Addition of Heterotrophic Nitrification and Aerobic Denitrification Bacterial Agents to Enhance Bio-Nests Treating Low Carbon-to-Nitrogen Ratio Municipal Wastewater

Abstract

Municipal wastewater with a low carbon-to-nitrogen (C/N) ratio presents challenges for conventional nitrogen removal processes, often requiring costly external carbon sources. This study investigated the enhancement of nitrogen removal in a simultaneous nitrification and denitrification (SND) system by incorporating heterotrophic nitrification and aerobic denitrification (HN-AD) bacterial agents (Klebsiella variicola L3, Acinetobacter beijerinckii W4, and Acinetobacter sp. Z1) with modified basalt fiber carriers. Three reactors were compared: mixed HN-AD strains (M), mixed strains with activated sludge (A+M), and activated sludge alone (A). Results demonstrated that the A+M reactor achieved superior performance, with median removal efficiencies of 82.2% for NH₄⁺-N, 52.9% for total nitrogen (TN), and 51.6% for COD, outperforming the M reactor (75.2%, 43.6%, and 51.6%) and the A reactor (63.2%, 29.3%, and 44.8%). The A+M reactor also exhibited a 40% reduction in COD consumption per unit TN removed (2.55 ± 1.75) compared to the control reactor A (4.25 ± 3.99). Microbial analysis revealed Acinetobacter sp. Z1 (6.1%) and K. variicola L3 (1.1%) as dominant species, with the A+M reactor showing higher microbial diversity (56.4% Proteobacteria, 10.2% Bacteroidota) and biological viability (VSS/SS ratio of 0.70 ± 0.01). Extracellular polymeric substance (EPS) content in A+M reached 242.26 ± 15.52 mg/g-VSS, with a protein-to-polysaccharide ratio of 2.77 ± 0.00, indicating robust biofilm activity. These findings highlight the potential of HN-AD bacterial agents to enhance nitrogen removal in low C/N wastewater treatment, offering a cost-effective and sustainable alternative to traditional methods by reducing reliance on external carbon sources and improving system efficiency.

1. Introduction

Currently, municipal wastewater entering wastewater treatment plants generally has a characteristic of low carbon-to-nitrogen ratio [1]. It primarily originates from residential life, including toilet, kitchen, laundry, and other types of wastewaters. These wastewaters have a high nitrogen content, mainly from urea and ammonia nitrogen in urine and feces, while the content of organic matter is relatively low, resulting in a low carbon-to-nitrogen ratio. The processes commonly used by most wastewater treatment plants to treat such wastewater are typically oxidation ditches, anoxic–oxic (A/O), and anaerobic-anoxic–oxic (A²/O). To meet municipal wastewater discharge standards, these processes usually require the addition of a large amount of carbon sources in the anoxic phase, including glucose, methanol, acetic acid, and so on [2]. According to engineering practice, the cost of carbon sources accounts for about 20% of the operating costs of wastewater treatment plants.

Some research and engineering practices have begun to explore the introduction of low-carbon or zero-carbon processes into wastewater treatment. The most typical process is anaerobic ammonia oxidation (Anammox). While Anammox offers a carbon-efficient alternative, its application is limited by strict control requirements (DO, temperature, sludge age) and incompatibility with ammonium-dominated wastewater [3].

In contrast, the simultaneous nitrification and denitrification (SND) process is more practical. It allows nitrification and denitrification to occur within the single reactor, utilizing the metabolic characteristics of microorganisms to reduce the need for additional carbon sources. Using the SND process, about 20% to 40% of the carbon source can be saved compared to the A²/O process [4]. But their performance is constrained by slow-growing autotrophic nitrifiers and competition from heterotrophs in biofilms [5]. Theoretically, the growth and metabolism of autotrophic nitrifying bacteria do not require organic carbon; however, the presence of organic carbon in municipal wastewater makes it difficult for them to become the dominant microbial community in the biofilm [6]. Additionally, the effective thickness of the biofilm is generally only about 2 mm, which further limits the species and quantity of autotrophic nitrifying bacteria, thereby affecting the nitrogen removal efficiency of the reactor.

Recently, an increasing number of studies have begun to focus on heterotrophic nitrification and aerobic denitrification (HN-AD) bacteria. These bacteria are capable of converting various forms of nitrogen into nitrogen gas in an aerobic environment through a heterotrophic mode. To date, eight genera of HN-AD bacteria have been identified, and their main characteristics are as follows: (1) they exhibit significantly higher growth rates than autotrophic nitrifying bacteria; (2) they have a wide range of utilizable carbon sources; and (3) they can be enriched and cultured in different environments [7]. However, most studies focus on single-strain inoculations [8], which lack the functional diversity needed for real wastewater matrices. Additionally, the long-term stability of HN-AD biofilms and their interactions with native activated sludge communities remain poorly understood. Prior work has not systematically evaluated how mixed HN-AD strains enhance carbon utilization efficiency in low C/N systems. The combined effects of HN-AD bacteria and activated sludge on biofilm dynamics, EPS production, and nitrogen pathway regulation are underexplored.

This study aims to evaluate the efficacy of a multi-strain HN-AD bacterial consortium (K. variicola L3, A. beijerinckii W4, and Acinetobacter sp. Z1) in enhancing nitrogen removal from low C/N municipal wastewater when integrated with activated sludge and modified basalt fiber carriers. By assessing pollutant removal performance, microbial community dynamics, and metabolic mechanisms, we seek to complete the following: (1) quantify improvements in nitrogen and COD removal efficiency compared to conventional systems; (2) elucidate the synergistic interactions between HN-AD bacteria and native sludge microbiota; and (3) provide a scalable, carbon-efficient solution to reduce reliance on external carbon dosing in real-world wastewater treatment. This work addresses critical gaps in bioaugmentation strategies for low C/N wastewater, offering insights for sustainable process optimization.

2. Materials and Methods

2.1. Seed Sludge, HN-AD Strains, and Wastewater

The seed sludge was collected from an aeration tank of a municipal treatment plant, and after cultivation under laboratory conditions, it was added to the reactor. The three HN-AD bacterial strains used in this study, K. variicola L3, A. beijerinckii W4, and Acinetobacter sp. Z1, were all isolated and identified by our laboratory. They have major differences in their physiological conditions, and co-culturing and adding them to the bioreactors could give full play to their respective advantages. Through synergistic effects, the nitrogen removal efficiency can be improved. The wastewater used in the experiment was real domestic wastewater, collected from the drainage outlet of an anaerobic fermentation tank in an urban community. The specific water quality parameters are shown in

Table 1. Water quality index and their concentrations in real domestic wastewater.

Index	NH3+-N	NO3−-N	NO2−-N	TN	COD
(mg/L)	70.0 ± 19.6	2.0 ± 1.8	0.3 ± 0.5	110.3 ± 34.2	203.2 ± 57.8

.

2.2. Experimental Reactor Setup and Operating Conditions

The reactor was constructed from acrylic plates with an effective volume of 15.7 L. Water entered the reactor from the bottom inlet and exited through the top outlet. On the side near the outlet, a carrier medium was suspended, which consisted of modified basalt fiber (MBF) twisted with double-strand titanium wires (15 cm in diameter and 3 cm in length). The aerator was placed on the side of the inlet, having the function of thoroughly mixing air, sludge, and water, and, by creating circulating flow conditions, it provided oxygen and other nutrients for the bio-nest formed on the MBF carrier medium (Figure 1). This study set up three reactors, which were the mixed strain control group (M), the mixed strain + activated sludge experimental group (A+M), and the activated sludge control group (A). Before starting the reactors, 3 L of amplified HN-AD mixed bacterial agents (with a volume ratio of 1:1:1) was added to the reactors M and A+M, followed by the addition of 12 L of synthetic wastewater for cultivation. After one week, a large number of HN-AD bacteria was observed to be attached to the MBF carrier. At this point, aeration was stopped in the A+M reactor for sedimentation, and after removing 3 L of supernatant, an equal volume of activated sludge was added (both mixed bacterial agents and activated sludge having the same mixed liquor suspended solids, ca. 6120 mg/L). Meanwhile, 3 L of activated sludge and 12 L of simulated wastewater were added to reactor A for cultivation. Subsequently, the effluent water quality was tested, and, when the effluent quality stabilized, real domestic wastewater was used for continued cultivation with dissolved oxygen (DO) maintained at 3~4 mg/L, HRT set at 36 h, temperature controlled between 25 and 30 °C, and pH value controlled between 7.0 and 7.5. The reactors were operated continuously for about 80 days. During the experiment, to ensure the species abundance of HN-AD bacteria, 900 mL of bacterial agents was added to reactor A+M every 10 days.

2.3. Biological Viability

To investigate the biological viability within the reactors, after the experiment was completed, the activated sludge from the bio-nest was mixed uniformly with the suspended sludge, then gently rinsed with sterile PBS (phosphate buffered saline) to remove impurities. A 100 μL sludge sample was fixed on a glass slide, and an equal volume of the mixed dyes (SYTO9 + PI) solution was added to the sample surface. After thorough immersion, the sample was gently rinsed with PBS to remove unbound dyes, followed by incubation in the dark for 15 min. After incubation, the stained samples were observed using a confocal laser scanning microscope (CLSM, TCS SP5, Leica Microsystems, Wetzlar, Germany). The excitation wavelength for SYTO9 was ~488 nm, and the emission wavelength was ~500–550 nm; the excitation wavelength for PI was ~535 nm, and the emission wavelength was ~600–650 nm. Fluorescent signals from SYTO9 (green, indicative of live cells) and PI (red, indictive of dead cells) were collected separately, and the images were processed and analyzed using the microscope’s accompanying software [9]. Furthermore, the mixed liquor suspended solids (MLSS) and mixed liquor volatile suspended solids (MLVSS) in the activated sludge were determined by the weight method. Their combination can be used to reflect the sludge activity in the reactor.

2.4. Extracellular Polymeric Substance Extraction

Extracellular polymeric substance (EPS) extraction was performed using the heating method. Fifty milliliters of sludge suspension was first collected and placed in a test tube, followed by centrifugation at 3000 rpm for 10 min, and then at 4000 rpm for an additional 3 min. The supernatant was subsequently removed, and 30 mL of phosphate buffer solution (PBS: 0.01 M) was added to the sample, which was then mixed thoroughly. The sample was placed in a water bath at a constant temperature of 80 °C and heated continuously for 1 h before being cooled to room temperature. Finally, the sample was centrifuged at a speed of 10,000 rpm for 30 min, and the supernatant was collected and filtered through a 0.22 μm membrane filter, and the filtered liquid was dried for 48 h by lyophilization [10]. The polysaccharides (PSs) and proteins (PNs) in the extracellular polymeric substances were measured for their content using the anthrone–sulfuric acid method and the Bradford method, respectively [11].

2.5. DNA Extraction, Metagenomic Sequencing, and Analyses

At the end of the experiment, the sludge on the bio-nests was thoroughly mixed with the suspended sludge in the reactors. An appropriate amount of the mixed sludge from each group was centrifuged to remove the supernatant, with the precipitate preserved at low temperature (−80 °C). Total genomic DNA was extracted using the FastDNA^® Spin Kit (MP Biomedicals, Irvine, CA, USA) following the manufacturer’s instructions. The purity and concentration of the extracted DNA were checked using the agarose gel electrophoresis and the Qubit 2.0 Fluorometer (Life Technologies, Carlsbad, CA, USA). The qualified DNA was fragmented into 350 bp segments using an ultrasonic disruptor (Covaris M220), followed by end repair, A-tailing, adapter ligation, purification, and PCR amplification to complete the library construction. The insert size of the library was checked using an Agilent 2100 bioanalyzer. All libraries were sequenced on an Illumina PE150 platform by Novogene Co, Ltd., (Tianjin, China). The resulting raw reads were assessed through quality control and host filtering to obtain clean data, which were assembled to scaftigs using MEGAHIT (v1.1.2). The open reading frames from scaftigs were predicted using MetaGeneMark, while the microbial taxonomy annotation was performed based on blast result against NCBI NR database (https://www.ncbi.nlm.nih.gov/ (accessed on 16 February 2024)) using DIAMOND (v2.1.8) software. The raw data are available in the NCBI Sequence Read Archive (SRA) repository under the accession numbers SUB15433103.

2.6. Analytical Methods

NH₄⁺-N, NO₂⁻-N, NO₃⁻-N, and total nitrogen (TN) were determined using Nessler’s method [12], N-(1-Naphthyl) ethylenediamine dihydrochloride method [13], phenol disulfonic acid method [14], and alkaline potassium persulfate method [15], respectively. Chemical oxygen demand (COD) was measured using the dichromate method [16]. All water samples needed to be filtered through a 0.22 μm membrane filter before measurement. DO and pH were measured using a portable multi-parameter meter (HQ30d, HACH, Loveland, CO, USA). Additionally, all the experiments were performed in duplicate, and the results were presented as mean ± standard deviation using Origin 2022.

3. Results

3.1. Pollutant Removal Capability and Carbon Source Utilization

After one month of acclimation, the actual wastewater was continuously fed to these reactors. The nitrogen source in the influent mainly existed in the form of ammonia nitrogen (70.94 ± 19.59 mg/L) and organic nitrogen (37.17 ± 36.42 mg/L), with almost no nitrate and nitrite nitrogen present. The carbon source content in the influent (expressed as COD) was only 203.16 ± 57.86 mg/L, resulting in a very low carbon-to-nitrogen ratio (2.01 ± 0.84) in the influent. The high content of organic nitrogen was due to the low microbial activity under winter conditions, which slowed down the ammonification process finished by ammonifying bacteria. The treatment results showed that the removal efficiency of NH₄⁺-N (expressed as the median) was highest in reactor A+M, reaching 82.2%, followed by reactor M (75.2%) and reactor A (63.2%). Correspondingly, similar differences were observed in the removal efficiencies of TN and COD: 52.9%, 43.6%, and 29.3%; 51.6%, 51.6%, and 44.8%, respectively (Figure 2A). Clearly, under the same MLSS concentration, the mixed bacteria system (M) exhibited significantly better nitrogen and carbon removal performance than the activated sludge system (A). Although the mixed system (A+M) was equivalent to adding an equal amount of mixed HN-AD bacteria to the activated sludge system (A), there was no superimposed effect on carbon and nitrogen removal. Previous research used the pure culture of Acinetobacter indicus strain ZJB20129 to treat municipal wastewater (NH₄⁺-N: 180.8 mg/L; TN: 319.2 mg/L), and the removal efficiencies of NH₄⁺-N and TN were only 42.6% and 30.9%, respectively. This result is significantly lower than that of reactors A+M and M. Therefore, the effect of treating wastewater by the mixed culture of different HN-AD bacterial agents is significantly better than that of the pure culture mode [17].

In the effluent of the three reactors (A+M, M, A), all four forms of nitrogen were detectable. Among them, there was no significant difference in the concentrations of NO₃⁻-N and NO₂⁻-N, which were as follows: 20.78 mg/L, 19.31 mg/L, and 16.28 mg/L; 6.43 mg/L, 9.71 mg/L, and 9.20 mg/L, respectively. However, there were significant differences in the concentrations of NH₄⁺-N and organic nitrogen, which were as follows: 13.44 mg/L, 16.81 mg/L, and 26.69 mg/L; 9.46 mg/L, 13.44 mg/L, and 17.95 mg/L, respectively (Figure 2B). The accumulation of nitrate and nitrite was likely related to the insufficient carbon source, which inhibited the denitrification process [18]. Since the HN-AD bacteria in the reactors A+M and M grew in the heterotrophic mode, their growth was significantly faster than the autotrophic nitrifying bacteria in reactor A, ensuring higher nitrification activity in reactors with HN-AD bacteria. Additionally, when both NH₄⁺-N and NO₃⁻-N were present, HN-AD bacteria generally exhibited the physiological characteristic of preferentially utilizing ammonia nitrogen before nitrate [19]. This explains the relatively lower ammonia nitrogen concentrations and higher nitrate concentrations in the first two reactors.

For nitrogen removal processes, an appropriate COD/N ratio ranges from 5:1 to 10:1 [20]. In fact, the COD/N ratio in the effluent had decreased from 2.01 ± 0.84 to 1.96 ± 0.96 (A+M), 1.83 ± 0.92 (M), and 1.76 ± 0.82 (A) (Figure 3A), suggesting that the carbon source in the wastewater was insufficient for the denitrification. However, for the reactors inoculated with the HN-AD mixed bacteria agent, the amount of COD consumed for the removal of per unit total nitrogen was significantly reduced (A+M, 2.55 ± 1.75; M, 2.90 ± 2.04; A, 4.25 ± 3.99) (Figure 3B), indicating an improved efficiency of carbon source utilization.

3.2. Influence of HN-AD Strain Agents on Biomass

At the end of the experiment, the biomass (including suspended and attached states) from the three groups of reactors was uniformly mixed, and the contents of MLSS and MLVSS were measured separately. The initial MLSS concentration in reactors M and A was approximately 3698.00 mg/L. After nearly 80 days of operation, the MLSS concentration in reactor M increased by 79.4%, while that in reactor A increased by 2.49 times. Reactor A, inoculated with activated sludge, primarily consisting of aerobic heterotrophic bacteria, had a faster growth rate, and a shorter doubling time under suitable conditions. In contrast, reactor M, inoculated with the HN-AD mixed bacteria, had a slower growth rate. Therefore, the MLSS in reactor A was significantly higher than that in reactor M (

Table 2. Biomass of activated sludge in different groups of reactors at the end of the experiments.

Group	SS (mg/L)	VSS (mg/L)	VSS/SS
RAW	3698.00 ± 154.65	1928.00 ± 148.95	0.52 ± 0.02
M	6633.33 ± 560.39	4673.33 ± 181.47	0.71 ± 0.03
A+M	10,520.00 ± 1075.13	7395.00 ± 838.50	0.70 ± 0.01
A	12,916.67 ± 1173.90	8968.33 ± 1029.44	0.69 ± 0.02

Note: Raw represents the initial SS inoculated in the A reactor or the A+M reactor.

). Since the initial MLSS concentration in reactor A+M was twice that of reactor A or M, and HN-AD bacteria were supplemented three times during the operation, the final MLSS concentration should have been higher than that of reactor A. However, it was only 10,520.00 ± 1075.13 mg/L, indicating that when aerobic heterotrophic bacteria and HN-AD bacteria coexisted, the system might experience substrate competition and metabolic product inhibition, leading to a reduction in the cell yield of both. Nevertheless, reactor A’s treatment capacity for NH₄⁺-N, TN, and COD was significantly lower than that of reactor A+M (Figure 2A), suggesting that MLSS was not the main factor determining the nitrogen and carbon removal capacity. Nevertheless, compared with the biofilm formed on conventional carriers, the bio-nest formed on modified basalt fibers has a huge biomass advantage. Previous research has proven that the bulk density of the bio-nest is approximately several times that of the biofilm [21]. This creates conditions for the efficient operation of the reactor. Overall, adding the HN-AD mixed bacteria to the activated sludge system could significantly enhance the reactor’s performance.

As an important component of biomass, the content of EPS in reactor A+M was slightly higher (242.26 ± 15.52 mg/g-VSS), followed by reactor A (225.00 ± 46.18 mg/g-VSS) and reactor M (218.51 ± 5.40 mg/g-VSS). As components of EPS, PN and PS showed similar trends in content across the three reactors (Figure 4). Due to the use of MBF carrier media, the contents of both are higher in the three reactors than in conventional fillers. Some research shows that under low carbon-to-nitrogen ratio conditions, the contents of PS and PN in the biofilm attached to polypropylene spherical carrier media are only 38.67 mg/g-VSS and 122.67 mg/g-VSS. Therefore, it is believed that MBF plays an important role in biomass fixation and the stability of the bio-nest [22]. However, the PN content was higher than PS in all groups, with PN/PS ratios of 2.77 ± 0.00 (A+M), 2.78 ± 0.02 (M), and 2.74 ± 0.01 (A), respectively. Typically, microorganisms attached to the surface of the carrier secrete a large number of PS to enhance adhesion and structural stability, resulting in a lower PN/PS ratio. In this study, despite the use of MBF carriers, the high PN content might be related to the low influent carbon-to-nitrogen ratio. Studies have shown that EPS in activated sludge cultivated in water with a low C/N ratio is mainly composed of PN [23].

3.3. Influence of HN-AD Strain Agents on Biological Viability

Considering that the addition of HN-AD bacterial agent to the activated sludge may alter biological activity, this study used the fluorescent dyes SYTO9 and PI to stain the biomass in the three reactors to observe this potential impact. The CLSM images showed that the sample from reactor A+M had the strongest green fluorescence, followed by reactors M and A. In terms of red fluorescence, all reactors seemed weak. The absence of yellow or orange areas after superimposing the two fluorescence images indicated a low number of dead cells within each reactor, and the green fluorescence was not quenched by PI (Figure 5). This is consistent with previous research findings; that is, biological activity can still be detected at a depth of 6 cm in the bio-nest, and the presence of dissolved oxygen can still be detected at a depth of 5 cm [21]. Therefore, adding HN-AD bacterial agent to the activated sludge does not impair the biological activity within the system but rather enhances it.

SS/VSS ratio is an important indicator reflecting the biological viability or activity and organic matter content in activated sludge. Generally, the ratio between 0.6 and 0.8 is considered to indicate good sludge activity, suggesting a higher content of organic matter and stronger microbial activity [24]. In this study, the highest ratio was found in reactor M (0.71 ± 0.03), followed by reactor A+M (0.70 ± 0.01), then reactor A (0.69 ± 0.02) (Table 2), all of which were close to 0.7. This indicates that the biomass in all three reactors had good biological activity; this agrees well with the fluorescence images mentioned above.

3.4. Microbial Community Differences in Reactors with and Without HN-AD Strains

After more than 80 days of operation, the microbial community composition in the reactors underwent significant changes. At the phylum level (Figure 6A), the dominant bacteria in reactors M, A+M, and A, and the raw sludge, were Proteobacteria (65.2%, 56.4%, 41.9%, 29.8%), followed by Bacteroidota (9.9%, 10.2%, 18.7%, 1.6%), and Chloroflexota (3.8%, 8.5%, 10.4%, 4.7%). Compared to the raw sludge, the proportions of these three bacterial phyla in reactor A increased significantly by 40.6%, 10.7-fold, and 1.2-fold, respectively. Because the three HN-AD bacterial strains, K. variicola L3, A. beijerinckii W4, and Acinetobacter sp. Z1, all belong to Proteobacteria, Proteobacteria accounted for nearly 100% of the microbial community in reactor M during the initial operation. By the end of the experiment, the proportion of Proteobacteria had decreased to 65.2%, implying that other phyla primarily originated from domestic wastewater.

At the species level (Figure 6B), Acinetobacter sp. Z1 and K. variicola L3 remained the dominant bacterial species in reactor M, accounting for 6.1% and 1.1% of the community, respectively. Although K. variicola is widely present in sewage treatment and plays an important role in the SND process, it is also a pathogenic bacterium, posing a potential risk to human health. Therefore, it is necessary to disinfect and kill it before the reuse of sewage [25]. The proportions of all other bacterial species were below 1.0%. This indicates that under low C/N conditions, heterotrophic microorganisms compete intensely for limited carbon sources. HN-AD bacteria, particularly Acinetobacter sp. Z1 and K. variicola L3, exhibited lower affinity constants for organic substrates, enabling them to sustain growth at low substrate concentrations. In contrast, most bacteria experienced severe nutrient limitation, resulting in extremely low abundances. Although A. beijerinckii W4 belongs to the same genus as Acinetobacter sp. Z1, it was hardly detected in the community, which might be due to its physiological characteristics being poorly adapted to such domestic wastewater.

Compared to reactor M, the microbial sources in reactor A+M included not only HN-AD bacterial agents but also activated sludge from a wastewater treatment plant. The proportion of Acinetobacter sp. Z1 decreased to 3.4% but remained the dominant species in the community. The next most abundant species were Rubrivivax sp. SCN 71–131 (1.9%), followed by Bacteroidota bacterium (1.5%), Chloroflexota bacterium (1.4%), Candidatus Promineofilum breve (1.3%), Flavobacteriales bacterium (1.1%), Anaerolineae bacterium (1.0%), and Gammaproteobacteria bacterium (1.0%). The proportions of all other bacterial species were below 1.0%, with the proportion of K. variicola L3 slightly decreasing to 0.9% compared to that in reactor M. In addition to Rubrivivax sp. SCN 71–131 having denitrification capabilities [26], the other mentioned bacterial species primarily participate in the secretion of carbohydrate-active enzymes, polysaccharide degradation, and the anaerobic fermentation process [27,28,29,30,31]. In reactor A, the three mixed bacterial strains were barely detectable. The dominant strain was Bacteroidota bacterium (2.8%), followed by Anaerolineae bacterium (1.8%), Chloroflexota bacterium (1.6%), Flavobacteriales bacterium (1.5%), and Candidatus Promineofilum breve (1.3%).

The significant differences in microbial community structure and relative abundance might be an important reason affecting the nitrogen removal efficiency of the three groups of reactors. As mentioned above, the reactors Aand M showed lower removal efficiencies for NH₄⁺-N and TN compared to the reactor A+M. The weak nitrogen removal capacity of reactor A might be related to the relatively low abundance of nitrifying and denitrifying bacteria in the bio-nest, while the weak nitrogen removal capacity of reactor M may be related to the preference of each HN-AD strain for carbon sources [7]. For example, Alcaligenes faecalis No. 4 prefers to utilize succinate, while Pseudomonas sp. DM02 prefers to utilize glucose [32]. These differences in utilization ultimately affect the denitrification rates of the HN-AD strains. In the present study, Acinetobacter sp. Z1 and K. variicola L3 were found to preferentially utilize acetate and pyruvate, respectively. Clearly, reactor M has limitations in carbon source utilization. The activated sludge in reactor A+M contained a large number of microorganisms involved in the degradation of carbohydrates, lipids, and amino acids, and its diverse metabolic products could provide the necessary organic substrates for each HN-AD strain. This explains why the nitrogen removal capacity of reactor A+M was higher than that of the other reactors.

3.5. Role of HN-AD Bacterial Stains in Nitrogen Conversion

Previous studies have confirmed that Acinetobacter sp. Z1 possesses only two complete ammonia transformation pathways: the conversion of urea to ammonia and the assimilation of nitrate into ammonia. However, annotation of its whole genome did not reveal any canonical nitrification or denitrification genes. This suggests that Acinetobacter sp. Z1 contains a large number of homologous genes [19]. In contrast, K. variicola L3 contains multiple nitrogen metabolism pathways, but only denitrification is relatively complete. Genes involved in denitrification included narGHI, nirK/S, norBC, and nosZ, while dissimilatory nitrate reduction to ammonia was annotated with only the gene nirBD, assimilatory nitrate reduction to ammonia with only the gene nasAB, and nitrification with only nxrAB. Gene sequencing of biological samples from reactor A+M revealed that at least four complete nitrogen metabolism pathways were identified, including nitrification, denitrification, and assimilatory/dissimilatory nitrate reduction to ammonia (Figure 7).

4. Discussion

Regarding the nitrification process (Figure 7A), the number of genes encoding ammonia monooxygenase (AmoCAB), hydroxylamine oxidoreductase (Hao), and nitrite oxidoreductase (NxrAB) was the lowest in reactor M and the highest in reactor A+M. Since none of the three HN-AD bacterial agents possess genes related to nitrification, the presence of the aforementioned three genes in reactor M should originate from wastewater. In reactor A+M, Acinetobacter sp. Z1 remained the dominant bacterium in the community, and its homologous proteins involved in nitrification likely played a reinforcing role in the system.

For the denitrification process (Figure 7B), a similar phenomenon was observed. The number of genes encoding membrane-bound nitrate reductase (NarGHI), periplasmic nitrate reductase (NapAB), copper-type nitrite reductase (NirK), cytochrome cd₁-type nitrite reductase (NirS), nitric oxide reductase (NorBC), and nitrous oxide reductase (NosZ) was lower in reactor M compared to reactors A+M and A. Although K. variicola L3 possesses a relatively complete denitrification gene system, its relative abundance in reactor M was low, contributing minimally. Therefore, these functional genes primarily originated from wastewater. Since Acinetobacter sp. Z1 lacks known denitrification genes but was the dominant bacterium in both reactors A+M, its inherent homologous proteins were likely to reinforce the nitrogen removal performance of reactor A+M.

Regarding whether the HN-AD bacteria contain genes encoding for dissimilatory nitrate reduction to ammonia, there are currently no relevant reports in the literature. Gene abundance analysis (Figure 7C) revealed that dissimilatory nitrate reduction genes (nirBD, nirAH) were least abundant in reactor M. Although numerous studies have demonstrated that small amounts of ammonia are produced in HN-AD bacterial cultures when nitrate is used as a substrate, this does not confirm whether the ammonia originates from cell lysis or dissimilatory reduction. Compared to the genes encoding other processes, the genes involved in the assimilatory reduction in nitrate to ammonia (narB, nr, nasAB, nirA, and nasBDE) were significantly less abundant in all reactors, with no notable differences observed (Figure 7D). Previous studies have shown that these genes are present in Acinetobacter sp. Z1 and are primarily used for cellular synthesis [19]. Overall, the presence of Acinetobacter sp. Z1 as a dominant bacterium, in either reactor A+M or reactor A, enhanced the nitrogen removal performance of the system.

5. Conclusions

This study demonstrates the effectiveness of HN-AD bacterial agents in enhancing nitrogen removal for low C/N municipal wastewater. Key findings include the following:

• Enhanced performance: the A+M reactor (HN-AD strains+activated sludge) achieved the highest removal efficiencies for NH₄⁺-N (82.2%), TN (52.9%), and COD (51.6%), with a 40% reduction in carbon consumption per unit TN removed compared to conventional systems.
• Microbial synergy: dominance of Acinetobacter sp. Z1 (6.1%) and K. variicola L3 (1.1%) in the A+M reactor, alongside increased biodiversity (Proteobacteria, Bacteroidota), confirmed the synergistic role of HN-AD bacteria with native sludge communities.
• Biofilm robustness: high EPS content (242.26 mg/g-VSS) and PN/PS ratio (2.77) in the A+M reactor indicated stable biofilm activity, facilitated by modified basalt fiber carriers.
• Metabolic flexibility: gene annotation revealed complementary nitrogen pathways (nitrification, denitrification, DNRA) in the A+M system, explaining its superior performance under low C/N conditions.
• Future research directions include the following:
• Long-term stability: evaluate the reactor performance over extended periods (>1 year) to assess microbial community shifts and potential loss of HN-AD strains.
• Pathogen control: investigate disinfection strategies (e.g., UV, chlorination) to mitigate risks from pathogenic strains like K. variicola L3 in effluent reuse.
• Full-scale validation: test the technology in pilot-scale reactors with real wastewater to optimize carrier design, aeration, and operational parameters.
• Carbon source alternatives: explore low-cost carbon supplements (e.g., agricultural waste) to further improve denitrification efficiency.
• Mechanistic modeling: develop kinetic models integrating HN-AD bacterial growth and nitrogen pathways for process prediction and control.