Sulfadiazine and Nitrogen Removal Performance and Mechanisms of a Saline-Resistant Strain of Acinetobacter sp. RT-6

Abstract

The main goal of this article is to elucidate the SDZ and TN removal characteristics and mechanisms of a saline-resistant strain of Acinetobacter sp. RT-6. The SDZ and TN removal characteristics indicated that SDZ and TN removal were significantly affected by SDZ concentration and salinity. The removal curves and kinetics of different typical amounts of nitrogen showed the shortcut nitrification and denitrification present in the ammonia-N removal process; the nitrogen-containing compound was mainly transformed into N₂, and approximately 19.05 ± 0.83 mM of the electrons was transferred to the nitrate respiratory chain. The intermediates of SDZ degradation were identified, and it was speculated that the main reaction sites for SDZ transformation were the N–C bond, N–S bond, pyrimidine ring, and benzene ring.

1. Introduction

Sulfonamide is one of the most commonly used medications for treating bacterial infections in humans and animals, and sulfadiazine (SDZ) is a commonly used sulfonamide antibiotic [1]. Due to the wide application of SDZ and the fact that conventional wastewater treatment plants cannot degrade it effectively, SDZ has been frequently detected in the environment [2]. Antibiotics entering the environment could lead to the disruption of the microflora ecosystem and the promotion of antibiotic resistance [3]. Therefore, SDZ degradation, along with the development and transmission of antibiotic-resistant genes (ARGs), has induced much concern [4].

Some physical and chemical methods have been proposed for the removal of antibiotics, such as adsorption [5], membrane filtration [6], the photo-Fenton process [7], ozonation [8], and photocatalysis [9]. However, these methods have the basic problems of high cost, secondary pollution, and so on [10]. As is well known, microorganisms play an important role in the process of antibiotic migration, transformation, and degradation [3]. Meanwhile, antibiotic biological treatment technology tends to be more compatible with the push toward more economic, green, and environmental treatments [10]. Several researchers have reported on antibiotic biological treatment technologies, such as the biofilm reactor and microbial function, which can effectively degrade antibiotics [11]. Cetecioglu et al. (2015) used a lab-scale anaerobic sequencing batch reactor to remove 40 mg·L⁻¹ of antibiotics and achieved a good removal efficiency [12]. Zhao et al. (2022) constructed two anaerobic sequencing batch reactors (ASBRs) for the degradation of sulfonamide antibiotics, and the SDZ and SMX degradation efficiencies were 37.18% and 86.26% [13]. Several researchers isolated pure functional microorganisms from acclimated activated sludge to degrade antibiotics and explore the degradation mechanism of antibiotics [14]. Reis et al. (2014) isolated a PR1 strain with high sulfonamide removal efficiency; strain PR1 can completely degrade sulfonamide antibiotics within 56 h [15]. Tappe et al. (2013) isolated two sulfonamide antibiotic degradation bacteria (SDZm4 and C448), and the sulfonamide antibiotic degradation rate increased when another carbon source was added [16].

Microbial nitrification and denitrification reactions form an important method for removing nitrogen from pharmaceutical wastewater [17]. However, studies have reported that pharmaceutical wastewaters may have high salinity and toxicity and contain multiple organic components, usually containing high concentrations of salts, ammonia-N, and antibiotics [18,19]. The presence of antibiotics and a high salt content (>1 wt%) in a reaction system has an important effect on the efficiency of biological ammonia-N removal [20]. Several researchers have reported the adverse or inhibitory effects of various antibiotic exposures on biological ammonia-N removal processes [21]. Zhang et al. (2013) indicated that copper(II) and antibiotics would inhibit anaerobic ammonium oxidation (ANAMMOX) activity and performance [22]. Liang et al. (2022) reported that sulfonamide antibiotics have selective inhibition on biological aerobic nitrogen removal [23]. Sudmalis et al. (2018) also indicated that the high osmotic pressure caused by high salinity may cause microorganisms to dehydrate, dissociate, and eventually die [24]. Unilateral studies of antibiotic degradation or nitrogen conversion cannot adapt to the complex conditions of pharmaceutical wastewater. Therefore, studies on synchronous SDZ degradation and nitrogen conversion in pharmaceutical wastewater are relatively scarce, and this area needs further study.

The main objectives of this study were (1) to isolate a pure strain of Acinetobacter sp. RT-6 and observe the basic morphological and physiological traits, (2) to investigate the effects of different factors (e.g., SDZ concentration, salinity, and C/N ratio) on nitrogen and SDZ removal and determine the boundary conditions of SDZ and TN removal in strain RT-6, and (3) to formulate a complete model for nitrogen removal processes with different nitrogen-containing compounds and deduce the SDZ degradation pathways via the detected intermediates.

2. Materials and Methods

2.1. Culture Medium

The enrichment medium (EM) used for microbial screening and experiments and the main components are shown in Table S1. The components of the basal medium (BM) (used for single-factor experiments) included CH₃COONa (1.0 g·L⁻¹), NH₄Cl (0.2 g·L⁻¹), KH₂PO₄ (0.1 g·L⁻¹), MgSO₄·7H₂O (0.05 g·L⁻¹), CaCl₂ (0.05 g·L⁻¹), NaCl (25 g·L⁻¹), trace element solution (2 mL), and SDZ (0.005 g·L⁻¹). The components of the trace element solution followed those of Liang et al. (2020) [25]. Before the addition of SDZ, the BM was sterilized at 121 °C for 30 min, and SDZ was added to the BM after it had cooled to room temperature. The initial concentrations of TOC and TN were 300 ± 15 mg/L and 50 ± 3 mg/L. Finally, strain RT-6 was cultured in the BM (5%, v/v), and the temperature and rotation speed of the constant-temperature oscillation incubator were adjusted to 30 °C and 120 rpm.

2.2. Isolation of Strain RT-6

The SDZ degradation and nitrogen removal bacterium, Acinetobacter sp. RT-6, was isolated from an industrial park wastewater treatment plant (Guangzhou, Guangdong, China). The methods of strain isolation were as follows: First, 15 mL of the activated sludge was added to BM containing 5 mg/L SDZ and cultured in a constant-temperature culture oscillator at 30 °C and 120 r/min for 7 days. After 7 days, 15 mL of the mixture was taken from the conical bottle and added to fresh BM with an SDZ concentration of 10 mg/L for 7 days. The concentration of SDZ was gradually increased to 40 mg/L using a continuous acclimation method with a gradually increasing SDZ concentration. After the above acclimation and enrichment, 1 mL of the mixture was obtained from the conical bottle and diluted 10⁻¹, 10⁻², 10⁻³, 10⁻⁴, 10⁻⁵, and 10⁻⁶ times with sterile water. Then, 0.2 mL of each diluted mixture was evenly smeared on a fresh BM plate with a sterilized triangle stick and then cultured inverted in a constant-temperature incubator at 30 °C for 7 days.

2.3. Identification of Strain RT-6 and nirK Gene Detection

The partial 16S rRNA gene of strain RT-6 was amplified via polymerase chain reaction (PCR). The primers were purchased from Sangon Biotech (Shanghai, China) Co., Ltd., and the primer series were 1492R (5′-TACGGTTACCTTGTTACGACTT-3′) and 27F (5′-AGAGTTTGATCATGGCTCAG-3′) [26]. The PCR amplification protocol consisted of the following steps: initial denaturation at 94 °C for 4 min; 30 cycles at 94 °C for 45 s, 55 °C for 45 s, and 72 °C for 60 s; and final extension at 72 °C for 10 min. The amplified products were analyzed on 1.0% agarose gel to determine the target DNA fragments and sequenced by Sangon Biotechnology Co., Ltd. (Shanghai, China). The closest relative sequences of strain RT-6 were retrieved from the NCBI database, and a phylogenetic tree was constructed using MEGA.7 [25]. The functional gene nirK is one of the important genes related to nitrite removal [27]. The amplification program of the nirK gene was as follows: 95 °C pre-denaturation 5 min; 30 cycles of amplification at 95 °C for 30 s, annealing at 45 °C for 40 s, and 72 °C for 40 s, wherein the annealing temperature was reduced by 0.5 °C per cycle to 44 °C and then maintained at 43 °C for the last 20 cycles; and final extension at 72 °C for 7 min. The product of amplification was sent to Sangon Biotechnology Co., Ltd. for analysis.

2.4. The Characteristics of SDZ and Ammonia-N Removal

To evaluate the SDZ degradation and TN removal performances of strain RT-6, 5% (v/v) bacterial suspension (OD₆₀₀ ≈ 1.0) was inoculated into sterilized BM; the initial SDZ and ammonium-N concentrations were 5 and 50 mg·L⁻¹. After inoculation, strain RT-6 was cultured in a shock incubator at a constant temperature (30 °C) and rotation speed (120 rpm) for 64 h. Samples were withdrawn every 8 h to determine the biomass concentration and pH; the supernatant was obtained by centrifugation and used to analyze the ammonia-N, nitrate-N, nitrite-N, TN, and SDZ concentrations.

2.5. The Suitable SDZ and Ammonia-N Removal Conditions of Strain RT-6

To determine the suitable conditions for ammonia-N and SDZ removal of strain RT-6, the initial SDZ concentrations, C/N ratio, and salinity were applied as variables. The influence of the SDZ concentration on TN removal and SDZ degradation was investigated at SDZ concentrations of 5 mg·L⁻¹, 10 mg·L⁻¹, 20 mg·L⁻¹, and 40 mg·L⁻¹. The effect of the C/N ratio on TN removal and SDZ degradation was determined by increasing the C/N ratio from 2 to 6 while maintaining the NH₄⁺-N concentration at 50 mg·L⁻¹. Additionally, to explore the effect of salinity (NaCl) on TN removal and SDZ degradation, the salinity was increased from 1 to 5 wt%. The above experiments were conducted with an inoculum proportion of 5% (v/v), a rotation speed of 120 rpm, and a temperature of 30 °C. Samples were withdrawn every 8 h, and the concentrations of TN and SDZ were determined.

2.6. The Ammonia-N Removal Mechanism of Strain RT-6

Different nitrogen source experiments were conducted to investigate the ammonia-N removal mechanisms of strain RT-6. A bacterial suspension (20 mL, OD600 ≈ 1.0) was added to 400 mL of sterile BM, and ammonia-N, nitrate-N, and nitrite-N at the concentration of 50 mg/L were added as nitrogen sources in groups I, II, and III, respectively. The other experimental conditions were as follows: (1) SDZ concentrations of 5 mg/L; (2) reaction temperature, initial pH, and shake speed controlled at 30 ± 2 °C, 7.0 ± 0.2, and 125 ± 10 rpm, respectively. Samples were withdrawn regularly at intervals of 8 h to detect the concentrations of NH₄⁺-N, NO₃⁻-N, and NO₂⁻-N.

2.7. SDZ and TN Removal Kinetic Model

To clarify the characteristics of SDZ degradation and TN removal by strain RT-6, zero-order, pseudo-first-order, and pseudo-second-order kinetic models were applied to fit the SDZ degradation and nitrogen removal data. The validity of calculations can be measured using the calculation of the standard deviation and the coefficient of determination R². Through comparison, the pseudo-first-order kinetic model was selected to fit the data of TN removal and SDZ degradation, and the rate constant was obtained. The kinetic equation is as follows [25]:

−ln(C_t/C₀) = kt

(1)

where k (1/h) refers to the degradation rate constant of TN/SDZ, C₀ (mg·L⁻¹) is the concentration of TN/SDZ at 0 (h), C_t (mg·L⁻¹) is the concentration of TN/SDZ at t (h), and t is the reaction time (h) (Liang et al., 2020) [20].

2.8. Analysis of Nitrogen Balance and Electron Flow Distribution

Different nitrogen components were measured to analyze the nitrogen balance, and the contents of different compounds before and after the reaction were used to analyze the electron flow distribution. The prepared bacterial suspension was added to 400 mL of sterile EM (v/v = 5%). Each reaction system was then aerated with oxygen for 15 minutes, and the bottles were sealed with a stopper immediately after aeration. The gas of the bottle’s headspace was collected after incubation (2 days), and the N₂ concentration was measured. Meanwhile, samples were withdrawn from each bottle to determine the biomass-N, ammonia-N, nitrite-N, nitrate-N, TN, and TOC [28]. The total nitrogen in the sample containing the suspension was recorded as TN1. The total nitrogen in the samples after centrifugation and membrane filtration was recorded as TN2. The biomass-N was then calculated by subtracting TN2 from TN1. The electron flow distribution with ammonia-N as a nitrogen source is calculated according to Equations (2)–(5) [29]:

$$\mathrm{N}{\mathrm{O}}_3^{-}+2{\mathrm{H}}^{+}+2{\mathrm{e}}^{-}\to \mathrm{N}{\mathrm{O}}_2^{-}+{\mathrm{H}}_{2}O$$

(2)

$$\mathrm{N}{\mathrm{O}}_{\mathrm{2}}^{-}+4{\mathrm{H}}^{+}+3{\mathrm{e}}^{-}\to 0.5{\mathrm{N}}_2+2{\mathrm{H}}_2\mathrm{O}$$

(3)

$${\mathrm{C}}_2{\mathrm{H}}_3{\mathrm{O}}_2^{-}+2{\mathrm{O}}_2\to \mathrm{C}{\mathrm{O}}_2^{}+\mathrm{H}\mathrm{C}{\mathrm{O}}_3^{-}+{\mathrm{H}}_2\mathrm{O}$$

(4)

$$2\mathrm{N}{\mathrm{H}}_4^{+}+5{\mathrm{C}}_2{\mathrm{H}}_3{\mathrm{O}}_2^{-}+3{\mathrm{H}}^{+}\to 2{\mathrm{C}}_5{\mathrm{H}}_7{\mathrm{O}}_2\mathrm{N}+6{\mathrm{H}}_2\mathrm{O}$$

(5)

2.9. Identification of Intermediates

The intermediates of SDZ degradation were detected via QExactive LC-MS/MS; the LC instrument used was a Thermo UltiMate™ 3000 BioRS system (Thermo Fisher Scientific Inc, Bremen, Germany), and the mass spectrometry instrument used was a Thermo QExactive^TM Plus mass spectrometer (Thermo Fisher Scientific Inc, Bremen, Germany). The mobile phase of QExactive LC-MS/MS used to detect intermediates consisted of two components: phase A: acetonitrile (30%); phase B: ultrapure water with 0.1% glacial acetic acid (70%). The intermediates in the nucleo-cytoplasmic ratio range of 50 to 600 were fully scanned. The following gradient elution mode was used: 10–50% B (0–2 min); 50–70% B (2–5 min); 70% B (5–6 min); 70–80% B (6–7 min); 80% B (7–9 min); 80–100% B (9–10 min); 100% B (10–13 min). Masslynx4.1 (Thermo Scientific) was used to process the mass spectrometry data. The intermediates were further identified via standard mass spectrometry (National Institute for Standard Technology library), and the structures were drawn using ChemDraw 8.0 (CambridgeSoft; Cambridge, MA, USA).

2.10. Analytical Methods

The water samples were centrifuged at 8000 rpm for 10 min to measure the concentrations of nitrate-N, ammonia-N, nitrite-N, and TN. The measurements were conducted according to the APHA method described in reference [28]. NH₄⁺-N, NO₂⁻-N, NO₃⁻-N, and TN were analyzed using salicylic acid–hypochlorite, N-(1-naphthyl)-ethylene diamine spectrophotometry, ultraviolet spectrophotometry, and alkaline potassium persulfate oxidation–ultraviolet spectrophotometry, respectively. The SDZ concentration was determined using a high-performance liquid chromatography (HPLC) instrument (Waters e2695) equipped with a detector (Waters 2998). A mixture of acetonitrile and purified water (15:85, v/v) was used as the mobile phase, with a flow rate of 0.2 mL/min. The oven temperature and UV detector wavelength were set at 30 °C and 270 nm, respectively. The total organic carbon (TOC) concentration was determined using a TOC detector (Vario TOC cube; Elementar; Germany). All experiments mentioned in this article were carried out at least in triplicate. The above experiments conducted to analyze the SDZ degradation and nitrogen removal performance were carried out under static conditions, and pure water was used as a control group.

3. Results and Discussion

3.1. Isolation and Identification of Strain RT-6

Fifty different bacteria were purified from activated sludge, and the TN removal and SDZ degradation efficiencies of each bacterium were evaluated. Strain RT-6 showed high TN and SDZ removal efficiencies and was selected as the target strain. Strain RT-6 belongs to the Gram-negative, short-rod bacteria category. Based on 16S rRNA gene sequence analysis and phylogenetic tree construction, strain RT-6 showed a high similarity (99%) with Acinetobacter calcoaceticus and was identified as an Acinetobacter species. To further explore the mechanism of nitrogen conversion, the nirK gene was amplified (Figure S1), as it has been reported to be involved in biological nitrogen removal [27].

3.2. Ammonia-N and SDZ Removal Performances of Strain RT-6

Batch experiments were conducted to evaluate the ammonia-N and SDZ removal performances of strain RT-6. As depicted in Figure 1, the concentrations of SDZ, TOC, and ammonia-N increased slightly in the adaptive phase (0–8 h), and the biomass of strain RT-6 grew slowly. The results indicated that a high-salinity environment inhibits the growth and pollutant degradation activity of strain RT-6 in the adaptive phase, and strain RT-6 can adapt to this environment within 8 h. Previous studies have also reported that cells in the lag phase make the appropriate adjustments to adapt to complex environments [30]. The growth and reaction curves of strain RT-6 entered the logarithmic phase after 8 h of incubation, the biomass of strain RT-6 sharply increased to 132.68 mg·L⁻¹, with the growth rate of 2.32 mg·L⁻¹·h⁻¹. Meanwhile, approximately 79.33% of NH₄⁺-N, 60.69% of TN, and 51.81% of SDZ were removed in the logarithmic phase, and the removal rates were 0.60, 0.48 and 0.04 mg·L⁻¹·h⁻¹, respectively. These results indicate that strain RT-6 had high ammonia-N, TN, and SDZ removal efficiencies in the logarithmic phase, and the removal efficiency of various pollutants (TN, ammonia-N, and SDZ) was positively correlated with the bacterium growth rate. Previous studies also reported that ammonium-N and TN removal are strongly related to cell growth and carbon source utilization [31]. Meanwhile, nitrite-N and nitrate-N started to accumulate at 8 h, reaching maximum concentrations of 4.66 mg·L⁻¹ and 6.89 mg·L⁻¹, respectively, and then decreased to a non-detectable level at the end of the experiment. As depicted in Figure 1b, the pH value increased from 7.08 (0 h) to 8.09 (32 h), probably because of the alkalinity produced by the biological denitrification process. Sun et al. (2016) also indicated that the pH value increased from 7.0 to 9.0 during the heterotrophic nitrification–aerobic denitrification process [31]. In the stationary phase, the biomass of strain RT-6 gradually increased from 132.68 mg·L⁻¹ (24 h) to 156.26 mg·L⁻¹ (48 h) and then decreased to 142.39 mg·L⁻¹. In addition, the TN, ammonia-N, and SDZ removal efficiencies also increased slowly in this phase. These phenomena can be explained by the fact that most of the organic carbon sources and electron donors are consumed during logarithmic growth. Consequently, there may not be enough carbon sources for biological growth and metabolism in the stationary phase, resulting in a reduction in the efficiency of the biological removal of TN and SDZ. Previous studies also suggested that SDZ biodegradation is consistent with the bacterial growth trend [32].

Overall, the average removal ratios for TN, ammonia-N, and SDZ were 95.48% (ammonia-N), 92.24% (TN), and 87.84% (SDZ). These removal ratios showed that strain RT-6 utilizes sodium acetate as a carbon resource for growth, performs simultaneous heterotrophic nitrification and aerobic denitrification, and degrades SDZ through cometabolism. Zhao et al. (2021) also isolated a novel Acinetobacter sp. that could effectively remove ammonia-N under aerobic conditions, with minimal nitrite-N and nitrate-N accumulation [33]. However, this strain only removed 55% of sulfonamides in 300 h [34,35]. When compared to the other research, strain RT-6 exhibits superior sulfonamide antibiotic degradation efficiency and is capable of removing ammonia nitrogen and total nitrogen simultaneously under aerobic conditions.

3.3. Effects of SDZ Concentration on SDZ Degradation and TN Removal

The SDZ degradation ratios were 87.31%, 88.74%, 77.00%, and 64.64% with the initial SDZ concentrations of 5, 10, 20, and 40 mg·L⁻¹, respectively (Figure S2). The SDZ degradation profiles at different concentrations were well correlated with the pseudo-first-order kinetic model (R² > 0.98). As shown in Figure 2a, the SDZ degradation rate constant gradually decreased from 0.043 h⁻¹ to 0.019 h⁻¹ as the initial SDZ concentration increased from 5 mg·L⁻¹ to 40 mg·L⁻¹, with the highest constant observed at an SDZ concentration of 10 mg·L⁻¹. These results illustrate the following: (1) Strain RT-6 has strong resistance and SDZ degradation efficiency. The highest degradation efficiency was 88.74%, and the SDZ degradation efficiency remained above 60% even at an SDZ concentration of 40 mg·L⁻¹. Wang et al. (2019) also indicated that aerobic nitrifiers can degrade SDZ through cometabolism in the presence of antibiotics [21]. (2) High concentrations of SDZ may affect the degradation activity of strain RT-6, with the optimal degradation efficiency observed at an SDZ concentration of 10 mg·L⁻¹. These results confirmed that strain RT-6 has a strong adaptability in practical applications compared to other strains [36,37].

The concentration of SDZ also impacts the TN removal activity of strain RT-6, and the TN removal ratios were 93.57%, 91.41%, 89.72%, 83.85%, and 80.26% with the SDZ concentration increasing from 0 to 40 mg·L⁻¹ (Figure S2). The TN removal curve follows a pseudo-first-order kinetic model (R² > 0.95) (Figure 2b), and the TN removal rate constants were 0.048 (0 mg·L⁻¹), 0.044 (5 mg·L⁻¹), 0.040 (10 mg·L⁻¹), 0.028 (20 mg·L⁻¹), and 0.02 h⁻¹ (40 mg·L⁻¹). These results indicate that strain RT-6 had strong HN-AD efficiency and could remove over 90% of TN within 56 h, maintaining TN removal efficiency above 80% under a high SDZ concentration. Wang et al. (2022) also indicated that sulfonamide antibiotics would make a bio-retention cell lose its water treatment capacity [38].

3.4. Effect of C/N Ratio on SDZ Degradation and TN Removal

Organic carbon compounds are essential carbon resources and electron donors for heterotrophic bacteria [33]. Therefore, it is crucial to determine the suitable C/N ratio for SDZ degradation and TN removal of strain RT-6. The SDZ degradation ratios achieved at different C/N ratios were 89.96% (C/N = 2), 91.14% (C/N = 3), 90.86% (C/N = 4), 87.98% (C/N = 5), and 86.87% (C/N = 6) (Figure S2). As shown in Figure 2c, the SDZ degradation profiles at different C/N ratios were well correlated with the pseudo-first-order kinetic model (R² > 0.97); the corresponding removal rate constants were 0.051 (C/N = 2), 0.046 (C/N = 3), 0.043 (C/N = 4), 0.042 (C/N = 5), and 0.042 h⁻¹ (C/N = 6). These results indicate that an increase in the C/N ratio slightly inhibits SDZ degradation, possibly because of competition between organic carbon sources and SDZ as electron donors. Excessive organic carbon sources provide sufficient electron donors for microbial growth and metabolic activities, thus inhibiting the SDZ degradation activity of strain RT-6. On the other hand, the TN removal efficiency increased with an appropriate increase in the C/N ratio, and the TN removal ratio reached 90% when the C/N ratio was increased to 5 (Figure S2). Correspondingly, the TN removal rate constant of strain RT-6 increased from 0.016 h⁻¹ to 0.060 h⁻¹ when the C/N ratio was increased to 6 (Figure 2d). These results indicate that the organic carbon source is an electron donor for the HN-AD process, significantly influencing TN removal efficiency. Ji et al. (2015) also reported that aerobic denitrifiers can operate efficiently when the C/N ratio is 5–10 [39]. The influence of the C/N ratio on biological TN removal is mainly due to the C/N ratio impacting the synthesis of cell internal storage products [40].

3.5. Effect of Salinity on SDZ Degradation and TN Removal

Salinity, as a stressor, directly affects microbial osmotic pressure metabolism, inhibiting enzyme activity and eventually resulting in cellular lysis [40]. Therefore, it is necessary to explore the influence of salinity on SDZ degradation and TN removal performance. The SDZ degradation efficiency remained between 85% and 90% as salinity increased from 1.0 to 3.0%, and the SDZ degradation efficiency sharply decreased to 36.28% when the salinity increased to 5.0% (Figure S2). This suggests that high salinity can significantly suppress the degradation activity of the strain towards antibiotics. Meanwhile, the TN removal efficiency also remained between 85% and 90% when salinity increased from 1.0 to 3.0%, and it sharply decreased to 45.68% when the salinity increased to 5.0% (Figure S2). Notably, the removal rate constants of both SDZ and TN decreased as the salinity increased to 5.0% (Figure 2e,f). These results indicated that strain RT-6 can adapt to the salinity conditions within the range of 1.0 to 3.0%. However, the TN and SDZ degradation efficiencies sharply decreased when the salinity reached 4.0 to 5.0%. Although salinity directly affects microbial osmotic pressure metabolism and enzyme activity [41], strain RT-6 showed resistance to salinity. Chen et al. (2022) indicated total nitrogen removal was significantly inhibited with the addition of 1 wt% salinity [42].

3.6. The Mechanisms of Biological TN Removal

Three group experiments (group 1, ammonia-N as nitrogen resource; group 2, nitrate-N as nitrogen resource; group 3, nitrite-N as nitrogen resource) were designed to investigate the ammonia-N pathways of strain RT-6. As shown in Figure 3a, the concentration of ammonia-N decreased from 48.59 mg·L⁻¹ (0 h) to 5.148 mg·L⁻¹ (40 h) and remained stable thereafter. This indicates that strain RT-6 can effectively remove ammonia-N under aerobic heterotrophic conditions. Simultaneously, the concentrations of nitrite-N and nitrate-N gradually increased, reached the maximum concentrations of 4.66 mg·L⁻¹ and 6.89 mg·L⁻¹, respectively, after 24 h, and then decreased to non-detectable levels. These results indicate that strain RT-6 transforms ammonia-N to nitrate-N and nitrite-N, with an ammonia-N removal ratio of 0.453 mg·L⁻¹·h⁻¹. Yao et al. (2013) indicated that HN-AD bacteria can oxidize ammonia-N to nitrite-N or nitrate-N and then immediately denitrify these products to N₂O and/or N₂ [43]. Figure 3b shows the biological nitrogen removal curve with nitrate-N as the sole nitrogen source. The nitrate-N removal efficiency reaches 98.97% within 48 h, indicating efficient nitrate removal by strain RT-6. The accumulated nitrite-N reaches 8.56 mg·L⁻¹ after a 32 h reaction and then decreases to a non-detectable level. The accumulated ammonia-N reaches 1.976 mg·L⁻¹ at 24 h and then remains stable at 2.0 mg·L⁻¹. These results suggest that most nitrate-N is ultimately converted to nitrogen gas, with a nitrate-N removal rate of 0.706 mg·L⁻¹·h⁻¹. Medhi et al. (2017) also reported that Paracoccus denitrificans ISTOD1 can remove nitrate without nitrite accumulation under aerobic conditions [44]. Nitrite-N was used as the sole nitrogen source in group 3, and the nitrite-N removal curve is displayed in Figure 3c. The concentration of nitrite-N decreased from 49.52 mg·L⁻¹ (0 h) to 0.84 mg·L⁻¹ (64 h), and a nitrite-N removal efficiency of 98.30% was achieved. The accumulation of nitrate-N reached 3.49 mg·L⁻¹ at 32 h and then decreased to 1.02 mg·L⁻¹. The ammonia-N concentration reached 1.37 mg·L⁻¹ and remained stable at approximately 1.0–1.5 mmg·L⁻¹. Kinetic modes were used to fit the experimental data obtained with different nitrogen sources. As shown in Figure 3d, the kinetic constants of ammonia-N (k₁), nitrate-N (k₂), and nitrite-N (k₃) were 0.077, 0.054, and 0.049, respectively. The kinetic constants of nitrate-N and nitrite-N were higher than the kinetic constant of ammonia-N, indicating that the rate-limiting step of the HN-AD process is heterotrophic ammonia-N oxidation. Furthermore, the kinetic constant of nitrate-N (k₂) was higher than that of nitrite-N (k₃), which is consistent with the other reports showing that the reduction of nitrate-N was more rapid than that of nitrite-N due to a higher redox potential [45]. The above results show that the removal rates of nitrate-N and nitrite-N were higher than the removal rate of ammonia-N, and nitrate-N was consumed more rapidly than nitrite-N. Moreover, neither nitrate-N nor nitrite-N accumulates when ammonia-N is a nitrogen resource. Therefore, shortcut nitrification and denitrification may have appeared in the HN-AD process.

3.7. Nitrogen Balance and Electron Flow Distribution

The initial and final distributions of nitrogen during the ammonia-N removal process are displayed in Figure 4a. The initial TN (51.10 mg·L⁻¹) concentration in the BM mainly included biomass-N (3.02%) and ammonia-N (96.58%). The concentration of TN sharply decreased to 23.25 mg·L⁻¹ during the ammonia-N removal process. This final TN composition primarily consisted of biomass-N, ammonia-N, nitrate-N, and nitrite-N. The concentrations of these nitrogen species at the end of the process were as follows: biomass-N, 20.23 mg·L⁻¹; ammonia-N, 2.61 mg·L⁻¹; nitrate-N, 0.36 mg·L⁻¹; and nitrite-N, 0.05 mg·L⁻¹. An analysis of these results reveals that 40.14% of ammonia-N was transformed to biomass-N via assimilation, and about 53.87% of ammonia-N was converted to N₂ via the biological nitrogen removal process. These results indicated that the primary nitrogen removal pathway in this system is microbial aerobic denitrification rather than assimilation. Zhang et al. (2012) also indicated that the nitrogen removal pathway of HN-AD is similar to that of heterotrophic denitrification [46].

The distribution of electron flow in the biological nitrogen removal process was determined using the equation presented in Section 2.6. Based on the carbon and nitrogen balance analysis, a portion of the organic carbon source is used for cell synthesis, while the remainder serves as electron donors for respiration and nitrogen removal. According to the TOC and ammonia-N concentrations, the detailed distribution of electron flow was calculated using the equations presented in Section 2.6. The results indicate that 70.01 ± 1.72 mM of the organic carbon source was utilized as electron donors for cell respiration, and about 48.92 ± 1.08 mM of electrons was provided for the synthesis of cellular materials. Meanwhile, about 18.35 ± 0.64 mM of electrons was utilized for the conversion of nitrate-N to gas-N, accounting for 12.66% of the total electrons (Figure 4b). Despite nitrogen and oxygen being high-quality electron acceptors in terms of thermodynamics, only a small amount of electrons flow to biological denitrification. As reported by Chen and Strous (2013), aerobic denitrification requires the synergistic action of multiple core components, whereas aerobic respiration only requires the participation of oxidase [29,47].

3.8. SDZ Degradation Pathway

The intermediates of SDZ degradation were identified via QExactive LC-MS/MS and analyzed on the basis of biology and chemistry principles. A total of nine intermediates were identified, namely IP1, IP2, IP3, IP4, IP5, IP6, IP7, IP8, and IP9 with the mass charge ratios (m/z) of 174, 142, 128, 207, 99, 105, 222, 100, and 191, respectively. The proposed elemental composition and chemical structure of each intermediate are provided in the Supplementary Material (Table S2). The identified TPs can be divided into three categories: benzene ring compounds, sulfanilamide ring compounds, and single-chain compounds. Three possible transformation pathways were proposed based on the detected TPs and biological characteristics (Figure 5). In Route I, IP1 is formed due to the elimination of pyrimidine groups from SDZ. Then, IP1 loses a SO₂ to form IP2, which is then further transformed to IP3 by losing a methyl group. IP3 can be hydroxylated at the location of -NH₂ to form IP4. Alternatively, the aniline ring of IP3 can be broken to form IP5, and the amino group of IP5 is then hydroxylated to form IP6, which is the final product of SDZ degradation. Other researchers also reported that the C-S bond may be one of the priority reaction sites of sulfa antibiotics [48]; the first bond-breaking step usually occurs between NH₂ and the pyridine ring, and SDZ-pyridine is obtained [49]. In Route II, the benzene ring of IP1 is hydroxylated to form IP7, and the ring of IP3 is further broken to form IP6. In Route III, IP9 is formed by the breaking of the bond between NH₂ and a pyrimidine ring, the pyrimidine ring of IP9 is then broken to form IP8, and IP8 is hydroxylated at the location of -NH₂ to form IP6. Therefore, it can be speculated that the main reaction sites for SDZ transformation include the N-C bond, N-S bond, pyrimidine ring, and benzene ring [49,50].

4. Conclusions

In this study, a novel bacterium with nitrogen removal and SDZ degradation abilities was isolated and identified. The boundary conditions of SDZ degradation were as follows: SDZ concentration of 5–10 mg·L⁻¹, C/N ratio of 2–6, and salinity of 1–3%. The boundary conditions of TN removal were as follows: SDZ concentration of 0–10 mg·L⁻¹, C/N ratio of 4–6, and salinity of 1–3%. The maximum ammonium-N, TN, and SDZ removal ratios of strain RT-6 were 95.48%, 87.84%, and 92.24%, respectively. The SDZ and TN removal curves fitted well with the pseudo-first-order kinetic model, indicating that the two pollutants conform to the same degradation rule, and were markedly affected by SDZ concentration and salinity. Nitrogen balance and electron flow distribution suggested that nitrogen was primarily converted to N₂, while only a small fraction of electrons was transferred to the nitrate respiratory chain. Moreover, the main sites of SDZ transformation were the N-C bond, the N-S bond, the pyrimidine ring, and the benzene ring.