Fe Salts Hinder and Fe Oxides Help: Divergent Mechanisms in Sewage Sludge Anaerobic Digestion

Abstract

Anaerobic digestion (AD) is an important method for sewage sludge (SS) stabilization and methane recovery. Fe compounds are widely present in SS because they are commonly used for phosphorus removal and organic matter (OM) capture in wastewater treatment plants. Endogenous Fe occurs in different forms, but the roles of these forms in SS AD remain unclear. This study systematically compared the effects of FeCl₃, Poly-FeCl₃, Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O on AD. The results showed that FeCl₃ and Poly-FeCl₃ decreased methane yield by 9.90% and 11.92%, respectively, whereas Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O increased it by 18.54%, 15.23%, and 15.09%. The analysis suggested that flocculating salts FeCl₃ and Poly-FeCl₃ groups increased sludge particle size, decreased SCOD concentrations by 10.21% and 12.41%, as well as F₄₂₀ by 16.88% and 28.63%, respectively, thereby inhibited the methanogenesis process. In contrast, Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O enhanced methane production by promoting OM hydrolysis, with SCOD concentrations increased by 12.71%, 8.99%, and 7.47%, respectively. XRD, CV, and EIS results showed that Fe₃O₄ likely promoted methanogenesis through a stable Fe(III)/Fe(II) cycle and electron transfer. Although FeOOH and Fe₅HO₈·4H₂O also underwent Fe(III)/Fe(II) conversion, their promoting effects were weaker than that of Fe₃O₄, possibly because the lack of a bulk mixed-valence structure reduced the efficiency of continuous electron transfer. This study highlights that the chemical form of Fe in SS fundamentally determines its effects on AD performance.

1. Introduction

Anaerobic digestion (AD) is a key method for sewage sludge (SS) reduction, stabilization, and resource recovery, as it can convert biodegradable organic compounds into biogas [1]. However, its application is constrained by insufficient release of organic matter (OM), slow hydrolysis, and incomplete methanogenesis [2,3,4]. It has been reported that only 30–40% of the OM in SS is converted into CH₄ [5]. Therefore, exploring effective methods to enhance OM release, accelerate intermediate conversion, and improve methane production is necessary. Methanogenic archaea drive terminal methane production, traditionally relying on interspecies hydrogen transfer (IHT) and interspecies formate transfer (IFT) to maintain syntrophic partnerships [6]. In IHT, H₂ acts as a diffusible electron carrier, yet the H⁺/H₂ redox potential (E⁰′ = −414 mV) is substantially more negative than those of NADH (−320 mV), FADH₂ (−220 mV), and ferredoxin (−398 mV) [7], resulting the H₂ generation being thermodynamically unfavorable unless its partial pressure in the system drops to a very low level (<10⁻⁴–10⁻⁵ atm) [8]. In recent years, direct interspecies electron transfer (DIET) has emerged as a more efficient alternative, which allows fermentative bacteria to transfer electrons directly to methanogens via conductive appendages (e.g., nanowires) [9,10]. Compared with the IHT/IFT, DIET has higher electron transfer efficiency and lower energy loss, and is thus considered a promising way to enhance anaerobic conversion of organic waste and improve methane production [11].

As the by-product of wastewater treatment, the compositions of SS are the key factors that affect following AD performance. In wastewater treatment plants, Fe is a common component of SS. Fe salt flocculants are widely used for chemical phosphorus removal, enhanced suspended wastewater OM removal. These applications result in a large amount of Fe entering the subsequent SS treatment process [12,13,14,15]. The average Fe content in SS reached 68.9 mg/g dry sludge [16]. In addition, Fe in SS usually does not exist in a single form but coexists as Fe phosphates, Fe oxides, Fe (oxy)hydroxides, and their transformation products [17]. Among these, Fe oxides (e.g., Fe₃O₄) have received particular attention because they can participate in electron transfer, regulate the redox environment, and affect microbial metabolism, being thus regarded as potential enhancers for AD [18].

Despite these observations, the mechanisms by which different Fe compounds affect AD remain poorly understood. Fe salt flocculants were found to decrease AD performance, as they form dense sludge flocs and reduce the hydrolysis of organic mass [19]. For example, Zhu et al. found that adding 40 g/kg TS poly-ferric chloride (Poly-FeCl₃) reduced methane production by 20.0%. Moreover, the dissimilatory iron reduction (DIR) process competed with methyl-CoM for electrons, which further lowered methane production [19]. In contrast, Yu et al. found that adding 200 mg/L ferric chloride (FeCl₃) to SS increased total gas yield by 79.6% [20]. On the other hand, certain Fe oxides have been reported to increase methane production rate and OM degradation efficiency. For example, Fe₃O₄ was reported to significantly promote the AD process and improved methane yield [18,21]. Hematite (Fe₂O₃) and Goethite (α-FeOOH) were also found to increase methane efficiency and facilitate OM degradation to some extent [22]. These positive effects are often attributed to the influence of Fe oxides on the redox environment, electron transfer ability and microbial metabolism pathways. However, a systematic comparison of different Fe compounds-ranging from soluble Fe salts to various Fe oxides is still lacking. Therefore, two common Fe salts (FeCl₃ and Poly-FeCl₃) were selected as the Fe-based additives, while Fe₃O₄, FeOOH, and Ferrihydrite (Fe₅HO₈·4H₂O) were chosen as representative Fe oxides, as these three compounds differ in stability, conductivity, Fe(II)/Fe(III) composition, and redox activity.

To address these knowledge gaps, this study systematically investigated the effects of five typical Fe compounds (FeCl₃, Poly-FeCl₃, Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O) on SS AD, including: (1) organic components conversion and methane production; (2) the evolution of Fe compounds and the underlying mechanisms linking Fe forms to AD outcomes. By systematically comparing these soluble Fe salts with Fe oxides, this study aims to clarify the specific effects of different Fe forms on OM conversion, electron transfer, and methane production. This study aims to clarify how different Fe compounds influence the efficient bioconversion of SS into biogas.

2. Materials and Methods

2.1. Materials and Sludge Characteristics

SS was collected from a secondary clarifier of a wastewater treatment plant in Shanghai, China. The SS was first passed through a 10–mesh sieve to remove debris and coarse particles, then concentrated by gravity settling for 12 h with supernatant decantation. The concentrated SS exhibited total solid (TS) of 20.22 ± 0.25 g/L, volatile solid (VS) of 14.07 ± 0.42 g/L, and a total chemical oxygen demand (TCOD) of 13.25 ± 0.63 g/L (

Table 1. Main features of concentrated SS and inoculum sludge.

Parameters	Units	Sewage Sludge	Inoculum
pH	/	6.44 ± 0.01	/
TS (g/L)	g/L	20.22 ± 0.25	69.44 ± 0.45
VS (g/L)	g/L	14.07 ± 0.42	55.68 ± 0.57
TCOD (g/L)	g/L	13.25 ± 0.63	/

). Anaerobic inoculum was obtained from industrial scale mesophilic AD operated at 37 ± 1 °C, with TS values of 69.44 ± 0.45 g/L and VS values of 55.68 ± 0.57 g/L.

Six groups of AD tests, including a control group (without Fe addition), a Fe₃O₄ group, a FeOOH group, a Fe₅HO₈·4H₂O group, a FeCl₃ group, and a Poly-FeCl₃, group were set up in this study. The Fe dosage of 40 mg Fe/g TS was chosen to encompass the typical range observed in chemically enhanced primary treatment residuals (2.648.3 mg Fe/g TS) [19,23].

2.2. Sludge Conditioning and Batch AD Tests

For Fe salt conditioning, FeCl₃ and Poly-FeCl₃ were dosed via wet addition to establish homogeneous distribution. Briefly, FeCl₃ was dissolved in deionized water to yield a stock solution of 100 mg Fe/mL. Calculated volumes of FeCl₃ or Poly-FeCl₃ stock solutions were separately pipetted into 200 mL SS aliquots to achieve the target Fe dosage. The mixtures were rapidly mixed at 120–150 rpm for 2 min, followed by slow stirring at 40–60 rpm for 15 min to facilitate floc maturation without shear disruption [24,25].

For the Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O groups, 0.22 g Fe₃O₄, 0.26 g FeOOH, and 0.28 g Fe₅HO₈·4H₂O were weighed, respectively. Each Fe compound was added to 200 mL of SS to obtain the same Fe dosage of 40 mg Fe/g TS. The mixtures were then stirred at 120–150 rpm for 10 min to ensure uniform suspension. The conditioned SS was then used for the subsequent batch AD tests. The pH value of all sludge samples was adjusted to 7.0 by using 4M HCl before AD.

The batch experiments were carried out in 120 mL serum bottles. Each bottle contained 10 mL of inoculum sludge, 50 mL of well-mixed Fe-containing sludge, and 5 g/L buffer solution (sodium bicarbonate). The headspace was purged with high-purity N₂ for 5 min to establish anaerobic conditions, after which the bottles were immediately sealed with butyl rubber stoppers and aluminum crimp caps. All bottles were incubated in a constant-temperature orbital shaker at 37 ± 1 °C and 180 rpm. Triplicate reactors were prepared for each treatment, alongside control reactors without Fe amendment. Biogas production was monitored periodically throughout the digestion period. All samples were prepared in triplicate, and the sludge without added Fe compounds was used as the experimental control group.

2.3. Analysis Methods

The methane content of the biogas was measured using a gas chromatograph equipped with a thermal conductivity detector (GC-TCD). Before analysis, the gas pressure in the injector was equilibrated to the same conditions as the headspace of the serum bottles. A 1 mL of the pressure equilibrated biogas was then withdrawn by using a gas-tight syringe and injected into the GC-TCD for analysis. A calibration curve was established using standard methane gas, and the methane volume fraction in each sample was calculated from the corresponding peak area. The methane volume in each bottle was then obtained by multiplying the methane volume fraction by the corresponding headspace volume. Finally, methane production was normalized to the TS content of the substrate and expressed as mL CH₄/g TS. Volatile fatty acids (VFAs), including acetic acid, propionic acid, and butyric acid, were measured using a gas chromatograph (GC, Shanghai Chromatograph) equipped with a flame ionization detector (FID). The measurements of total solid (TS), volatile solid (VS), soluble chemical oxygen demand (SCOD), and soluble protein were conducted according to the standardized methods established by the American Public Health Association [26]. Dissolved organic matter (DOM) was measured by an excitation-emission matrix (EEM) fluorescence spectrometer (F-7000, Hitachi, High-Tech Corporation, Tokyo, Japan) [27]. The concentration of Fe²⁺ and Fe(II) was determined by the 1,10-phenanthroline method [28]. For total Fe(II) measurement, samples were extracted with 0.5 M HCl at an extractant/sample ratio of 1:9 (v/v) under dark conditions at 25 °C for 24 h. The extract was centrifuged at 8500 rpm for 5 min, and the supernatant was filtered through a 0.22 μm aqueous membrane. The filtrate was then mixed with 200 mM 4-(2-hydroxyethyl)-1-piperazineethanesulfonic acid (HEPES) buffer containing 0.1% 1,10-phenanthroline at a ratio of 1:1 (v/v), reacted in the dark for 10 min, and measured at 510 nm to determine the Fe(II) concentration. The solid-phase Fe(II) concentration was calculated by subtracting the liquid-phase Fe²⁺ concentration from the HCl-extractable total Fe(II) concentration. X-ray diffraction (XRD) was used to identify the crystalline phases and mineralogical transformation of Fe-containing solids after AD. The particle size distribution of the sludge samples was measured using a laser particle size analyzer (MASTERSIZER 2000, Malvern Instruments Ltd., Malvern, UK) to evaluate the effect of different Fe compounds on sludge particle size. Before measurement, the sludge samples were gently mixed to ensure homogeneity without destroying the floc structure. The median particle size, expressed as D₅₀, was used to compare the particle size changes among different treatment groups.

The electrochemical redox behavior of the AD samples was characterized by cyclic voltammetry (CV), using a three-electrode electrochemical workstation (Chenhua, CHI660,Shanghai, China) [29]. The prepared graphite sheet was used as the working electrode, with an Hg/Hg₂Cl₂ electrode as the reference electrode and graphite as the counter electrode. The measurement was carried out in 10× PBS, and the AD liquid was directly added into the buffer solution for measurement. The scan range was from −0.6 V to 0.8 V, with a scan rate of 50 mV/s. Electrochemical impedance spectroscopy (EIS) was conducted using the same three-electrode system. The frequency range was set from 1 × 10⁵ Hz to 0.01 Hz.

Protease activity was measured according to the method described by Maeda et al. [30]. The samples were centrifuged at 10,000 rpm for 10 min, and the supernatants were collected and passed through a 0.2 μm membrane filter. The filtered supernatants were then subjected to protease activity determination. Coenzyme F₄₂₀ was measured according to the method of P. J. Reynolds [31]. All experiments were performed in triplicate, and the results were reported as mean ± standard deviation values.

3. Results

3.1. Effects of Fe Compounds on the Evolution of OM During AD

3.1.1. Effects of Different Iron Compounds on Methane Production

The effects of Fe salts and Fe oxides on SS AD were clearly different, with Fe salts reducing methane production and Fe oxides promoting it. Figure 1 shows the changes in cumulative methane production in each group. Specifically, the methane yield of the control group reached 151.49 mL CH₄/g TS, while the yield was reduced by 9.90% and 11.92%, respectively when FeCl₃ and Poly-FeCl₃ were added. In contrast, the Fe₃O₄ group achieved the highest cumulative methane yield of 179.30 mL CH₄/g TS (an increase by 18.54%). The methane yields in the FeOOH and Fe₅HO₈·4H₂O groups also increased by 15.23% and 15.09%, respectively. These results indicate that different Fe compounds exert divergent effects on SS AD, fundamentally determined by their chemical forms and underlying reaction mechanisms.

The inhibitory effect observed for Poly-FeCl₃ is consistent with the findings of Zhu et al., who reported that adding 40 g/kg TS of Poly-FeCl₃ reduced methane yield by 20.0%, attributing this inhibition to the formation of dense sludge flocs that physically encapsulated OM and restricted hydrolysis [19]. Our work further confirmed the inhibitory tendency for Poly-FeCl₃ and FeCl₃ for SS AD. It should be noted that the inhibitory effect of FeCl₃ observed in this study differs from the result of Qin et al. [32], who reported that 200 mg/L FeCl₃ increased methane yield by 28.9% during high–solids sludge AD. This difference may be attributed to the different Fe dosage, sludge properties, and AD conditions. At lower FeCl₃ dosage, Fe(III) may provide beneficial trace elements or electron–accepting capacity for anaerobic microorganisms. In contrast, the promoting effects of Fe oxides observed in this study confirm previous reports that solid Fe-based minerals can enhance methane production during AD. Zhao et al. found that Fe₃O₄ addition increased methane production from 1261.70 mL to 1638.80 mL by improving hydrolysis, acidification, and subsequent conversion processes [33]. In summary, the subsequent sections further examine how different Fe compounds affect the hydrolysis, acidification, and methanogenesis stages of SS AD, in order to clarify why Fe salts inhibited methane accumulation whereas Fe oxides promoted it.

3.1.2. Effects of Different Fe Compounds on OM Degradation

The divergent effects of Fe compounds on methane production were fundamentally rooted in their distinct influences on OM solubilization and transformation throughout the sequential stages of AD. Hydrolysis is the rate limiting initial step of SS AD. In this study, all Fe compounds increased SS particle size to varying degrees before AD (Figure 2a). The D₅₀ value increased from 45.99 μm in the control to 78.68, 60.81 μm in the Poly-FeCl₃ and FeCl₃ groups, respectively. In contrast, Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O exerted milder effects on particle size enlargement (D₅₀: 57.15, 46.86, and 54.84 μm, respectively). These particle size results confirm that the Fe salts caused stronger sludge flocculation than the Fe oxides under the same Fe dosage.

The particle size in the Fe salt groups could be increased because the released Fe³⁺ could neutralize the negative charges on colloidal particles and EPS in the sludge [34]. This promoted the aggregation and flocculation of small particles, leading to the formation of sludge flocs and an increase in particle size. In contrast, the particle size increases caused by Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O were negligible. In particular, FeOOH is a stable solid–phase mineral with low solubility and cannot rapidly release large amounts of Fe(III) hydrolysis products into the liquid phase. Therefore, its charge neutralization and bridging abilities are much weaker. The disparate particle size distributions suggested that Fe salts and Fe oxides established fundamentally different sludge matrices, which would consequently govern OM accessibility and mass transfer efficiency during subsequent AD.

During the early hydrolytic stage, different groups showed contrasting OM release patterns. SCOD is an overall indicator of soluble OM. Figure 2b shows that the SCOD concentration in all groups during AD first increased and then decreased over time. The highest SCOD concentration appeared on day 8, ranging from about 3900 to 5100 mg/L. Among all groups, the Fe₃O₄ group showed the highest SCOD concentration of 5086.33 mg/L, which was 12.71% higher than the control (4513.38 mg/L). FeOOH and Fe₅HO₈·4H₂O also promoted OM release, with SCOD values of 4919.34 and 4850.88 mg/L, respectively. In contrast, SCOD in the FeCl₃ and Poly-FeCl₃ groups decreased to 4052.34 and 3953.89 mg/L, which were 10.21% and 12.40% lower than the control. Soluble protein showed a trend consistent with SCOD. The Fe₃O₄ group achieved the highest value of 2255.23 mg/L, surpassing the control by 13.03%, whereas the Poly-FeCl₃ group exhibited the lowest peak at 1624.55 mg/L, representing an 18.60% reduction. FeCl₃ decreased by 10.63% compared with the control group (Figure 2c). This difference indicates that the dense flocs formed by Poly-FeCl₃ and FeCl₃ physically encapsulated particulate OM and restricted the effective contact between extracellular hydrolytic enzymes and substrates, thereby attenuating protein solubilization. This interpretation is corroborated by Zhu et al. [19], who reported that Poly-FeCl₃ promoted SS aggregation and hindered organic solubilization through hydroxy polymer bridging. Conversely, the elevated SCOD and soluble protein concentrations in the Fe₃O₄ group suggest that this mixed-valence Fe oxide facilitated OM release, potentially through enriching hydrolytic microorganisms and stimulating extracellular enzyme secretion at the solid-liquid interface. This interpretation is supported by the findings of Peng et al., who reported that Fe₃O₄ can strengthen the hydrolysis stage of SS AD by adsorbing and enriching functional microorganisms and increasing extracellular enzyme secretion [35]. These results show that the inhibitory effect of Fe salts and the promoting effect of Fe oxides on methane production can be traced back to the hydrolysis stage.

3D-EEM fluorescence spectra provided complementary evidence for the differential OM transformation trends (Figure 2d,i). From day 0 to day 15, three groups exhibited intensified fluorescence signals, indicating that sludge flocs were gradually broken during AD and that particulate OM was released into the liquid phase. Notably, the control group displayed prominent fluorescence peaks on day 15, suggesting the accumulation of soluble metabolites. In contrast, the FeCl₃ group showed attenuated fluorescence intensity, reinforcing the conclusion that flocculation constrained OM liberation. The Fe₃O₄ group, however, presented distinct protein-like and humic-like fluorescence peaks with enhanced intensity, demonstrating its superior capacity for OM release and transformation.

The VFAs profiles further suggest that the different OM release patterns influenced the subsequent acidification stage. As shown in Figure 2j–l, the three representative groups reached peak total VFAs concentrations around day 8 (1649–2094 mg/L). The Fe₃O₄ group attained the highest VFAs concentration of 2094.91 mg/L, which was 15.40% higher than the control group (1815.39 mg/L). In contrast, the FeCl₃ group showed the lowest VFAs concentration of 1649.33 mg/L, representing a decrease of 9.14%. Acetic acid dominated the VFAs composition across all treatments, accounting for over 40% of the total. Furthermore, a clearer distinction was observed during the subsequent VFAs consumption phase: the Fe₃O₄ group exhibited the lowest residual propionic acid concentration at the end of AD (64.79 mg/L), FeCl₃ showed the highest residual propionic acid concentration at the end of AD (295.39 mg/L). The efficient propionate degradation is particularly significant since propionate oxidation is thermodynamically unfavorable [36]. These results demonstrate that Fe₃O₄ played a dual role in SS AD by promoting acidification products formation in the early stage and facilitating intermediate VFAs conversion in the later stage. Conversely, FeCl₃ led to VFAs accumulation and inhibited further consumption.

3.1.3. Effects of Different Fe Compounds on Activity of Key Enzymes During AD

To further decipher the mechanistic basis for the divergent OM transformation, the activities of two key enzymes were analyzed. Protease, as the principal enzyme initiating protein hydrolysis, serves as a sensitive indicator of OM hydrolysis capacity in SS. The Fe₃O₄ group exhibited a 19.23% increase in protease activity relative to the control, whereas the Poly-FeCl₃ and FeCl₃ groups showed substantial reductions of 35% and 26.12%, respectively (Figure 3a). In addition, the FeOOH and Fe₅HO₈·4H₂O groups showed increases of about 13.54% and 11.09%, respectively. This enzymatic pattern directly explains the observed soluble protein variation: the dense floc structures induced by Fe salts physically sequestered particulate proteins from protease attack, while Fe₃O₄ likely provided favorable surface microenvironments for hydrolytic bacteria and enhanced the retention of extracellular enzymes.

The coenzyme F₄₂₀ content, which reflects methanogenic metabolic vigor, also showed a trend consistent with methane yields (Figure 3b). Compared with the control group (46.8 U/L), the F₄₂₀ content in the FeCl₃ and Poly-FeCl₃ groups decreased by 16.88% and 28.63%, respectively. However, the Fe oxide groups showed higher F₄₂₀ contents overall, with increases of 22.43%, 15.61%, and 11.96% in the Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O groups, respectively. The different F₄₂₀ levels directly correspond to the relevant methane yields, further suggesting Fe salts inhibited F₄₂₀ activity, and Fe oxides were more favorable for sustaining methanogenic activity. Previous studies have reported that Fe oxides provide a more advantageous redox environment and surface conditions for methanogens [37].

3.2. Evolution of Fe and the Mechanism on Sludge AD

3.2.1. Transformation of Different Fe Compounds

The contrasting AD performances and OM conversion across different groups (Section 3.1) are fundamentally linked to the distinct transformation pathways and redox activities of the added Fe compounds. To elucidate this connection, the dynamics of soluble Fe²⁺ and solid-phase Fe(II) were monitored, and the electrochemical properties and mineralogical evolution of the Fe phases were further characterized.

As shown in Figure 4a, soluble Fe²⁺ concentration in all Fe groups increased during the initial 8 days of AD, in which DIR may have been involved, where Fe(III) species serve as terminal electron acceptors for microbial metabolism [38]. Subsequently, soluble Fe²⁺ concentrations declined, suggesting re-oxidation, adsorption onto sludge flocs, or incorporation into secondary minerals. Concomitantly, the solid-phase Fe(II) content (Figure 4b) generally increased over time, implying a net accumulation of reduced Fe in the solid matrix.

Notably, the FeCl₃ and Poly-FeCl₃ groups exhibited relatively high soluble Fe²⁺ and Fe(II) throughout the process (the concentrations were 39.23 and 103.82 mg/L, and 36.39 and 80.35 mg/L, respectively). However, this larger soluble Fe(II) pool did not translate into enhanced methanogenesis. Instead, these groups showed suppressed CH₄ yields along with reduced SCOD, soluble protein and protease activity and decreased VFAs accumulation (Figure 1, Figure 2 and Figure 3). This suggests that while ferric Fe from soluble salts is readily bioavailable for DIR, this process competes with methanogenesis for electrons, and the simultaneous formation of dense flocs physically hinders OM hydrolysis, overriding any potential benefit from DIR-derived soluble Fe(II).

In contrast, the Fe₃O₄ group achieved the highest CH₄ production, although it showed only moderate soluble Fe²⁺ and Fe(II) accumulation of 25.34 and 44.62 mg/L, respectively, which were 23.91% and 23.70% higher than the control. This suggests that the promoting effect of Fe₃O₄ was not mainly due to the release of large amounts of free Fe²⁺. Instead, as a mixed-valence and conductive mineral, Fe₃O₄ likely enhanced electron transfer through surface Fe(III)/Fe(II) redox cycling. This pathway may allow electron transfer between acetogens and methanogens without strong Fe dissolution, thereby reducing electron competition caused by DIR [36,39]. In comparison, FeOOH and Fe₅HO₈·4H₂O released more soluble Fe²⁺ than Fe₃O₄, reaching 28.34 and 32.44 mg/L, respectively, but their CH₄ yields were lower. This may be because they lack intrinsic mixed valence and have lower conductivity, which limited sustained electron transfer compared with Fe₃O₄.

3.2.2. Electrochemical Evidence and Mineralogical Transformation of Fe

To explore the electron transfer capability of the different samples, CV was performed. As illustrated in Figure 4c, all Fe groups exhibited more pronounced redox peaks than the control, confirming that Fe addition enhances the overall redox activity. The CV curves of the FeCl₃ and Poly-FeCl₃ groups displayed strong redox peaks, consistent with their active soluble Fe²⁺ release. However, the imbalance between oxidation and reduction currents may imply that the Fe redox cycle is not fully coupled to methanogenesis. Conversely, the Fe₃O₄ group showed a more symmetrical and stable redox response, suggesting a reversible and sustainable Fe(III)/Fe(II) surface redox process. This balanced electrochemical behavior might help Fe₃O₄ act as a stable electron mediator, potentially substituting for c-type cytochromes to enhance electron transfer. In addition, the electrochemical impedance spectroscopy (EIS) results (Figure 4d) showed that Fe compounds had a clear effect on the system’s impedance. The Fe₃O₄ group had the lowest impedance, while the FeCl₃ group had the highest. Taken together with the cumulative methane production and CV curves, the low impedance in the Fe₃O₄ group indicated more efficient electron transfer and stronger interfacial redox activity. These features favored microbial methane production and may also have promoted OM hydrolysis and acidification. In contrast, electron transfer was limited in the high-impedance FeCl₃ group, leading to a significant decrease in methane production.

The mineralogical fate of the added Fe compounds after AD provides further evidence for these dynamic transformations (Figure 4e). Characteristic peaks for Fe₃O₄ appeared in the Poly-FeCl₃ group, indicating the neo-formation of Fe₃O₄ due to Fe(III) reduction and subsequent mineralization during AD. Similarly, the Fe₅HO₈·4H₂O treatment also showed the formation of Fe₃O₄, indicating a transformation from Fe(III)-rich phases toward more reduced and crystalline Fe-bearing minerals. The FeCl₃ group showed the formation of FeOOH and FeO phases. These mineralogical changes, together with the temporal variations in soluble Fe²⁺ and solid-associated solid-phase Fe(II), indicate that the added Fe compounds underwent continuous redistribution among dissolved, adsorbed, and mineral-bound Fe pools during AD. Such secondary mineralization is an important sign of Fe transformation in the biogeochemical Fe cycle. Fe(III)-bearing minerals may be involved in this process as redox-active intermediates, participating in Fe transformation and OM conversion through reversible Fe(III)/Fe(II) transformation, while also providing potential interfaces for extracellular electron exchange.

3.2.3. Mechanism of Different Fe Compounds and Prospect for Application

Figure 5 summarizes the mechanisms of different Fe compounds on AD of SS. Overall, Fe did not remain static. Instead, it went through a continuous process of Fe(III) reduction, Fe²⁺ release, liquid-phase migration, and solid-phase adsorption or precipitation, thereby affecting both electron transfer and OM conversion. The different effects of Fe compounds were mainly related to their influences on sludge structure, OM hydrolysis and release, OM degradation, and Fe(III)/Fe(II) transformation.

Based on the distinct transformation pathways (Section 3.2.1) and electrochemical properties (Section 3.2.2) of the added Fe compounds, the mechanisms underlying their different effects on AD can be elucidated in Figure 5. For the Fe salt groups (FeCl₃ and Poly-FeCl₃), inhibition was mainly caused by floc formation. Fe(III) hydrolysis products formed dense sludge flocs (D₅₀ increased to 60.81 and 78.68 μm, respectively; Figure 2a) that limited OM hydrolysis. Consequently, protease activity was suppressed (decreased by 26% and 35% in Figure 3a), resulting in the reduction of soluble protein and SCOD release (Figure 2b,c), and ultimately lowering CH₄ yields by 9.90% and 11.92% (Figure 1). Second, soluble Fe(III) was readily reduced via DIR and competed with methanogenesis for electrons. Although secondary mineralization (e.g., Fe₃O₄ formation) was observed (Figure 4e), the overall effect remained inhibitory. For the FeOOH and Fe₅HO₈·4H₂O groups, moderate enhancement was observed. These pure Fe(III) oxides did not induce strong flocculation (D₅₀: 46.86 and 54.84 μm; Figure 2a). Their reductive dissolution released soluble Fe²⁺ (Figure 4a) and drove Fe(III)/Fe(II) redox cycling, as indicated by enhanced CV redox peaks (Figure 4c). Consequently, CH₄ yields increased by 15.23% and 15.09% (Figure 1). However, electron transfer relied on dissolution–reprecipitation, which is less efficient than solid-state conduction.

For the Fe₃O₄ group, the strongest enhancement was achieved (CH₄ yield increased by 18.5% in Figure 1). Several mechanisms contributed: (i) Fe₃O₄ caused only mild flocculation (D₅₀: 57.15 μm; Figure 2a), thereby preserving greater OM accessibility; (ii) as a mixed-valence conductive mineral, Fe₃O₄ might have directly participated in extracellular electron transfer via surface Fe(III)/Fe(II) redox cycles, consistent with the balanced CV response (Figure 4c) and moderate soluble Fe²⁺ levels (Figure 4a); and (iii) Fe₃O₄ likely facilitated DIET by substituting for c-type cytochromes and working with conductive pili, thereby promoting propionate degradation, as evidenced by the lowest residual propionate (Figure 2i). Additionally, higher protease activity and coenzyme F₄₂₀ content (Figure 3a,b) indicated enhanced hydrolysis and methanogenesis.

The results also have practical implications for wastewater treatment plants. Fe salts, which are commonly used as coagulants, may adversely affect sludge AD because they promote sludge flocculation and limit OM solubilization. Therefore, the application of Fe-based coagulants should be carefully managed when sludge is subsequently used for AD. In contrast, Fe oxides, especially Fe₃O₄, enhanced methane production under the tested conditions, suggesting their potential as functional additives for improving methane recovery from SS. This study can provide theoretical guidance on the effects of Fe compounds on sludge AD. In addition, in terms of practical application, the Fe dosage applied in this study was within the same order of magnitude as the Fe levels commonly reported in sludge and those used in practical sludge treatment, indicating that the selected dosage is relevant to real-world applications.

To further clarify the relationship between our results and previous reports, a comparison between this study and relevant studies on Fe-assisted sludge AD is summarized in

Table 2. Comparison of this study with previous studies on Fe compounds during SS AD.

Fe Compounds	Dosage	Effect on Cumulative Methane Production	Reference
FeCl3	200 mg/L	Methane yield increased by 28.9%.	[32]
Poly-FeCl3	40 g/kg TS	Methane yield decreased by 20.0%.	[19]
Fe3O4	120 mg/L	Methane yield increased by 69.6%.	[40]
FeOOH	1675 mg Fe/L	Methane yield increased by 37.33%.	[41]
Fe5HO8·4H2O	4545 mg Fe/L	Methane yield decreased by 23.0%.	[42]
This study	40 mg Fe/g TS	FeCl3 and Poly-FeCl3 decreased methane yield by 9.90% and 11.92%, respectively. Fe3O4, FeOOH, and Fe5HO8·4H2O increased methane yield by 18.54%, 15.23%, and 15.09%, respectively.	/

. Previous studies mainly focused on a single Fe compound or a specific type of Fe additive, whereas the present study systematically compared soluble Fe salts and solid-phase Fe oxides under the same Fe dosage and digestion conditions. Overall, Table 2 indicates that the apparently contradictory effects of Fe compounds reported in the literature may be partly explained by differences in Fe chemical form, dosage, sludge properties, and digestion conditions, while the present study demonstrates under identical experimental conditions that Fe salts mainly inhibit AD, whereas Fe oxides promote AD.

4. Conclusions

This study demonstrated that different Fe compounds in SS had quite different effects on following AD. FeCl₃ and Poly-FeCl₃ increased sludge particle size and formed compact flocs, which restricted OM release and reduced protease activity. As a result, methane yields decreased by 9.90% and 11.92%, respectively. In contrast, Fe₃O₄, FeOOH, and Fe₅HO₈·4H₂O promoted OM release, with SCOD increasing by 12.71%, 8.99%, and 7.47%, respectively, and supported Fe(III)/Fe(II) cycling. As a result, methane production increased by 18.54%, 15.23%, and 15.09%, respectively.