Comparison of aerobic and anoxic-oxic sequential batch reactors for treating textile wastewater

ABSTRACT The discharge of untreated textile dyeing wastewater containing azo dyes has significant ecological consequences. The incomplete biodegradation of these dyes can produce harmful and carcinogenic aromatic amines (AAs), which pose a threat to the environment. This study evaluates the treatment of real textile dyeing wastewater containing azo dyes by two different systems: an anoxic sequential batch reactor (A/O-SBR) and an aerobic sequential batch reactor (A-SBR). The effects of hydraulic retention time (HRT) on both systems were investigated in terms of chemical oxygen demand (COD), total nitrogen (TN), color, and AAs removal. Ranging between 85% and 92%, the COD removal efficiency was almost similar in the A/O-SBR and A-SBR systems, suggesting that the aerobic phase is primarily responsible for COD removal. However, A/O-SBR showed higher removal of color (90%) than A-SBR (75%) due to the presence of an anoxic phase. HRT has a significant impact on color removal in both systems. However, there was no significant effect on COD removal when HRT exceeded 12 h. According to LC-MS/MS analysis, increasing the HRT from 12 h to 24 h increased the AAs concentration in the anoxic phase, from 253 ng/L to 261 ng/L and slightly decreased the AAs concentration in the aerobic phase from 107 ng/L to 102 ng/L. This resulted in an overall improvement in both dye and AAs removal. The presence of aerobic phase in A/O-SBR and A-SBR demonstrates strong AAs removal capacity. GC-MS/MS analysis indicated that as the HRT of A/O-SBR and A-SBR increased, the number of by-products increased.


Introduction
Azo dyes are the dominant type of dyes utilized in textile processing industries, making up over 50% of the total.They are also the most frequently encountered synthetic dyes released into the environment (Albahnasawi et al. 2022).Azo dyes are defined by the presence of one or more azo bonds (-N=N-) that connect aromatic rings.These dyes exhibit a wide range of structural diversity and versatility due to various substitutions on the aromatic nucleus.However, this structural complexity also contributes to their resistance to degradation and classification as xenobiotic compounds (Journal and Reuse 2018).It is estimated that 7 × 10 5 metric tons of azo pigments are used in industry each year (Ilić Đurđić et al. 2020).Moreover, the yearly azo dye effluent release into the environment is estimated at around 50,000 tons (Benkhaya, M'rabet, and el Harfi 2020).
The application of dyes often leads to waste and discharge into aquatic ecosystems, resulting in the water body taking on a dark color.This coloration hampers sunlight penetration, negatively impacting aquatic life.Furthermore, the presence of high levels of organic matter and nitrogen in these discharges can contribute to eutrophication, causing further harm to the ecosystem (Pratheba, Balasundaram, and Preethi 2023;Reddy and Osborne 2022).Another concern arises when azo dyes partially degrade in aquatic bodies.While these dyes are generally stable in their natural state, under anaerobic conditions, the azo bond can spontaneously reduce, or degradation can occur through photocatalysis.This process leads to the creation of colorless aromatic amines that exhibit mutagenic and carcinogenic properties (Zhang et al. 2021).Due to their widespread environmental distribution and associated toxicity, aromatic amines have recently gained significant attention (González et al. 2023).These compounds have been linked to cancer, mutagenicity, and hemotoxicity in humans, animals, and plants, underscoring their potential risks (Wu, Huang, and Huang 2022).Consequently, it is crucial to assess the environmental risks associated with aromatic amines and implement effective management strategies to safeguard ecosystems (Rathi, Kumar, and Vo 2021).
Various methods have been utilized to remove dyes and their metabolites from textile effluents.These approaches include physical methods such as adsorption (Teixeira et al. 2022), coagulation, electrocoagulation (Louhichi et al. 2022), precipitation (Li et al. 2022), and filtration (Mahlangu, Mamba, and Mamba 2023), as well as chemical methods such as oxidation (Fenton's oxidation) (Brillas 2023), ozonation (Torres et al. 2023), and ion exchange (Türk, Çiftçi, and Arslanoğlu 2022).Biological methods (Khan et al. 2023) have also been employed.Electrocoagulation consumes high energy and generates sludge as a by-product.It also relies on pH adjustment and may not be suitable for all types of dyes or wastewater compositions (Mousazadeh et al. 2021).Adsorption can be limited by saturation of the adsorption media, requiring regeneration or replacement.It may also have selectivity issues and require careful optimization for effective performance (Wong et al. 2020).Ozonation, while effective in some cases, can be costly due to ozone generation requirements and may not efficiently degrade all types of dyes.It can also generate by-products and has a short half-life (Venkatesh et al. 2017).These drawbacks highlight the need for a comprehensive evaluation of the specific operating conditions and dye characteristics to determine the most appropriate treatment approach.
Biological techniques are the simplest, most effective, and least expensive of the methods available.They can treat large amounts of effluent without causing severe harm to the ecosystem (Touliabah et al. 2022).Moreover, biological processes can completely detoxify effluents and prevent secondary contamination that could result from the excessive use of chemical agents needed to further mineralize dyes and their aromatic amine compounds (Taheri, Fallah, and Nasernejad 2023).Technologies for biological treatment include anoxic, aerobic, and anaerobic processes (Bidu et al. 2023).The use of bioreactors for the treatment of textile effluents containing azo dyes offers several advantages over physical and chemical methods.Bioreactors are environmentally friendly and cost-effective since they use natural microbial processes to break down pollutants.Additionally, bioreactors can handle a wide range of dyes and their related metabolites, and they can operate under different conditions such as anaerobic, anoxic, and aerobic.Moreover, bioreactors can achieve high removal rates, up to 90% or more, making them a promising method for dye removal from textile effluents (Louhichi et al. 2022).
Combining multiple bioreactors has emerged as an effective approach for treating complex industrial effluents with specific treatment objectives, as highlighted by Cinperi et al. (2019).However, the presence of certain aromatic amines poses significant challenges for conventional biological treatment methods due to their persistence and toxicity to microorganisms, which can result in habitat risks (Zhu et al. 2023).Unfortunately, a notable research gap exists in our understanding of the biodegradability, fate, and environmental exposure of aromatic amines.This lack of crucial information hinders the development of comprehensive risk assessments for these compounds (Zhou et al. 2020).Despite the growing interest in combining bioreactors for complex industrial effluent treatment, there is limited research focused on the treatment of aromatic amines within these systems.This research gap stems from a lack of comprehensive understanding regarding the biodegradability, fate, and exposure of aromatic amines in the environment.This knowledge gap impedes the development of effective risk assessments and targeted treatment strategies for these persistent and toxic compounds.Addressing this research gap is crucial to advance our understanding of the fate and behavior of aromatic amines during treatment processes and to develop more sustainable and efficient treatment solutions.This study aims to evaluate the efficacy of a pilot-scale A/O-SBR and A-SBR in treating real textile wastewater.The study compares the performance of these reactors in removing COD, TN, color, and AAs-emerging pollutants of concern.By utilizing advanced analytical instruments such as liquid chromatography mass spectrometry (LC-MS/MS) and gas mass spectrometry (GC-MS/MS).Through this research, the results contribute to the development of sustainable and effective treatment strategies for the textile industry's wastewater challenges.

Materials and methods
The textile wastewater for this study was collected from a dye factory located in Istanbul, Turkey.To preserve the characteristics of the wastewater, it was stored at 4°C to prevent biological degradation before being used in the experiments.The wastewater was characterized before storage as shown in Table 1.The same analysis was performed every 14 days over a storage period of 100 days to monitor any changes in the wastewater composition over time.To assess the effectiveness of color removal, the concentration of color was measured using the American Dye Manufacture's Institute Color Unit (ADMI) (Shen et al. 2018).Samples were collected from the influent, anoxic phase effluent, and aerobic phase effluent in the A/O-SBR system, while in the A/SBR system, samples were collected from the influent and aerobic phase effluent.

Characterization of real textile wastewater
The concentrations of COD and BOD were measured according to standard methods (American Public Health Assocation APHA 1997).The TOC and TN analysis was done using the IL550 and IL530 TOC-TN analyzers (Hach, Germany).The color concentrations were analyzed by a DR 5000 spectrophotometer, as stated in 2120 F (Federation 1999).The analysis of pH, conductivity, alkalinity, chloride, sulfate, and TP was performed according to standard methods reported in Federation (1999).To ensure accuracy and reproducibility of the results, all measurements in this study were conducted in triplicate, and the average value was calculated and reported.Furthermore, analyticalgrade chemicals were utilized to maintain the quality and consistency of the experiments.

Aromatic amines analysis
The analysis of aromatic amines (AAs) was carried out using advanced analytical instruments (LC-MS /MS) for both aqueous and sludge samples collected from the reactors.The selection of 23 specific AAs was based on the existing literature (Ning et al. 2015).The chemical characteristics of the targeted AAs are presented in Table S1.

Aqueous phase extraction
To prevent biological degradation, the collected samples from A/O-SBR and A-SBR were stored in a 1.0-liter polypropylene bottle at 4°C.The AAs were extracted from the aqueous samples using the following steps: 1) 0.45 mm glass papers were used for filtration before extraction, and 2) solid-phase extraction (SPE) columns C18 (DIONEX SOLEX, 6 mL, 1000 mg, Thermo Fisher, USA) were utilized for SPE extraction.The details of the AAs extraction process (conditioning, extraction, and elution) were previously discussed in a study by (Albahnasawi et al. 2020).The column (C18) conditioning process involved two steps.First, 5 ml of methanol was used to condition the column, followed by an additional 5 ml of ultra-pure water.This process ensures proper preparation of the column for efficient separation and analysis.After conditioning, the wastewater samples were passed through the column.The column serves as a medium to selectively retain and separate the desired analytes from the wastewater matrix.For elution, 10 ml of methanol was used to extract the AAs from the C18 column.Elution is the process of removing the retained analytes from the column, allowing them to be collected for further analysis.To concentrate the eluted analytes, nitrogen gas was used to evaporate the methanol.This step removes the solvent and leaves behind a concentrated residue of the AAs.Subsequently, a 1 ml mixture consisting of 75% ultra-pure water and 25% methanol was added to the residue.This step serves to reconstitute the analytes in a suitable solvent for subsequent analysis.The final mixture, prepared after reconstitution, was ready for injection into LC-MS /MS and GC-MS/MS instruments.

Sludge phase extraction
The sludge samples collected from the A/O-SBR and A-SBR systems underwent several steps for analysis.First, the samples were subjected to centrifugation to separate and remove excess water.
The resulting sludge was then stored at a temperature of −20°C to preserve its integrity.To prepare the sludge for extraction, it was dried and powdered, resulting in 1 gram of dry sludge.This dry sludge was subjected to a sequential extraction process using methanol as the solvent.An ultrasonic solvent extraction method was employed, wherein 10 mL of methanol was added to the powdered sludge.The extraction process was carried out for 10 minutes at a temperature of 40°C to facilitate the extraction of target compounds.After extraction, the methanol was evaporated using nitrogen gas, leaving behind dried extracts containing the target compounds from the sludge.
To dissolve these dried extracts, they were reconstituted in 1 mL of a mixture consisting of 25% methanol and 75% ultra-pure water.The final aliquots, or representative portions of the dissolved extracts, were transferred to 2 mL amber vials.These vials were then chilled to a temperature of −20°C to maintain the stability of the extracted compounds.This preparation step was carried out in anticipation of subsequent analyses using LC-MS/MS and GC-MS/MS techniques (Albahnasawi et al. 2020).

LC-MS/MS analysis
The LC-MS/MS analysis of AAs was performed using the UHPLS System 3000 coupled with a quadrupole mass spectrometer (TSQ Fortis triple).The analysis method for AAs was qualitative, following the guidelines provided by Macherey-Nagel ( 2018).The recovery rates, separation, and detection methods by LC-MS/MS were described in Table S2 and Table S3.The relative intensity vs. retention time of targeted AAs is presented in Figure 1.

GC-MS/MS analysis
Gas chromatography-mass spectrometry (GC-MS/MS) was utilized as a method to assess the ability of the cultures to eliminate organic compounds.This analytical technique combines the separation power of gas chromatography with the identification and quantification capabilities of mass spectrometry.For the GC-MS/MS analysis, a state-of-the-art instrument, specifically Thermo Scientific's TSQ 8000 Evo Triple Quadrupole GC-MS/MS from the United States, was employed.This instrument is designed to deliver high sensitivity and accuracy in the detection of organic compounds.To facilitate the analysis, Thermo Scientific's TG-624SilMS GC Columns (30 m x 0.25 mm x 1.4 m), were utilized.These columns are specifically designed for full scan analysis, enabling a comprehensive examination of the organic compounds present in the samples.For further information on the specific techniques employed in the GC-MS/MS analysis, please refer to Table S4, which provides a detailed description of the methodology used in this study.HRT on system performance, both the A/O-SBR and A-SBR systems were operated with the same overall HRT.In the A-SBR system, for instance, with a total HRT of 12 hours, the HRT for the anoxic and aerobic phases were each 6 hours, to a total of 12 hours of HRT.Table 2 presents the detailed operation conditions for A/O-SBR and A-SBR.In the A/O-SBR system, the dissolved oxygen concentration was maintained at 0.5 mg/L for the anoxic phase and 3-3.5 mg/L for the aerobic phase, while the recycle ratio from aerobic to anoxic was kept at 1.In contrast, the dissolved oxygen concentration was controlled at 3-3.5 mg/L in the A-SBR system.Aeration was employed in both systems to regulate the dissolved oxygen concentration in the aerobic phase.Throughout the experiments, the pH of the textile wastewater was consistently maintained at 7.8, ensuring a stable and controlled environment.The temperature of the experiments was set at room temperature, ranging between 20 and 25 degrees Celsius (20-25°C).

Reactors setup and operational setup
The seed sludge used in this study was obtained from the central wastewater treatment plant in Dilovasi Industrial area, Turkey.Both the A/O-SBR and A-SBR systems were operated for 100 days, divided into five phases, as shown in Table 3.

Reactors stability and performance
The performances of A/O-SBR and A-SBR are summarized in Table 4.The TOC concentrations of A/ O-SBR effluent were around 20 mg/L for all investigated HRTs, where the TOC concentrations of A-SBR was around 25 mg/L, which indicated the insignificant effect of HRT on COD and TOC removal for both systems.Işık and Sponza (2007) reported small increase in COD removal when HRT of upper SBR treating textile wastewater increased from 2.5 h to 10 h (Işık and Sponza 2007).In terms of nitrogen removal, the TN concentration in A/O-SBR was 20 mg/L, with removal efficiency of 60%, due to the nitrification process in aerobic phase and denitrification process in anoxic phase.On the other hand, the TN concentration in A-SBR effluent remained around 40 mg/L due to absence of anoxic conditions.The same results were reported by Zhou et al. (2022) who investigated the performance of anoxic/oxic sequential batch reactor for dyeing and finishing wastewater treatment (Zhou et al. 2022).Table 5 shows that the MLSS concentration was highly dependent on the HRT.The biomass concentration in all phases decreased as the HRT increased, indicating that HRT has a significant effect on the number of bacteria present.As the HRT was extended, there was a reduction in the amount of organic matter present.This decrease meant that fewer microorganisms received nutrients, which consequently affected the growth and development of the bacteria.These findings align with the research conducted by Muda et al. (2011), who observed a decrease in the organic loading rate (OLR) with an increase in HRT.Consequently, a decline in the concentration of mixed liquor suspended solids (MLSS) can be observed at low OLR levels.This decline can be attributed to a decrease in various factors, including the oxygen uptake rate (OUR), the overall specific biomass growth rate, the endogenous decay rate (kd), and the biomass yield (Muda et al. 2011).

COD and TN removal
In all the HRTs studied, the COD concentration in the influent was around 1000 mg/L, whereas COD concentration ranged from 107-77 mg/L for A/O-SBR effluents and 150-100 mg/L for A-SBR effluents.The detailed COD concentrations in influent and effluent for the investigated systems and different HRTs are presented in Table 4.The comparison of COD removal between A/O-SBR and A-SBR during 100 experiment days and 5 HRTs is given in Figure 4.As can be seen, the average COD removal by A/O-SBR ranged from 89.3 to 92.3%,

Parameter Influent
Phases (A-SBR) Phases (A/O-SBR) whereas the COD removal efficiency by A-SBR ranged from 85% to 90%.As seen in Figure 3, the overall COD removal by A/O-SBR was higher than A-SBR, despite the using the same HRT The difference in COD removal efficiencies for HRT at 24, 16, 12, 36, and 24 h was between 2 and 6%.The superiority of the A/O-SBR system was attributed to the different metabolic mechanisms of the microbial consortium.The same results were observed by Jena et al. (2016), who reported that using anoxic conditions resulted in different bacteria that increased COD removal (Jena et al. 2016).However, the difference between the two systems was not significant in terms of COD removal, which revealed that the aerobic condition was responsible for COD removal during the decolorization of textile wastewater.The same results were reported by Kong et al. (2022), who investigated the performance of an integrated system with an anaerobic bio-electrochemical system and aerobic moving bed biofilm reactor for dye removal.The authors reported that, in terms of COD removal, aerobic operating conditions played the main role and the removal efficiency reached about 87%, which is consistent with the results of a parallel study by Kong et al. (2022).Remarkably, the HRT duration was not a significant factor affecting the COD removal by both A/O-SBR and A-SBR.For instance, when the HRT increased from 24 h to 48 h, a slight increase in COD removal was observed (2%) in A/O-SBR and (5%) in A-SBR.
In other words, the HRT of 12 h was sufficient to achieve higher COD removal (more than 85%) and no significant increase in COD removal with any further increase in length of HRT.This resonates with the findings of (Al-Amrani et al. 2014), who investigated the performance of A/O-SBR for decolorization of mixed azo dyes at different HRT (6,10 and 16 h).The authors reported that the COD removal efficiency ranged from 90% to 92% for the investigated HRTs without notable improvement in COD removal at higher HRTs (Al-Amrani et al. 2014).The measured concentrations of TN, NH 3 -N, and NO 3 -N in real textile wastewater were 45, 35, and 5 mg/L, respectively.The removal of TN occurred in the A/O-SBR system as NH 3 -N oxidized in the aerobic phase and was recycled to the anoxic compartment, where the denitrification of NO 3 -N to N 2 occurred.The concentrations of TN, NH 3 -N, and NO 3 -N in the A/O-SBR system effluent were 15, 0.5, and 15 mg/L.On the other hand, the concentration of TN in A-SBR effluent was 35 mg/L.This is attributed to the absence of anoxic environment.Looking at HRT effect on TN removal, the results indicate no increase in TN removal by A/O-SBR when HRT increased from 12 h to 48 h.This agrees with the findings by Yakamercan and Aygün (2020) who found no decrease in nitrogen effluent concentration from Anaerobic/aerobic sequential batch reactor when HRT was increased (Yakamercan and Aygün 2020).

Color removal
In this section, the performance of the investigated systems concerning color removal is presented.Table 4 provides the influent and effluent color concentrations for both the A/O-SBR and A-SBR systems, considering various Hydraulic Retention Times (HRTs).The color removal efficiency of the A/O-SBR ranged from 80.4% to 90.7% for HRTs between 12 and 48 hours.On the other hand, the A/SBR system achieved color removal efficiencies ranging from 65.5% to 77.9% at the same HRTs.Comparative results for color removal between the two systems are illustrated in Figure 5.The findings clearly demonstrate that the A/O-SBR system offers several advantages over the A/ SBR system in terms of color removal.The former system exhibited a consistently higher color removal efficiency of 10% to 18% compared to the latter, despite both systems having the same total HRT.This difference can be attributed to the presence of a reduction condition in the A/ O-SBR, which enhances dye degradation and color removal.Moreover, it is well-known that the combination of reduction environments (anaerobic/anoxic) and oxidation environments (aerobic) leads to increased color removal during the biological treatment of textile wastewater (Xu et al. 2021).This supports the notion that the A/O-SBR system's enhanced color removal performance is a result of its ability to create favorable conditions for both reduction and oxidation processes, thereby providing more effective treatment for textile wastewater.The observed phenomenon can be attributed to the breaking of azo bonds under reduction conditions, particularly in environments such as anoxic reactors.This process results in the production of colorless by-products.In a reduction environment, a reduction reaction requires the presence of an electron acceptor.Interestingly, the azo bond itself can act as an electron acceptor, facilitating the reduction reaction.As a result, the azo bond undergoes reduction, leading to the degradation of color compounds and the subsequent generation of colorless by-products.This mechanism highlights the role of the azo bond as an electron acceptor, enabling the occurrence of reduction reactions during the treatment process (L.Zhou et al. 2022).In addition, dye's structure has electron-withdrawing groups that are undegradable under aerobic conditions and degradable under anoxic conditions (Kapoor et al.

2021
).However, it is preferable to complete the anoxic treatment of textile wastewater with an aerobic process (Khehra et al. 2006;Yuan et al. 2023).Moreover, the enhancement of dye removal using A/O-SBR could be attributed to the unique microbial culture that developed between the two environments (anoxic and aerobic), resulting in the production of diverse enzymes that increase the decolorization (Jagaba et al. 2022;Ji, Li, and Ni 2022).The color removal by A-SBR was discussed in detail in our previous work (Albahnasawi et al. 2022).To sum up, the superiority of the A/O-SBR system is attributed to the diversity of microbial cultures shared between two different cultures (anoxic and aerobic) in this study, which reached up to 90%.Despite the lengthy HRT (48 hours), A-SBR removed 77% of the color.

Removal of aromatic amines
The performance of A/O-SBR and A-SBR in terms of AAs removal was investigated in this study.The sophisticated analytical technique LC-MS/MS was used to measure the concentration of 23 AAs, while GC-MS/MS was used to assess the by-products.Three HRTs (12, 24, and 48 h) were selected to study the effect of environmental conditions and HRT on the removal of AAs and organic matter.In addition, the analysis was conducted on both aqueous and sludge samples.

Aqueous phase
After the acclimatization of the sludge in the A/O-SBR and A-SBR, the HRT was controlled to study its effect on the AAs removal.The results showed that the AAs removal rate varied with different HRTs in the two investigated system.Samples for chromatography analysis were taken from effluents and sludge to evaluate the performance of A/O-SBR and A-SBR.The LC-MS/MS analysis of AAs pointed that the ∑AAs in influent was 217 ng/L.For HRT 12 h and A/O-SBR, the ∑AAs in anoxic tank was 253 ng/L and in aerobic tank was 149 ng/L indicating that an increase of AAs concentration in anoxic phase due to the cleavage of azo dyes under reduction environment and removal of the produced AAs in the aerobic phase.For example, the influent concentrations of 2,4-Dimethylaniline, 2,6-Dimethylaniline, and 2-Naphthylamine were 5.9, 21, and 60 ng/L, whereas the concentrations of these AAs in anoxic phase were 9.9, 37.1, and 89.3 ng/L, respectively.The same results were reported by Gadow and Li (2020), who reported the formation of AAs under reduction environment (anaerobic) and its removal in the following aerobic stage (Gadow and Li 2020).In another study, Yuan et al. (2023) reported complete decolorization of azo dyes and by-product using integrated reductionoxidation environment (Yuan et al. 2023).The mechanisms of azo dye cleavage could be attributed to the following paths: 1) NH 2 deamination from aromatic ring; 2) azo bond (R-N=N-R) cleavage; 3) ring open of benzene and naphthalene; 4) -SO 3 -desulfonation from sulfonated aromatic amine and 5) carbon bond cleavage for complete mineralization of carbon dioxide and water (Ali, Al-Tohamy, and Sun 2022).An increase in HRT, specifically from 12 to 24 hours and 48 hours, had a significant impact on the concentration of AAs in both the anoxic and aerobic phases of the treatment process.As the HRT increased, the AAs concentration in the anoxic phase increased, while the concentration in the aerobic phase decreased.This shift in AAs distribution indicated that the overall removal of both dye and AAs was enhanced by the longer HRT.The increase in HRT allows for extended contact time between the wastewater and the treatment process, providing more opportunities for biological reactions to occur (Friha et al. 2015).In the anoxic phase, the longer HRT facilitates the breakdown of azo dye, leading to the release of AAs.On the other hand, in the aerobic phase, the longer HRT allows for improved degradation of AAs through biological oxidation (Muda et al. 2011).By increasing the HRT, the treatment system can achieve higher AAs removal efficiency and subsequent improvement in dye removal (Lun et al. 2022).This is because the extended HRT provides sufficient time for the microorganisms to metabolize and degrade AAs and dye molecules effectively.Therefore, the findings indicate that increasing the HRT positively influences the removal of both dye and AAs in the system, as it enhances the treatment process by facilitating the degradation and transformation of these azo dye (Xie et al. 2023).
Table S5 , Table S6, and Table S7 show the concentration of AAs in the A/O-SBR and A-SBR effluents at HRTs of 12, 24, and 48 h, respectively.The ∑AAs in the studied phase was evaluated at three HRTs: 12, 24, and 48 h.Consequently, the highest concentration of AAs was observed in the anoxic phase at HRT 24 h where the high removal was obtained at HRT of 24 h in the aerobic phase.
To illustrate, at HRT 48 h no more AAs were generated at anoxic phase as no sufficient biomass could store at high HRT which decrease the dye cleavage to produce AAs.On the other hand, long HRT in aerobic phase increase the AAs removal by biological degradation.Boonnorat et al. (2019) reported that HRT play a significant role in micropollutant removal: too long HRT contributes to low daily throughput and high treatment cost, and too short HRT results in low micropollutants removal efficiency (Boonnorat et al. 2019).To sum up, when conduct A/O-SBR system, the HRT of anoxic phase must not to be long, to maintain sufficient and homogenous bacterial culture.The optimal conditions for A/O-SBR system could be anoxic phase 24 h and aerobic phase 48 h.The effect of HRT on A-SBR performance in terms of AAs removal was discussed in detail in our previous work (Albahnasawi et al. 2022).
The presence of the aerobic phase in the A/O-SBR and A-SBR demonstrated high removal of AAs, However, the AAs were still detected in the effluents revealing that AAs were not completely removed under aerobic conditions.Many previous studies reported the incomplete removal of AAs (Ning et al. 2015;van der Zee and Villaverde 2005;Xu et al. 2021).

Sludge
The handling of sludge produced from textile wastewater treatment plants is challenging due to presence of toxic chemicals, which are difficult to biodegrade and endanger human health (Lai et al. 2023;Silva Júnior et al. 2022).In this study the extraction of AAs from sludge samples was performed.The sludge samples were taken from anoxic phase and aerobic in A/O-SBR system, and aerobic phase in A-SBR system at different HRTs (12, 24, and 48 h).For A/O-SBR system, the ∑AAs in anoxic phase at HRTs 12, 24, and 48 h were 948.1, 749.7, and 722 ng/g, respectively; where the ∑AAs in aerobic phase were 421.9, 412.8, and 433 ng/g, at HRTs 12, 24, and 48 h, respectively.These results indicated that the produced AA in anoxic phase tends to sorb into sludge due to its low degradability under reduction environment.The presence of active groups such as (hydroxyl, carboxyl, amine, and sulfhydryl) on the cell surface increase the binding between AAs and organisms which makes the adsorption process possible, especially, when low degradability by bacteria is the case (Didier de Vasconcelos et al. 2021).In anoxic phase, the ∑AAs decreased by 20% when the HRT increased from 12 h to 24 h, and remained fix (around 722 ng/g) despite of HRT increase to 48 h.These results indicated that at HRT 12, the anoxic tank was subjected to high AAs load resulting in high AAs absorbed into sludge.The results of ∑AAs in aerobic phase (A/O-SBR) indicated that HRT did not have a significant effect on the AAs adsorption into sludge as the main mechanism was biodegradation of AAs.Remarkably, in terms of A-SBR system, HRT has a significant effect on the accumulated AAs on sludge.The accumulated AAs on sludge decreased by 13% when HRT increased from 12 to 48 h.This could be attributed to AAs loading rate as long HRT affords more time for bacteria to degrade the AAs.Table S8 shows the concentration of targeted AAs in the sludge samples.

GS-MS/MS full scan analysis
The analysis of GC-MS/MS scan results revealed an interesting trend regarding the impact of HRT on the generation of by-products in both the A/O-SBR and A-SBR systems.As the HRT increased, there was a corresponding increase in the number of byproducts formed.For example, when the total HRT was 12 hours, the A/O-SBR effluent contained 34 by-products, while the A-SBR effluent had 38 by-products.However, when the total HRT was extended to 48 hours, the number of byproducts increased to 37 in the A/O-SBR effluent and 42 in the A-SBR effluent.This indicates that a longer HRT in both systems can lead to higher formation of by-products, although the A/O-SBR system consistently had a lower number of by-products compared to the A-SBR system at all analyzed HRTs.The observed increase in the number of by-products with longer HRT can be attributed to the cleavage of recalcitrant organic matter by bacteria, which requires more time to break down these complex organic compounds.Consequently, a longer HRT provides a greater opportunity for the generation of diverse by-products.Figures S1-S5 present the GC-MS/MS analysis results for both the A/O-SBR and A-SBR systems, providing visual representations of the identified compounds.In line with our findings, Balapure, Bhatt, and Madamwar (2015) conducted a study on the metabolic products of azo dyes using a microaerophilic fixed film reactor and reported similar results.They found that increasing the HRT resulted in further degradation of azo dyes and their by-products, ultimately leading to the detection of more organic compounds.These findings support the notion that prolonged HRT plays a crucial role in the breakdown and transformation of complex organic pollutants (Balapure, Bhatt, and Madamwar 2015).

Conclusion
In this comparative study, we evaluated the performance of A/O-SBR and A-SBR systems for real wastewater treatment.The effect of HRT on COD, TN, and color removal was studied.In addition, chromatography analysis was conducted to study the AAs removal and by-product organic occurrence in both systems.The HRT significantly affects color removal in both treatment systems.The color removal using A/O-SBR and A-SBR at HRT of 12 h and 48 h were 78.5, 90.7, 66.8, 77.9%, respectively.The COD removal is not significantly affected by the HRT, and the average COD removal were 91.63% and 88.2% using A/O-SBR and A-SBR, respectively.Overall, the data shows that increasing the HRT led to a decrease in the accumulation of AAs, with the sludge at an HRT of 12 hours having a greater accumulation of AAs than that at 48 hours.Moreover, the study found that the A/O-SBR process was more effective in removing TN, color, and AAs compared to the A/SBR process, indicating its potential as a treatment option.However, no significant difference was observed between the two processes in terms of COD removal.These findings provide valuable insights into the factors affecting the efficiency of wastewater treatment processes and can inform the development of more effective and sustainable treatment strategies.Moving forward, further research and development efforts can focus on optimizing the A/O-SBR process to enhance its performance in removing TN, color, and AAs from textile wastewater.Additionally, exploring advanced treatment technologies and integration with renewable energy sources can pave the way for more eco-friendly and economically viable solutions in the field of textile wastewater treatment.

Figure 2 and
Figure 2 and Figure 3 presents the A/O-SBR and A/O-SBR system setups.The A/O-SBR system consisted of both anoxic and aerobic phases, each with a working volume of 15 L and a volume exchange rate of 28%.Similarly, the A-SBR system had a working volume of 15 L for the aerobic phase with the same volume exchange rate of 28%.To investigate the effect of

Figure 2 .
Figure 2. Schematic diagram of the A/O-SBR.

Figure 3 .
Figure 3. Schematic diagram of the A-SBR systems.

Figure 4 .
Figure 4. Effect of HRT on COD removal for A/O-SBR and A-SBR.

Figure 5 .
Figure 5. Color removal efficiency by A/O-SBR and A-SBR.

Table 1 .
Characteristics of real textile wastewater used in this study.

Table 2 .
Operating Conditions in Phases for A/O-SBR and A-SBR.

Table 3 .
Operating cycles for A/O-SBR and A-SBR.

Table 4 .
Performance of A/O-SBR and A-SBRin various phases.

Table 5 .
MLSS concentration of both A/O-SBR and A-SBR in various phases.