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Article

Enhanced Remediation of Lead and Cadmium by the Co-System of Phosphate-Solubilizing Bacteria Immobilized on Goethite-Modified Biochar

by
Gongduan Fan
1,*,
Junhou Zhou
1,
Xingfeng Cao
1,
Wu You
2,3,
Chen Lin
1,
Jing Luo
4,
Jianyong Zou
5,
Kai-Qin Xu
1,* and
Quanda Luo
3
1
College of Civil Engineering, Fuzhou University, Fuzhou 350116, China
2
School of Life Sciences, Fujian Agriculture and Forestry University, Fuzhou 350002, China
3
Fujian Agricultural Ecological Environment and Energy Technology Extension Station, Fuzhou 350001, China
4
Fujian Jinhuang Environmental Sci-Tech Co., Ltd., Fuzhou 350002, China
5
Anhui Urban Construction Design Institute Co., Ltd., Hefei 230051, China
*
Authors to whom correspondence should be addressed.
Water 2024, 16(13), 1917; https://doi.org/10.3390/w16131917
Submission received: 30 May 2024 / Revised: 24 June 2024 / Accepted: 26 June 2024 / Published: 5 July 2024
(This article belongs to the Special Issue Application of Biochar and Activated Carbon in Water Treatment)
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Versions Notes

Abstract

:
Bioremediation has drawn widespread concern in passivating heavy metals, but the intense toxicity of heavy metals to biological cells limits the application of functional strains. Herein, goethite-modified biochar (GMB) was chosen as the carrier to immobilize phosphate-solubilizing bacteria (PSB) of strain L1 for lead and cadmium remediation. Batch experiments showed that the GMB-L1 possessed excellent adsorption performance with a maximum adsorption of 496.54 and 178.18 mg/g for Pb and Cd, respectively. Moreover, adding GMB-L1 in contaminated soil converted heavy metals (Pb and Cd) into more stable fractions and reduced TCLP-extracted heavy metal concentrations (73.24% of Pb and 57.25% of Cd). The GMB-L1 was proved to accomplish Pb and Cd remediation via the process of chemical precipitation, surface complexation, electrostatic attraction, and biomineralization, which was accompanied by the transformation of heavy metals into a more stable crystal structure, such as Pb5(PO4)3OH and Cd5(PO4)3OH. Therefore, the co-system of GMB and strain L1 could be regarded as a prospective option for efficiently remedying environmental heavy metal pollution.

1. Introduction

Environmental pollution caused by heavy metals is regarded as one of the most severe global problems owing to their high toxicity, difficult degradation, and extended persistence [1,2]. Human endeavors involving mining and smelting, industrial operations, and agricultural production exacerbate heavy metal pollution in the environment [3]. In the case of lead (Pb) and cadmium (Cd) pollution, it has long been considered a significant environmental issue. The overabundance of Pb in the human body can result in irreversible hazards on the nervous, circulatory, immunological, and kidney systems [4]. Heavy exposure to Cd could cause kidney, bone, and pulmonary damage [5,6]. Furthermore, the coexistence of multiple heavy metals in the environment will inescapably contribute to higher levels of toxicity. Therefore, the adverse impacts associated with heavy metal pollution are tremendous, and an efficient and ecological method is urgently needed for the remediation of heavy metal pollution.
Up to now, a variety of methods have been devised to attenuate the heavy metal contamination, such as physical remediation, chemical remediation, and biological remediation [7]. Among these methods, bioremediation has attracted great concern because of its economical, operational, and environmentally friendly characteristics [8,9]. Microorganisms can use metabolic activities such as biosorption, biotransformation, and biomineralization to achieve the adsorption, enrichment, and transformation of pollutants, thereby diminishing the harm of heavy metals [10]. As one of most extensively used functional strains, phosphate-solubilizing bacteria (PSB) exhibit the ability to transform insoluble phosphate to soluble phosphate by producing organic acids and phosphatases. The interaction between the soluble phosphate released by PSB and heavy metals can form stable phosphorus compounds, which in turn reduces the mobility and bioavailability of heavy metals [11]. Although many microorganisms have been used to remediate heavy metal contamination, the presence of elevated levels of heavy metal in the environment can severely inhibit the activity of microorganisms and even result in their death [12]. For instance, excessive lead can adversely affect microorganisms by causing cell membrane disruption, enzyme inactivation, and DNA damage [13]. Hence, it is critically important to explore techniques to preserve microorganisms from heavy metal toxicity.
To improve the resistance of microbes to elevated levels of heavy metals, the immobilization of bacteria on carrier materials is an effective technique. Compared to free bacteria, microorganisms immobilized on carrier materials provide numerous advantages, both in terms of shielding microorganisms from harmful substances and avoiding the loss of microorganisms, which in turn maintain their high activity and stability [14,15]. The use of carrier materials in immobilizing microbes has been reported to provide protection for microbes from the adverse effects induced by contaminants, unsuitable environmental conditions, and competition with indigenous microbial communities in contaminated sites [16,17]. The porous carrier materials have a strong ability to adsorb and stabilize the contaminants in its surface, thus mitigating the toxic effects of contaminants on bacteria [18].
Recently, biochar derived from waste biomass under oxygen-limited conditions has emerged as a popular medium for carrying free bacteria due to its large specific surface area, remarkable porosity, and high-activity surface functional groups [19,20]. Not only can biochar adsorb heavy metals to reduce their biotoxicity through pore adsorption, chemical precipitation, surface complexation, and electrostatic interaction, but its well-developed porous structure can also offer an appropriate shelter for microorganisms to enhance biomineralization [21,22]. The application of biochar to assist PSB in alleviating heavy metal stress might be a strategy to boost the microbial remediation efficiency. However, the original biochar as an immobilization carrier is incapable of effectively alleviating the toxicity on microorganisms at higher heavy metal concentrations due to its restricted adsorption capacity. Therefore, it is particularly important to design multifunctional immobilization matrices to improve the biomineralization of PSB. Researchers in recent years have proved that iron-based nanomaterials are expected to be promising in the remediation of heavy metals due to their relatively low cost and excellent adsorption performance [23,24,25]. Goethite is one of the natural minerals widely found in soil and is considered attractive for the adsorption of various pollutants including heavy metals because of its higher surface area and richer pore structure [26]. Nevertheless, the effectiveness of bare goethite nanoparticles is significantly limited due to their easy aggregation properties caused by the higher surface energy arising from a strong van der Waals force or magnetic force [27]. Multiple studies have confirmed that the introduction of biochar can not only successfully inhibit the agglomeration of goethite nanoparticles, but also bring together the benefits of each component, resulting in superior adsorption performance. Compared to the original biochar, the goethite-modified biochar exhibited an enormous boost in adsorption capacity, with a remarkable enhancement of 135% and 6279% for Cd(II) and As(III), respectively [28]. The biochar after goethite modification could make up the deficiencies of conventional biochar and achieve the simultaneous remediation of Cd(II) and As(V) co-polluted soil [29]. Therefore, the combination of goethite-modified biochar and PSB has a very high potential to remediate toxic heavy metals. However, few studies were focused on the effective immobilization of PSB on the goethite-modified biochar for releasing phosphate to remediate heavy metals, and the remediation mechanism of toxic heavy metals has not been deeply investigated.
Herein, the purposes of this research were (1) to synthesize a biochemical compound (GMB-L1) combining the advantages of goethite-modified biochar (GMB) and phosphate-solubilizing bacteria (PSB); (2) to investigate the adsorption behavior of heavy metals by GMB-L1 under different conditions, including adsorption kinetics and isotherms; (3) to examine the remediation effect of GMB-L1 on heavy-metal-polluted soil; and (4) to explore possible heavy metal passivation mechanisms of the co-system of GMB and PSB. The results of this study provide insights into the development of green and cost-effective inoculants for remediating heavy-metal-contaminated soil.

2. Materials and Methods

2.1. Materials

The surface layer of soil (0–20 cm) was collected from Fujian Province, China, and was dried naturally, pre-treated to eliminate stones and leaves, and then crushed and sieved with a 16-mesh sieve. Soil contamination was simulated by adding Pb(NO3)2 and Cd(NO3)2·4H2O solutions and aged for 60 days at room temperature for the subsequent studies. The physicochemical characteristics of experimental soil are displayed in Table S1.
The strain L1 employed for immobilization was screened from agricultural soil located in Fujian Province, China. Firstly, strains with Pb(II) and Cd(II) tolerance were preliminarily isolated from soil. Subsequently, their phosphorus-solubilizing ability was examined to finally obtain PSB with outstanding resistance to heavy metals (detailed description can be seen in Text S1).

2.2. Performance of Strain L1

An amount of 2% of strain L1 was inoculated in 50 mL of Ca3(PO4)2 medium with Pb(II)/Cd(II) concentrations of 0, 50, 100, 200, and 400 mg/L, and incubated for 72 h at 30 °C to evaluate its phosphate-solubilizing performance. And the solubilized phosphate concentration was detected through the phosphomolybdate method [30].
To determine the heavy metal tolerance performance of strain L1, 2% of strain L1 was inoculated and cultured in 100 mL of LB medium with Pb(II)/Cd(II) concentrations of 0, 50, 100, 200, and 400 mg/L. The growth curve of strain L1 was plotted by monitoring the variation in OD600 values during the period of incubation.

2.3. Synthesis of Goethite-Modified Biochar and Immobilization of PSB

The biochar was obtained from ground dry cow manure (through 16-mesh sieve) that was pyrolyzed in a muffle furnace at 700 °C for 3 h with a low heating rate (5 °C/min) under oxygen-limited conditions. The produced biochar is denoted as BC. The preparation of goethite-modified biochar was based on previous research [26]. An amount of 3.2 g of BC was first dissolved in 100 mL of 5 mol/L of KOH solution and mixed with 50 mL of 1.0 mol/L of Fe(NO3)3 solution. Afterwards, the mixture was put into a capped container and aged at 70 °C for 60 h. Finally, the mixture was dried and ground through a 100-mesh sieve to obtain goethite-modified biochar and is marked as GMB.
To immobilize the PSB, 1.0 g of GMB was added to 20 mL of strain L1 incubation solution (OD600 = 1.0) and shaken for 2 h (30 °C, 150 r/min), and then added to an amount of SA solution and mixed thoroughly so that the final SA concentration was 2% (w/v). The resulting mixture was dropped into 2% (w/v) CaCl2 solution through a syringe and cross-linked at 4 °C for 24 h to prepare GMB-L1. Ultimately, GMB-L1 was washed several times with sterile saline to remove of calcium chloride and lyophilized and stored for further experiments.
Various immobilization parameters including temperatures (20, 25, 30, 35, and 40 °C), bacterial incubation solution volumes (5, 10, 20, 30, and 40 mL), SA contents (0.5, 1.0, 2.0, 3.0, and 4.0%), and cross-linking times (2, 4, 8, 12, and 24 h) were optimized by maintaining all factors at a constant level, except the variable under study. The optimization of the immobilization conditions was achieved by testing the number of immobilized bacteria as well as the amount of heavy metal adsorption capacity.

2.4. Aqueous Solution Adsorption Experiment

To explore the heavy metal removal capacity of various materials, free strain L1, GMB, and GMB-L1 were added to 100 mL conical flasks involving 50 mL of LB medium (containing heavy metals) at a dosage of 1.0 g/L, and then incubated for 24 h at 30 °C. The adsorption performance of GMB-L1 on Pb(II) and Cd(II) was further assessed by adding 0.5 g/L of GMB-L1 to 50 mL of LB medium containing Pb(II) (200 mg/L) or Cd(II) (100 mg/L). In addition, the influence of reaction conditions including temperature (20, 25, 30, 35, and 40 °C), initial pH (2, 3, 4, 5, 6, and 7), composite dosage (0.25, 0.5, 1, 2, and 4 g/L), and initial Pb(II)/Cd(II) concentration (25, 50, 100, 200, 300, and 400 mg/L) was investigated. The adsorption process of GMB-L1 was fitted by adsorption kinetics and isothermal adsorption models. Details about the adsorption kinetics and isothermal adsorption models are supplied as Text S2.

2.5. Soil Incubation Experiment

Four treatments were carried out by implementing diverse materials to the polluted soil: control check (CK), 1% of free PSB (T1), 1% of GMB (T2), and 1% of GMB-L1 (T3). Each pot was stuffed with 100 g of soil and incubated for 30 days at 60% of soil water—holding capacity and room temperature. Each treatment was carried out in triplicate and the average values are reported. The fractions of Pb and Cd in soil were assessed by the modified Community Bureau of Reference (BCR) method [31], and the mobility of Pb and Cd in soil was evaluated through the Toxicity Characteristic Leaching Procedure (TCLP) [32]. Moreover, the Nemerow index (NI) and potential ecological risk index (RI) methods were used to estimate the ecological risk of heavy metals in soil, and the relationship between relevant indices and the heavy metal pollution categorization is listed in Table S2 [30,33].

2.6. Characterization

Strain L1 was identified by Sangong Bioengineering (Shanghai, China) Co., Ltd. Field emission scanning electron microscopy (SEM, FEI Nova NanoSEM 230, FEI Czech Republic s.r.o., Brno, Czech Republic) was carried out to observe the surface morphologies of samples. The crystalline structures of the samples were analyzed by X-ray diffraction (XRD, Rigaku Ultima IV, Rigaku Co., Tokyo, Japan). Fourier transform infrared spectroscopy (FTIR, Nicolet iS50, Thermo Fisher Scientific, Waltham, MA, USA) was conducted to detect the functional groups of samples. The elemental compositions of samples were characterized by X-ray photoelectron spectroscopy (XPS, ESCALAB 250XI, Thermo Fisher Scientific, Waltham, MA, USA).

2.7. Analytical Methods

The number of immobilized bacteria was determined by the plate counting method, GMB-L1 was disintegrated in 50 mL of buffer solution containing 0.05 mol/L of Na2CO3 and 0.02 mol/L of citric acid, and then diluted and coated on plates [34]. The OD600 and phosphate were analyzed using a UV spectrophotometer (T6, Beijing Purkinje, Beijing, China). The zeta potential was measured using a laser particle analyzer (Zetasizer Nano ZS 90, Malvern, Malvern, UK). The concentrations of heavy metals were determined using an inductively coupled plasma optical emission spectrometer (ICP-OES, OPTIMA 8000, PerkinElmer, Waltham, MA, USA).

3. Results and Discussion

3.1. Characterization of Strain L1

3.1.1. Identification of Strain L1

The phylogenetic tree was generated and is presented in Figure S1. The sequence homology between strain L1 and Burkholderia sp. strain EF471225.1 was 99%, which implied that strain L1 was identified as Burkholderia sp.

3.1.2. Phosphate-Solubilizing Ability of Strain L1

Figure 1a,b depicts the phosphate-solubilizing ability of strain L1 under different concentrations of Pb(II)/Cd(II). Under the condition without heavy metals present, the phosphate-solubilizing ability of free strain L1 could reach 117.05 mg/L, but it decreased continuously with the rise in Pb(II)/Cd(II) concentration. This phenomenon may be explained by the increase in Pb(II)/Cd(II) concentration deteriorating the condition of the medium that failed to accommodate more microorganisms, leading to the lower phosphate-solubilizing ability of strain L1.

3.1.3. The Growth Curve of Strain L1

The growth of strain L1 in the medium with varied concentrations of Pb(II)/Cd(II) are shown in Figure 1c,d. In the absence of Pb(II)/Cd(II), the biomass of strain L1 dramatically increased during the initial 24 h, and then plateaued. The addition of heavy metals caused a significant reduction in the biomass of strain L1, and the inhibition effect of Pb(II)/Cd(II) on bacterial growth was further enhanced under higher concentrations because of the toxicity. Nevertheless, it was noteworthy that strain L1 showed excellent heavy metal tolerance, and the growth curve still maintained a positive trend even at the initial Pb(II)/Cd(II) concentration up to 400 mg/L. Therefore, L1 has a high potential for use in the remediation of Pb/Cd pollution.

3.2. Optimization of Immobilization Conditions

3.2.1. Temperature

It can be clearly observed from Figure 2a that temperature significantly affected the immobilized strain L1. With the gradual increase in temperature, the number of immobilized bacteria first increased and then decreased, and then reached a maximum value of about 7.95 × 108 CFU/mL at 30 °C. The lower temperature resulted in the difficulty of bacteria attaching to biochar for growth due to the reduction in extracellular polymers produced by them, whereas higher temperatures could damage the enzymes and proteins required for the growth of microorganisms, which is equally unfavorable for their immobilization on biochar [34]. Meanwhile, the trend in heavy metal adsorption was similar to that of the immobilized number of strain L1, which was due to the rapid growth and metabolism of microorganisms in the appropriate temperature range, and the removal ability of heavy metals was also improved. Based on the above analysis, 30 °C was selected as the optimum immobilization temperature.

3.2.2. Bacterial Volume

As illustrated in Figure 2b, the immobilized number of bacteria increased with the increase in the volume of the bacterial incubation solution within a certain range under the condition of a certain GMB amount. However, the increase in immobilized strain L1 number became slow and tended to decrease when the bacterial solution exceeded a certain volume, which may be attributed to the limited immobilized sites of GMB. In addition, the growth and multiplication of excessive strains caused an increase in the internal pressure of the system, which in turn damaged the reticulation formed by the SA and caused strain leakage. The adsorption of Pb(II) and Cd(II) also partly depended on the immobilized amount of L1. Comprehensively considering the heavy metal removal as well as the microbial utilization efficiency, the 1:20 (w/v) ratio of GMB and bacteria was the most suitable choice.

3.2.3. SA Content

The influence of different SA contents on GMB-L1 was examined and the results are shown in Figure 2c. When the SA content gradually increased, the number of immobilized bacteria presented a tendency of increasing and then decreasing. The reticular structure formed by SA became denser and more stable with the increase in SA content, and the mechanical strength of the immobilized strain also became higher, thereby reducing strain leakage. However, excessive SA content would lead to worse mass transfer and consequently slower diffusion rates of oxygen and reaction substrates, which was detrimental to the growth of the strain. Additionally, the higher SA concentration would increase the viscosity of the bacterial solution, making it difficult to evenly mix with the biochar or drop into the CaCl2 solution, thus increasing the difficulty in the preparation of immobilized bacteria. The adsorption of Pb(II) and Cd(II) also exhibited a tendency of increasing and then decreasing with the increase in SA content, which was mainly due to the variation in strain L1 activity. Therefore, 2% SA content was used to prepare biochar immobilized bacteria for subsequent experiments.

3.2.4. Cross-Linking Time

As depicted in Figure 2d, the number of immobilized strain L1 showed an increasing and then decreasing trend with the continuous extension of cross-linking time, which can be attributed to the change in SA during the cross-linking process. In the early stages of cross-linking, the reticular structure of SA was relatively loose, which made the bacteria susceptible to leakage. As the cross-linking time increased, the embedding of SA became tighter, resulting in an elevated number of immobilized bacteria. However, the bacterial number decreased when the cross-linking time reached 24 h. This phenomenon was attributed to the fact that the overly dense SA impeded the exchange of substances between the microorganisms and the environment, which caused a decrease in the activity of strain L1 [35]. At the same time, the variation in heavy metal adsorption and immobilized bacteria number with cross-linking time was consistent. Based on the above conclusions, 12 h as the cross-linking time was a reasonable choice for preparing better biochar-immobilized bacteria with better performance.

3.3. Removal of Heavy Metals from Aqueous Solution

3.3.1. Removal of Heavy Metals in Different Treatments

The removal capacity of free strain L1, GMB, and GMB-L1 for Pb(II) and Cd(II) was investigated, and the results are shown in Figure S2. The free strain L1 exhibited a relatively low removal efficiency of Pb(II) and Cd(II), which was attributed to the harmful effects of excessive heavy metal ions in solution weakening the activity of strain L1. Compared to free strain L1 and GMB, the co-system of strain L1 immobilized on GBC exhibited a higher remediation ability, and the removal of Pb(II) and Pb(II) could reach 80.87% and 72.58%, respectively. The above results indicated that strain L1 and GMB exhibited complementary potential in remediating heavy metal pollution. In other words, biochar can preferentially immobilize pollutants by adsorption to reduce the toxicity of Pb(II)/Cd(II) on strain L1, which allowed the microorganisms to enter the growth stage more smoothly and achieve better remediation performance [36]. Moreover, the SA covered on GMB-L1 can protect strain L1 from direct contact with heavy metal ions and alleviate the biotoxicity [37].

3.3.2. Effect of Different Factors on Heavy Metal Removal

The reaction conditions were optimized for further improving the removal of heavy metal ions by the co-system of strain L1 immobilized on GMB. The influence of temperature on adsorption can be seen from Figure 3a. The adsorption of heavy metals by GMB-L1 presented a rising trend followed by a declining trend as the temperature increased, and the highest adsorption capacity was obtained at 30 °C. This may be explained by the fact that the growth and metabolism of microorganisms accelerate within a specific temperature range, but excessively high temperature can lead to a sharp decline in cell function and even death [38].
During the remediation of heavy metals, pH as a crucial factor could simultaneously alter the active site and microbial activity of GMB-L1. As noticed in Figure 3b, the adsorption of heavy metal ions by GMB-L1 greatly depended on the solution initial pH values. The adsorption capacity increased when the initial solution pH was raised from 2.0 to 5.0 and reached a plateau at pH 6.0. Furthermore, the zeta potential of GMB-L1 became more electronegative as the initial solution pH was elevated from 1.0 to 8.0 (Figure S3a). At lower pH, the competitive adsorption of H+ as well as the repulsive resistance limited the contact between metal ions and GMB-L1 [39,40]. With the increasing pH, the deprotonation of functional groups offered the opportunity to coordinate with metal ions, leading to a higher removal efficiency. In addition, it could be clearly seen that the maximum adsorption capacity was attained at pH 6, which was also the optimal pH for the growth of strain L1 (Figure S3b). While the pH was above 6.0, the decreasing trend in adsorption capacity might be attributed to the formation of soluble hydroxyl complexes [41].
In order to concurrently achieve highly efficient and cost-effective contaminant removal, an appropriate composite dosage will be required. The impact of various dosages on adsorption capacity was examined, and the findings are demonstrated in Figure 3c. When the dosage of GMB-L1 was raised from 0.25 to 0.50 g/L, the adsorption capacity of Pb(II) and Cd(II) increased from 277.41 and 105.83 mg/g to 297.19 and 122.70 mg/g, respectively. However, the adsorption capacity exhibited a falling trend when the dosage was further increased. The increase in heavy metal adsorption with higher dosage can be due to increased GMB-L1 providing more adsorption sites and functional groups for metals removal [42,43]. The adsorbent aggregation produced by excessive GMB-L1 dosage was responsible for the decrease in adsorption capacity [15,44].
Figure 3d depicts the effect of initial concentration on the adsorption of heavy metals onto GMB-L1. It could be clearly observed that the adsorption capacity of both Pb(II) and Cd(II) increased with growing solution concentration. The metals ions existed on the outer surface of GMB-L1 independently under a low initial concentration and entered the internal structure with the increase in the solution concentration, which improved the contact probability between GMB-L1 and metal ions and led to a higher removal capacity [41,45].

3.3.3. Adsorption Kinetics

The adsorption process of Pb(II) and Cd(II) onto GMB-L1 was fitted with pseudo-first-order models and pseudo-second-order models (Figure 4a), and the kinetic parameters of each model are presented in Table S3. The R2 values of the pseudo-second-order model were higher than those of pseudo-first-order model, and the experimental values of equilibrium adsorption capacity for Pb(II) (295.265 mg/g) and Cd(II) (103.435 mg/g) were much closer to the theoretical values through the pseudo-second-order model. The results indicated that the pseudo-second-order kinetic model could better describe the adsorption behavior of Pb(II) and Cd(II) on GMB-L1. Therefore, it can be concluded that the adsorption of heavy metal ions was dominated by chemisorption [46]. In addition, the values of R2 for both models were greater than 0.96, suggesting the coexistence of physisorption and chemisorption [47].

3.3.4. Adsorption Isotherm

Four acknowledged isotherm models consisting of Langmuir, Freundlich, Redlich-Peterson, and Sips models were then offered to quantitatively fit isothermic data, and the corresponding fitting results and isotherm parameters are shown in Figure 4b and Table S4. Langmuir and Sips models approximated well the adsorption isothermic data according to the higher correlation coefficients, which revealed that the uptake of Pb(II) and Cd(II) was monolayered at the adsorption sites of GMB-L1 [11]. Furthermore, the maximal Pb(II) and Cd(II) uptake on GMB-L1 was approximated to be 496.54 and 178.18 mg/g, respectively, revealing the preeminent performance of GMB-L1 for Pb(II) and Cd(II) passivation.

3.4. Immobilization of Heavy Metals in Soil

3.4.1. Fractions of Heavy Metals in Soil

The BCR method classifies the heavy metals in soil into four fractions, including acid-soluble (F1), reducible (F2), oxidizable (F3), and residual (F4) fractions. The acid-soluble and reducible heavy metals in soil can be easily adsorbed by plants, while the oxidizable and residual fractions are difficult to adsorb and utilize because of their low bioavailability [48].
Figure 5a presents the percentage of Pb and Cd fractions under different treatments in soil. The Pb and Cd in soil mainly existed in the form of acid-soluble and reducible fractions, which together accounted for more than 80% of the total, while the proportions of oxidizable and residual fractions were relatively low. Under the condition of adding strain L1, GMB, and GMB-L1, the content of acid-soluble Pb dropped from 24.99% to 21.92%, 20.52%, and 11.244% after 30 days, respectively. The proportion of reducible Pb decreased by 1.26%, 2.45%, and 6.17% under the treatment of T1, T2, and T3, respectively. Compared with CK, there was no significant variation in the percentage of oxidizable Pb, only increasing by 0.45%, 1.02%, and 1.43%, while the percentage of residual fractions of Pb remarkably rose by 3.88%, 5.90%, and 18.49% with the addition of strain L1, GMB, and GMB-L1, respectively. It can be clearly seen that the changes in different Cd fractions exhibited a similar tendency. After the addition of strain L1, GMB, and GMB-L1, the proportion of acid-soluble Cd decreased by 1.92%, 4.05%, and 10.33%; the reducible Cd only reduced by 0.44%, 0.83%, and 1.52%; and the percentage of residual Cd increased by 1.93%, 4.12%, and 10.41%, respectively. The oxidizable fractions of Cd remained essentially stable, between 12% and 13%. The results mentioned above proved that the addition of strain L1, GMB, and GMB-L1 all promoted the transformation of soil heavy metals from acid-soluble and reducible fractions to oxidizable and residual fractions, and GMB-L1 exhibited a higher stabilization efficiency. The enhanced passivation of heavy metals with GMB-L1 may be ascribed to the involvement of bacteria in the biosorption and degradation of heavy metals through various metabolic activities [49]. Moreover, GMB can provide strain L1 and indigenous microorganisms with a safe habitat for their growth [50].

3.4.2. TCLP-Extracted Heavy Metals and Ecological Risk Assessment

The TCLP method was usually adopted to estimate the leaching toxicity and mobility of heavy metals in soil [51]. As shown in Figure 5b, the concentrations of TCLP-extracted Pb and Cd in soil supplemented with strain L1, GMB, and GMB-L1 were lower than those in untreated soil. The TCLP-extracted Pb and Cd concentrations in CK were 69.54 and 4.76 mg/kg, respectively. In comparison with the CK treatment, the concentration of TCLP-extracted Pb with strain L1, GMB, and GMB-L1 treatments was decreased by 14.21%, 45.66%, and 73.24%, respectively. Similarly, the addition of strain L1, GMB, and GMB-L1 reduced the TCLP-extracted Cd concentration by 13.86%, 32.77%, and 57.25%, respectively. In conclusion, the leaching toxicity of heavy metals in soil supplemented with GMB-L1 was much lower than that of the other treatments, which implied the superior performance of GMB-L1 in decreasing the toxicity and mobility of heavy metals in soil.
The ecological risk was assessed according to the Nemerrow index and potential ecological risk index methods, with the calculation results illustrated in Table S5. The Pi values of Pb and Cd in CK treatment were, respectively, 13.91 and 9.52, suggesting that the tested soil was heavily contaminated. After the addition of strain L1, GMB, and GMB-L1, the Pi values of Pb and Cd declined obviously, especially under the treatment of GMB-L1. The PI values reflected the comprehensive pollution status of two heavy metals, which reduced from 11.92 (CK) to 10.24 (T1), 7.00 (T2), and 3.90 (T3), respectively. With the addition of strain L1, GMB, and GMB-L1, the Ei values of Pb decreased from 69.54 to 59.66, 37.79, and 18.61, and the Ei values of Cd decreased from 385.60 to 246.00, 192.00, and 122.10, respectively. Furthermore, the EI values fell from 1355.14 (CK) to 305.66 (T1), 229.79 (T2), and 140.71 (T3), respectively. These results suggested that the three treatments of T1, T2, and T3, particularly the incorporation of GMB-L1, were beneficial in reducing the ecological risk of heavy metals in soil.

3.5. Remediation Mechanism

As found in Figure 6a, GMB featured a distinct porous structure and rough surface, which offered an adequate foundation for the immobilization of free bacteria. Figure 6b depicts that the surface of GMB-L1 was covered by strain L1, which explicitly indicated the successful immobilization of strain L1 onto GMB.
Numerous vibrational peaks could be found in the FTIR spectrum of strain L1 (Figure 7a). The prominent and wide peak at 3306 cm−1 was assigned to the -OH stretching vibration, while the peaks at around 2928 and 2859 cm−1 can be attributed to the asymmetric and symmetric C-H stretching of the -CH2 and -CH3 groups [52]. The characteristic peaks between 1730 and 1400 cm−1 were mainly generated by the stretching vibration of aromatic C=O/C=C [28]. The peak at 1238 cm−1 corresponded to the C-O stretching vibration and the peak at 1061 cm−1 represented the C=O-C stretching vibration [53,54]. The characteristic peak at approximately 960 cm−1 in GMB was indicative of the Fe-O/Fe-OH bending vibration in α-FeOOH, and this peak of GMB-L1 obviously shifted [55,56]. Moreover, the characteristic peak of PO43- (590 cm−1) can be obviously observed in GMB-L1 [57]. The functional groups of GMB-L1 after adsorption are presented in Figure 7b. The characteristic peaks associated with -OH and C-O decreased or disappeared after the reaction, which might be due to the interaction between heavy metals and -OH/C-O. In addition, the weakening of the Fe-OH peak after adsorption confirmed its important role in the removal of heavy metal ions.
The XRD patterns of GMB and GMB-L1 are presented in Figure 7c, with the distinct broad peak at approximately 25° corresponding to the (002) crystal plane of graphite carbon [58]. The peaks of GMB that appeared at 21.22°, 33.24°, 36.65°, 53.24°, and 68.40° were associated with the (110), (130), (111), (221), and (301) crystal planes of goethite (PDF#29-0713), respectively, suggesting that α-FeOOH has been successfully attached to biochar. After immobilization of strain L1, these peaks witnessed an obvious reduction or even disappeared. Additionally, there was the apparent generation of heavy metals on GMB-L1 after the reaction with heavy metals (Figure 7d). The peaks of Pb(OH)2 and Cd(OH)2 at 13.42° and 47.35° were recognized on GBCL1 after the reaction. In addition, the peaks located at 23.96°, 30.12°, 39.84°, and 67.32° were related to the Pb-phosphate [Pb5(PO4)3OH, Pb3(PO4)2] minerals, and the peaks at 30.22° and 33.15° corresponded to Cd5(PO4)3OH. This phenomenon might be attributed to strain L1, with phosphate-solubilizing ability, participating in the heavy metal biomineralization through the secretion of organic acids or phosphates [59].
The XPS wide-scan survey of GMB-L1 before and after the reaction proved that the biochemical composite can effectively adsorb lead and cadmium (Figure S4). As shown in Figure 8a, there were three characteristic peaks located at 284.80, 286.24, and 288.70 eV that could be recognized in the C 1s spectrum, which were assigned to C-C, C-O, and C-C=O, respectively [60]. The proportion of C-O decreased from 29.17% to 26.27% and 27.55% following lead and cadmium removal, respectively, indicating the possible participation of C-O in the reaction. The O 1s spectrum (Figure 8b) could be separated into three peaks situated near 530.50, 531.80, and 533.32 eV, which were ascribed to metal oxide (M-O), hydroxyl bonded to metal (M-OH), and hydroxyl on the GBCL1 surface (C-OH), respectively [61]. It was obviously noticed that the area ratio of C-OH dropped from 66.80% to 40.61% and 36.99% after the reaction of Pb and Cd, respectively, which suggested that these heavy metal ions complexed with C-OH. In addition, the area ratio of M-OH and M-O rose upon reaction with Pb (21.59% to 40.83% for M-OH, 11.62% to 18.56 for M-O) and Cd (21.59% to 47.54 for M-OH, 11.62% to 15.47% for M-O), which might be contributed by the presence of Pb-OH/Pb-O and Cd-OH/Cd-O on the surface of GMB-L1 [62]. The XPS of Fe 2p in Figure 8c presented main peaks at binding energies of 710.88 eV for Fe 2p3/2 and 724.34 eV for Fe 2p1/2 with two satellite peaks, which is consistent with the characterization of Fe(III) in α-FeOOH [62]. After adsorption, a slight shift in binding energy to a higher position in Fe 2p could be clearly observed, which could be elucidated by the generation of Fe-O-Pb/Fe-O-Cd complexes [63]. In the Pb 4f spectrum (Figure 8d), two peaks located at 138.21 and 143.11 eV can be assigned to Pb(OH)2 and Pb-O, respectively [64]. And the characteristic peaks at 138.92 and 143.82 eV belong to Pb-phosphate substances in different orbits [10]. From the Cd 3d spectrum presented in Figure 8e, the characteristic peaks of 405.78 and 412.52 eV, respectively, correspond to Cd(OH)2 and Cd-Fe hydroxide [65], while the peaks at 406.99 and 412.52 eV represented Cd-phosphate [66].
In light of the aforementioned findings, the remediation mechanism of heavy metals can be concluded in Figure 9. The GMB not only possesses a well-developed porous structure that allowed microbes to colonize, but can also greatly reduce the physiological toxicity of heavy metals and safeguard the conditions in which microorganisms flourish through functions of chemical precipitation, complexation, and electrostatic attraction. Under the conditions of GMB protection, the positive feedback of strain L1 can be activated through the process of releasing organic acids to solubilized inorganic phosphates on biochar [67]. Subsequently, Pb2+ and Cd2+ combined with phosphate to form stable minerals (Pb/Cd-phosphate), which optimized the remediation of heavy metal ions.

4. Conclusions

In this study, a biochemical composite called GMB-L1 was successfully prepared and employed for heavy metal passivation. The adsorption experiments demonstrated that the heavy metal removal performance of GMB-L1 outperformed the individual components, with the maximal adsorption capacities of 496.54 mg/g for Pb and 178.18 mg/g for Cd. During the remediation of heavy-metal-contaminated soil by GMB-L1, the conversion of Pb and Cd in contaminated soil from unstable to stable fractions was found, with the corresponding decline in the TCLP-extracted content and ecological risks associated with heavy metals. Moreover, various characterizations substantiated that GMB not only adsorbs heavy metals by chemical precipitation, complexation, and electrostatic attraction, but also creates an appropriate shelter for strain L1 to promote its biomineralization. In all, this study proposed an effective and environmentally friendly method for Cd and Pb remediation.

Supplementary Materials

The following supporting information can be downloaded at https://www.mdpi.com/article/10.3390/w16131917/s1. Text S1 Screening process of strain L1. Text S2 Adsorption kinetics and isothermal adsorption models. Table S1. The physicochemical properties of experimental soil. Table S2. Relationship between Nemerrow, potential ecological risk index and classification of heavy metal pollution. Table S3. Kinetic fitting parameters of Pb(II) and Cd(II) adsorption onto GMB-L1. Table S4. Isotherm fitting parameters of Pb(II) and Cd(II) adsorption onto GMB-L1. Table S5. Nemerrow and Potential ecological risk index of heavy metals in soil. Figure S1. The phylogenetic tree of strain L1. Figure S2. The removal efficiency of free L1, GMB and GMB-L1 for (a) Pb2+ and (b) Cd2+ (experiment conditions: Pb2+ = 400 mg/L, Cd2+ = 200 mg/L, dosage = 1.0 g/L, T = 30 °C and initial pH= 6). Figure S3. Zeta potential of GMB-L1 (a), and OD600 of strain L1 (b) at different pH values. Figure S4. XPS survey spectra of GMB-L1 before and after reacting with Pb(II)/Cd(II). Refs. [68,69] are cited in Supplementary Materials.

Author Contributions

G.F.: Investigation, Data Analysis, Validation, Writing—Original Draft Preparation, Writing—Review and Editing, Funding Acquisition. J.Z. (Junhou Zhou): Investigation, Data Analysis, Validation, Writing—Original Draft Preparation. X.C.: Data Analysis, Validation, Writing—Review and Editing. W.Y.: Investigation, Data Analysis, Validation. C.L.: Investigation, Data Analysis, Validation. J.L., J.Z. (Jianyong Zou), K.-Q.X. and Q.L.: Validation, Writing—Review and Editing. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Natural Science Foundation of Fujian Province in China (Nos. 2021N0022, 2021Y3002, 2023J02006), the Science and Technology Project of Fuzhou City (No. 2022-P-013), and the Science and Technology Project of Anhui Province (Nos. 2021-YF26, 2021-YF27).

Data Availability Statement

Data are contained within the article and Supplementary Materials.

Conflicts of Interest

Author Jing Luo was employed by the company Fujian Jinhuang Environmental Sci-Tech Co., Ltd. Author Jianyong Zou was employed by the company Anhui Urban Construction Design Institute Co., Ltd. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Figure 1. The phosphate-solubilizing ability of strain L1 under different concentrations of (a) Pb2+ and (b) Cd2+, and the growth curve of strain L1 under different concentrations of (c) Pb2+ and (d) Cd2+.
Figure 1. The phosphate-solubilizing ability of strain L1 under different concentrations of (a) Pb2+ and (b) Cd2+, and the growth curve of strain L1 under different concentrations of (c) Pb2+ and (d) Cd2+.
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Figure 2. Effect of (a) temperature, (b) bacterial volume, (c) SA content, and (d) cross-linking time on bacterial count and heavy metal adsorption by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
Figure 2. Effect of (a) temperature, (b) bacterial volume, (c) SA content, and (d) cross-linking time on bacterial count and heavy metal adsorption by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
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Figure 3. Effect of (a) temperature, (b) initial pH, (c) dosage, and (d) initial heavy metal concentration on Pb2+ and Cd2+ adsorption by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
Figure 3. Effect of (a) temperature, (b) initial pH, (c) dosage, and (d) initial heavy metal concentration on Pb2+ and Cd2+ adsorption by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
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Figure 4. Adsorption kinetics (a) and adsorption isotherm model (b) for Pb2+ and Cd2+ by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
Figure 4. Adsorption kinetics (a) and adsorption isotherm model (b) for Pb2+ and Cd2+ by GMB-L1 (experimental conditions: Pb2+ = 200 mg/L, Cd2+ = 100 mg/L, dosage = 0.5 g/L, T = 30 °C, and initial pH = 6).
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Figure 5. The fractions (a) and TCLP-extracted concentrations (b) of Pb and Cd in soil under different treatments.
Figure 5. The fractions (a) and TCLP-extracted concentrations (b) of Pb and Cd in soil under different treatments.
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Figure 6. SEM images of (a) GMB and (b) GMB-L1.
Figure 6. SEM images of (a) GMB and (b) GMB-L1.
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Figure 7. FTIR spectra of (a) strain L1, GMB, and GMB-L1, and (b) GMBL1 before and after adsorption; XRD patterns of (c) GMB and GMB-L1, and (d) GMBL1 before and after adsorption.
Figure 7. FTIR spectra of (a) strain L1, GMB, and GMB-L1, and (b) GMBL1 before and after adsorption; XRD patterns of (c) GMB and GMB-L1, and (d) GMBL1 before and after adsorption.
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Figure 8. XPS spectra of GMB-L1 before and after reacting with Pb(II)/Cd(II), including (a) C 1s, (b) O 1s, (c) Fe 2p, (d) Pb 4f, and (e) Cd 3d.
Figure 8. XPS spectra of GMB-L1 before and after reacting with Pb(II)/Cd(II), including (a) C 1s, (b) O 1s, (c) Fe 2p, (d) Pb 4f, and (e) Cd 3d.
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Figure 9. The mechanism analysis of GMB-L1 to remediate heavy metals.
Figure 9. The mechanism analysis of GMB-L1 to remediate heavy metals.
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MDPI and ACS Style

Fan, G.; Zhou, J.; Cao, X.; You, W.; Lin, C.; Luo, J.; Zou, J.; Xu, K.-Q.; Luo, Q. Enhanced Remediation of Lead and Cadmium by the Co-System of Phosphate-Solubilizing Bacteria Immobilized on Goethite-Modified Biochar. Water 2024, 16, 1917. https://doi.org/10.3390/w16131917

AMA Style

Fan G, Zhou J, Cao X, You W, Lin C, Luo J, Zou J, Xu K-Q, Luo Q. Enhanced Remediation of Lead and Cadmium by the Co-System of Phosphate-Solubilizing Bacteria Immobilized on Goethite-Modified Biochar. Water. 2024; 16(13):1917. https://doi.org/10.3390/w16131917

Chicago/Turabian Style

Fan, Gongduan, Junhou Zhou, Xingfeng Cao, Wu You, Chen Lin, Jing Luo, Jianyong Zou, Kai-Qin Xu, and Quanda Luo. 2024. "Enhanced Remediation of Lead and Cadmium by the Co-System of Phosphate-Solubilizing Bacteria Immobilized on Goethite-Modified Biochar" Water 16, no. 13: 1917. https://doi.org/10.3390/w16131917

APA Style

Fan, G., Zhou, J., Cao, X., You, W., Lin, C., Luo, J., Zou, J., Xu, K. -Q., & Luo, Q. (2024). Enhanced Remediation of Lead and Cadmium by the Co-System of Phosphate-Solubilizing Bacteria Immobilized on Goethite-Modified Biochar. Water, 16(13), 1917. https://doi.org/10.3390/w16131917

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