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Article

Comparing Different Strategies for Cr(VI) Bioremediation: Bioaugmentation, Biostimulation, and Bioenhancement

by
Xiao Yan
1,2,3,4,5,†,
Zhenghao Yan
6,†,
Xuezhe Zhu
1,2,3,4,7,
Yupin Zhou
1,2,3,4,
Guoying Ma
1,2,3,4,
Shuangquan Li
1,2,3,4,
Xingyu Liu
1,2,3,4 and
Mingjiang Zhang
1,2,3,4,*
1
National Engineering Research Center for Environment-Friendly Metallurgy in Producing Premium Non-Ferrous Metals, Beijing 100088, China
2
GRINM Resources and Environment Tech. Co., Ltd., Beijing 100088, China
3
General Research Institute for Nonferrous Metals, Beijing 100088, China
4
Beijing Engineering Research Center of Strategic Nonferrous Metals Green Manufacturing Technology, Beijing 100088, China
5
GRIMAT Engineering Institute Co., Ltd., Beijing 101407, China
6
School of Materials Science and Engineering, Henan University of Science and Technology, Luoyang 471000, China
7
School of Metallurgy, Northeastern University, Shenyang 110819, China
*
Author to whom correspondence should be addressed.
†
These authors have contributed equally to this work and share first authorship.
Sustainability 2023, 15(16), 12522; https://doi.org/10.3390/su151612522
Submission received: 12 June 2023 / Revised: 25 July 2023 / Accepted: 1 August 2023 / Published: 17 August 2023
Browse Figures
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Abstract

:
Unchecked releases of industrial waste, including chromium smelting slag (CSS), have resulted in disastrous effects on the environment for human use. Considering the problems of environment, efficiency, and sustainability, the present research was designed to evaluate the potential feasibility of Cr(VI) bioremediation by different strategies of natural attenuation (NA), bioaugmentation (BA), biostimulation (BS), and bioenhancement (BE). Results showed the BE was the best strategy for Cr(VI) removal and reached 86.2% in 84 days, followed by the BA, BS, and NA. The variation of Eh values indicated all systems translated the oxidation state into reduction continuously except for NA and BS during the bioremediation process. After bioremediation, the Tessier sequential extraction analyzed in the BE showed stable chromium levels up to 97%, followed by BA (89~93%), BS (75~78%), and NA (68%), respectively. Moreover, High-throughput sequencing was also used to assist in revealing the differences in microbial community structure between the different strategies. Stenotrophomonas, Ochrobactrum, and Azomonas, as the bioremediation microbes, were enriched in the BE in comparison with the others. This provided a new enhancement strategy for bioremediation microbes colonized in a new environment to achieve sustainable removal of Cr(VI).

Graphical Abstract

1. Introduction

Owing to improper disposal of chromium smelting slag (CSS), chromium (Cr) pollution has caused a series of environmental issues [1]. Especially Cr(VI) is highly toxic because of its easy migration, high bioavailability, mutagenicity, and carcinogenicity effects [2]. In a way, Cr(VI) accumulation entered and moved along the food chain from the lower ecosystem to the higher ecosystem by means of biomagnification, leaving its impact throughout [3,4]. Previous studies have found that Cr(VI) is more toxic (10–100 times) than Cr(III) [5,6]. Therefore, an effective, eco-friendly, and sustainable strategy for remediation of Cr(VI) has been considered an urgent task.
Traditional methods for removal/detoxification of Cr(VI), including the tillage method and soil washing of physical engineering, chemical stabilization through the addition of a metal passivator, and membrane filtration, were more efficient and faster [7,8]. However, the disadvantages of traditional methods, such as high cost, poor sustainability, and easy secondary pollution, were outstanding [9,10]. Bioremediation as a promising technology with eco-friendly, efficient, and lasting stability has attracted more attention [5,11]. A study reported that biostimulation (BS) and bioaugmentation (BA) are the classical strategies of bioremediation [12].
The BS is a strategy to promote indigenous microbial growth through the addition of sufficient nutrients (carbon and nitrogen sources) [13,14]. The amounts of indigenous microorganisms can enhance Cr(VI) removal efficiency by biosorption or bioreduction [10]. However, indigenous microbial species have a significant effect on Cr(VI) removal efficiency; several microbes can reduce Cr(VI) to Cr(III); in contrast, some microbes are able to dissolve chromium compounds to Cr(VI) by some acidic substances during their growth and metabolism [6,15]. Ultimately, the content of soluble Cr(VI) in the bioremediation system increased [16]. Thus, the BS strategy alone for remediating chromium contamination made it difficult to achieve the desired objectives.
When indigenous microbes have a relative low capacity to remove Cr(VI), the introduction of functional microbes was considered another method for remediating chromium contamination, named the BA strategy [17]. For instance, sulfate-reducing bacteria (SRB) are widely used for bioremediation of heavy metals in contaminated soil or defunct mine tailings [18,19]. Liu et al. performed bioremediation by SRB to treat acid mine drainage and reported that the removal efficiency of copper and iron was 60.95% and 97.83% within 3 days, respectively [20]. In addition, Yan et al. [18] found that Cr(VI) was reduced by SRB to Cr(III) up to 51% within 73 h and highlighted the potential application of SRB to promote bioremediation of chromium-contaminated sites [18]. However, the adaptability and competitiveness of these introduced microorganisms in comparison with the native microbial communities are highly dependent on the environmental conditions (pH, temperature, nutrition, initial Cr(VI) concentration) [21,22]. In addition, the introduction of exogenous microorganisms may increase the risk of ecological problems [23]. To overcome these limitations, indigenous microbial species after domestication, which have the function of removing Cr(VI), have been considered as an alternative method [17,24]. Numerous indigenous microorganisms have been isolated and described as possessing high Cr(VI) removal potential, such as Geobacter sulfurreducens, SRB, Bacillus sp., Stenotrophomonas maltophilia, and Pseudomonas gessardii [10,18,25,26,27]. Particularly, the mixed native bacterial consortium (Pseudomonas stutzeri L1 and Acinobacter baumanni L2) was reported to be able to reduce Cr(VI) to over 95% at 37 °C in a Cr(VI) of 1000 mg/L contaminated environment [28].
However, the adverse microbial community structure and poor nutrition may greatly limit the universal application of the single remediation strategy of BS or BA at the chromium-contaminated sites. Unfortunately, studies on the application of the integration of BS and BA in the chromium sites, named the bioenhancement strategy (BE), are rather insufficient, let alone the comparison and evaluation of different bioremediation strategies for Cr(VI) removal effectiveness from the aspects of long-term stability and ecological safety. In this study, different treatments of the natural attenuation (NA), BS, BA, and BE were conducted for bioremediation Cr(VI) in the CSS. The specific objectives are as follows: (i) compare the removal effects of Cr(VI) by different bioremediation strategies; (ii) determine the perdurable stability of chromium in different systems after bioremediation; and (iii) evaluate the feasibility of different bioremediation strategies.

2. Materials and Methods

2.1. The CSS Samples

The CSS samples used were obtained from a chromate factory (latitude 36°54′ N, longitude 101°02′ E) in Qinghai province, China. The samples were collected at a depth of 20 cm beneath the surface, then sealed in polyethylene bags and stored in the refrigerator at 4 °C for physico-chemical property analysis and bioremediation experiments immediately. The original CSS samples have a higher pH, a redox potential (Eh) value of 9.25, 381 mV, and moisture content of 29.1 ± 1.7%. The main chemical components were analyzed by X-ray Fluorescence Spectrometer (XRF), and results are presented in Table 1. The Cr2O3 level of CSS samples was 8.96%.

2.2. Microorganisms and Medium

In our previous study, the Stenotrophomonas, Ochrobactrum, Pseudomonas, and Bacillus isolated from the CSS were reported to remove Cr(VI) [29]. In this study, Stenotrophomonas, Ochrobactrum, Pseudomonas, and Bacillus, as native microbes with a high capacity of removing Cr(VI), were inoculated in the bioremediation systems. Before inoculation, native microbes were initially enriched in the Luria-Bertani medium (LB) in a shaker at 30 °C with 140 rpm until the exponential phase. The SRB, another Cr(VI)-removing strain, was provided by the National Engineering Laboratory of Biohydrometallurgy. The compositions of all mediums were listed in the Supporting Information Tables S1–S4. Finally, the optical density of cells (OD600) was monitored during the experiment process.

2.3. Bioremediation Experiments

In order to study the effect of different strategies for Cr(VI) bioremediation on the CSS, six experimental groups of the NA, BS, BA, and BE were carried out; the detailed information is shown in Table 2. All treatments were run in triplicate at 30 °C for 84 days. The Cr(VI) concentration, pH, Eh, and OD600 in the leaching solutions were measured at a regular interval of 7 days. After bioremediation, the chemical fractions of chromium, the leaching toxicity, and the microbial community structure under different treatments were evaluated.

2.4. Analytical Methods

2.4.1. Detection of Physicochemical Properties

The mineral phase composition of the CSS sample was detected by X-ray diffraction (XRD) (Rigaku-Smart Lab, Tokyo, Japan). The pH and redox potential (Eh value) in the leaching solutions were quantified by a pH meter (S470, Seven Excellence, Zurich, Switzerland) with a glass electrode and an oxidation-reduction potential meter (PC-320, Shanghai Honff, Shanghai, China) with a glass electrode. The OD600 was determined at the wavelength γ = 600 nm by UV/Vis Spectrophotometer (TU-1810 and Puxitech, Shanghai, China).

2.4.2. Total Chromium, Cr(VI), and Cr(III) Determination

To remove bacterial cells and metabolites, the extraction suspension was centrifuged and filtered by membranes (0.22 μm). The reaction mixture contained 1 mL of sample, 0.5 mL of 50% H2SO4 (v/v), 0.5 mL of 50% H3PO4 (v/v), 2 mL of diphenylcarbazide (5 mg/mL), and distilled water to complete a final volume of 50 mL. The Cr(VI) was measured at the wavelength γ = 540 nm using a UV/Vis Spectrophotometer (TU-1810 and Puxitech, Hangzhou, China). Additionally, an Atomic Absorption Spectrometer instrument (Thermo Electron Co., Manchester, UK) was used to determine the total chromium [30]. The Cr(III) was determined by measuring the reoxidation of reduced Cr(VI) after adding an acid mixture (sulfuric acid, phosphoric acid), 1% AgNO3, and ammonium persulfate [29]. After heating to boiling for 3 min, phenylo-aminobenzoic acid was used as an indicator, and the reaction system was titrated with 0.1 mol/L standard ferrous ammonium sulfate. The concentration of Cr(III) was calculated as follows:
C C r ( III ) m g / L = C i × V × 1.73 × 10 4
where C C r ( III ) is the concentration of Cr(III) at a certain time,   C i is molarity of the standard solution of ammonium ferrous sulfate, and the V is the consumption of ammonium ferrous sulfate.

2.4.3. The Chemical Fraction of Chromium Determination by Sequential Extraction

The chromium chemical fractions were divided into six parts by modifying the sequential extraction of Tessier [31], including water-solution fraction (F1), exchangeable fraction (F2), carbonate fraction (F3), Fe-Mn oxides fraction (F4), organic fraction (F5), and residual fraction (F6). A total of 2 ± 0.0001 g of sample from different systems were weighed out after drying and stored in a 45 mL polycarbonate centrifuge tube. The sequential extraction steps of different chromium fractions are shown in Table 3.
After finishing every successive extraction, the samples were centrifuged at 8800 rpm for 15 min, the supernatant was filtered by a 0.22 μm filter membrane for analysis Cr(VI) concentration, the residual precipitation was prepared for the next extraction after washing three times with deionized water.

2.5. The Leaching Toxicity Experiments after Bioremediation

After bioremediation, leaching toxicity was investigated to assess the lasting stability of the CSS. According to the standard of solid waste-extraction procedure for leaching toxicity (HJ/T299-2007), 100 g of the CSS samples were dried and loaded into 2.5 L anti-reverse bottles with a solid to solution ratio of 1:10 (w:v). The solution was composed of 0.2 H2SO4, 0.1 HNO3, and then diluted with deionized water to 1000 mL with adjusting pH to 3.2 ± 0.25 by 1 M NH3·H2O. The samples were fixed on a reciprocating oscillator at 30 ± 2 rpm for 18 ± 2 h. After leaching, the supernatant was filtered for analysis Cr(VI) concentration.

2.6. Analysis of Microbial Community Structure

After bioremediation, the samples from different treatments were analyzed by an Illumina HiSeq-250. DNA of samples was extracted using a DNA Extraction Kit (OMEGA, M5635-02, Shanghai, China). All DNA concentrations and quality were confirmed by using a NanoDrop 2000 spectrophotometer (Thermo Scientific, Waltham, MA, USA). Totally, all extracted genomic DNA samples were sent to high-throughput sequencing (Sangon Biotech Co., Ltd., Shanghai, China). For bacterial community structure analysis, the primers of 16s rRNA gene [27F(50-agagtttgatcctggctcag-30) and 1492R (50-ggttaccttgttacgactt-30)] were selected to amplify the V3–V4 region as reported by Mala et al. [32].

2.7. Statistical Analysis

All statistical analysis was displayed using the Origin pro 8.5.1 software. The pH, Eh, removal efficiency of Cr(VI), and leaching toxicity are presented as the mean ± standard deviation (SD).

3. Results

3.1. The CSS Samples Characterization

The main mineral compositions of the CSS samples are shown in Figure 1. The results showed that the directions of XRD patterns match well with those of quartz (SiO2), periclase (Mg(OH)2), and some insufficient crystal structures (Ca2O(Cr, Fe3+)2O3). This was consistent with the analysis of XRF (Table 1), which indicated that samples had higher magnesium oxide, calcium salt, iron sulfide, and chromium content. Therefore, it was relatively reasonable that the main crystal structures were determined as Ca2O(Cr, Fe3+)2O3, Mg(OH)2, and SiO2 in the process of long-term accumulation.

3.2. Comparison of the Variation of pH and Eh in Different Bioremediation Systems

Figure 2 showed that a downward tendency in pH was present in all systems. The pH decreased in the BA1, BA2, and BE from 9.5, 9.20, and 9.25 to 8.01, 7.43, and 8.21, respectively. This result may be due to the fact that microbes in the CSS secrete some acidic substances during their growth and metabolism [5]. The pH in the NA and BS were relatively stable. A slightly decreasing trend may be attributed to the increase of acid-producing oxidizing microbes in systems with sufficient nutrients [33]. Above all, the results indicated that the original alkaline environment was slowly weakening in all bioremediation systems.
A significant difference in Eh values in different systems is shown in Figure 3. A significantly decreased BA1, BA2, and BE in comparison with the NA and BS. The Eh in the BA1, BA2, and BE decreased from 386 mV to below zero in the first 21 days and then reached the minimum values of −292, −284, and −358 mV in 42 days, respectively. The variation of Eh indicated that the system changed from oxidation to reduction conditions. Studies have shown a certain correlation between bioremediation and Eh; the bioremediation efficiency of heavy metals in the culture system increased with the decrease of Eh [15]. Finally, it was interesting that a slight increase in Eh occurred at the end of BA. This result explained that nutrients consumed over time lead to the mass mortality of bioremediation microbes. One difference was that the Eh value of the BE group was lower than that of the BA group, implying that the functional microbes in the BE group rapidly reconstructed the relative stronger reductive condition. However, for the NA and BS groups, the systems Eh was about 400 mV, belonging to the strong oxidation environment, possibly because of the existence of endogenous acidophilic microorganisms, which helped dissolve Cr(VI) from the CSS samples and increased the redox potential of the systems, leading to worse results [33].

3.3. Comparison of the Removal of Cr(VI) in Two Bioremediation Systems

The bacterial growth and Cr(VI) removal in different systems are shown in Figure 4. The removal rates of Cr(VI) were higher in the BA and BE than in the NA and BS. The removal rates of Cr(VI) in the BA1, BA2, and BE were 17.3%, 10.1%, and 18.1% in the first 14 days, and then reached the maximum values of 62.2%, 44.5%, and 86.2% at 84 days. Studies reported that the introduced functional microorganisms and the enriched nutrients played a greater role in the bioremediation of heavy metals [14,19]. Therefore, the removal rates of Cr(VI) were much better in the BE in comparison with the BA. Especially, The removal rates of Cr(VI) in the BE were over 24% compared with the BA1. For the NA and BS, a slight increase in removal rates may be due to the biosorption of indigenous microorganisms. In addition, the OD600 values increased obviously with time in all systems except for the NA (Figure 4b). Suggesting that the microbial population can be improved by the addition of nutrients or the inoculation of microbes. Just like the BE, the OD600 values were up to 0.82 in 84 days. In the BA, the inoculation of numerous native microbes and SRB resulted in an increase in OD600 values to over 0.61. The variation of OD600 values in the BS was similar to that in the BA, with the difference that a decreasing trend occurred in the late stage of the BS. This finding may be due to the fact that nutrient species influence bacterial growth. In addition, competitive relationships between microbes also lead to several microbial decays. The above results of bacterial growth and Cr(VI) removal indicated that BE was the best strategy in comparison with the others.

3.4. The Distribution of Chromium Fractions in Different Bioremediation Periods

The percentage of different chemical forms of chromium was sequentially extracted by Tessiers’ method [31]. Due to its high mobility, easy bioavailability, and strong toxicity, the F1 was widely concerned during the experiment process [26,34]. The F2 and F3 with bioavailability were accepted as the less stable states, and the F4, F5, and F6 were considered the more stable states [35].
To evaluate the perdurable stability of the CSS after different strategies, the variation of chromium chemical fractions was shown in Figure 5. First, the percentages of F1, F2, F3, F4, F5, and F6 in the original CSS samples were about 1%, 14%, 16%, 6%, 4%, and 58%, respectively. According to an analysis of the chemical stability of chromium, about 32% was measured as unstable or less stable. After 84 days of bioremediation by different strategies, it can be found that the proportion of F1, F2, and F3 decreased with the increase of F4 and F5. This result suggested that the chemical forms of chromium transformed into stable states after bioremediation. After bioremediation, the F1 was hardly detected in the BA and BE treatments in comparison with the NA and BS. The chromium content of the F2 and F3 in the BS1, BS2, BA1, BA2, and BE decreased to 10%, 12%, 3%, 5%, 1%, 11%, and 12%, 4%, 5%, and 2%, respectively. Therefore, the sum of unstable chromium in different systems was only 22%, 25%, 7%, 11%, and 2%, respectively. Due to the addition of nutrients and microbes, the percentage of stable fractions increased significantly; F4 occupied 17%, 15%, 26%, 20%, 26%, and F5 about 12%, 8%, 21%, 17%, and 23%, respectively. The total stable chromium in bioremediation increased to 78%, 75%, 93%, 89%, and 97%, respectively. The above result implied that bioremediation is beneficial to promote the transition of chromium from a relatively unstable chemical form to a stable form. Further analysis indicated that the proportion of stable chromium in BE was highest, followed by BA and BS.

3.5. The Leaching Toxicity in the CSS after Bioremediation

The leaching toxicity of the CSS in the different strategies was measured at reaction times of 84 days to investigate the stabilization of chromium after bioremediation. Figure 6 shows that Cr(VI) concentration in leaching solution varied greatly in conjunction with the difference in treatment methods. Cr(VI) concentration was relative lower in the BA and BE than the NA and BS. Especially, the Cr(VI) concentration in the BE was lowest at 3.6 mg/L, which was far below the maximum allowable emission standard of 5 mg/L. Cr(VI) concentrations in the BA were 12.99 and 21.16 mg/L, respectively. The 12.99 and 21.16 mg/L of Cr(VI) concentration in the BA were followed by the BS (79.77 mg/L and 82.57 mg/L), and the Cr(VI) concentration was relatively low in the BA in comparison with the BS. But the above result indicated that it was difficult to detoxify chromium slag with BS or BA alone. This finding was consistent with the variation of chemical forms. After bioremediation, the chromium chemical form was the most stable in the BE, followed by the BA and BS.

3.6. The Variation of Microbial Community Structure during Bioremediation Process

The high-throughput sequencing showed that the relative abundance presented a significant difference between the different strategies (Figure 7). The Acinetobacter, Bacillus, Pseudomonas, and Azomonas were the dominant genera in the CSS, occupying 11.81%, 10.00%, 15.09%, and 11.74%, respectively, with the others reaching 51.15%. After bioremediation, the addition of nutrients (BS) changed the Actinobacteria level slightly. While the prevalence of Pseudomonas, Bacillus, Pseudomonas, Geobacillus, and Azomonas decreased greatly. Moreover, the addition of native microbes and SRB in the BA and BE changed the structure of the microbial community in the CSS. Obviously, the relative abundance of Bacillus (26.9%) increased significantly in the BE treatments and was much higher than that in BA1 (22.9%) and BA2 (4.13%). Functional microbes, including Stenotrophomonas, Ochrobactrum, and Azomonas, were enriched in the BE treatment. It increased to 10.99, 6.38, and 12.03%, respectively, whereas the relative abundances of Acidiphilium and Hydrogenophaga were significantly lower in the BE (0.01% and 0.21%, respectively) than those in the BA (4.99%, 2.45%, and 10.99%, respectively). As for rare bacteria, the BE stimulated the growth of Escherichia and Desulfovibrio, resulting in their higher relative abundances (1.66% and 0.54%) compared with the control (0.95% and 0.02%). No special microbial species were detected in the BE treatment compared with the NA. In addition, from the perspective of the OD600 in different strategies, the BE was highest at 0.82, followed by BA1 (0.68), BA2 (0.50), BS1 (0.39), BS2 (0.22), and NA (0.11). It can be found that, consistent with the bioremediation efficiency, the removal rate of Cr(VI) increased with the increase of the OD600 in the different systems.

4. Discussion

4.1. Comparison of Bioremediation Effects from Different Strategies

The bioremediation effects of different strategies varied greatly in this study. We found that the removal rate of Cr(VI) was as follows: BE (86.2%) > BA1 (62.2%) > BA2 (44.6%) > BS1 (6.4%) > BS2 (5.9%) > NA treatment (4.0%) This finding was consistent with the variation of OD600 values in different systems (Figure 4). Studies have reported a certain correlation between the removal rate of Cr(VI) and OD600 values and concluded that the removal rate of Cr(VI) in the system increased with the increase of OD600 values in the same system. Due to the addition of functional microbes in the BA and BE, the OD600 values increased rapidly, which then promoted the removal of Cr(VI). For the BS, the addition of nutrients can stimulate the growth of indigenous microbes and then improve the removal rate of Cr(VI). However, obviously, a slightly improved Cr(VI) removal efficiency with time was detected in the BS. One explanation was that the addition of nutrients to generate energy allowed all indigenous microbial species to reproduce in a relatively open environment [36,37]. Several microbial species without the capacity to remove Cr(VI) not only competed for nutrients but also produced acids, causing the re-dissolution of Cr(VI) [37,38]. This is one of the key reasons why it was too hard to enhance the bioremediation effect through the addition of nutrients alone in the BS. In addition, due to the poor competitiveness of the BS, the lower abundance of functional microbes leads to a lower removal rate of Cr(VI) during the process of bioremediation [39,40]. Compared with the NA, the removal rate of Cr(VI) in the BS was relatively high. From the perspective of OD600 values, this may be attributed to the biosorption of numerous microbes by the stimulation of nutrients [41].
It was widely reported that the inoculation of functional microbes can improve the removal/degradation rates of heavy metals and organic pollutants [23,42]. Mahbub et al. concluded that 60% of Hg can be removed by augmentation of Sphingobium SA2 for 7 days in a field with 280 mg/kg Hg contamination [43]. Another study has reported that the inoculation of Acinetobacter tandoii LJ-5 increased the degradation to 85% in wastewater contaminated with polycyclic aromatic hydrocarbons [17]. Our studies indicated that the inoculation of functional microbes (Stenotrophomonas, Ochrobactrum, Pseudomonas, Bacillus, and SRB) in the BA presented a high removal efficiency in comparison with the BS and NA. Due to competitive advantages in nutrition and interspecies competition, numerous functional microbes quickly colonized harsh environments and then became the dominant bacteria. Ultimately, the above-mentioned dominant bacteria can remove Cr(VI) to combat the adverse environment in the process of their growth and metabolism [44,45]. The difference in the removal rate of Cr(VI) in BA1 and BA2 may be related to the microbial populations of the systems. Studies have shown that OD600 values have a positive effect on the bioremediation of Cr(VI) in the systems [6,44]. Prabhakaran et al. [46] investigated that the Cr(VI) of 4 mg/L was completely removed within 2 h when the population of C. paurometabolum was determined to be approximately 3.4 × 1010 CFU/mL [46].
In our study, from the perspective of the removal rate of Cr(VI) and OD600 values, the BE was the most promising strategy, and the removal rate (86.2%) and OD600 values (0.82) in 84 days were the highest among all strategies. From the study of the BS and BA strategies, it can be seen that the individual addition of nutrients or functional microbes was difficult to achieve the promising goal for bioremediation of Cr(VI). The addition of both sufficient nutrients and numerous functional microbes simultaneously may be the best way to solve the problems. Studies have reported that the synergy of sufficient nutrients and numerous functional microbes can not only decrease the bioremediation time but also enhance the bioremediation efficiency of Cr(VI) [5,10]. A study concluded that the mixed native bacterial consortium (Pseudomonas stutzeri L1 and Acinobacter baumanni L2) in the optimal medium can reduce Cr(VI) to over 95% in the Cr(VI) of 1000 mg/L [28]. In addition, several factors like loss of competition, temperature, type of nutrients, and Cr(VI) concentration could impact the effect of bioremediation [47,48].

4.2. The Perdurable Stability of Chromium after Bioremediation

After bioremediation, the perdurable stability of chromium compounds played a key role in ecological security and sustainable development [6,49]. A study reported that the bioremediation of heavy metals aimed to transform the dissolved state into a stable state [49]. As we know, the perdurable stability of heavy metals after bioremediation can be assessed by means of the variation of the Eh value, the chemical fractions, and the leaching toxicity [48,50].
The variation of Eh values has been widely known as a key sign involved in the transformation of oxidation state and reduction state in systems, Especially for variable-value metals like chromium and uranium [50,51]. From the variation of Eh value in this study, the Eh in the BA and BE presented a trend of decreasing firstly to about −300 mV in 42 days and then increasing slightly to −100 mV. The trend of Eh suggested that the system was gradually changing to a reducing state and then benefiting from the conversion of Cr(VI) to stable Cr(III) [6,52]. This is a promising and developing technology for the future to realize the efficient bioremediation of Cr(VI) [6,53]. Further studies reported that this approach was defined as a reverse process of biohydrometallurgy, which was effectively limiting the re-dissolution of heavy metals from the contaminated sites, achieving a conversion of the vicious circle into a beneficent cycle in the bioremediation process of heavy metals [11,19]. Therefore, the Eh level variation of systems played a key role in the bioremediation process of Cr(VI) [6].
After bioremediation, the different chromium fractions were sequentially extracted from the samples to assess the stability of chromium [31]. In this study, it could be observed from Figure 5 that the chromium gradually increased to the stable fractions (F4, F5, and F6), which were classified as having no risk to the surrounding environment. The chromium stable fractions reached 68–97% after the different treatments. This difference was attributed to the fact that the formation of different stable states was closely related to microbial species [54]. The addition of nutrients in the BS treatments stimulated the growth of oxidizing microbial species and caused the dissolution of stable chromium complexes in the process of acid production. For the BA treatments, the inoculation of functional microbes aided in their transformation into stable chromium compounds. Owing to both the introduced functional microbes and the enriched nutrients, the proportion of chromium in its stable state was the most prominent in the BE treatments. Studies have reported a certain correlation between the variation of chromium chemical fractions and the toxicity, mobility, and bioavailability of chromium [6,55,56]. It can be found that the toxicity, mobility, and bioavailability of chromium in the system decreased with the increase in chromium stability. The further studies were summarized as follows:
The variation of the Fe-Mn oxide fraction may be related to bioreduction products of Cr(VI) [41]. The chromate was smelted from chromite with irons, and the larger number of irons provided a great opportunity to build the formation process of Fe-Mn compounds under the action of microorganisms (Table 1). The inoculation of native microorganisms with reduction ability could transform Fe3+ into reduction (Fe2+, Fe0) by transporting free electrons during microbial respiration [57]. In traditional chemical methods, Fe2+ and Mn2+, which possess high reduction abilities, were the common reductants applied for the removal of Cr(VI) [58]. Therefore, an increase in Fe-Mn complexes was relatively reasonable after bioremediation. A similar result on the removal of Cr(VI) by Microbacterium sp. Y2 was also supported by He et al. [59]. In comparison with BS treatments, our study showed that the percentage of Fe-Mn compounds was higher in BA treatments. When both nutrients and functional microorganisms were added by BE treatments, the percentage of Fe-Mn fractions increased to 26%. The explanation was that the addition of nutrients and microorganisms contributed to improving the formation of Fe-Mn compounds (Figure 4).
As for the organic fraction, the variation of these had a close relationship with the bioreduction and biosorption during the bioremediation process [60]. The percentage of organic fraction was increased to 23% under the BE treatment. One possible reason might be attributed to the stimulation of nutrients for functional microorganisms, consequently producing various reductases that could use cofactors (NADPH or NADH+) to induce Cr(VI) reduction under aerobic conditions [32], such as the ChrR and YieF from Bacillus and Pseudomonas associated with the cofactors to catalyze the reduction of Cr(VI) to Cr(III)-organic matters (organic fraction) [3,61]. In addition, due to the numerous microbes and metabolites, including functional groups, proteins, and organic acids, in the BE system, these can provide more chemisorption sites for Cr(VI), so as to realize the efficient solidification of Cr(VI). This is consistent with the previous report that the increase in OD600 has a positive relationship with the variation of the organic fraction [5]. The increase in organic fraction in the BS may be due to the addition of rich nutrients, which provide an adsorption site for Cr(VI) directly and rapidly in comparison with those in the BA. Studies reported that rich nutrients, including complexant/chelating agents, were considered a promising and efficient method in the process of removing Cr(VI) [9,10].
The residual was considered the most stable state due to its hardly dissolved, migrated, and bioavailability in the natural environment [62]. Biomineralization is an important process that influences the formation of residual fractions and consequently reflects on the stability of chromium during the process of bioremediation [63]. In this study, the residual fraction in all systems was dominant. However, the relative percentage of residual fraction was different with the different bioremediation strategies. An explanation was that the total chromium of the residual fraction in different strategies had changed after bioremediation (Figure S1). For example, in the BE, the conversion content of unstable chromium was high in comparison with the BA and BS, causing an increase in total chromium in the residual fraction. This suggested that a portion of the chromium in the residual fraction was transformed from unstable chromium in the system. Moreover, under the condition of sufficient nutrition, the functional microbes with higher activity and a larger population would work for a longer period, promoting an increase in the percentage of chromium in the residual continuous by means of biomineralization [64]. The above result indicated that bioremediation was beneficial in increasing the perdurable stability of chromium.
From the perspective of leaching toxicity, the concentration of Cr(VI) in the NA and BS was much higher than the standard value in comparison with those of the BA and BE. It was attributed to the following factors: First, due to the addition of acid solution (leaching solution) causing a decrease in pH in the systems, it provided a reaction environment to promote the dissolution of chromium in the unstable condition. It was widely accepted that an acidic pH significantly increased the leaching of toxic metals (i.e., Zn, Cr, and Cu) [65]. Furthermore, owing to the stimulation of nutrients, several microbes that produce acid had a positive effect on the re-dissolution of Cr(VI) in the BS. For the BA, the analysis of leaching toxicity showed that the inoculation of functional microbes effectively enhanced the stability of chromium. Especially in the BE, the Cr(VI) concentration was less than the discharge standard (5 mg/L) in GB5085.3-2007 (Identification Standard for Hazardous Waste Identification). These findings suggested that CSS would be more stable and safer after the bioremediation of BE.
Based on a comprehensive analysis of Eh, chemical fraction, and leaching toxicity, it is found that the BE was the best strategy to realize the perdurable stability of chromium in the CSS.

4.3. The Feasibility of the Bioremediation Cr(VI) from the CSS

Due to poor nutrients, a harsh environment, hazardous intermediate products, and a complicated microbial community structure in the field of CSS, it is difficult to reach a bioremediation effect with single microbes [6]. In this study, we inoculated functional microbes and nutrients against adverse environmental conditions with the aim of selecting the best strategy for bioremediation in the CSS. For the BS strategy, the addition of nutrients stimulated the growth of several microbes that produced acid, leading to the re-dissolution of Cr(VI). This is consistent with previous studies, which reported that oxidative microbial species were able to leach the heavy metals as well as being involved in the process of producing acids [66,67]. Furthermore, the addition of nutrients significantly decreased the bacterial diversity, which was not beneficial to the ecological balance of the soil [68]. So, the BS strategy is not suitable for bioremediation of the CSS.
The BA strategy involved restructuring the microbial community structure via inoculating preadapted functional microbes to improve the removal capacity of Cr(VI). In this study, the composition of the microbial community in the BA indicated that functional microbes, including members of the genera Acinetobacter, Azomonas, Enterobacter, Stenotrophomonas, Pseudomonas, Bacillus, and Desulfovibrio, have gradually become the dominant flora. Among them, Enterobacter, Stenotrophomonas, Pseudomonas, Bacillus, and Desulfovibrio were the functional genera of Cr(VI) removal with high metabolic activities in the contaminated soil. For example, Stenotrophomonas, as an oligotrophic bacteria, not only survives in the harsh environment but also secretes massive adhesives (functional groups, proteins, and secondary metabolites) to remove Cr(VI) by means of bioreduction and biosorption [27]. Moreover, Bacillus sp., as the common species in the contaminated sites, had been investigated by Das et al. [69], who concluded that it exhibited high tolerance (900 mg/L) and maintained a bioreduction rate of 2.22 mg Cr(VI)/L/h for a long time. This was not alone; Pseudomonas sp. isolated from electroplating sludge was found to work for a long time by transforming Cr(VI) into Cr(III); meanwhile, the abundance of microbial biomass (functional groups, the solution protein, microorganisms, etc.) had the capacity to adsorb Cr(VI) during microbial respiration [70]. In addition, it was interesting to note that the functional microbes in the study were previously isolated from the CSS and belonged to native microbial species.
Generally, the BA could effectively remove Cr(VI) in comparison with the BS. Nevertheless, poor nutrition was a great obstacle for the activity and population of functional microbes, which made it difficult to achieve long-term sustainability of the bioremediation [71,72]. The BS process had excellent potential for microbial growth but led to the re-dissolution of Cr(VI) and a decrease in microbial diversity. According to the analysis of the removal effect and the perdurable stability of chromium after bioremediation, BE was considered the most feasible bioremediation strategy. More importantly, because the functional microbes in the BE were isolated from the CSS, they were inoculated into the territory and gradually became the dominant flora. This provided evidence for reducing the risk of ecological invasion for bioremediation strategies [73].
In general, according to the results of the analysis of Eh level, chromium chemical fractions, leaching toxicity, and microbial community succession, it was concluded that the BE treatment provided new insights into the removal of Cr(VI) over the long term.

5. Conclusions

Different strategies of the NA, BS, BA, and BE for bioremediation of Cr(VI) were assessed by analysis of the removal rate, the perdurable stability, and ecological safety. The BE was considered a promising and feasible strategy for chromium-contaminated sites. The removal rate of Cr(VI) in the BE reached 86.2% in 84 days, followed by the BA, BS, and NA. After bioremediation, the proportion of chromium in the stable state was highest at 97% in the BE in comparison with the NA, BS, and BA. The result of leaching toxicity indicated that the concentration of Cr(VI) was over 5 mg/L (the discharge standard in the Identification Standard for Hazardous Waste Identification) in all systems except for the BE. In addition, microbial community succession has shown that functional microbes with good colonization ability from the CSS gradually become the dominant microbes in the BE, avoiding the risk of ecological invasion.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/su151612522/s1.

Author Contributions

Conceptualization, X.Y. and M.Z.; Methodology, Z.Y. and X.Z.; Validation, X.Z.; Formal analysis, Y.Z.; Investigation, S.L.; Data curation, G.M. and X.L.; Project administration, M.Z. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the National Natural Science Foundation of China [grant numbers 51974279], the National Key Research and Development Program of China [grant numbers 2018YFC18018, 2018YFC18027], KeJunPing [2018] No. 159, and the Guangxi Scientific Research and Technology Development Plan [grant numbers GuikeAB16380287 and GuikeAB17129025].

Informed Consent Statement

Not applicable.

Data Availability Statement

The original contributions presented in the study are included in the article. Because part of this batch of original data involves other research topics and paper publication, further inquiries can be directed to the corresponding author.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. X-ray diffraction (XRD) patterns of chromium slag sample.
Figure 1. X-ray diffraction (XRD) patterns of chromium slag sample.
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Figure 2. The variation of pH values with time in the CSS by different treatments.
Figure 2. The variation of pH values with time in the CSS by different treatments.
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Figure 3. Variation of Eh values in the CSS by different treatments.
Figure 3. Variation of Eh values in the CSS by different treatments.
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Figure 4. The removal rates of Cr(VI) and OD600 levels with time in the CSS by different treatments. (a) The variation of Cr(VI) removal rates; (b) The variation of OD600 level.
Figure 4. The removal rates of Cr(VI) and OD600 levels with time in the CSS by different treatments. (a) The variation of Cr(VI) removal rates; (b) The variation of OD600 level.
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Figure 5. The percentage distribution of Cr fractions for different treatment systems of the CSS following the sequential extraction method.
Figure 5. The percentage distribution of Cr fractions for different treatment systems of the CSS following the sequential extraction method.
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Figure 6. Leaching concentration of Cr(VI) from the CSS samples after different treatments. The MCL is the maximum contaminant level determined by the standard for solid waste extraction procedures for leaching toxicity (HJ/T299-2007). The green circle means that the value is below MCL.
Figure 6. Leaching concentration of Cr(VI) from the CSS samples after different treatments. The MCL is the maximum contaminant level determined by the standard for solid waste extraction procedures for leaching toxicity (HJ/T299-2007). The green circle means that the value is below MCL.
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Figure 7. The relative abundance of bacterial communities in the CSS by different treatments on phylum bar.
Figure 7. The relative abundance of bacterial communities in the CSS by different treatments on phylum bar.
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Table 1. The main chemical components of the CSS (wt %).
Table 1. The main chemical components of the CSS (wt %).
OxideMass Percentage (wt %)OxideMass Percentage (wt %)
MgO30.08NiO0.15
CaO26.06TiO20.11
SO310.95Cl0.1
Fe2O39.35P2O50.09
Cr2O38.96V2O50.08
SiO27.33Co3O40.03
Al2O35.23ZnO0.03
Na2O0.71K2O0.02
Table 2. The information of experimental design in different systems.
Table 2. The information of experimental design in different systems.
Bioremediation StrategiesMedium (mL)CSS (g)MicroorganismInoculation Size (v/v)
NADeionized water (300)180--
BS1BMA (300) --
BS2BMB (300) --
BA1Deionized water (300)180native microbes10
BA2Deionized water (300)180SRB10
BEBMA (300)180native microbes10
Table 3. The detailed extraction procedures of modified Tessier sequential extraction.
Table 3. The detailed extraction procedures of modified Tessier sequential extraction.
Extraction ProceduresExtraction ConditionsChemical Forms
16 mL Deionized waterOscillation140 rpm for 18 hF1
16 mL of 1 mol/L MgCl2 (pH = 7)Oscillation 140 rpm for 1.5 hF2
16 mL of 1 mol/L NaOAc (pH = 5)Oscillation 140 rpm for 6 hF3
16 mL of 0.04 mol/L NH2OH·HCl dissolved in 25% (m/v) HOAcIntermittent oscillation 120 rpm in a thermostatic equipment (97 °C) for 5 hF4
10 mL of 0.01 mol/L HNO3 and 8 mL of 30% (m/v) H2O2, and adding 5 mL of 3.2 mol/L NH4OAc in the 20% HNO3Oscillation 120 rpm in a thermostatic equipment (87 °C) for 4 h, then, oscillation 120 rpm for 30 minF5
15 mL HCl + 5 mL HNO3 + 2 mL H2SO4Digestion for 3 hF6
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MDPI and ACS Style

Yan, X.; Yan, Z.; Zhu, X.; Zhou, Y.; Ma, G.; Li, S.; Liu, X.; Zhang, M. Comparing Different Strategies for Cr(VI) Bioremediation: Bioaugmentation, Biostimulation, and Bioenhancement. Sustainability 2023, 15, 12522. https://doi.org/10.3390/su151612522

AMA Style

Yan X, Yan Z, Zhu X, Zhou Y, Ma G, Li S, Liu X, Zhang M. Comparing Different Strategies for Cr(VI) Bioremediation: Bioaugmentation, Biostimulation, and Bioenhancement. Sustainability. 2023; 15(16):12522. https://doi.org/10.3390/su151612522

Chicago/Turabian Style

Yan, Xiao, Zhenghao Yan, Xuezhe Zhu, Yupin Zhou, Guoying Ma, Shuangquan Li, Xingyu Liu, and Mingjiang Zhang. 2023. "Comparing Different Strategies for Cr(VI) Bioremediation: Bioaugmentation, Biostimulation, and Bioenhancement" Sustainability 15, no. 16: 12522. https://doi.org/10.3390/su151612522

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