Aerobic biodegradation pathways of pentabromobiphenyl ethers (BDE-99) and enhanced degradation in membrane bioreactor system

Polybrominated diphenyl ethers (PBDEs) are a class of brominated flame retardants with high flame retardant efficiency and low cost, and are widely used in industry (Stiborova et al. 2015a, 2015b; Liu et al. 2019). The general chemical formula of PBDEs is C₁₂H_(0-9)Br_(1-10)O. According to the number and position of bromine atoms on the benzene ring, PBDEs can be divided into 209 homologous compounds. Pentabromobiphenyl ether (BDE-99), decabromobiphenyl ether (BDE-209), tetrabromobiphenyl ether (BDE-47) and hexabromobiphenyl ether (BDE-153) are the highest pollution concentration of the four monomers (Tian et al. 2015; Katima et al. 2017). However, there are few studies on BDE-99. Research shows that PBDEs are a kind of ubiquitous persistent organic pollutant (POP) (Sibiya et al. 2019). They are volatile at room temperature, have low water solubility and a lipophilic and cumulative nature, and are a serious threat to human health and the environment (Song et al. 2015). BDE-99 is the most toxic contaminant and accounts for about 12% of the total PBDEs (Dubowski et al. 2018). Toxicity can be induced at very low doses. Many countries and regions have begun to limit the use of PBDEs (Corsolini et al. 2019).

Due to the serious threat of PBDEs to human health, the research on the removal of PBDEs has attracted much attention. The treatment technology of PBDEs is mainly reductive debromination, including the photodegradation method (Wang et al. 2018a, 2018b), zero-valent iron degradation method (Yang et al. 2017), heat treatment method (Nose et al. 2007) and biological degradation (Panda & Manickam 2019). The reaction rate of PBDE treated by photodegradation method depends on many factors including the size of light source, the number of PBDEs bromine substitutions and the type of catalyst. BDE-99 belongs to low brominated diphenyl ethers, and the photodegradation rate is slow. It is necessary to use a catalyst to promote degradation (Wang et al. 2018a, 2018b). Decabromodiphenyl ether (BDE-209) did not degrade under visible light irradiation, and made use of Fe₃O₄-g-C₃N₄ hybrid to enhance the activity of photocatalytic debromination of BDE-209 and improve the degradation efficiency (Shao et al. 2018). Zero-valent iron treatment is simple in operation and low in cost, but the formation of iron oxides reduces the reducibility of the reaction, which needs to be modified to improve its degradation effect (Wang et al. 2017). The magnetic Fe⁰@Fe₃O₄ nanocomposite could achieve complete ring opening of PBDEs, but the efficiency of reduction and debromination was not high. Nose and colleagues studied the degradation of BDE-209 by using the water heat treatment method. The results showed that the degradation was not complete. Low brominated biphenyl ethers were hardly degraded (Nose et al. 2007). Biodegradation is one of the economic and effective methods for the treatment of organic pollutants, which converts toxic substances into non-toxic substances through the action of microorganisms (Panda & Manickam 2019). Aerobic degradation is completed by hydroxylation of diphenyl ether (DE), which is eventually opened and transformed by co-metabolism under the action of aerobic microorganisms (Tang et al. 2016; Ganci et al. 2019). The degradation cycle is relatively short and no toxic intermediates are produced (Lv et al. 2016). Most studies on the degradation of BDE-99 have been carried out in reaction bottles, while few studies on the sewage treatment of BDE-99 and related genes have been reported in bioreactors. The biodegradation pathways of PBDEs from gradual debromination can be determined by measuring the functional genes encoding the catalytic enzymes of PBDEs degradation (Zhao et al. 2018), which is very useful in studying biochemical degradation pathways of BDE-99.

Bioaugmentation is a method of adding a bioactive population to a biotreatment system to remove a particular class of harmful substances and to enhance the pollutants removal (Zeneli et al. 2019). Using the acid-tolerant methanogenic enrichment technology, the bioaugmentation test of anaerobic digestion and overload recovery was carried out, and the methane production was significantly increased after bioaugmentation (Li et al. 2018). Although bioaugmentation technology has been widely reported in recent years for the removal of nitrogen and phosphorus in wastewater (Yu et al. 2019), there has been no report on the application of this technology in BDE-99 wastewater treatment. In addition, as a refractory pollutant, BDE-99 usually needs some enhanced measures. The effect of surfactant and co-metabolism on its degradation is not clear. Furthermore, the associated microbial process in the bioaugmentation has not been reported.

In this study, the aerobic biodegradation pathways of BDE-99 degradation were studied. The effects of surfactants and co-metabolic substrates on BDE-99 degradation were investigated. The bioaugmentation technique was further applied to the membrane bioreactor (MBR) system to investigate the enhanced degradation effect of BDE-99. At the same time, the associated functional genes and microbial community structures in the MBR system were analyzed. The results from this study are expected to provide reference for the treatment of BDE-99 wastewater.

BDE-99 (99%), chromatographic grade methanol and toluene were purchased from Tianjin Kemiou Chemical Reagent Company. The composition of LB fluid nutrient medium was yeast extract (0.5 g), peptone (5 g), NaCl (0.25 g) and H₂O (500 mL). The pH was adjusted to 7.0. Minimal salt medium (MSM) composition was MgSO₄·7H₂O (0.4 g), CaCl₂·2H₂O (0.05 g), K₂HPO₄ (1.80 g), KH₂PO₄ (1.50 g), FeSO₄·7H₂O (0.05 g), NaNO₃ (0.50 g), (NH₄)₂SO₄ (1.00 g) and FeCl₃ (0.15 g). Twenty grams of agar was added to 1,000 mL inorganic liquid medium to make the corresponding solid medium.

The activated sludge samples for isolating BED-99 degrader were collected from the secondary sedimentation tank of Chengdu Xinjin county municipal sewage treatment plant (30°24′49″ N, 103°49′11″ E). Initially, 10 mL of activated sludge was placed in an MSM with a mass concentration of 0.5 mg/L BDE-99. The suspension was incubated in a shaking table at 140 r/min and 25 °C for 7 days. After that, the gradient pressure acclimation method was used for the strains' isolation. The BED-99 concentration in the solution was gradually increased (1, 2, 3, 4 mg/L) to 5 mg/L. After four consecutive transfers, the degradation effect was tested. The final enrichment solution was streaked onto MSM agar plates containing BDE-99 (5 mg/L) and cultured at 25 °C. Strains with good growth were selected for degradation tests to verify their degradability in MSM with BDE-99. The isolates that could effectively degrade BED-99 were selected and slant cultured at 4 °C for further study. Gram's stain reaction and cell morphology were determined under a light microscope. Strain growth (OD₆₀₀) was measured by a UV-1880 UV-visible spectrophotometer (Shanghai Mepuda Instrument Co., Ltd.).

DNA kit (TRANGEN BIOTECH) was used to extract genomic DNA from the samples, using 27F forward primer 5′-AGAGTTTGATCCTGGCTCAG-3′ and 1492R reverse primer 5′-TAGGGCTACCTTGTTACGACTT-3′ amplified its 16S rDNA by polymerase chain reaction (PCR) to identify the isolated strain. The PCR reaction system contained PCR Mix 15 μL, primers 0.3 μL and DNA templates 2 μL, and 12.7 μL double-distilled water (ddH₂O) was supplemented. PCR amplification procedure is: heated to 95 °C and held for 3 min; then modified to 94 °C and held for 1 min; hybrid annealing at 56 °C for 2 min; then heated to 72 °C and held for 3 min. This was carried out for 35 cycles, and finally, product purification and recovery was obtained at 4 °C.

The PCR product was purified and connected with pGM-T Vector. Then the DH-5α competent cells of Escherichia coli were extracted. The strain was cultured at 37 °C and screened for positive plasmid. Sequencing was carried out by Shanghai Sengong Biological Engineering Company, whose results were compared with the gene bank of the National Center for Biotechnology Information (NCBI) in the United States. The related strains were searched by BLAST in GenBank. Phylogenetic trees were constructed in MEGA 5.0 (version 6.0, USA) software.

The isolated strain was enriched and cultured at 25 °C for 7 days on a rotary shaker rotating at 150 r/min, centrifuged for 10 min and rinsed with phosphate buffer (pH = 7.4) three times. In order to obtain more comprehensive and accurate information of the degradation pathway, different concentrations of BDE-99 (0.5, 1, 2 and 3 mg/L) were set. Dissolved oxygen was maintained at normal aerobic levels. Samples were collected every 12 hours and the final analysis was based on multiple samples. The intermediates of BED-99 biodegradation were determined by gas chromatography mass spectrometry (GC-MS), Agilent 6890N, whose chromatographic column was HP-5 (60 m × 0.25 mm × 0.25 μm). Column temperature heating procedure was: the initial temperature was 90 °C, maintained for 2 min, and then increased to 210 °C at the speed of 25 °C/min, maintained for 1 min, and then increased to 275 °C at the speed of 10 °C/min, held for 10 min, then increased to 330 °C at 25 °C/min for 10 min. The carrier gas was argon, and the flow rate was 1.2 mL/min. There was no split injection, and the injection volume was 1 μL. Mass spectrometry conditions were: EI source, electron energy 38 eV, source temperature 280 °C, using the quantitative external standard curve method. The intermediate products were identified by comparing peak weight determination time with a standard database.

In the formula: C is the mass concentration of BDE-99, mg/L; K is the kinetic constant of the degradation rate of BDE-99, h⁻¹; A is A constant; t_1/2 is half-life, h.

Co-metabolism is to add another substrate to organic pollutants as a common carbon source, which can be used by microorganisms to promote growth and meanwhile promote pollutant degradation (Nsenga Kumwimba & Meng 2019). The determination process was the same as described in the section ‘Effect of surfactants on degradation’. Different co-metabolites (toluene, biphenyl ether, ethanol, ethyl acetate, glucose, and benzoic acid) of 10 mg/L were added into the bacterial solution at a 5% concentration, and were cultured in an inorganic medium of 1 mg/L BDE-99 in an oscillating manner for 7 d. The growth rate (OD₆₀₀ value) and degradation rate of the bacterial strain were measured within 7 d. All experiments were repeated three times. The inactivated strain NLPSJ-22 was used as the CK. The OD₆₀₀ value was determined by the UV-1880 visible spectrophotometer.

In the experiment, the integrated MBR reactor was used. The activated sludge was collected from the secondary sedimentation tank of the municipal sewage treatment plant in Xinjin. Inoculation volume was 0.6 L. The MBR reactor had an effective volume of 2.8 L, and was made of polymethyl methacrylate with a built-in hollow fiber membrane. The membrane assembly was made of polyvinylidene difluoride with a pore diameter of 0.1 m². The reactor did not discharge mud during operation. The reaction device is shown in Figure 1.

Artificial wastewater was used in the experiment and pH was adjusted to 7.0; (NH₄)₂SO₄ was used as the nitrogen source, and KH₂PO₄ as the phosphorus source. The trace elements in synthetic wastewater were FeSO₄·7H₂O 5,000 mg/L, CuSO₄·5H₂O 100 mg/L, ZnCl₂ 1,000 mg/L, MnCl₂·4H₂O 500 mg/L, CoCl₂·6H₂O 200 mg/L and KCl 5,000 mg/L. The composition of synthetic wastewater was based on a previous study about the growth characteristics of PBDEs degradation bacteria (Zhang et al. 2017a, 2017b). According to relevant studies (Tahhan et al. 2011; Hamjinda et al. 2017), the inoculation amount of biodegradable bacteria of the BDE-99 was set as 7% of the volume of sludge. The sludge and the degradation bacteria of the logarithmic growth period were put into the MBR system simultaneously. The control reactor without biodegradable bacteria of BDE-99 was used as the CK. According to the characteristics of BDE-99 degrading bacteria, the inlet water concentration of BDE-99 at the domestication stage was set as 3 mg/L. Continuous water was used in the process of treatment, dissolved oxygen was 3–4 mg/L, aeration was 0.5 m³/h, operating temperature was 25 °C and hydraulic retention time was 48 h.

To investigate the functional genes encoding BDE-99 degradation, samples from reactors were selected for q-PCR on an ABI 7500 real-time PCR system instrument (Applied Biosystems, CA, USA). The purified PCR products were ligated into pMD18-T vectors (TaKaRa, Japan) and transformed into competent Escherichia coli DH5a (TaKaRa). Plasmid DNA was extracted using a plasmid extraction mini kit (Transgen). The 20 μL reaction mixtures contained 10 μL of SYBR Premix Ex Taq^TM II (Takara), 0.6 μL of each primer and 1 μL of genomic DNA and 7.8 μL ddH₂O. The plasmid standard curve was obtained with 10-fold serial dilutions of the plasmid DNA. All real-time PCR standards and samples were treated in triplicate. The reaction procedure was described in our previous study (Zhang et al. 2017b). All the samples and standards for real-time PCR were analyzed in triplicate.

The PCR primer pairs encoding PBDE biodegradation enzymes were used in this study: C12OF (5′-GGCACCAAGAGCATCGAGGGCCCGTACTAC-3′), C12OR (CAGGTGCAGGTGCGCCCGCCACGGATGGCC); C23OR (5′-CCAGCAAACACCTCGTTGCGGTTGCC-3′), C23Of (AAGAGGCATGGGGGCGCACCGGTTCGATCA); C34OF (5′-CTCACGCAGCACGACATCGACCT-3′), C34OR (CCGGGCGCGACTGTCGATCGTGGT); NidAR (5′-TCAAGCACGCCCGCCGAATGCGGGAG-3′), NidAf (ATGACCACCGAAACAACCGGAACAGC); Rrl (5′-TGTTCCCGAACTTGTCCTTC-3′), Rf1 (AGGGATCCCCANCCRTGRTANSWRCA) were used for function genes amplification (Chou et al. 2016).

To explore the microbial community structure of activated sludge in different operation periods of the reactor, samples were collected every 7 days for determination by an Illumina MiSeq high-throughput sequencing instrument. And genomic DNA of sludge samples were extracted by a DNA extraction kit (PowerSoil). The extracted genomic DNA was used as the template for sample PCR amplification. The primers were V3-V4 universal primers of the Miseq sequencing platform: 515F primer 5′-GTTTCGGTGCCAGCMGCCGCCGTAA-3′ and 806r primer 5′-GCCAATGGACTACHVGGGTWTCTAAT-3′ (Chou et al. 2016). The amplification products were loaded into 1.2% agarose gel electrophoresis for purification. The DNA concentration and mass were measured by a NanoDrop 2000. The Illumina MiSeq was applied to read the microbial sequences by Shanghai Viki biotechnology company. Sequence clustering was made into an OTU operation unit for effective data with 97% similarity; sequence clustering annotation analysis was performed. Community abundance was calculated with the Mothur method. And histograms of bacterial species composition at different classification levels were drawn with Origin (2017).

All of the data were analyzed using the SPSS statistical software package (version 22.0). One-way analysis of variance followed by least significant difference tests was employed to identify differences between groups with a p < 0.05. The graphical work was conducted using Origin 8.0.

After domestication under the pressure of a BDE-99 gradient, a bacterial strain designated as NLPSJ-22, capable of utilizing BDE-99 as source of carbon and energy, was isolated. The strain NLPSJ-22 showed the more effective degradation of BDE-99 compared to other isolates. Therefore, this strain was selected for subsequent investigation. Microscopic examination revealed the strain is rod-shaped without spores and formed yellow glossy colonies. The strain was identified as Gram-negative bacteria by Gram's stain. The 16S rDNA gene sequence was determined. The phylogenetic relationship of strain NLPSJ-22 and its closest species is shown in Figure 2. Strain NLPSJ-22 was homologous with Pseudomonas asplenii and had 100% similarity. Based on its morphological and 16S rRNA gene sequence analysis, strain NLPSJ-22 was identified as Pseudomonas sp. The strain NLPSJ-22 was deposited in GenBank with the accession number of MK793699.

The intermediates of BDE-99 biodegradation by NLPSJ-22 were monitored by GC-MS (Agilent 6890N). The major m/z fragments were detected and analyzed after BDE-99 biodegradation through full scan mode GC-MS. The monitored ion fragments included 72.7, 797, 81.4, 161, 208.7, 354.3 and 427.8. According to the analysis of degradation products and previous studies (Stiborova et al. 2015a, 2015b; Liu et al. 2019), it was speculated that BDE-99 was debrominated into two tetrabromodiphenyl ethers (BDE-49 and BDE-47) (Figure 3). BDE-49 was converted to BDE-17 by contraposition substitution, while BDE-47 was converted to two tribromodiphenyl ethers (BDE-28 and BDE-17) by interposition substitution. BDE-28 and BDE-17 were meta-transformed into BDE-15 and BDE-4 respectively and ended up in DE. Aerobic degradation is completed by hydroxylation of DE, which is eventually opened and transformed by co-metabolism under the action of aerobic microorganisms. Under the action of 2,3-dihydroxy diphenyl dioxygenase, DE changed into 1,2-dihydroxy diphenyl ether. And under the action of 1,2-dioxygenase, the ring was opened and cleaved to phenol. Finally, under the condition of 2-hydroxy muconic acid into the tricarboxylic acid (TCA) cycle.

The genes encoding degradation enzymes associated with biodegradation of BDE-99 were tested by q-PCR. Because the chemical structure of BDE-99 consists of two benzene rings, dioxygenase may initiate a series of benzene ring cleavage reactions under aerobic conditions. Based on previous studies (Brezna et al. 2003; Kumar & Khanna 2010; Stiborova et al. 2015a, 2015b; Chou et al. 2016; Petrova et al. 2019), functional genes similar to those for aerobic biodegradation of POPS were selected in this study. As shown in Table 1, only two functional genes were detected, namely catechol 1,2-dioxygenase (C12O) and catechol 2,3-dioxygenase (C23O). However, catechol 3,4-dioxygenase (C34O), the terminal dioxygenase NidA polypeptide (NidA) and a-subunit gene fragments of dioxygenase (Rf) were not detected. The number of degradation genes C12O and C23O increased gradually with the detection time, and gradually decreased after reaching the highest value (5.9 × 10⁴ copies/L and 6.4 × 10⁴ copies/L) on the fifth day.

To investigate the degradation ability of the strain to BDE-99 under different concentrations, the initial concentrations of 0.5, 1, 2, 3, 4 and 5 mg/L were set. The degradation experiments were conducted at room temperature, 150 r/min and pH = 7. According to the data obtained in the experiment, the degradation curve was drawn as shown in Figure 4.

By fitting the curve in Figure 4, it was found that the biodegradation of BDE-99 by strain NLPSJ-22 conforms to the first-order reaction kinetic equation.

The kinetic equation and half-life were arranged as shown in Table 2. When the initial mass concentration of BDE-99 was 3 mg/L, the degradation rate of BDE-99 was the fastest, and its half-life period was the shortest at 8.3 h in the measured concentration. With the increase of initial concentration to 2 mg/L, and from 3 mg/L to 5 mg/L, the degradation rate constant decreased and the half-life became longer. The results showed that the degradation half-life of strain NLPSJ-22 to BDE-99 was 8.35–10.04 h.

To investigate the effect of surfactants on the degradation of BDE-99 by NLPSJ-22, different concentrations of triton, rhamnolipid, tea saponin, Tween-80, SDS and sucrose fatty acid ester were added to the medium, and no surfactant was added as the CK. The effect of surfactant on degradation efficiency under different dosage is shown in Figure 5. In general (CK), the degradation efficiency of BDE-99 was 84%. The addition of all surfactants at a concentration of 5% promoted the degradation of BDE-99, especially SDS, where the degradation efficiency was as high as 93%. In addition, when the concentration of SDS was 1, 2 and 3%, the degradation of BDE-99 was inhibited. The higher the concentration, the lower the degradation rate; the lowest was 42%. Triton, rhamnolipid, tea saponin, and Tween-80 promoted degradation at concentrations of 0.5, 1 and 2%, and inhibited degradation at concentrations of 3%. When the concentration of rhamnolipid was 1%, the degradation efficiency was as high as 94%. Sucrose fatty acid ester promoted the degradation of bacterial strains when the concentrations were 0.5 and 1%, and inhibited the degradation of bacterial strains when the concentrations were 2 and 3%.

The CSH of bacteria is one of the factors that determine the non-specific adhesion of bacteria to biological and abiotic surfaces, which affects the ability of microorganisms to absorb organic compounds (Feng et al. 2019). As can be seen from Figure 6, anionic surfactants (rhamnolipid and SDS) had the most obvious effect on the degradation of BDE-99, with the highest CSH. The degradation rate was up to 94.1% and the CSH increased by 23% under the effect of rhamnolipid. The degradation rate and CSH were still improved under the effect of four nonionic surfactants (triton, tea saponin, Tween-80, sucrose fatty acid ester), but were slightly lower than those of the two anionic surfactants. In this experiment, the enhanced effect of anionic surfactants on the degradation of pollutants was better than that of nonionic surfactants. The degradation rate of BDE-99 was positively correlated with the CSH.

Biodegradation of BDE-99 can be improved by adding carbon sources (co-metabolism substrates) (Feng et al. 2019). In this study, the effects of six co-metabolism substrates of toluene, diphenyl ether, ethanol, ethyl acetate, glucose and benzoic acid on the degradation of BDE-99 by NLPSJ-22 were investigated. It can be seen from Figure 7 that the various co-metabolism substrates all had impacts on the degradation efficiency of BDE-99 and the growth of the strain (OD₆₀₀ value). The degradation rate in the CK system was 84.7%. After adding toluene, diphenyl ether, ethanol and ethyl acetate, the degradation rate of BDE-99 was above 85%. In particular, the promotion of degradation by diphenyl ether was significant and the degradation efficiency was as high as 96%. However, glucose and benzoic acid did not show the enhanced degradation of BED-99. Glucose was the first to be used for microbial growth; therefore the degradation of BED-99 decreased. Probably because of the toxicity of benzoic acid, the growth of the strain was restricted, and the lowest OD₆₀ was observed when benzoic acid was added., The other five co-metabolic substrates all promoted the growth of the bacterial strain and the OD₆₀₀ showed an upward trend.

To evaluate the performance of bioaugmentation, the removal rates of BDE-99 in the MBR reactor was measured every day. During the start-up and operation of the reactor, the concentration of BDE-99 was set at 3 mg/L in both the bioaugmented reactor and the control reactor. And hydraulic retention time was 48 hours. The degradation efficiency of BDE-99 in each period of reactor operation is shown in Figure 8. The degradation trend of the two reactors is roughly the same. The start-up time was long in the control reactor, which was not stable at the initial stage (1–5 days). In the middle stage (6–47 days), the degradation efficiency started to increase gradually. The bioaugmented reactor entered the stable operation stage from the 48th day. The stability means that the fluctuation of degradation efficiency was very small and tended to be the same. After domestication, the degradation rate gradually increased to 60% in the control reactor, whereas the removal rate of BDE-99 in the bioaugmented reactor was above 80% after 35 d. The degradation efficiency was reduced in the bioaugmented reactor on days 55–60 because the reactor was affected when the experiment was about to end and the reactor was moved. The degradation trend of the two reactors is roughly the same. The degradation rate of BDE-99 in the bioaugmented reactor was constantly higher than in the control reactor.

To understand the relationship between the change of community structure and BDE-99 removal during operation, the activated sludge in the bioaugmented reactor (Figure 9(a)) and the controlled reactor (Figure 9(b)) was collected at different operating periods. As shown in Figure 9, there were obvious microbial community shifts at the phylum level during different periods. The general trends in microbial change in the two bioreactors were similar. Chloroflexi, Clostridia and Planctomycetes constituted a relatively stable proportion throughout the trial. The relative abundance of Actinobacteria, Firmicutes, γ-Proteobacteria, β-Proteobacteria and α-Proteobacteria gradually increased, whereas Fusobacteria, Acidobacteria, Bacteroidetes, ɛ-Proteobacteria and δ-Proteobacteria progressively decreased. In addition, the relative abundance of γ-Proteobacteria in the control reactor increased from 4% to 9%. After bioaugmentation, it increased from 4% to 18%. This indicated that bioaugmentation promoted the growth of γ-Proteobacteria. Proteobacteria is the most abundant phylum in the two reactors, accounting for 48% (bioaugmented reactor) and 44% (control reactor) of the total bacterial population, indicating that it is likely to be the main phylum involved in the BDE-99 aerobic biodegradation process.

The bacterial community dynamics at the genus level is shown in Figure 10. The original activated sludge mainly consisted of Micropruina, Thauera, Nitrospira, Nitrosomonas and Acinetobacter. Their relative abundances were above 10%. As the reaction progressed, the relative abundance of the five genera gradually decreased in the two reactors. The difference in the composition of microbial communities between the two reactors was not significant. Thauera and Acinetobacter were gradually eliminated from the bioaugmented system (Figure 10(a)), whereas these genera persisted in the control reactor (Figure 10(b)). The relative abundance of Comamonas gradually decreased to 0% in the control reactor, but was eliminated early in the bioaugmented reactor (30–40 d). In contrast, Rhodococcus, Bacillus, Pseudomonas, Burkholderia and Sphingobium progressively increased in both reactors. In addition, Pseudomonas increased most obviously in the bioaugmented reactor, from 5% to 24%, which is about twice that of the control system. Bacillus increased most distinctly in the control reactor, from 4% to 19%, which was more significant than Pseudomonas in the control reactor. These data indicated that Pseudomonas is closely related to the biodegradation of BDE-99.

In this study, Pseudomonas NLPSJ-22 was isolated from activated sludge, with the ability to degrade BDE-99. Pseudomonas strains exist widely in the environment and can effectively degrade environmental pollutants. In the anaerobic environment, researchers (Chang et al. 2016) detected Pseudomonas, which played an important role in the biodegradation of BDE-209 during the whole process using organic compost. The crude enzyme extracted from Pseudomonas aeruginosa was used to biodegrade BDE-209, with a degradation rate of 92.77% (Liu et al. 2015). The bacteria that can degrade BDE-99 reported in the past study mainly belong to genera of Stenotrophomonas (Wu et al. 2018) and Bacillus (Lu et al. 2013; Yu et al. 2019). Therefore, the degradation mechanism of BDE-99 by Pseudomonas needs to be further studied.

The biochemical degradation pathways of BDE-99 using strain NLPSJ-22 were studied. The strain could gradually debrominate BDE-99 into DE. DE transformed to phenol under the action of dioxygenase, and finally entered the TCA cycle under the action of 2-hydroxyl muconic acid. BDE-49 was only detected from BDE-99 by Huang et al. (2014), but BDE-47 was not detected. Lv and colleagues found that BDE-47, BDE-15 and BDE-3 could gradually enter the TCA cycle through angular dioxygenation (Lv et al. 2016). In our experiment, it was speculated that tetrabromodiphenyl ethers (BDE-49, BDE-47) were all the possible intermediate products of BDE-99 degradation, which indicated that the degradation of BDE-99 could be debrominated in various ways. BDE-49 was converted to BDE-17 by para-substitution, and then to BDE-4 by inter-substitution. BDE-47 could be converted to BDE-17 or BDE-28 by inter-substitution, and BDE-28 could be converted to BDE-15 by inter-substitution. Only BDE-3 could be converted to DE in the end. The process of DE ring opening was similar to a previous study (Waaijers & Parsons 2016). DE was decomposed by Pseudomonas NLPSJ-22 into the homologues of phenol and catechin, and finally entered the TCA cycle. The degradation process of PBDEs is generally: reduction and debromination, hydroxylation, ring opening. It can be concluded that the degradation pathways of BDE-99 are diverse. Therefore, further studies on the metabolic mechanism of BDE-99 biodegradation are needed to improve the removal of BDE-99 in various environments.

In this study, only two types of degradation genes were detected during the biodegradation process. It can be inferred that the degradation genes C120 and C23O in Pseudomonas NLPSJ-22 can cleave the benzene ring, which improved the biodegradation of BDE-99. In another study (Chou et al. 2016), only C23O was detected to decompose catechol, which is in line with this experiment. C23O was detected in Pseudomonas, and gene C23O was more abundant than gene C12O. It is speculated that C23O may be the main degradation gene of Pseudomonas promoting BDE-99 aerobic degradation.

The rhamnolipid concentration of 1% is most effective for the degradation of BDE-99, but SDS almost inhibited the degradation of BDE-99. This was similar to the case of Liu et al. (2018) who used Phanerochaete chrysosporium to remove BDE-47 efficiently. The promotion of low concentration rhamnolipid was the most obvious. The reason is that rhamnolipid affects the growth of degrading bacteria and enzyme activity (Liu et al. 2018). SDS (inhibition) and rhamnolipids (promotion) were anionic surfactants that had a significant impact on degradation. Some studies had shown that the effect of surfactants on biodegradation is related to CSH (Petrova et al. 2019). In this study, anionic surfactants were superior to nonionic surfactants in enhancing the degradation of contaminants. This is different from the results of Zhang's research on surfactant degradation of dioctyl phthalate (DOP), indicating that surfactants had different degradation effects on different pollutants (Zhang et al. 2017a, 2017b). It is speculated that surfactants with different properties have different arrangement of hydrophilic groups on different pollutants, leading to different degradation effects. Theoretically, as the concentration of BDE-99 increases, the half-life increases (Zhang et al. 2017a, 2017b). But in our experiment, when the concentration was 2.0 mg/L, it had the longest half-life. The possible reason is that a large number of inactivated bacteria were found in this group of experiments.

When different co-metabolism substrates were added, glucose and benzoic acid had an inhibitory effect on degradation. The reason is that glucose is easy to use; therefore, the strain preferentially uses glucose and reduces the use of BDE-99, which leads to a decrease in the degradation rate of BDE-99 (Lv et al. 2016; Zhang et al. 2017a, 2017b). In addition, due to experimental defects, the concentration of the co-metabolism substrate was 10 mg/L. Experiments with other concentrations of co-metabolism substrates were not completed. Therefore, it is unclear whether the degradation rate and growth rate of the strain will be better at the other concentrations. Therefore, further research is needed to fully understand the effect of co-metabolism on BDE-99 degradation.

The strain NLPSJ-22 was used in the MBR system to enhance the degradation of BDE-99. From the perspective of system start-up time, the bioaugmented system shortened the reactor start-up time. In the first 10 days, both reactors were unstable and the rate of degradation of the control reactor was higher than that of the bioaugmented reactor. The reason may be that the strain is in the adaptation period, and the removal of BDE-99 is mainly based on adsorption. When it comes to bioaugmentation in the MBR, how to keep it working over long periods of time was the key. Although this experimental strain (NLPSJ-22) showed stability during the whole experiment, the long-term maintenance has yet to be further verified. Bioaugmentation has been gradually applied in various organic degradation experiments. Hamjinda and colleagues used Pseudomonas to bio-enhance ciprofloxacin in a two-stage MBR, and the degradation rate was increased by 30% and had a better nitrogen removal effect (Hamjinda et al. 2017). By studying the effect of continuous inoculation of bioaugmented degrading bacteria on the degradation of petroleum hydrocarbons, it was found that the total amount of petroleum hydrocarbon removal increased by more than 30%. Continuous bioaugmentation was better than pure bioaugmentation (Tahhan et al. 2011). However, some research results showed that the addition of degrading bacteria can only accelerate the start-up of the bioaugmented reactor (Yu et al. 2010). And the removal of pollutants was not obvious due to the loss of strains in the later stage of operation. Therefore, maintaining the number of strains in the reactor is the key to the stable operation of the system.

Microorganisms are the main drivers for pollutants removal in wastewater treatment. And the stable operation of the reactor depends on the structure of the microbial community (Huang et al. 2016). During the operation of the bioaugmented reactor, Proteobacteria and Bacillus were the dominant bacteria. It is speculated that they have obvious correlation with refractory compounds such as BDE-99 and DOP (Zhang et al. 2017a, 2017b). In this study, the introduction of BDE-99 resulted in the disappearance of Comamonas. In the process of bioaugmented degradation, Comamonas disappeared 20 days earlier than in the control reactor. In the bioaugmented system, the three genera of Comamonas, Thauera and Acinetobacter disappeared. Most of these bacteria are nitrogen and phosphorus removal microorganisms (Yu et al. 2019). It showed that the nitrogen and phosphorus removal bacteria in the system were gradually decreasing. Therefore, whether the reactors still have the ability to remove nitrogen and phosphorus needs to be investigated in the following study. Although our results indicated that the bioaugmentation technique could significantly improve the removal rate of BDE-99, it has potential in the treatment of BDE-99 wastewater. However, in engineering application, it is a great challenge to keep the functional bacteria in the reaction system in the long run. We will also carry out further pilot experiments.

A bacterial strain NLPSJ-22 was isolated and identified as Pseudomonas sp. BDE-99 can be completely mineralized by NLPSJ-22. The strain gradually debrominated BDE-99 into DE and underwent ring-opening cleavage under the action of dioxygenase, and finally entered the TCA cycle. Both C12O and C23O degradation genes were detected during BDE-99 degradation by NLPSJ-22. Rhamnolipid as surfactant and diphenyl ether as co-metabolic substrate significantly promoted BDE-99 degradation, while SDS and glucose did not promote the removal of BDE-99. Compared with the control system, bioaugmentation could effectively shorten the start-up time of the bioreactor and significantly improve the degradation rate of BDE-99. Rhodococcus, Bacillus, Pseudomonas, Burkholderia and Sphingobium were the main genera involved in the degradation of BDE-99. This study indicated that the strain NLPSJ-22 and bioaugmentation technique had potential in BDE-99 wastewater treatment.

This work was supported by the National Natural Science Foundation of China (No. 51808363), and the Development Project of Science and Technology benefitting the Public in the Science & Technology Bureau of Chuengdu City (No. 2015-HM01-00325-SF).

All relevant data are included in the paper or its Supplementary Information.