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Article

Investigating the Potential of Microbially Induced Carbonate Precipitation Combined with Modified Biochar for Remediation of Lead-Contaminated Loess

by
Pengli He
1,*,
Jinjun Guo
1 and
Shixu Zhang
2
1
School of Civil Engineering, Luoyang Institute of Science and Technology, Luoyang 471023, China
2
School of Civil Engineering, Xi’an University of Architecture and Technology, Xi’an 710055, China
*
Author to whom correspondence should be addressed.
Sustainability 2024, 16(17), 7550; https://doi.org/10.3390/su16177550 (registering DOI)
Submission received: 2 August 2024 / Revised: 22 August 2024 / Accepted: 28 August 2024 / Published: 31 August 2024
(This article belongs to the Special Issue A Study on Sustainable Geological Disaster Prevention and Control in Engineering)
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Abstract

:
Lead (Pb) contamination in loess poses a significant environmental challenge that impedes sustainable development. Microbially induced carbonate precipitation (MICP) is an innovative biomimetic mineralization technology that shows considerable promise in remediating soil contaminated with heavy metals. However, the toxicity of lead ions to Bacillus pasteurii reduces the efficiency of mineralization, subsequently diminishing the effectiveness of remediation. Although biochar can immobilize heavy metal ions, its adsorption instability presents a potential risk. In this study, we first compared the pH, electrical conductivity (EC), unconfined compressive strength (UCS), permeability coefficient, and toxicity leaching performance of lead-contaminated loess specimens remediated using biochar (BC), red mud (RM), red-mud-modified biochar (MBC), and MICP technology. Additionally, we evaluated the mechanism of MICP combined with varying amounts of MBC in remediating lead-contaminated loess combing Zeta potential, X-ray diffraction (XRD) analyses, and scanning electron microscopy (SEM) tests. The results showed that MICP technology outperforms traditional methods such as RM, BC, and MBC in the remediation of lead-contaminated loess. When MICP is combined with MBC, an increase in MBC content results in a higher pH (8.71) and a lower EC (232 us/cm). Toxic leaching tests reveal that increasing MBC content reduces the lead leaching concentration in loess, with optimal remediation being achieved at 5% MBC. Microscopic analysis indicates that the remediation mechanisms of MICP combined with MBC involve complexation, electrostatic adsorption, ion exchange, and precipitation reactions. The synergistic application of MICP and MBC effectively adsorbs and immobilizes lead ions in loess, enhancing its properties and demonstrating potential for pollution remediation and engineering applications.

1. Introduction

Over the past few decades, extensive metallurgical activities associated with rapid urbanization have resulted in widespread heavy metal contamination of soil and groundwater. While some of these metals are essential trace elements for biological processes, their excessive concentrations pose significant threats to both the surrounding environment and human health, primarily due to their bioaccumulative and non-degradable properties [1]. Elevated levels of lead in human blood can lead to lead (Pb) poisoning, which damages the hematopoietic system in bone marrow and the nervous system, adversely affecting the intellectual development of children [2,3,4]. Soil heavy metal contamination is one of the most pervasive and detrimental environmental problems today, attracting significant attention due to its low mobility, prolonged persistence, and resistance to degradation [5]. The current state of heavy metal contamination in the loess regions of Northwest China is particularly severe [6]. In light of this, various remediation techniques, such as soil replacement, electrokinetic remediation, and soil washing, have been extensively employed for the remediation of lead-contaminated soils [7,8,9,10,11,12,13,14]. However, traditional remediation methods present numerous drawbacks, including the high costs associated with physical remediation and the limited environmental friendliness of chemical remediation [11]. Conversely, biological remediation methods demonstrate significant environmental friendliness and align with sustainable development goals, offering promising prospects for future development [15]. In recent years, microbially induced carbonate precipitation (MICP) technology has emerged as a focal point in biomineralization due to its ability to provide numerous nucleation sites. Compared to enzymatic-induced carbonate precipitation (EICP) technology, MICP typically exhibits higher carbonate precipitation and immobilization efficiency [16].
As an alternative to traditional remediation methods, MICP has attracted considerable attention from scholars due to its efficiency, cost-effectiveness, and lower risk of secondary pollution. The main principle of MICP technology is shown in Figure 1 [17]. Bacteria decompose organic matter as an energy source, releasing urease, which hydrolyzes urea to release carbonate ions. These carbonate ions combine with heavy metal ions to form carbonate precipitates, thereby reducing their migration and diffusion risk [18]. Some studies also indicate that the negatively charged bacterial surface can adsorb heavy metal cations, and the nucleation sites for carbonate precipitation are the biofilm and bacterial cell walls, which aid in the aggregation of carbonate precipitates [19,20,21,22]. However, the encapsulation of bacteria by such aggregates leads to premature bacterial death, deteriorating the remediation effect. This points to the potential combination of MICP technology with other techniques to reduce the adsorption of heavy metal ions on bacterial surfaces, delay bacterial death, and enhance mineralization.
Furthermore, traditional solidification/stabilization (S/S) technologies incorporate remediation materials into contaminated media through forced mixing or injection [21]. These processes leverage physical and chemical interactions to enhance the physicochemical properties of the contaminated media and reduce pollutant mobility, ensuring the efficient remediation and safe reuse of heavy-metal-contaminated soils. Grounded in the principles of green and sustainable remediation, there is vigorous research and development of low-cost, efficient, and environmentally friendly solidification/stabilization (S/S) materials globally. Modified industrial waste residues, composite phosphate-based solidifiers, alkali-activated materials, and other high-performance adsorbent materials are employed for the solidification and stabilization of pollutants in heavy-metal-contaminated soils [22,23]. On the one hand, biochar (BC) is often regarded as a renewable “carbon-neutral” material [24]. In contrast to non-renewable solidifiers like cement, BC can effectively stabilize heavy metals in contaminated soils, enhancing the soil’s physicochemical properties and promoting plant growth. Consequently, it has extensive applications in soil remediation. BC remediates contaminated soils through adsorption-fixation mechanisms, including physical adsorption, complexation, precipitation reactions, and ion exchange [25,26]. However, the adsorption-fixation efficacy of raw BC is constrained by the structure and composition of its raw materials. In complex environmental pollution scenarios, raw biochar often falls short of achieving the anticipated remediation outcomes. Conversely, red mud (RM), an industrial by-product of alumina extraction from bauxite, possesses a unique oxide-rich structure (notably iron and aluminum oxides), surface hydroxyl groups, and porosity, which confer it with a notable potential for heavy metal adsorption, making it suitable as an environmental remediation material. Studies have also demonstrated that pyrolyzed modified red mud (MBC) exhibits an enhanced adsorption of heavy metal ions [27].
Existing studies suggest that relying solely on microbial remediation for contaminated soils may have limited effectiveness and could potentially increase soil metal concentrations [28]. Therefore, the integration of biomineralization technology with other remediation techniques has gained significant attention in research. MICP has emerged as an eco-friendly remediation technology, often combined with other ecological materials to meet environmental standards [29]. While previous studies have explored the use of MICP in conjunction with biochar (BC) to improve soil strength, there is a lack of information on their application for remediating heavy-metal-contaminated loess. It is essential to carefully select appropriate ecological materials to enhance the efficiency of MICP technology. Therefore, this study explores the feasibility of combining MICP with MBC for the remediation of lead-contaminated loess. This study aims to (1) assess the effectiveness of various materials in solidifying lead-contaminated loess, (2) examine the impact of MBC content on the joint remediation of lead-contaminated loess using MICP and MBC, and (3) elucidate the mechanisms involved in the combined remediation of lead-contaminated loess through MICP and MBC.

2. Testing Materials

2.1. Tested Soil

The Chinese Loess Plateau is the world’s most extensive and thickest loess deposit area [30]. Due to its continuous deposition, it contains climate and environmental data from the sedimentary process and is listed as one of the three major global paleoclimate archives, along with deep-sea sediments and polar ice cores. The Guanzhong area has the most widespread loess distribution, covering an area of 6.4 × 105 square kilometers, with well-developed loess landforms [6]. Loess terraces, alluvial plains, and flatlands are common. This study focused on the Q3 loess in Yaozhou District, Tongchuan City, Shaanxi Province, the Guanzhong Plain, collecting loess samples from about 4.5 m below the surface, which were sealed and transported to the laboratory for preservation [31,32,33]. The authors retrieved a series of block samples from depths of 4.5 m, where loess/paleosol profiles resulting from intensive eluviation and reprecipitation induced by the East Asian Monsoon and Siberian High are distinct. Considering that Q3 loess distributes at shallower depths, it is deemed as the most vulnerable to chemical contaminants. These results lead us to retrieve the samples at such depths. According to the “Standard for Geotechnical Testing Method (GB/T 50123-2019) [34],” the loess specimens were analyzed for particle size distribution, specific gravity, porosity ratio, dry density, moisture content, and liquid–plastic limits. Additionally, the chemical composition of the loess was analyzed using inductively coupled plasma mass spectrometry (ICP-MS). The test results, as shown in Table 1 and Figure 2, indicate that the loess used in this study is low-plasticity clay [35], with the main component being SiO2, followed by Al2O3, CaO, and Fe2O3. It is important to note that the soil samples used in this experiment do not contain lead ions.

2.2. Red Mud, Biochar, and Modified Biochar

The RM utilized in this experiment was prepared via the Bayer process [36]. Following air drying, the red mud was sieved through an 18-mesh standard sieve for subsequent use. The primary chemical components and their concentrations in the red mud are detailed in Table 2.
The method for preparing MBC in this study involves the co-pyrolysis of peanut shells loaded with red mud particles. A specific quantity of red mud was added to a beaker containing 500 mL of distilled water and stirred with a magnetic stirrer for 30 min to achieve a stable red mud suspension. Subsequently, 10 g of peanut shell powder was added to the suspension and stirred for an additional hour. Upon completion of stirring, the mixture was filtered and dried in an oven at 80 °C for 12 h to obtain the preliminarily treated peanut shells with red mud. The red-mud-treated peanut shell powder was then uniformly placed into a quartz boat. The tube furnace was set to a final carbonization temperature of 700 °C, with a heating rate of 10 °C/min, and a final temperature hold time of 2 h under a nitrogen atmosphere. Following pyrolysis, nitrogen was continuously passed through the furnace until the biochar cooled to room temperature. The red-mud-modified biochar was subsequently weighed, bagged, sealed, and stored in a cool, dry place. The raw biochar was produced by directly pyrolyzing the peanut shell powder in the tube furnace under identical conditions as those used for the red-mud-modified biochar.

2.3. Bacteria and Cementation Solutions for MICP Treatment

Bacteria play two primary roles in the MICP process, as follows: producing urease to hydrolyze urea and providing nucleation sites for calcium carbonate precipitation. Different bacterial strains have varying extracellular polymeric substances (EPS) on their cell surfaces, which can affect mineral nucleation and crystal growth. Consequently, the morphology and crystal types of calcium carbonate induced by different urease-producing bacteria during mineralization vary. The most commonly used bacterium in MICP research, both domestically and internationally, is Sporosarcina pasteurii, a common alkaliphilic aerobic bacterium found in soil. It is characterized by non-aggregating cells, a high specific surface area, a high urease production capacity, and good environmental adaptability. The bacterium used in this study is Sporosarcina pasteurii (CGMCC1.3687), purchased from the China General Microbiological Culture Collection Center [6]. The composition of the liquid culture medium comprised urea (20 g/L), peptone (5 g/L), yeast extract (3 g/L), and manganese sulfate (0.01 g/L). A 10% NaOH solution was used to adjust the pH to 7.0, followed by sterilization of the mixture at 121 °C in an autoclave for 20 min [37]. After allowing the medium to cool, the bacteria were inoculated at a ratio of 1:100 and incubated in a constant temperature shaker set to 30 °C at 180 rpm for 48 h under aerobic conditions. The bacterial concentration was assessed by measuring the optical density at 600 nm (OD600), while the urease activity was evaluated through the average change in conductivity over 5 min, reported as U (mM urea hydrolyzed/min). The OD600 of the cultured bacterial solution reached approximately 1.74, and urease activity was about 4.02 mM urea hydrolyzed/min, determined using a spectrophotometer and a conductivity meter. Furthermore, the cementation solution applied in this experiment consisted of a combination of urea and calcium chloride dissolved in deionized water at a 1:1 ratio [38]. This aligns with multiple prior studies showcasing the efficacy of MICP. In this process, urea acts as a nitrogen source, whereas calcium chloride supplies the necessary calcium. To avoid urea volatilization, the cementation solution was employed within one hour post preparation [39].

3. Testing Methods

3.1. Specimen Preparation

The loess was dried, ground, and sieved for subsequent use. A specific quantity of Pb(NO3)2 powder was dissolved in deionized water to prepare a Pb(NO3)2 solution, which was then sprayed onto the loess and mixed thoroughly. The mixture was subsequently transferred to a sealed bag to maintain a constant moisture content and left in a cool, dark place for 14 days. After this period, it was air dried, ground, and sieved through a 10-mesh sieve for further use. Considering that the control value for lead in the Soil Environmental Quality Risk Management Standards for agricultural land is 240 mg/kg, the pollution concentration was set higher than this control value to explore the remediation mechanisms and behaviors under higher concentrations. This process resulted in contaminated loess with a lead concentration of 3000 mg/kg, which is also 1.2 times that of the regulatory limit for second-class construction land, as stipulated in the ‘Soil Environmental Quality/Risk Control Standard for Soil Contamination of Development Land (GB 36600-2018)’ [40]. The prepared lead-contaminated loess was air dried at room temperature for one month to obtain the heavy-metal-lead-contaminated loess ready for treatment.

3.2. Remediation Process for Lead-Contaminated Loess

A certain mass of lead-contaminated soil and varying proportions of remediation materials were placed in acrylic soil boxes measuring 150 mm × 150 mm × 50 mm. Deionized water was added and uniformly mixed to 60% of the soil’s maximum field capacity. A layer of plastic wrap with ventilation holes was placed over the top of the soil boxes. The soil boxes were then placed in a cool area and weighed every 1–2 days, with deionized water being added to maintain the soil moisture content at 60% of the field capacity. After a specified remediation period, the soil in the boxes was air dried, sieved through a 2 mm mesh, and used for the determination of physicochemical properties and the evaluation of remediation effectiveness. The experimental design for studying the effects of remediation materials, remediation time, and incorporation ratios on the remediation of lead-contaminated loess is shown in Table 3. Here, the incorporation ratio is defined as the mass ratio of the solid components of the MICP treatment agent to the mass of the dry soil. Given that the dry unit weight of the loess and its maximum measured as 13.8 kN/m3 and 17.8 kN/m3, respectively, their ratio gives a degree of compaction of the loess as approximately 80%. For this reason, the authors chose the degree of compaction of 0.80. Furthermore, the authors handled each specimen on the EK reactor in three batches. This ensures that the given degree of compaction of the specimens was accurately attained.

3.3. Physicochemical Properties of Soil

3.3.1. Measurement of Soil pH

According to a soil-to-deionized-water ratio of 1:2.5 (W/V), 10 g of soil to be tested was weighed into a 50 mL centrifuge tube. A total of 25 mL of deionized water was added and mixed. The mixture was shaken at 25 °C and 180 rpm for 30 min [34]. After shaking, it was left to stand for 10 min, and the pH of the supernatant was measured using a pH meter (SX-620 portable pH meter).

3.3.2. Measurement of Soil EC

According to the Standard for Geotechnical Testing Method (GB/T 50123-2019) [34], 5 g of soil to be tested was weighed into a 50 mL centrifuge tube at a soil-to-deionized-water ratio of 1:5 (W/V). A total of 25 mL of deionized water was added and mixed. The mixture was shaken at 25 °C and 180 rpm for 30 min. After shaking, it was left to stand for 10 min, and the conductivity of the supernatant was measured using a conductivity meter (HANNA, HI 2550, Gothenburg, Sweden).

3.4. Measurement of Physical and Mechanical Properties of Soil

3.4.1. Unconfined Compressive Strength (UCS) Test

The bearing capacity of the soil is a key factor affecting its engineering applications. The unconfined compressive strength (UCS) of the samples was tested using a universal testing machine (AG2000A; Shimadzu, Kyoto, Japan). The displacement rate during the test was set to 1 mm/min, and data on time–displacement–force changes and peak stress were recorded [34]. The peak stress of the sample corresponds to its unconfined compressive strength, while the stress–strain relationship is calculated from the force and displacement measurements.

3.4.2. Permeability Test

Permeability is a key indicator for evaluating the safe disposal and application of heavy-metal-contaminated soil, and the permeability coefficient directly reflects the material’s permeability. This study followed the constant head test standards of the ‘Standard for Geotechnical Testing Method (GB/T 50123-2019)’ [34]. The test was conducted using a flexible wall permeameter (GDS Company, Hook Hampshire, UK). The test sample was placed between filter paper and permeable stones and wrapped with a flexible rubber membrane and rubber rings to prevent water from flowing out of the sample’s sidewalls. Subsequently, the confining pressure was set to 50 kPa, the back pressure to 120 kPa, and the bottom back pressure to 130 kPa for saturation treatment. After the sample was fully saturated, the back pressures were removed, and a permeability pressure difference of 100 kPa was applied. Once the flow rate had stabilized, the outflow was measured under the corresponding pressure, and the permeability coefficient was calculated.

3.5. Toxic Leaching Test of Soil

In this study, we employed a custom-designed leaching apparatus to investigate the behavior of the contaminated soil under simulated infiltration conditions. This setup facilitated the percolation of both neutral and acidic solutions through soil samples using a peristaltic pump, which ensured precise control over the flow rate, consistently maintained at 30 mL/h. The leaching height remained constant throughout the process, and the duration of each leaching experiment was standardized to 20 h, as depicted in Figure 3. To achieve accurate control of the leaching process, the vacuum pump played a crucial role by regulating the flow rate and maintaining the integrity of the sealing layer. Specifically, the vacuum pump was set to provide a negative pressure of −10 kPa, enhancing control over the infiltration of the leaching solution. This precise control was essential for accurately simulating real-world conditions. Previous research by He et al. (2023) [6] elucidated the formation of acid rain, attributing it to the reaction of nitrogen oxides or sulfur dioxide with atmospheric water molecules, resulting in acidic solutions that lower the pH of rainwater. Based on these findings, we prepared an acidic solution with a pH of 5.6 by mixing sulfuric acid and nitric acid in a 1:1 molar ratio. This mixture was designed to mimic the toxic leaching characteristics experienced under acid rain conditions, providing a realistic scenario for our experiments. Upon completion of the leaching process, the leachate was carefully collected and analyzed to determine the concentration of lead ions. This analysis was performed using a flame atomic absorption spectrophotometer, a method known for its precision and reliability in detecting metal ions at low concentrations. This analytical technique enabled us to quantify the extent of lead ion leaching and assess the effectiveness of our experimental setup in replicating environmental conditions.

3.6. Microstructure Characteristics of Soil

3.6.1. Zeta Potential Measurement

Charged soil particles suspended in an aqueous solution have the capacity to attract ions of opposite charge to their surfaces [6], as described by He et al. (2023) [6]. Ions in close proximity to the soil particles are strongly attracted, forming relatively stable associations known as the Stern layer. In contrast, ions located farther from the soil particles, beyond the Stern layer, experience weaker attractions and form less stable associations, referred to as the diffuse layer. Heavy metal ions adsorb onto the surfaces of clay minerals through processes such as hydration and ion exchange. These interactions modify the thickness of the diffuse double layer, subsequently affecting the soil structure and its hydraulic properties. The Zeta potential is a critical parameter for characterizing the thickness of the diffuse double layer, providing insights into the electrochemical stability of soil suspensions. In preparation for the experiment, the loess samples were pretreated by passing them through a 50 μm sieve to ensure uniform particle size. Following this, 100 mg of the pretreated loess was transferred into a glass beaker containing 100 mL of various chemical solutions. The mixture was then agitated using a magnetic stirrer for 30 min to ensure thorough interaction between the loess particles and the solutions. Subsequently, the Zeta potential of the loess particles in the presence of different heavy metal solutions was measured using a Zeta potential analyzer. This analytical technique is essential for assessing the electrokinetic behavior of soil particles in various chemical environments. To prevent cross-contamination and ensure measurement accuracy, the microelectrophoresis cell was meticulously cleaned with deionized water before and after each measurement. This rigorous approach not only ensures the reliability of the data, but also provides a comprehensive understanding of how heavy metal ions influence the electrochemical properties of loess, which is vital for predicting soil behavior in contaminated environments.

3.6.2. XRD Tests

The mineralogical composition of both lead-contaminated loess and modified loess specimens was meticulously analyzed using a Bruker AXS X-ray diffractometer (D8 Advance). This analysis aimed to elucidate the interactions between the Microbially Induced Calcite Precipitation (MICP) process and lead-contaminated loess. By comparing the X-ray diffraction (XRD) patterns of the specimens before and after modification, we were able to identify the mineral phases and assess the impact of the MICP treatment. The XRD analysis was conducted at a scanning rate of 2° per minute with a step size of 0.02°, ensuring high-resolution detection of mineralogical changes. These parameters were selected to provide detailed and accurate diffraction patterns, facilitating a comprehensive comparison of the mineral phases present in the samples. The identification of mineral phases was achieved by matching the observed XRD patterns with standard reference patterns. This comparative approach enabled the detection of any new mineral phases formed as a result of the MICP process, as well as alterations to existing minerals due to interactions with lead ions.

3.6.3. SEM Tests

Scanning electron microscopy (SEM) is a widely employed technique for both qualitative and quantitative analysis of soil microstructure. In this study, SEM was utilized to elucidate the seepage characteristics and mechanisms of heavy-metal-contaminated loess. The SEM testing details and equipment are identical to those used in our previously published study [6]. By qualitatively and quantitatively evaluating the microstructure of the loess before and after seepage, we aimed to gain insights into the effects of contamination and subsequent remediation processes. For a comprehensive description of the soil microstructure characterization process, readers can refer to our previous study [6]. SEM images of the loess at various magnifications reveal significant variations in the microstructural characteristics, as depicted in Figure 4. This image presents a detailed comparison of scanning electron microscopy (SEM) observations of geotechnical materials at varying magnifications, specifically aimed at illustrating the relationship between the magnification level and the observable microstructure. At low magnifications (≤200×), wide observation fields are captured, offering a broad overview of the sample’s general texture without revealing microstructural details. At medium magnifications (200X–1000×), the microstructural features, such as particle cementation, become distinguishable, though fine contact relationships between particles remain ambiguous. At high magnifications (1000×–4000×), significant structural details emerge, allowing for a clear observation of particle skeletons, contact relationships, porosity, and cementation processes. Finally, extra-high magnifications (>4000×) focus on single-point observations, offering exceptional detail on minute surface features while losing the broader context of the sample’s overall structure. To effectively capture these variations and understand the evolution of the soil microstructure before and after the remediation of lead-contaminated loess using curing materials, a primary magnification of 1000× was employed. Additionally, based on the findings of Li et al. (1988) [41], magnifications of 500× and 1000× were selected to investigate the microstructure of loess. These magnifications were chosen to provide a detailed view of the microstructural changes, enabling a thorough analysis of the effects of heavy metal contamination and the efficacy of the remediation process. The SEM analysis yielded valuable qualitative and quantitative data on the microstructural evolution of loess, contributing to a better understanding of the mechanisms underlying soil stabilization and remediation. These insights are crucial for optimizing remediation strategies and enhancing the overall effectiveness of soil treatment methodologies.

4. Results and Discussion

4.1. Response of Soil pH and EC Values

The pH value of the environment is a critical factor influencing the speciation and mobility of lead in soil. Extensive studies have demonstrated that soil pH directly affects the solubility and speciation of heavy metals. Specifically, higher soil pH levels result in the decreased mobility of heavy metal ions, while lower pH levels increase their mobility. Consequently, soil pH is considered a pivotal determinant of the bioavailability of contaminants, exerting a direct influence on the physicochemical parameters of heavy metals, including leachate concentration and bioavailability. Furthermore, the electrical conductivity (EC) of the soil is closely related to the concentration of soluble ions within the soil matrix. Elevated EC values typically indicate increased concentrations of soluble ions, which can affect various soil properties and processes. Understanding the interplay between soil pH and EC is essential for predicting the behavior of heavy metals in contaminated soils and developing effective remediation strategies. This study underscores the importance of maintaining optimal soil pH levels to minimize the mobility and bioavailability of heavy metals, thereby reducing their environmental and health impacts. By closely monitoring and adjusting soil pH and EC, we can enhance the efficacy of soil remediation efforts and protect ecosystems from heavy metal contamination.

4.1.1. Soil pH

The results of pH changes in lead-contaminated loess under different influencing factors are shown in Figure 5. The pH value of untreated lead-contaminated loess (CK) is 8.14, indicating a slightly alkaline condition. The addition of RM, BC, MBC, and MICP in the four experimental groups increased the pH values. RM increased the pH value of lead-contaminated loess more significantly than the two types of biochar (BC and MBC), reaching 8.31. BC and MBC showed only slight increases in pH compared to CK. Notably, the MICP experimental group achieved the highest pH value, reaching 8.40. As depicted in Figure 5b, the soil pH value increased with the increasing proportion of MBC. After a 30-day remediation period, the pH values of lead-contaminated loess treated with MICP combined with 1% MBC, 3% MBC, 5% MBC, and 7% MBC were recorded at 8.46, 8.57, 8.63, and 8.71, respectively. These values represent increases of 3.9%, 5.3%, 6.0%, and 7.0%, respectivley, compared to CK. The observed pH elevation can be attributed to several factors. Firstly, the MICP process generates urease, which catalyzes the hydrolysis of urea, releasing a substantial amount of OH− ions and thereby raising the pH value. Secondly, MBC, characterized by high ash content and fewer acidic functional groups, contributes to increased alkalinity. Furthermore, the mineral components present in RM also exhibit alkaline properties.

4.1.2. Soil EC

The changes in EC values of lead-contaminated loess exposed to various remediation methods are shown in Figure 6. The EC value of untreated lead-contaminated loess (CK) is 456 μs/cm. The addition of RM, BC, MBC, and MICP in the four experimental groups all reduced the EC values. Red mud (RM) significantly reduced the EC value of lead-contaminated loess to 389 μS/cm and did so more effectively than the two types of biochar (BC and MBC). BC and MBC showed only slight reductions in EC compared to CK. Additionally, the changes in EC values are negatively correlated with the changes in pH values. As shown in Figure 6b, the EC value of lead-contaminated loess gradually decreases with the increase in the proportion of MBC. Research indicates that during the initial stages of the MICP mineralization process, a large number of mobile ions are released, increasing the EC value of lead-contaminated loess. However, the formation of carbonate precipitates at the end of mineralization reduces the soil’s EC value. Additionally, MBC contains a large number of mineral components and base ions. When MBC is incorporated into lead-contaminated loess, it releases conductive ions such as OH−, Na+, Ca2+, Al3+, and CO32− into the loess. The free heavy metal ions in the loess and the dissolved Ca2+ combine with these conductive ions to form precipitates and other insoluble substances, thereby reducing the concentration of soluble ions in the loess and lowering the EC value.

4.2. Response of Physical and Mechanical Properties

In recent years, due to the need for sustainable development in civil engineering and environmental pollution control, biochar has been widely applied in geotechnics, environmental science, agriculture, and construction. Therefore, research on using MICP as an engineering application material has become increasingly important. From the perspective of strength, MICP is more feasible for engineering applications in soil. In geotechnical engineering, mechanical strength is one of the most important indicators, and permeability significantly affects the migration of heavy metal ions in contaminated soil. Therefore, this section briefly examines the impact of solidification materials and enhancement methods on the basic physical properties of lead-contaminated loess, specifically unconfined compressive strength and permeability. This provides a theoretical basis for addressing sustainable engineering environmental issues in the future.

4.2.1. UCS

Figure 7 shows the unconfined compressive strength of lead-contaminated loess after 30 days of standard curing with different remediation measures. MICP has the highest unconfined compressive strength, reaching 715 kPa, due to its ability to generate carbonate precipitates that increase inter-particle cementation [42,43,44,45]. When MICP is combined with MBC, the unconfined compressive strength of the samples increases gradually with the proportion of MBC. This indicates that the combination of MICP and MBC can significantly enhance the bearing capacity of lead-contaminated loess.

4.2.2. Permeability

The permeability coefficient of loess is closely related to the particle size, mineral composition, internal pores, saturation, and soil structure. Figure 8 shows the permeability coefficient variations of lead-contaminated loess after 30 days of standard curing with different remediation measures. It is evident from the figure that the permeability coefficient of lead-contaminated loess treated with MBC is lower than that treated with BC, indicating that modified BC can provide a more alkaline environment and form a denser pore structure. MICP can independently complete mineralization, producing a large amount of calcium carbonate, which increases the density of the pore structure, thus having the lowest permeability coefficient [45,46,47,48]. When MICP technology is combined with MBC, the resulting denser structure enhances the bearing capacity of the samples, requiring higher permeability pressure to increase the permeability coefficient. In other words, the permeability coefficient decreases under the same permeability pressure. As the proportion of MBC increases, the permeability coefficient gradually decreases. The changes in the permeability coefficient are not significant at 5% and 7% MBC incorporation ratios, possibly due to the stable structure of MBC. A large amount of MBC encasing soil particles reduces the permeability pathways, and calcium carbonate precipitates formed by MICP further reduce these pathways, effectively resisting the permeability pressure of the samples, leading to more stable changes [49]. Therefore, the combination of MICP and MBC can alter the permeability of loess, enhancing the permeability of lead-contaminated loess. In other words, although the UCS gradually increases with the addition of MBC, a closer comparison reveals that, when the MBC content exceeds 5%, the growth trend of UCS becomes less significant.

4.3. Response of Lead Ion Leachate Concentration

When contaminated soil comes into contact with a liquid phase, lead ions in the soil migrate to the liquid phase. The concentration of ions in the liquid phase is used to evaluate the effectiveness of the materials in remediating lead-contaminated loess [50]. This liquid phase is commonly referred to as the leachate of leaching toxicity. In actual contaminated sites, rainwater infiltration is usually the main pathway for toxicity diffusion. To evaluate the remediation effectiveness of lead-contaminated loess, the concentration of lead ions in the leachate was extracted using a self-developed leaching apparatus, as shown in Figure 9. Under neutral rainwater conditions, the leachate concentration of MICP was 54.8% lower than that of CK. Notably, although MBC showed significant differences from MICP in the aforementioned physical and mechanical tests, the difference in leachate concentration between MBC and MICP was minimal and almost negligible. This is because MBC has a large specific surface area and dense pore structure, providing sufficient adsorption sites for lead ions in the loess, significantly enhancing the leachate resistance of lead-contaminated loess. MICP technology releases a large number of hydroxide and carbonate ions, which facilitate the precipitation of lead ions [51,52,53,54]. Additionally, although BC can increase the pH of loess, MBC enhances it to a greater extent than ordinary BC. When MICP is combined with MBC, the pH of the soil increases, creating an alkaline environment. This has two effects, as follows: First, the alkaline environment enhances the negative charge of MBC, promoting the hydrolysis of acidic functional groups on the MBC surface and increasing the density of cation-exchange sites on MBC. This leads to more lead ions undergoing ion exchange and complexation reactions with MBC. Second, the negative charge on colloids such as iron and manganese oxides and mineral components in the loess increases, promoting the adsorption of lead ions and reducing the active lead ions in the loess, thereby decreasing the concentration of soluble lead ions in the loess. Furthermore, a detailed comparison of the experimental results reveals that, although the permeability coefficient gradually decreases with the increasing MBC content, when the MBC content exceeds 5%, the permeability coefficient shows almost no further change. This phenomenon suggests that, while MBC encapsulates soil particles and reduces soil porosity, excessive amounts of MBC may lead to agglomeration, preventing any further reduction in porosity. The synergistic effect of MICP and 5% MBC content is optimal.

4.4. Analysis of Microscopic Results

4.4.1. Zeta Potential Test Results

Heavy-metal-induced cation exchange and adsorption can increase leachate ion concentration, reduce soil strength, and alter soil microstructure. It is well known that this effect is based on soils containing clay minerals. Loess contains abundant clay minerals, such as montmorillonite, illite, and kaolinite. Therefore, cation exchange and adsorption must be considered in the leaching process of lead-contaminated loess. The intrusion of lead ions disrupts the original charge balance of clay minerals, causing Na+, K+, Ca2+, and Mg2+ to adsorb onto the clay minerals, resulting in cation exchange. This is one of the reasons for the increased leachate ion concentration. Heavy metal ions in the exchange phase can also reduce the thickness of the diffuse double layer (DDL), promote the flocculation of clay particles, and increase pore space. Additionally, the DDL effect needs to be considered when the clay content exceeds 10%. In this study, the clay content in the loess was 12.47%, indicating that the DDL effect on loess during leaching cannot be ignored. This effect reduces the DDL thickness and promotes the formation of aggregate structures. This phenomenon has also been observed in other types of heavy-metal-contaminated soils, indicating that the results of this study are widely recognized. The thickness of the DDL can be evaluated by the Zeta potential [6]. As shown in Figure 10, with the use of solidification materials, the alkaline environment gradually intensifies, and the Zeta potential values of both the solidification materials and the loess surface increase, enhancing the adsorption of lead ions. However, higher values cannot explain the formation of aggregate structures, as an increase in Zeta potential indicates a tendency for soil particles to form a dispersed structure, which is inconsistent with the mechanism of MICP technology [6,52,53]. The reason for this is that the combination of MICP and MBC increases soil density, providing more cementing particles and enhancing compaction.

4.4.2. XRD Test Results

The XRD patterns of lead-contaminated loess and lead-contaminated loess remediated with a combination of MICP and MBC are presented in Figure 11. The XRD pattern of the lead-contaminated loess reveals a distinct peak for PbCO3 (cerussite), indicating that the adsorption of lead ions in the loess samples primarily occurs through ion exchange mechanisms. In this process, cations in carbonate minerals, such as calcite, are replaced by lead ions, which subsequently form lead carbonate precipitates with carbonate ions. In the XRD pattern of the lead-contaminated loess samples treated with the combination of MICP and MBC, the presence of lead carbonate peaks and the emergence of basic lead carbonate peaks provide compelling evidence of MICP mineralization. This indicates that MICP facilitates the formation of lead carbonate, while MBC, in close contact with the soil, effectively adsorbs and immobilizes lead ions in the loess. The synergistic action of MICP and MBC in the remediation process is evident from the observed changes in the XRD patterns. The vibrations of the lead carbonate peaks and the emergence of new peaks associated with basic lead carbonate confirm the successful mineralization and stabilization of lead ions. These results demonstrate that the combined application of MICP and MBC not only promotes the formation of stable lead carbonate minerals, but also enhances the overall immobilization of lead ions, thereby achieving effective remediation of lead-contaminated loess. This finding highlights the potential of integrating MICP and MBC for the sustainable remediation of heavy-metal-contaminated soils. These findings provide a robust theoretical and practical basis for developing advanced soil remediation technologies that leverage microbial processes and biochar amendments to mitigate environmental contamination and improve soil health [42,43].

4.4.3. SEM Test Results

Figure 12 illustrates the microstructure of lead-contaminated loess samples both before and after remediation using a combination of MICP and MBC. The untreated lead-contaminated loess, polluted by lead nitrate solution, displays a relatively rough surface morphology and a compromised surface structure. Additionally, a significant presence of network-like aggregates is observed, attached to the surface of the loess particles. These network structures are primarily the result of the interface precipitation of calcite minerals in the loess and lead ions, leading to the formation of substantial amounts of cerussite. Following remediation with the combination of MICP and MBC, the microstructure of the loess exhibits noticeable changes. The copious precipitates generated by MICP mineralization, in conjunction with MBC, effectively fill the soil pores, significantly reducing the pore volume. The introduction of MICP facilitates the mineralization of free lead ions, resulting in the formation of lead carbonate precipitates. This process markedly enhances the immobilization of lead within the soil matrix [52,53,54]. Moreover, the addition of MBC plays a crucial role in the adsorption of Pb2+ ions. MBC’s porous surface structure allows Pb2+ ions to diffuse into its pores, facilitating effective adsorption. Additionally, MBC can adsorb a significant amount of Pb2+ through complexation with the functional groups present on its surface. The negatively charged surface of MBC also contributes to the adsorption of Pb2+ ions via electrostatic interactions. These combined effects of MICP and MBC not only stabilize the lead ions through precipitation and adsorption, but also improve the overall soil structure by reducing pore spaces and enhancing soil integrity.

4.5. Discussion

The interactions among MICP, MBC, loess, and lead ions are quite complex. The introduction of MICP and MBC directly and significantly affects the chemical interactions of loess with lead ions, as shown in Figure 13. Urease produced by MICP can promote the hydrolysis of urea, contributing to the formation of an alkaline environment and increasing the soil pH [52,53]. It also forms calcium carbonate or lead carbonate precipitates, which, together with MBC, fill the soil pores, enhancing unconfined compressive strength and reducing the permeability coefficient. Additionally, the mechanisms by which the combination of MICP and MBC reduces leachate concentration are analyzed as follows:
(1)
MBC solidification of Pb2+: The solidification of Pb2⁺ by MBC occurs through several distinct mechanisms. Firstly, the alkaline nature of biochar elevates the pH of the soil, promoting the reaction between Pb2⁺ and hydroxide ions (OH⁻), which leads to the formation of lead hydroxide (Pb(OH)₂), an insoluble compound that significantly reduces the leaching of lead [52]. Secondly, MBC possesses a large specific surface area, enabling it to adsorb substantial amounts of free Pb2⁺. This adsorption capacity effectively immobilizes the lead ions within the soil matrix, further decreasing the leachate concentration of Pb2⁺.
(2)
Influence of MBC on MICP solidification of Pb2+: In the microbially induced calcite precipitation (MICP) process, the effectiveness of solidifying Pb2⁺ is often hindered by the lack of sufficient nucleation sites for free urease enzymes. This can result in the independent formation of calcium carbonate (CaCO₃) within pore spaces, which prevents the full integration of CaCO₃ with soil particles, thus limiting the solidification efficiency of metal ions [53]. MBC, due to its highly porous structure and strong adsorption capacity, introduces additional binding sites by adsorbing both the enzymes and the substrates needed for enzymatic reactions. This adsorption enhances the catalytic decomposition of urea by urease, allowing for more efficient CaCO₃ formation. Moreover, certain metal ions are adsorbed within the pores of MBC, where mineralization reactions occur, leading to localized high concentrations of metal ions. This environment promotes the timely solidification of metal ions through the MICP mineralization process. Other metal ions are either adsorbed within the porous network of MBC or are integrated into CaCO₃ precipitates formed during mineralization [53,54]. These multiple adsorption and mineralization pathways synergistically contribute to the effective immobilization and solidification of metal ions, resulting in successful remediation of Pb2⁺ contamination.
The present study further compares the results with state-of-the-art research in terms of treatment duration, soil type, UCS, and leaching concentration [55,56,57,58,59,60]. As reported by Hadi et al. (2022) [59], a UCS of 512 kPa was achieved through microbially induced calcite precipitation (MICP) for the remediation of Pb contamination over a 28-day treatment period (see Table 4). Under the same treatment conditions, the UCS for the synergistic remediation combining MICP and MBC was 1.59 times higher, likely due to the enhancement provided by MBC. Furthermore, a novel remediation process required a treatment period 1.33 times longer than that used for the combined MICP and MBC remediation, despite having a higher leaching concentration [55]. Additionally, the UCS achieved was 1.29 to 3.54 times higher, and the initial concentration was 7.5 to 30 times greater compared to that of MICP alone, despite the longer treatment time [56,57,58,59,60]. These results suggest that the combined MICP and MBC method demonstrates superior performance in terms of UCS and the remediation of lead-contaminated loess. However, material selection for the enhancement of MICP is crucial, as it directly impacts both the mechanical properties and the leaching concentration of lead ions in the treated loess. Therefore, materials that improve mechanical strength and reduce lead ion leaching are considered the main criteria for selecting appropriate enhancements for MICP remediation. The synergistic remediation approach not only enhances MICP’s effectiveness, but also broadens its applicability in the remediation of contaminated sites.

5. Conclusions

This study compared the remediation effects of different materials on lead-contaminated loess and further investigated the combined use of MICP technology and MBC. Based on the results and discussion, some main conclusions can be drawn, as follows:
(1)
Compared to traditional RM and BC methods, MBC releases more hydroxide and carbonate ions, resulting in a higher pH, lower EC, and reduced leachate concentration. In comparison, MICP introduces even more carbonate ions through biomineralization, leading to further increases in pH and reductions in EC. Additionally, the large amounts of carbonate precipitates produced by MICP increase the density of lead-contaminated loess, enhancing UCS and lowering permeability, while also reducing leachate concentration.
(2)
The combination of MICP and MBC in lead-contaminated loess remediation leads to higher pH levels. Carbonate precipitates, along with MBC, fill loess pores, improving UCS and reducing permeability. As the MBC proportion increases, the soil becomes denser, enhancing its bearing capacity while further reducing permeability. This combined method effectively adsorbs and immobilizes lead ions, significantly decreasing their concentration in the leachate and achieving successful remediation.
(3)
The combined use of MICP and MBC elevates soil pH, creating an alkaline environment. This enhances the negative charge of MBC, promotes the hydrolysis of acidic functional groups on its surface, and increases cation-exchange site density. These changes facilitate ion exchange and complexation reactions between MBC and lead ions. Additionally, the increased negative charge in lead-contaminated loess promotes the further adsorption of lead ions, reducing their mobility and lowering the concentration of soluble lead ions in the loess.
(4)
The present study primarily focuses on the synergistic remediation of microbially induced calcite precipitation (MICP) and microbial biochar (MBC) in lead-contaminated loess. Consequently, the results are applicable solely to the behavior observed during the designated period of MICP remediation. The effects of remediation age and potential enhancements through the integration of other technologies have not been addressed in this work, which may limit the applicability of the findings to contaminated sites.

Author Contributions

P.H.: Investigation, Formal Analysis, Writing—Reviewing and Editing. J.G.: Investigation, Formal Analysis, Writing—Reviewing and Editing. S.Z.: Conceptualization, Methodology, Resources, Writing—Original draft preparation. All authors have read and agreed to the published version of the manuscript.

Funding

This research received no external funding.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data used to support the findings of this study are available from the corresponding author upon request.

Conflicts of Interest

The authors declare that there are no conflicts of interest regarding the publication of this paper. All of the authors have read and approved this version of the article, and due care has been taken to ensure the integrity of the work. I would like to declare on behalf of my co-authors that the work described is original research that has not been published previously and is not under consideration for publication elsewhere, in whole or in part. We declare that all authors have no actual or potential conflicts of interest, including and financial, personal, or other relationships with other people or organizations.

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Figure 1. The process of MICP (EPS: extracellular polymeric substances) (Modified from [17]).
Figure 1. The process of MICP (EPS: extracellular polymeric substances) (Modified from [17]).
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Figure 2. (a) Particle size distribution of Q3 loess and (b) liquid limit and plastic index.
Figure 2. (a) Particle size distribution of Q3 loess and (b) liquid limit and plastic index.
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Figure 3. Schematic illustration of the leaching filtration device.
Figure 3. Schematic illustration of the leaching filtration device.
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Figure 4. Features of SEM images at different magnifications.
Figure 4. Features of SEM images at different magnifications.
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Figure 5. (a) Variations of soil pH against material types and (b) relations of the soil pH versus the proportion of the MBC.
Figure 5. (a) Variations of soil pH against material types and (b) relations of the soil pH versus the proportion of the MBC.
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Figure 6. (a) Variations of soil EC against material types and (b) relations of the soil EC versus the proportion of the MBC.
Figure 6. (a) Variations of soil EC against material types and (b) relations of the soil EC versus the proportion of the MBC.
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Figure 7. (a) Variations of UCS against material types and (b) relations of the UCS versus the proportion of the MBC.
Figure 7. (a) Variations of UCS against material types and (b) relations of the UCS versus the proportion of the MBC.
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Figure 8. (a) Variations of permeability against material types and (b) relations of the permeability versus the proportion of the MBC.
Figure 8. (a) Variations of permeability against material types and (b) relations of the permeability versus the proportion of the MBC.
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Figure 9. (a) Variations of leaching concentration of lead against material types and (b) relations of the leaching concentration of lead versus the proportion of the MBC.
Figure 9. (a) Variations of leaching concentration of lead against material types and (b) relations of the leaching concentration of lead versus the proportion of the MBC.
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Figure 10. (a) Schematic illustration of the diffuse double-layer model and (b) variations of Zeta potential against material types.
Figure 10. (a) Schematic illustration of the diffuse double-layer model and (b) variations of Zeta potential against material types.
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Figure 11. Results of X-ray diffraction for the specimens (a) before and (b) after remediation.
Figure 11. Results of X-ray diffraction for the specimens (a) before and (b) after remediation.
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Figure 12. Results of SEM for the specimens (a) before and (b) after remediation.
Figure 12. Results of SEM for the specimens (a) before and (b) after remediation.
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Figure 13. Schematic illustration of the mechanisms of lead-contaminated loess remediation using MICP and MBC.
Figure 13. Schematic illustration of the mechanisms of lead-contaminated loess remediation using MICP and MBC.
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Table 1. Physical properties of the loess.
Table 1. Physical properties of the loess.
Physical Index.Data
Fines (%)91.18
Sand (%)8.82
Gravel (%)0
Specific gravity, Gs2.72
Void ratio, e0.88
Dry density, ρdmax/(g/cm3)1.78
Initial water content, ѡn/%16.4
The Atterberg limit
  Liquid limit, ωL/%33.42
  Plastic limit, ωP/%20.43
  Soil classificationCL
Table 2. Chemical element composition of the loess specimen and red mud.
Table 2. Chemical element composition of the loess specimen and red mud.
Chemical ElementLoess (%)Red Mud (%)
SiO255.0831.14
CaO14.4718.31
Al2O313.5824.32
Fe2O37.739.41
MgO2.631.25
K2O3.360.64
Na2O1.419.60
SO3-0.18
Table 3. Design of the remediation experiments.
Table 3. Design of the remediation experiments.
CategoryMaterialsIncorporation Ratio (%)Ration of MBC (%)Treatment Time (d)Influencing Factors
CKCK3/30Material type
RMRM3/30
BCBC3/30
MBCMBC3/30
MICPMICP3/30
MICP + 1%MBCMICP3130Ration of MBC
MICP + 3%MBCMICP3330
MICP + 5%MBCMICP3530
MICP + 7%MBCMICP3730
Table 4. Comparison of this study with the state-of-the-art research about treatment time, soil type, UCS, and leaching concentration.
Table 4. Comparison of this study with the state-of-the-art research about treatment time, soil type, UCS, and leaching concentration.
ContaminantInitial Concentration (mg/kg)Treatment Time (d)Soil TypeUCS (kPa)Leaching Concentration (mg/L)References
Zn100010XinJiang soil625105[55]
//1Sand730/[56]
Mn836.6245Pyrite tailings Sand587.8613.68[57]
Cd100021Red clay229.70.002[58]
Pb40028Sand512/[59]
Pb10014/6300.5[60]
Pb300030Loess81246.4This study
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He, P.; Guo, J.; Zhang, S. Investigating the Potential of Microbially Induced Carbonate Precipitation Combined with Modified Biochar for Remediation of Lead-Contaminated Loess. Sustainability 2024, 16, 7550. https://doi.org/10.3390/su16177550

AMA Style

He P, Guo J, Zhang S. Investigating the Potential of Microbially Induced Carbonate Precipitation Combined with Modified Biochar for Remediation of Lead-Contaminated Loess. Sustainability. 2024; 16(17):7550. https://doi.org/10.3390/su16177550

Chicago/Turabian Style

He, Pengli, Jinjun Guo, and Shixu Zhang. 2024. "Investigating the Potential of Microbially Induced Carbonate Precipitation Combined with Modified Biochar for Remediation of Lead-Contaminated Loess" Sustainability 16, no. 17: 7550. https://doi.org/10.3390/su16177550

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