Bioleaching Process of Sewage Sludge and Anaerobically Digested Sludge via Indigenous Sulfur-Oxidizing Bacteria to Improve Dewaterability and Reduce Heavy Metal Content

Abstract

This study investigated the role of indigenous inoculum (primarily sulfur-oxidizing Acidithiobacillus thiooxidans and other acidophilic bacteria) in heavy metal removal from sewage sludge (SS) and anaerobic digested sludge (ADS). Four treatments were evaluated: inoculum + elemental sulfur (S/ADS + E), inoculum alone (S/ADS + B), elemental sulfur alone (S/ADS + S), and a control with no additives. After 7 days of bioleaching, SS and ADS exhibited comparable heavy metal removal rates on Ni (92–98%) and Pb (88–92%), which were significantly more mobilized than Cu (30–44%) and Cr (63–73%). After bioleaching treatment, residual metals in both sludge types were predominantly sequestered in the oxidizable (F3) and residual (F4) fractions, markedly reducing their environmental mobility and pollution risk during land application. The dewaterability performance, assessed via capillary suction time (CST), reached the optimal values in S + E and ADS + E within 24–48 h, after which CST increased alongside rising extracellular polymeric substances and dissolved organic carbon. While the S/ADS + B configuration exhibited marginally reduced Cu, Ni, and Pb removal efficiencies relative to S/ADS + E, it demonstrated superior dewaterability characteristics under equivalent reaction durations. These results suggest that limiting the sulfur (S₀) supply to moderate the growth and activity of autotrophic A. thiooxidans can maintain the bioleaching pH within 2.0–3.0, striking a balance between effective heavy metal removal and favorable dewatering performance.

1. Introduction

The management and final disposition of sewage sludge (SS) constitutes the most technologically challenging and cost-intensive component within wastewater treatment infrastructure. Sludge stabilization through anaerobic digestion represents the predominant approach for processing organic-rich sludge materials. Organic matter is decomposed anaerobically, and as the sludge is stabilized, a large amount of high-calorific biogas is produced as energy to make the sludge resource [1]. However, the heavy metal content in sludge can only be fixed/accumulated in the anaerobically digested sludge (ADS) rather than reduced, which will significantly affect the reuse and subsequent disposal of digested sludge, such as composting and land application [2,3].

Meanwhile, ADS needs to be dewatered for further utilization or disposal; thus, the dewaterability of ADS is of concern for a sustainable sludge treatment process [4]. The anaerobic digestion process induces the degradation of dissolved extracellular polymeric substances (EPSs), comprising proteins and polysaccharides, thereby compromising the dewatering characteristics of ADS. Thus, it is a big challenge to realize the dewatering of ADS [5]. Sludge dewatering is usually achieved by combining inorganic coagulants or synthetic polyelectrolytes (such as polyacrylamide) before filter pressing or centrifugation [6]. The dewaterability of sludge can also be enhanced by chemically oxidizing EPS and disintegrating sludge cells [7]. Nevertheless, apart from the high cost, secondary pollutants may also be introduced into the dewatered sludge before land application.

Contemporary research has identified bioleaching as an economically viable dual-purpose technology for simultaneous sludge dewaterability enhancement and heavy metal extraction [8]. The significant decrease in pH value of bioleaching sludge is due to the microbial oxidation of elemental sulfur by the inoculated sulfur bacteria. In the bioleaching process, two acidophilic chemolithoautotrophic bacteria are commonly involved, i.e., Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans [9]. The bioleaching of sewage sludge has been shown to efficiently remove heavy metals under mild conditions [10]. Concurrently, bioleaching often enhances sludge dewaterability (e.g., a 66.9% decrease in sludge filtration resistance after bioleaching) [11]. Recent work also highlights the role of microbial consortia, e.g., Khidr et al. (2025) found that mixed indigenous cultures outperformed monocultures, achieving metal removal efficiencies of up to ~62% [12]. The precise biochemical pathways underpinning this synergistic enhancement in both sludge dewaterability and heavy metal removal in ADS are not yet fully elucidated. In particular, most acidophilic sulfur-oxidizing bacteria operate strictly as autotrophs, deriving energy solely from inorganic substrates; any influx of organic carbon—such as low-molecular-weight peptides, amino acids, or polysaccharides released during cell lysis—increases the dissolved organic carbon (DOC) fraction in solution. Elevated DOC levels can complex with protonated metal ions, altering their speciation and redox potential, and may also form a protective colloidal layer on sludge particles, thereby inhibiting the attachment, biofilm formation, and sulfidogenesis activity of the acidophilic consortia. Together, these effects can diminish proton-driven metal solubilization rates and reduce EPS degradation, ultimately dampening both dewatering performance and metal leaching efficiency [13]. For example, Fang et al. find that when the dissolved organic matter concentration of the sludge is greater than 150 mg DOC/L, the bioleaching of heavy metals is inhibited due to the reduced activity of Acidithiobacillus species [14]. Attributed to the soluble organic compounds, pig manure extract is reported to inhibit the process of iron oxidation and metal solubilization [15]. Acetic acid and propionic acid are found to be the main inhibitors during the bioleaching of sludge [16,17]. This metabolic complementarity explains the documented superiority of co-cultured heterotroph–autotroph systems over axenic cultures in mineral oxidation applications [18]. Until now, the risks of the land application of sludge are yet to be well investigated. Sludge conditioning treatment provides a guaranteed technology for enhancing the safety of sludge land utilization. To develop high-efficiency bioleaching inoculum, researchers have isolated and purified acidophilic autotrophic strains with selected partner (heterotrophic) bacteria such as Rhodotorula mucilaginosa R30 [19] and Mucor circinelloides ZG-3 [20].

This can be tremendously simplified if indigenous bacteria with bioleaching potential are utilized [21]. In this study, the feasibility of fulfilling the bioleaching of SS and ADS by triggering the indigenous sulfur-oxidizing inoculum was explored. The performance of heavy metal removal rates and the changes in dewaterability of sludge slurry have been estimated. The results showed that with the addition of energy source S₀ and cultured inoculum, the remaining concentrations of various heavy metals almost fulfill the land application, and the dewaterability of tested sludges could be improved to a degree with the optimized condition of bioleaching.

2. Materials and Methods

2.1. Preparation of Indigenous Inoculum for Bioleaching

The secondary sludge was taken from a gravity thickening tank after the secondary sedimentation tank of the Shanghai Songshen Sewage Plant. The collected sludge sample was immediately transported to the laboratory and stored at 4 °C for further use. Sludge was used to enrich sulfur-oxidizing bacteria by using the modified Starkey medium. The medium consists of the following basic salts (g/L): (NH₄)₂SO₄ 2.0, MgSO₄·7H₂O 0.5, K₂HPO₄ 0.5, KCl 0.1, and CaCl₂·2H₂O 0.25. The pH of the medium was adjusted to 5.0 using 2 mol/L of sulfuric acid. During the enrichment process, in a 250 mL Erlenmeyer flask, 12.5 mL of activated sludge was inoculated into 125 mL of medium supplemented with sulfur powder S₀ (10 g/L), in which S₀ was used as the energy source. Microbial cultivation occurred under controlled conditions (28 °C thermostatic water bath, 180 rpm orbital agitation). During the enrichment process, the pH of culture broth was determined every day (used as an indicator of microbial activity). When the pH was below 2, 12.5 mL of the culture was re-transferred to 125 mL of fresh culture medium. The enrichment was repeated until the sulfur oxidation capacity of the community was stable. Before the bioleaching experiment, four batches of sequential culturation were finished and the pH of the culturing solution decreased to <2.0 within 48 h, and then the culture was used as the inoculum for the bioleaching experiment.

2.2. Bioleaching Experiment

Two kinds of substrates were used in the bioleaching experiment, i.e., fresh SS and ADS. The fresh SS was the secondary sludge taken from the Shanghai Songshen Sewage Plant. ADS was the effluent taken from the lab-scale semi-continuously operated anaerobic digester fed with sewage sludge and pig manure. The total suspended solids of the feedstock were maintained at 5% and the ratio of sludge to pig manure was 1.76 (vs/vs), while the hydraulic retention time was 20 days. The fresh swine manure was taken from a farm located at Pinghu of Zhejiang province. The total suspended solids, volatile suspended solids, and DOC were quantified following established analytical protocols. The characteristics of SS and ADS are presented in

Table 1. Characteristics of fresh sewage sludge and anaerobically digested sludge.

Parameters	SS	ADS
pH	7.24 ± 0.55	8.08 ± 0.36
TS (%)	2.31 ± 0.20	4.20 ± 0.14
VS (%)	1.33 ± 0.08	2.10 ± 0.10
SCOD (mg/L)	110.0 ± 4.67	1656.7 ± 22.11
Ni (mg/kg TS)	147.8 ± 3.26	50.5 ± 1.94
Pb (mg/kg TS)	76.4 ± 2.11	52.5 ± 2.34
Cu (mg/kg TS)	431.9 ± 14.84	786.3 ± 24.13
Cr (mg/kg TS)	363.1 ± 19.51	196.8 ± 9.41

.

To investigate the feasibility of indigenous sulfur-oxidizing bacteria on bioleaching heavy metals from SS and ADS, inoculum and energy source S₀ were supplemented as compared to the control reactors. Four treatments were set for each substrate, i.e., S + E (10% inoculum + 2 g/L S₀), S + CK (no inoculum, no S₀), S + S (no inoculum + 2 g/L S₀), and S + B (10% inoculum+ no S₀) for fresh SS, and ADS + E (10% inoculum + 2 g/L S₀), ADS + CK (no inoculum, no S₀), ADS + +S (no inoculum + 2 g/L S₀), and ADS + B (10% inoculum+ no S₀) for ADS. The pH of the reaction solution was initially adjusted to 5.0 by using 1.0 M sulfuric acid. The bioleaching experiment was conducted in 250 mL flasks containing 150 mL of sludge and 15 mL of inoculum. The flasks were incubated at 28 °C on a rotatory shaker at 180 rpm. At the intervals of 0, 1, 2, 3, 5, and 7 days, 10 mL of slurry samples was collected to characterize the changes in pH, ORP, capillary suction time (CST), heavy metal concentrations, and EPS contents. All the treatments were carried out in triplicates.

2.3. Analytical Methods

The microbial community of enriched inoculum was analyzed by using high-throughput sequencing on the Illumina platform (Illumina Miseq PE300, San Diego, CA, USA). Amplicon libraries were constructed by using primers 515F (5′-GTG CCA GCM GCC GCG G-3′) and 806R (5′-GGACTA CHV GGG TWT CTA AT-3′) to amplify the V4–V5 region of the 16S rRNA gene [22]. The sequences were clustered into operational taxonomic units by setting a 97% identity threshold in USEARCH v 7.0. Operational taxonomic unit information such as phylum, class, family, and genus levels was identified for taxonomic classification using RDP classifier v 2.2.

The TS and VS contents were determined by measuring the weight loss at 105 °C and 24 h, and 550 °C and 4 h, respectively. pH was determined using a digital pH-meter (pHS-3C, WTW, Xylem lnc, Charlotte, NC, USA). CST values of sludge samples were determined by a capillary suction timer (Model 304M, Triton, Wetherby, UK). For the determination of total heavy metal content, the sludge samples were subjected to di-acid digestion (HNO₃-HCl-HClO₄) and the heavy metals in the digested liquid were determined with an inductively coupled plasma optical emission spectrometer (ICP-OES, Optima 2100DV, Perkin Elmer, Waltham, MA, USA). To determine the morphological changes of heavy metal in sludge samples, the BCR’s sequential extraction method was performed to obtain four fractions, i.e., F1 (exchangeable fraction), F2 (reducible fraction), F3 (oxidizable fraction), and F4 (residual fraction) [23].

In order to study the effect of EPS content on sludge dewatering, the total amount of EPS and the amount of protein and polysaccharide in EPS were determined. The EPS fractions of SS and ADS samples were extracted using a modified extraction method. An amount of 10 mL of sludge was first centrifuged at 550× g for 15 min, and the organic matter of the supernatant was regarded as the slime fraction. Then, the sludge particles were re-suspended with 0.01 mol/L of NaCl, sonicated (120 W, 2 min), and centrifuged (9000 r/min, 15 min), and the collected supernatant was regarded as loosely bonded EPS (LB-EPS). Finally, the particles were re-suspended with 0.01 mol/L of NaCl, heated (70 °C, 30 min), and centrifuged (20,000 r/min, 20 min), and the collect supernatant was regarded as tightly bonded EPS (TB-EPS). The contents of protein and carbohydrate were analyzed for each fraction.

2.4. Data Analysis

Statistical analysis was performed using IBM-SPSS Software v21. The data presented are the mean and standard deviation of three independent replicates. The variation between different treatments (Tukey’s Test) and the correlation (Pearson’s correlation) between different parameters were considered statistically significant at a confidence level of p < 0.05.

3. Results and Discussions

3.1. Microbial Community in Bioleaching Culture

High-throughput sequencing elucidated structural modifications in microbial consortia during acclimatization culturing, and the results are presented in Figure 1. After culturing, the diversity of the microbial community (e.g., α-diversity) underwent significant changes, decreasing from 5.87 to 2.66. There are four kinds of bacteria enriched in the sulfur dosing acidophilic culture, i.e., Acidithiobacillus ferrooxidans (20.8%), Acidiphilium multivorum AIU301 (28.8%), Acidithiobacillus thiooxidans (0.24%), and unclassified Acidithiobacillus (2.85%). Usually, there are four kinds of bacteria included in the enriched acidophilic culture [18], i.e., the autotrophic bacteria related to bioleaching, the heterotrophic bacteria that may be related to bioleaching, the facultative heterotrophic bacteria, and the bacteria group unrelated to bioleaching. In the present study, acid-tolerant heterotrophs dominated the enriched consortium, comprising 52.7% of operational taxonomic units, and the remaining were facultative heterotrophic bacteria, such as Clostridium (5.27%), Turicibacter (3.67%), and Alicyclobacillus ferripilum (2.23%). Clostridium, the branches of Firmicutes, displayed positive and significant correlations with volatile fatty acids [24]. It is expected that the acidity will be generated via the joint action of autotrophic bacteria and facultative bacteria, using sulfur oxidation intermediates and organic matters, respectively [25].

3.2. pH and ORP Changes During Bioleaching Treatment

For the effective solubilization of metals, the maintenance of sludge pH at approximately 2.0 during bioleaching, driven by bio-acidification, is critical. Thus, pH 2.0 was chosen as the endpoint in this study. As presented in Figure 2, the pH reduction rate varied depending on the components supplemented. In the absence of an energy substrate and inoculum, the pH of SCK and PCK remained stable, indicating the suppressed activity of endogenous acid fermentation bacteria. In contrast, the pH of ADS + CK and S + CK increased slightly from 5.0 to 5.12 and 5.85, respectively, which might be associated with the ammonification process driven by indigenous heterotrophic bacteria in the sludge. The trends of pH changes in S + E and S + B were quite similar, but the pH level of S + E was marginally lower than that of S + B throughout the process.

In the reactors added with energy source S₀, the pH of SS decreased to around 4.0 after Day 2, whereas the pH of >4.0 persisted to Day 3 in ADS. By the end of the 7-day process, the pH of both reactors decreased to around 2.0, indicating the triggered activity of autotrophic sulfate-oxidizing bacteria; nevertheless, the initiation of autotrophic bacteria was subjected to varying degrees of inhibition. In the reactors with inoculum and energy source S₀, the pH considerably decreased to ~2.0 and ~2.6 on the second day for SS and ADS, respectively.

The changes in ORP (Figure 2c,d) were closely related to pH dynamics. Along with the formation of SO₄⁻ and HSO₄²⁻, the ORP in bioleaching reactors increased significantly from ~100 mV to ~300 mV within three days. In reactors supplemented with inoculum (S + B and ADS + B), ORP similarly rose to 260 mV and 250 mV, respectively, indicating the activity of sulfur-oxidizing bacteria. In reactors amended with S₀ alone, the final ORP approached levels observed in inoculated systems, and a short lag phase period appeared at the beginning. As shown in Figure 2, the ORP of S + S increased after Day 2, and the ORP of ADS + S showed a delayed increase starting at Day 3, whereas the ORP of the control reactors (S + CK, ADS + CK) were maintained at the original level. These pH and ORP trends demonstrate that sufficient inoculum accelerates bioleaching initiation, while S₀ serves as a critical substrate to activate sulfur-oxidizing bacteria.

3.3. Heavy Metal Solubilization During Bioleaching Treatment

3.3.1. Differentiated Removal Rates for Cr/Cu and Ni/Pb in Bioleaching Process

The removal rates of four kinds of heavy metals during the bioleaching process are presented in Figure 3 and Figure 4. The initial concentrations of Cu and Cr in the substrates accumulated to a high level, which were 431.9 and 363.1 mg/kg TS in SS, and 786.3 and 196.8 mg/kg TS in ADS. The “Control Standards of Pollutants in Sludge for Agricultural Use (GB 4284-2018)” [26] applied in China stipulate that Cu and Cr concentrations in Grade A sludge products must not exceed 500 mg/kg TS. Similarly, the “Organic Media for Greening use (LY/T 1970-2011)” issued by the State Forestry Bureau of China mandates stricter thresholds, i.e., Cu < 150 mg/kg TS (Grade I) or <300 mg/kg TS (Grade II), and Cr < 70 mg/kg TS (Grade I) or <300 mg/kg TS (Grade II) (Table S1). Given these regulatory requirements, reducing heavy metal concentrations in both sludge types is imperative.

As evidenced in Figure 3, the maximum removal rates of heavy metals were found in treatments with inoculum and S₀. The removal rates of Ni and Pb were quite high, which reached up to 89–98%. Comparatively, the removal rates of Cr and Cu were lower than those of Ni and Pb. The removal rates of Cr and Cu in S + E and ADS + E were (44.4%, 73.7%) and (29.3%, 64.0%), respectively. Thus, it can also be found that the solubilization rates of Cr and Cu in SS were slightly higher than those in ADS. The differences could be associated with the speciation of each heavy metal in the sludge slurry.

In comparison, the exchangeable, carbonate-bound, and Fe/Mn-oxide-bound fractions exhibit higher bioavailability. Conversely, oxidizable and residual fractions demonstrate low or negligible bioavailability and are thus regarded as less environmentally hazardous [27,28]. As shown in Figure 4a, the combined exchangeable fraction (F1) and reducible fraction (F2) of Pb and Ni were significantly higher than those of Cr and Cu, contributing to their greater solubilization rates during bioleaching. In contrast, Cu—which has a strong affinity for organic matter—was predominantly present in the oxidizable fraction (74.8–77.8%) in both SS and ADS, resulting in limited dissolution efficiency. This preference for organic binding aligns with prior findings, where Cu’s stabilization in organic complexes hindered its mobilization [28]. Similarly, Cr primarily exists as trivalent hydroxide (an inorganic precipitate) in sludge, as reported by Wang et al. [29], explaining its consistently low dissolution efficiency during bioleaching.

3.3.2. Comparative Removal Rates for Different Test Groups

In sulfur-based bioleaching systems, the dissolution of metal sulfides is achieved through two distinct mechanisms: direct and indirect pathways. During direct bioleaching, Acidithiobacillus thiooxidans physically contacts and directly oxidizes metal sulfides (e.g., NiS, CuS, ZnS) in sludge, converting them into soluble metal sulfates via the reaction described in Equation (1) [30]. The indirect mechanism involves sulfur-oxidizing bacteria oxidizing elemental sulfur or reducing sulfur compounds in sludge to sulfuric acid (Equation (2)), which significantly lowers the pH of the sludge medium and subsequently enhances metal ion dissolution through acid dissolution (Equation (3)) [31,32].

Experimental data reveal (Figure 3) that Cu removal rates in the inoculum groups (S + B and ADS + B) were comparable to those in the bioleaching groups (S + E and ADS + E), with Ni exhibiting similar removal trends. A similar trend was found for Ni. This phenomenon indicated that the dissolution of metals like Cu and Ni in sludge systems primarily relies on the microbial direct oxidation pathways, i.e., the direct pathways [28]. It is noteworthy that the heavy metal removal efficiency in S₀-supplemented groups was generally lower than both bioleaching and inoculum groups. Although the final system pH reached ~2.5, the incomplete metal dissolution may be attributed to the metabolic lag phase of sulfur-oxidizing bacteria during initial adaptation. This dissolution hysteresis suggests that in practical engineering applications, the influence of the microbial metabolic kinetics on leaching efficiency requires comprehensive consideration.

$$\mathrm{M}\mathrm{e}\mathrm{S}+{2\mathrm{O}}_2\to \mathrm{M}\mathrm{e}\mathrm{S}{\mathrm{O}}_4$$

(1)

$${\mathrm{S}}^0+{\mathrm{H}}_2\mathrm{O}+1.5{\mathrm{O}}_2\to {\mathrm{H}}_2{\mathrm{S}\mathrm{O}}_4$$

(2)

$${\mathrm{H}}_2{\mathrm{S}\mathrm{O}}_4+\mathrm{s}\mathrm{l}\mathrm{u}\mathrm{d}\mathrm{g}\mathrm{e}-\mathrm{M}\mathrm{e}\to \mathrm{s}\mathrm{l}\mathrm{u}\mathrm{d}\mathrm{g}\mathrm{e}-2\mathrm{H}+\mathrm{M}\mathrm{e}\mathrm{S}{\mathrm{O}}_4$$

(3)

In the group with sole inoculum, no substance was added, and the removal efficiency of heavy metals was low. Except for SCK, the removal rate of Ni in the experimental group and the control group was similar. The removal rate of Cu is related to the content of A. thiooxidans. According to the control standards of pollutants in sludge for agricultural use in GB 4284-2018 (Ni < 100 mg/kg, Cu < 500 mg/kg, Pb < 300 mg/kg, Cr < 500 mg/kg), the sludge after bioleaching meets the Class A standard for agricultural application.

3.4. Changes in Sludge Dewaterability

CST is a widely used indicator of sludge dewatering ability, where higher CST values correlate with the poorer dewaterability of sludge [33]. As the solid content of SS and ADS is different, the normalized CST was calculated and is presented in Figure 5. Pre-adjusting the sludge pH to 5.0 reduced CST values by 24.5% and 31% in SS and ADS, respectively. During the bioleaching process, CST values in all treatment groups (excluding controls) generally reached their lowest levels by Day 1 or 2, though the extent of reduction varied across treatments. Notably, S + B and S + S exhibited lower CST values than S + E, a trend mirrored in ADS treatments. Subsequently, CST values gradually increased until the process concluded, likely due to declining pH. Prior studies suggest optimal dewatering at pH 2.67 [3]. However, excessive acidification (pH < 2.5) reversed this effect, increasing CST and impairing dewaterability [34], which was inconsistent with the observed trends in this study.

Furthermore, the dewatering performance of sludge is intrinsically linked to the dynamic changes in EPS content within sludge flocs [35,36]. EPS, composing proteins, polysaccharides, lipids, and humus substances, serves as the structural matrix governing sludge floc aggregation and architectural stability [37]. Structurally, these biopolymers establish a three-dimensional network through electrostatic interactions and hydrogen bonding. Functionally, their characteristic hydroxyl-rich chemical composition enables substantial bound water retention through hydrogen bonding, while simultaneously strengthening inter-floc cohesion forces [38]. Therefore, the elevated EPS concentrations inversely correlate with sludge dewaterability.

As shown in Figure 5, the ADS exhibited significantly impaired dewatering characteristics compared to the SS, and the initial CST value increased from 60 s (SS) to 157 s (ADS), paralleled by a dramatic EPS content elevation from 4.5 to 160 mg/g TS. The normalized CST of ADS was 37.5 s·L/gTS as compared to 26.15 in SS. This correlation demonstrates that the post-digestion deterioration of dewaterability originates from EPS accumulation during anaerobic processes [39].

As seen in Figure 6, the EPS content of ADS + E decreased during the first 24 h but subsequently rose gradually, reaching 153 mg/g TS by the end of the process, while its DOC concentrations increased concurrently. A similar upward trend in EPS was found in ADS + S. These results indicated that the death and dissolution of autotrophic bacteria did occur in ADS + E and ADS + S, likely releasing intracellular materials such as proteins, carbohydrates, lipids, DNA, and RNA, thereby leading to the increase in total EPS content [40]. In contrast, EPS levels in ADS + B declined after 24 h and a similar trend was found in SS + B. This reduction may be attributed to the release of EPS under acidic pH conditions and the consumption by certain acid-resistant heterotrophic bacteria in the absence of energy source S₀ [19,41], or by microorganisms with certain enzymes (such as protease, amylase, DNase, and RNase) [42]. Consequently, the CST value of ADS + B remained at the lowest level among all treatments. Correlation analysis (

Table 2. Correlations of CST with various parameters of sludge.

CST	Coef.	St. Err.	t-Value	p-Value	95% Conf.	Interval	Significance
pH	−85.047	24.672	−3.45	0.001	−134.433	−35.662	***
ORP	−1.304	0.414	−3.15	0.003	−2.134	−0.475	***
sCOD	0.006	0.008	0.74	0.465	−0.01	0.021
EPS-slime	−0.402	0.575	−0.7	0.487	−1.554	0.749
EPS-LB	0
EPS-TB	−3.44	1.219	−2.82	0.007	−5.88	−1	***
EPS-slime-polysaccharide	3.791	1.729	2.19	0.032	0.329	7.252	**
EPS-slime-protein	0
EPS-LB-polysaccharide	−4.533	1.405	−3.23	0.002	−7.345	−1.721	***
EPS-LB-protein	−1.089	0.961	−1.13	0.262	−3.014	0.835
EPS-TB-polysaccharide	21.283	4.924	4.32	0	11.427	31.138	***
EPS-TB-protein	0
EPS-slime-protein/polysaccharide	1.072	10.454	0.1	0.919	−19.854	21.998
EPS-LB-protein/polysaccharide	5.732	4.371	1.31	0.195	−3.018	14.482
EPS-TB-protein/polysaccharide	0.024	2.921	0.01	0.994	−5.824	5.871
Total-EPS	0.423	0.905	0.47	0.642	−1.389	2.235

*** p < 0.01, ** p < 0.05.

) revealed a significant relationship between the total EPS content and the dewatering performance. These observations are consistent with prior studies [43], which reported that sludge with a higher organic content generally tends to exhibit elevated EPS levels, a lower density, and a more negative zeta potential, which hinder the particle aggregation and flocculation of sludge flocs.

Previous studies have found that sludge DOC is toxic to A. Thiobacillus, because the presence of some low-molecular-weight organic acids in sludge DOC will significantly reduce the activity of A. Thiobacillus [17,44]. It has been reported that the tannery-sludge-dissolved organic matter presented an inhibition effect on A. thiooxidans TS6 growth and led to the prolongation of the lag phase; the heterotrophic bacteria (P. spartinae D13) played a key role in eliminating the dissolved organic matter toxicity to Acidithiobacillus spp. It can be seen from Figure 6c,d that the DOC of S + E was greatly increased from the initial 0.3 mg/g TS to 5.47 mg/g TS, and the DOC of ADS + E rose from 21.8 mg/g TS to 28.3 mg/g TS after the bioleaching process. Comparatively, the rate of pH decrease in S + E was slightly higher than that in ADS + E, which might be attributed to the lower level of DOC in SS. Recent studies reported that bioleaching by the co-inoculation of A. ferrooxidans and A. thiooxidans showed an improvement in the dewaterability of activated sludge compared to other physical treatments [9]. Recent studies demonstrate the superiority of indigenous consortia: for instance, Khidr et al. found that co-cultures of native strains greatly exceeded individual strains in heavy metal reduction [12,45]. Similarly, our work leverages the natural sulfur-oxidizing community rather than a single purified strain, which is expected to enhance robustness and metal solubilization (consistent with the improved outcomes seen in consortium-based bioleaching). However, the co-presence of sulfur-oxidizing bacteria and iron-oxidizing bacteria extremely acidified the sludge pH (pH < 2) and deteriorated the sludge dewaterability during bioleaching. Thus, the A. thiooxidans-mediated acidification of sludge may be suitable for metal leaching rather than sludge dewatering [9]. In the present study, the inoculation of indigenous sulfate-oxidizing bacteria was also suitable for metal leaching; nevertheless, the improved dewatering capacity could also be compromised by optimizing the period of bioleaching.

4. Conclusions

The bioleaching of SS and ADS by indigenous sulfur-oxidizing consortia proved both highly effective and economical: under optimized conditions (10% inoculum, limited S₀), pH was stably maintained between 2.0 and 3.0, enabling a >90% removal of Ni and Pb and 30–70% removal of Cu and Cr. Simultaneously, sludge dewaterability improved by up to 30% (CST reduction) without chemical additives. Direct and indirect leaching pathways worked in concert to solubilize metals, while moderate acidification prevented excessive EPS release and CST rebound. This one-step, low-cost approach holds strong promise for the safe, resource-efficient conditioning of both fresh and digested sludges.