Investigation of TiO₂ Nanoparticles Added to Extended Filamentous Aerobic Granular Sludge System: Performance and Mechanism

Abstract

The widely utilized TiO₂ nanoparticles (NPs) tend to accumulate in wastewater and affect microbial growth. This work investigated the impacts of prolonged TiO₂ NP addition to filamentous aerobic granular sludge (AGS) using two identical sequencing batch reactors (SBRs, R1 and R2). R1 (the control) had no TiO₂ NP addition. In this reactor, filamentous bacteria from large AGS grew rapidly and extended outward, the sludge volume index (SVI₃₀) quickly increased from 41.2 to 236.8 mL/g, mixed liquid suspended solids (MLSS) decreased from 4.72 to 0.9 g/L, and AGS disintegrated on day 40. Meanwhile, the removal rates of COD and NH₄⁺-N both exhibited significant declines. In contrast, 5–30 mg/L TiO₂ NPs was added to R2 from day 21 to 100, and the extended filamentous bacteria were effectively controlled on day 90 under a 30 mg/L NP dosage, leading to significant reductions in COD and NH₄⁺-N capabilities, particularly the latter. Therefore, NP addition was stopped on day 101, and AGS became dominant in R2, with an SVI₃₀ and MLSS of 48.5 mL/g and 5.67 g/L on day 130. COD and NH₄⁺-N capabilities both increased to 100%. Microbial analysis suggested that the dominant filamentous bacteria—Proteobacteria, Bacteroidetes, and Acidobacteria—were effectively controlled by adding 30 mg/L TiO₂ NPs. XRF analysis indicated that 11.7% TiO₂ NP accumulation made the filamentous bacteria a framework for AGS recovery and operation without NPs. Functional analysis revealed that TiO₂ NPs had stronger inhibitory effects on nitrogen metabolism compared to carbon metabolism, and both metabolic pathways recovered when NP addition was discontinued in a timely manner. These findings offer critical operational guidance for maintaining the stable performance of filamentous AGS systems treating TiO₂ NP wastewater in the future.

1. Introduction

Over the past decade, TiO₂ nanoparticles (NPs), as successful engineered nanomaterials, have been extensively applied in industrial production processes such as paper manufacturing, electronics, and photocatalytic materials [1]. Consequently, NPs at concentrations ranging from ng/L to mg/L have been frequently detected in aquatic environments [2,3]. NPs exhibit significant toxicity to microbial communities, including bacteria, algae, and protozoa, through multiple mechanisms, including cellular membrane damage, genetic material alteration, and disruption of energy metabolism [4,5,6]. The inevitable accumulation of NPs in wastewater systems poses dual threats: while short-term exposure may enhance sludge settling performance via improving flocculation [7], chronic exposure critically impairs essential nutrient removal functions, particularly nitrogen and phosphorus metabolisms [8]. This paradoxical effect creates substantial operational challenges for conventional activated sludge (CAS) when treating NP wastewater.

Aerobic granular sludge (AGS) has been widely implemented in wastewater treatment systems, from lab- to full-scale bioreactors, over the past decades, owing to its excellent settleability, abundant biomass, and high resistance loading [9,10,11]. However, maintaining long-term operational stability remains a significant challenge, with filamentous bacterial overgrowth emerging as a critical limiting factor. Previous studies have identified multiple parameters influencing filamentous bacterial proliferation, including sludge size, dissolved oxygen (DO), pH, and nutrient type [12,13,14]. Subsequently, various mitigation strategies have been developed to control filamentous bacterial growth. These approaches include controlling sludge size with a spiny aeration device or removing oversized granules [15]; improving organic loading rate, shear force, or salinity [16]; adding chemicals (NaClO, Cl₂, or H₂O₂), ballasting agents and coagulants, particulate substrates, and sludge micropowder [17,18,19]. Although these strategies have demonstrated efficacy, current research predominantly addresses filamentous proliferation in CAS systems rather than AGS. Recent studies have systematically investigated the effects of NP addition on normal AGS systems, focusing on granulation dynamics, operational stability, organic removal efficiency, ecotoxicological impacts, and microbial evolution. Results demonstrated that ZnO and CuO NPs had negative effects on nitrogen transformations [20,21] and phosphorus removal [22,23]. Adding TiO₂ NPs could improve sludge settleability and decrease nitrogen removal, and SiO₂ would decrease sludge settleability and have no effect on nitrogen removal [3]. Meanwhile, the co-existence of CuO and ZnO NPs at a low dosage of 1 mg/L had no significant effect on nitrogen removal performance, suggesting that single NPs possess higher toxicity than the mixtures [24]. Nano iron, nano zero valent iron (nZVI), and magnetic nano iron have also been employed to evaluate normal AGS response. Adding iron NPs enhances sludge granulation through microbial metabolic activity, extracellular polymeric substance (EPS) secretion, electron transfer characteristics, and bacterial community shifts [25,26,27], mitigating membrane fouling in the AGS-membrane system via iron oxide adsorption [28]. Remarkably, AGS has exhibited exceptional resilience to toxic metalloids (Se^IV/Te^IV), achieving simultaneous nutrient removal and biosynthesis of Se⁰/Te⁰ nanostructures [29,30] or AuNPs [31]. However, co-exposure to NPs and antibiotics caused synergistic inhibition of nutrient removal through combined damage to cell membranes, EPS matrices, and microbial consortia, ultimately altering nitrogen transformation pathways [24,32]. Despite these advances, the investigation of filamentous AGS exposed to 5–30 mg/L TiO₂ NP wastewater remains poorly understood.

The primary objective of this experiment was to investigate the effects of adding 5–30 mg/L TiO₂ NPs to extended filamentous AGS in SBRs using artificial wastewater. Filamentous bacterial inhibition, sludge characteristics, organic degradation, microbial community, and metabolic pathways were also examined in this study. NP enrichment in sludge was also detected using X-ray fluorescence (XRF) spectroscopy. It is hoped that this study will provide a reference for controlling filamentous bacteria in an AGS system using NP wastewater in the future.

2. Materials and Methods

2.1. Preparation of TiO₂ NPs

White TiO₂ NPs with a mean size of 40–50 nm (Figure 1A) were commercially procured from Xianfeng Nanomaterials Technology Co., Ltd. (Nanjing, China). The TiO₂ NP solution was prepared using the following method: Precise quantities of TiO₂ NPs (0.15, 0.3, 0.6, and 0.9 g) were individually weighed and dispersed into 300 mL of Milli-Q water, followed by 30 min of ultrasonication (25 °C, 100 W, 40 kHz) to obtain homogeneous suspensions each time (Figure 1B). The NP solution was immediately transferred to the plastic tank and diluted to 30 L with artificial wastewater. Finally, the concentrations of TiO₂ NPs in the experiment were 5, 10, 20, and 30 mg/L each time, respectively.

2.2. Operation of the SBRs

Two identical SBRs (R1 & R2) made of plexiglas with effective volumes of 3.6 L (H × D = 60 cm × 10 cm) were used in this work (Figure 2). R1 (the control) had no TiO₂ NP addition, whereas TiO₂ NPs were added to R2 at concentrations of 0–30 mg/L throughout the experiment. Two fine air pumps (PL60, RESUN, Chaozhou, China) were used to supply aeration to the bottom of the SBRs through air diffusers with a uniform airflow rate of 0.5 m³/h and a 50% volumetric exchange ratio. Each 240-min cycle consisted of the following steps: feeding for 5 min, aeration for 180 min, settling for 2 min, discharge for 5 min, and idle time for 48 min. All pumps and valves were controlled by programmable logic controllers (PLCs) in this experiment. The indoor temperature was maintained at 20 ± 5 °C using industrial-grade air conditioning (Haier, Qingdao, China), and the pH of wastewater was maintained at 7.0–7.5 using an appropriate amount of 10% HCl.

2.3. Wastewater and Seeding Sludge

R1 and R2 were fed with acetate-containing synthetic wastewater. Artificial wastewater is mainly composed of sodium acetate (NaAc), NH₄⁺-N, and KH₂PO₄, serving as carbon, nitrogen, and phosphorus sources, respectively. The detailed composition of the synthetic wastewater was as follows: NaAc 600 mg/L, NH₄Cl 100 mg/L, KH₂PO₄ 60 mg/L, FeSO₄·7H₂O 8 mg/L, CaCl₂ 21 mg/L, MgSO₄·7H₂O 102 mg/L, and trace element solution 0.5 mL/L. Subsequently, influent COD, NH₄⁺-N, and total phosphorus (TP) were maintained at 400–500 mg/L, 20–25 mg/L, and 10–15 mg/L, respectively. The wastewater was manually prepared once every three days and then stored in a sealed plastic tank with a pump for stirring. The composition of the trace elements can be found in an early work [33]. The seeding sludge was AGS with an average particle size of 3 mm, obtained from our lab-scale SBR (Figure 3A,E).

2.4. Analytical Methods

2.4.1. Analysis of Water Quality and Sludge Properties

COD, NH₄⁺-N, ML(V)SS, and SVI (SVI₅ and SVI₃₀) values were measured using standard methods [34]. Sludge morphological changes were observed using an optical microscope (NE600, GREEN, Nanjing, China) and smartphone (Magic 6pro, Honor, Shenzhen, China). EPS were extracted using the formaldehyde–NaOH method [35]. Polysaccharide (PS) concentration was determined using the phenol–sulfuric acid method with glucose as the standard [36]. The protein (PN) concentration was determined using Coomassie brilliant blue with bovine serum albumin (BSA) as the standard [37]. Sludge samples S0, S2, S3, S4, and S6 were first oven-dried at 105 °C for 2 h, then manually ground into sludge powders and put into 10 mL centrifuge tubes. These powders were used to detect elemental composition using XRF.

2.4.2. Analysis of Microbial Community

Seven sludge samples (S0–S6) from R2 were regularly collected on day 0, 20, 40, 60, 80, 100, 110, and 130. Each sample was randomly collected three times and then mixed into one tube (10 mL). All tubes were promptly stored in a refrigerator at −20 °C for microbial analysis. All sludge samples were sent to Major Bio-pharm Technology Co., Ltd. (Shanghai, China) for microbial analysis. DNA extraction was performed using the E.Z.N.ATM Mag-Bind Soil DNA Kit (OMEGA, Norcross, GE, USA) according to the manufacturer’s instructions. Microbial analysis was conducted at this company using the Illumina Nextseq2000 platform (Illumina, San Diego, CA, USA). About 10 ng DNA of the sludge sample was used for 16S rDNA sequencing targeting the hypervariable V3–V4 regions. The data were processed using the Quantitative Insights into Microbial Ecology (Qiime) (v 1.91) pipeline (http://qiime.org/ (accessed on 26 October 2024)) and Mothur (v1.30.2). Bacterial analysis was carried out based on high-quality sequences with a similarity threshold of 97%. Bioinformatic analysis of the sludge was carried out using the Major-bio cloud platform (https://cloud.majorbio.com (accessed on 26 October 2024)). Operational taxonomic units (OTUs) and coverage were calculated using Mothur (v1.30.2). Microbial dynamic analysis was also conducted on the phylum and genus levels. The similarity among the microbial communities in different samples was determined using principal coordinate analysis (PCoA) and non-metric multi-dimensional scaling (NMDS) based on the OTUs using an unweighted unifrac distance. Furthermore, microbial metabolic pathways in the S0–S6 samples were predicted using PICRUSt2 (v2.2) with reference to the Kyoto Encyclopedia of Genes and Genomes (KEGG) database.

3. Results and Discussion

3.1. Variations in Sludge Characteristics

Figure 3 shows the morphological evolution of sludge in R1 and R2. Both SBRs were initially inoculated with lab-cultured AGS with a mean diameter of 3 mm, showing filamentous structures (Figure 3A,E). Filamentous bacteria grew quickly in R1 due to the easily biodegradable carbon resource NaAC, coinciding with an increase in AGS size (Figure 3B). Subsequently, excessive filamentous growth proliferated and extended outward from large AGS on day 20 (Figure 3C), and in the absence of effective action, the filamentous bacteria became dominant, leading to AGS disintegration on day 30 (Figure 3D). These results confirm that NaAC promotes filamentous bacterial growth [16,19]. Conversely, R2 was fed with synthetic wastewater containing 5–30 mg/L TiO₂ NPs from day 21 to 100. While TiO₂ NP concentrations of 5–20 mg/L showed limited effect on filamentous bacterial growth (Figure 3F,G), increasing the dosage to 30 mg/L on day 80 resulted in filamentous bacterial suppression and gradual AGS recovery on day 90 (Figure 3H). After NP supplementation was terminated on day 101, the extended filamentous bacteria were eliminated, and well-structured AGS re-emerged on day 130 (Figure 3I). This demonstrated that 30 mg/L TiO₂ NPs effectively inhibited extended filamentous bacterial overgrowth, maintaining AGS stability. These experimental results clearly established that filamentous bacteria are sensitive to high dosages of TiO₂ NPs [3,28].

Correspondingly, MLSS and SVI were systematically monitored in both SBRs (Figure 4). In R1, the uncontrolled proliferation of extended filamentous bacteria (Figure 3A–D) was correlated with dramatic changes in MLSS and SVI₃₀. MLSS declined from 4.72 to 0.61 g/L and SVI₃₀ increased from 41.2 to 236.8 mL/g between day 0 and 40 (Figure 4A). Ultimately, a large amount of sludge was washed out from R1, leading to AGS collapse and indicating that R1 failed. This was well in agreement with previous studies [14,23]. In order to maintain operational stability, R2 was fed with artificial wastewater containing 5–30 mg/L TiO₂ NPs from day 20 to 100 (Figure 4B). The initial 5 mg/L NP supplementation failed to prevent SVI₃₀ elevation or MLSS reduction from day 0 to 40; specifically, SVI₃₀ increased from 41.5 mL/g to 114 mL/g, and MLSS slowly decreased from 4.75 g/L to 4.46 g/L in stages a and b (Figure 4B). With increasing NP dosage, SVI₃₀ presented a downward trend and MLSS showed an upward trend in stages c and d (Figure 4B). When the NP concentration reached 30 mg/L, the excessive proliferation of the extended filamentous bacteria was restrained, and AGS gradually recovered on day 90 (Figure 3H). After TiO₂ NP addition was stopped, AGS with a compact structure and regular boundaries appeared again on day 130 (Figure 3I). In the end, both MLSS and SVI₃₀ in R2 stabilized at 5.67 g/L and 48.5 mL/g (Figure 4B). Interestingly, the curve of the MLVSS/MLSS ratio showed a descending trend from day 0 to 100 and then an ascending trend from day 101 to 130 (Figure 4B). These findings demonstrated that low concentrations of NPs (<30 mg/L) were ineffective in suppressing filamentous bacteria, whereas higher concentrations (≥30 mg/L) successfully inhibited filamentous bacterial overgrowth. However, the increased NP dosage negatively impacted the MLVSS/MLSS ratio, likely due to the sedimentation of a portion of NPs in the sludge [3,26].

The EPS secreted by bacteria plays an important role in maintaining the AGS structure for stable reactor performance. The concentrations of EPS, PN, and PS were quantified throughout the experiment (Figure 5A,B). Generally, EPS, PN, and PS in R1 still decreased from 64.9, 30.1, and 34.8 mg/(g·VSS) to 40, 12.3, and 37.7 mg/(g·VSS) on day 40, respectively, and the PN/PS ratio decreased from 0.86 to 0.67 (Figure 5A), coinciding with filamentous bacterial overgrowth (Figure 3B–D), a quick increase in SVI₃₀, and a rapid decrease in MLSS (Figure 4A). Finally, this imbalance accelerated sludge disintegration [19,37]. In contrast, when 5–30 mg/L TiO₂ NPs were added to R2, EPS, PN, and PS all increased from 64.8, 30.2, and 34.6 to 106.3, 32.8, and 73.5 mg/(g·VSS) on day 90, respectively (Figure 5B). This was due to microbes producing more EPS to protect themselves from the toxic effects of TiO₂ NPs [38]. After the extended filamentous bacteria were restrained (Figure 3H), NP addition was stopped on day 101. EPS gradually decreased and then stabilized at 92.4 mg/(g·VSS) on day 130. Meanwhile, the PN/PS ratio reached 0.98 (Figure 5B). PN provided a steric structure, and PS as a biogel facilitated AGS structural reintegration [3,38]. Subsequently, AGS with an integrated structure recovered well on day 130 (Figure 3I). The study revealed that NP dosage stimulated microbial EPS secretion as a protective response against TiO₂ NP toxicity [38,39].

The elements of the sludge samples (S0, S2–S4, S6) from R2 were analyzed using XRF. As shown in Figure 5C, the seeding sludge (S0) contained the main elements of Fe (22.13%), Ca (22.7%), Zn (13.12%), Mn (4.83%), S (4.70%), Cl (4.61%), and Mg (1.95%). Notably, the proportion of Ti quickly increased from 0.83% (S0) to a peak of 33.39% (S4) following the addition of 5–30 mg/L TiO₂ NPs between day 21 and 100 (Figure 3). This value finally decreased to 11.7% (S6) with the cessation of NP dosing between day 101 and 130 (Figure 5C). This confirmed that TiO₂ NPs sedimented in the sludge [3], resulting in a lower MLVSS/MLSS ratio in stages b–e (Figure 4B). Concurrently, Fe, Ca, and Zn content declined from 22.13%, 22.7%, and 13.12% (S0) to 19.92%, 11.69%, and 0.27% (S4), and then rose to 24.74%, 15.31%, and 0.39% (S6), respectively. These elements contributed to aggregation and EPS production, facilitating AGS structural recovery [19]. This indicated that NPs reduced Fe, Ca, and Zn metal deposition in AGS. In summary, these findings demonstrated that TiO₂ NPs induced significant Ti sedimentation while suppressing Fe, Ca, and Zn accumulation. High NP dosage altered metal sedimentation in the AGS system [3,30,40], highlighting the need for NP dosage optimization in practice.

3.2. Performance of Organic Degradation in Both SBRs

The performance of COD and NH₄⁺-N degradation processes in both SBRs was systematically monitored throughout the experiment (Figure 6). Generally, the influent COD and NH₄⁺-N in the two SBRs remained stable within the range of 420–490 mg/L and 21–25 mg/L, respectively. Effluent COD and NH₄⁺-N showed similar changes from day 0 to day 20, but significant differences from day 21 to day 130 (Figure 6). Specifically, the effluent COD and NH₄⁺-N in R1 increased rapidly from 0 mg/L to 180 and 11.9 mg/L on day 40 (Figure 6A). Correspondingly, the removal rates of COD and NH₄⁺-N also quickly decreased from 100% to 58.5% and from 100% to 43.6% because the uncontrolled filamentous bacteria grew excessively in R1 (Figure 3A–D), resulting in impaired sludge settling performance. Subsequently, a large amount of biomass with high SVI₃₀ was washed out, and MLSS greatly reduced in R1 (Figure 4A). These findings indicate that extended filamentous bacteria have negative impacts on sludge properties and organic degradation [16,17,19].

Compared to R1, R2 demonstrated superior organic degradation performance (Figure 6B). Both effluent COD and NH₄⁺-N in R2 showed an increasing tendency under the dual actions of filamentous bacterial over-proliferation and the addition of 5–30 mg/L NPs, reaching peak values of 71.2 and 10.8 mg/L on day 100. With the termination of NP addition on day 101, both values decreased to 0 mg/L (Figure 6B), and AGS recovered well on day 130 (Figure 3I). Correspondingly, COD and NH₄⁺-N removal rates in R2 decreased from 100% on day 0 to 83.6% and 54.4% on day 100, and then increased again to 100% on day 130 (Figure 6B). High concentrations of NPs (≥30 mg/L) restrained extended filamentous bacterial overgrowth, thereby preventing sludge deterioration (Figure 3E–I). However, the accumulated TiO₂ NPs (Figure 5C) exhibited toxic effects on the microbial community, especially on nitrifying or denitrifying bacteria [3,22]. This experiment revealed that long-term exposure to TiO₂ NPs had a more suppressive effect on ammonia nitrogen degradation than on carbon reduction [22,24,41].

3.3. Microbial Composition and Functional Groups

Seven sludge samples (S0–S6) from R2 were analyzed using 16S rRNA to characterize the microbial community on the phylum and genus levels in this experiment. A total of 54,088–58,518 high-quality reads with a mean length of 411–421.6 bp were obtained, which were clustered into 213–290 OTUs with 0.99 coverage. As shown in Figure 7A, seven phyla were consistently detected in all sludge samples. The predominant phyla in the seeding sludge (S0) were Firmicutes, Proteobacteria, Bacteroidetes, and Actinobacteria, with relatively high abundances of 0.56, 0.21, 0.07, and 0.06, respectively. These microbes are commonly found in AGS and CAS and are known for their roles in EPS production and organic degradation [19,22,33]. The percentage of Proteobacteria initially rose to 0.37 (S1) in the absence of TiO₂ NPs from day 0 to 20, then fell to 0.21 (S4) with the addition of 5–30 mg/L NPs from day 21 to 100 (Figure 7A). Following NP withdrawal from day 101 to 130, its proportion increased and stabilized at 0.51 (S6), as extended filamentous bacterial overgrowth was inhibited and AGS recovered completely on day 130 (Figure 3I). This was consistent with previous reports [3,19,21]. Similar trends were observed for Actinobacteria, Bacteroidetes, and Verrucomicrobia. Proteobacteria, Acidobacteria, and Bacteroidetes are recognized as typical filamentous bacteria and commonly associated with sludge bulking [16,19]. They were initially controlled by the timely addition of TiO₂ NPs and subsequently served as structural frameworks for AGS regranulation (Figure 3H,I) [3,20]. In contrast, Firmicutes abundance rapidly decreased to 0.16 in the absence of NPs from day 0 to 21(S1), rebounded to 0.48 (S4) with increasing NP dosage (5–30 mg/L) from day 21 to 100, and then sharply declined, stabilizing at 0.01 (S6) as AGS recovered on day 130 (Figure 7A). These results demonstrate that the predominant filamentous bacteria Proteobacteria, Acidobacteria, and Bacteroidetes were sensitive to TiO₂ NP addition [3,22].

The top 37 genera were found in all sludge samples in this work (Figure 7B). Exiguobacterium, which belongs to Firmicutes, became the dominant genus in the seeding sludge with a relative abundance of 0.54 (S0), then decreased following the addition of 5–30 mg/L TiO₂ NPs from day 21 to 100 (Figure 7B). The lowest abundance was 0.0009 (S4) when the NPs dosage reached 30 mg/L on day 100. When NP addition stopped, its abundance increased to 0.46 (S5) and remained at a low level (S6) after large AGS had recovered with a compact structure. This indicated that NP dosage and sludge structure played a crucial role in Exiguobacterium proliferation [22]. Abditibacterium abundance showed a progressive increase during NP addition (S0–S4) from day 21 to 100, and decreased to 0.08 (S6) after AGS recovered on day 130. Conversely, Leucobacter (a genus of Actinobacteriota) exhibited a gradual decline upon exposure to 5–30 mg/L TiO₂ NPs (S0–S4), then increased and stabilized at 0.19 (S6) after NP addition was suspended on day 101. This was consistent with the literature reports of its prevalence in industrial wastewater [22,24,42]. Flavobacterium had a relatively low abundance of 0.02 in the seeding sludge (S0), and this value decreased progressively upon exposure to 5–30 mg/L NPs (S1–S4). Finally, it remained at 0.006 (S6) when NP addition stopped and AGS recovered on day 130. Flavobacterium, as a typical filamentous bacterium, is often detected in sludge bulking and is extremely sensitive to TiO₂ NPs [19,20].

Beta diversity analysis was also carried out using both PCoA and NMDS based on unweighted unifrac distance (Figure 8A,B). As shown in Figure 8A, the PCoA results revealed significant microbial community differences, with principal components PC1 and PC2 explaining 45.73% and 18.39% of the total variance, respectively. Notably, the inoculated sludge (S0) displayed a remarkable gap compared with the other samples (S1–S6). While sludge was exposed to 5–20 mg/L TiO₂ NPs, the S1–S3 samples clustered relatively close together. When the NP dosage increased to 30 mg/L, S4 exhibited marked divergence. Interestingly, even after NP addition ceased on day 101, the S5 and S6 samples still remained significantly distinct (Figure 8A). Correspondingly, the NMDS analysis (Figure 8B) yielded patterns consistent with the PCoA results (Figure 8A). The seeding sludge S0 exhibited significant separation from the subsequent samples, reflecting distinct microbial community structures shaped by extended filamentous bacterial proliferation (S2) and TiO₂ NP addition (S3, S4). Even after NP cessation on day 101, the S4 and S5 samples demonstrated persistent divergence. This suggested that TiO₂ NP exposure induced irreversible alterations in bacterial community structure [42]. Consequently, the S0–S6 samples were divided into five distinct clusters I (S0), II (S1–S3), III (S4), IV (S5), and V (S6)—according to the PCoA and NMDS results (Figure 8).

Figure 9 presents the heatmap analysis of functional metabolisms related to carbon and nitrogen degradation pathways. The ranges of hydrocarbon degradation functional groups involving aliphatic non-methane, aromatic, and other hydrocarbons increased from 173–343 (S0) to a peak value of 5716–5721 (S4) as TiO₂ NP concentration rose from 0 to 30 mg/L during days 21–100. However, the values decreased and stabilized at 1233–1234 (S6) after NP cessation. Notably, other hydrocarbon degradation functional groups related to aerobic chemoheterotrophy and general chemoheterotrophy showed consistent increases throughout the experiment (S0–S6). In contrast, nitrite reduction functional groups (nitrate/nitrite/nitrous oxide denitrification, nitrate/nitrite/nitrous respiration, and nitrate) initially declined from 724–766 (S0) to 74–124 (S3), and then rose to the range of 1696–1988 (S6). These findings revealed that TiO₂ NPs exerted more pronounced inhibitory effects on NH₄⁺-N removal than on COD removal [3,24,41]. These metabolic patterns were corroborated by the COD and NH₄⁺-N degradation profiles shown in Figure 6B.

3.4. Proposed Model of TiO₂ NPs for Extended Filamentous AGS

This study proposed a model for controlling extended filamentous bacterial overgrowth to prevent AGS disintegration by adding TiO₂ NPs (Figure 10). The filamentous bacteria rapidly grew in large AGS fed with artificial wastewater containing easily degradable substances (NaAC) (Figure 3A–C), coinciding with deteriorating sludge characteristics (MLSS, SVI, EPS) in the absence of TiO₂ NPs (Figure 4A and Figure 5A). Finally, AGS disintegrated (Figure 3D), leading to poor organic degradation (Figure 6A) [12,13,16,19]. The dominant filamentous bacterial phyla—Proteobacteria, Acidobacteria, and Bacteroidetes—all increased as 5–30 mg/L TiO₂ NPs were added in a timely manner to the bioreactor between days 21 and 100. TiO₂ NP sedimentation reached 11.7% (Figure 5C), making these phyla a framework for granular sludge recovery [3,26,27,40], even though NP addition terminated on day 101. Meanwhile, Ca, Fe, Mg, and Mn accumulation (Figure 5C) enhanced sludge reaggregation in this study [33,38,39]. Importantly, TiO₂ NPs demonstrated stronger inhibition on nitrogen transport pathway than on carbon metabolism (Figure 9), resulting in worse NH₄⁺-N removal performance compared to COD removal (Figure 7B) [22,23,24,25]. After the cessation of NP addition, AGS recovered with a regular shape (Figure 3I), better sludge properties relating to MLSS, SVI, EPS (Figure 4B and Figure 5B), and enhanced organic removal performance on day 130 (Figure 6B). This was because nitrogen and carbon metabolisms enhanced after NP addition was stopped between days 101 and 130 (Figure 9). However, microbial community analysis confirmed irreversible structural alterations in the presence of TiO₂ NPs in the experiment (Figure 7 and Figure 8).

4. Conclusions

On day 30, the filamentous bacteria proliferated excessively and extended outward from large AGS fed with artificial wastewater. But they were effectively inhibited by adding 30 mg/L TiO₂ NPs. Correspondingly, the abundances of predominant filamentous bacteria—Proteobacteria, Acidobacteria and Bacteroidetes—decreased, and AGS with a regular shape gradually recovered on day 130. Notably, 11.7% TiO₂ NP sedimentation enabled the filamentous bacteria to serve as a framework for granular sludge reformation. The accumulated metal ions Ca, Mg, Fe, and Mn enhanced sludge regranulation after NP addition stopped on day 101. COD and NH₄⁺-N removal rates both showed decreasing trends upon exposure to 5–30 mg/L TiO₂ NPs; in particular, the latter showed a significant change under 30 mg/L NPs. This is because high concentrations of TiO₂ NPs exerted stronger suppression on nitrogen metabolism than on carbon metabolism. Cessation of NP addition restored metabolic functions, improving organic degradation. Briefly, TiO₂ NPs positively controlled filamentous bacterial overgrowth and negatively impacted organic degradation. Therefore, one needs to comprehensively consider filamentous bacterial inhibition and treatment efficiency when applying the filamentous granular sludge process to TiO₂ NP-containing wastewater in the future.