21966028) and 2021 Key Talent Project of Gansu, 2021 Science and Technology Project of Gansu (Grant number: 21YF5GA062).