Effectiveness of tetrahydrofuran at enhancing the 1,4-dioxane degradation ability of activated sludge lacking prior exposure to 1,4-dioxane Open Access

1,4-dioxane (DX) is a synthetic organic compound that is widely used as an extraction solvent as well as a cleaning agent in various chemical industries (Zenker et al. 2003). DX is also inevitably formed as an undesirable by-product during the manufacture of ethylene oxide, ethylene glycol, and detergents, such as alkyl ether sulfate (Zenker et al. 2003). Consequently, wastewater generated in specific industries contains certain levels of DX, in the range of 0.85–730 mg/L (Lee et al. 2015; Isaka et al. 2016; Mahmoud et al. 2017). DX is a suspected human carcinogen and can persist in water environments for a long period due to its physicochemically stable characteristics (Godri Pollitt et al. 2019). Therefore, DX is a contaminant of emerging concern in water environments, and appropriate treatment of DX-containing wastewaters is strongly desired to mitigate its risks.

Although DX is demonstrated to be highly recalcitrant from biodegradation, bacterial strains that can degrade DX metabolically or co-metabolically have been reported in the last two decades (Zhang et al. 2017; Li et al. 2018). According to the evidence, biological treatment using DX-degrading bacteria can be a promising strategy for treating DX-contaminated wastewater. There have been several successful applications of specific DX-degrading strain isolated elsewhere for the treatment of DX-contaminated wastewater (Isaka et al. 2016; Myers et al. 2018). However, the use of a pure culture of DX-degrading bacteria might not be effective owing to the difficulty of avoiding contamination worsening the DX-degradation ability, and the possibility of incomplete mineralization of DX and accumulation of dead-end intermediates (Vainberg et al. 2006). Thus, an effective DX treatment strategy using microbial consortia is expected for the practical treatment of DX-contaminated wastewater.

The enrichment of DX-degrading bacteria indigenous to activated sludge in a wastewater treatment plant is a potential strategy for enhancing the intrinsic DX-degrading ability. According to previous studies, the DX-degrading microbial consortia can be enriched using DX as the sole substrate from various microbial communities (Nam et al. 2016; He et al. 2018; Inoue et al. 2020; Ramalingam & Cupples 2020). However, the use of DX to enrich DX-degrading bacteria is not practically desirable owing to its low cell yield (Yamamoto et al. 2018b) and high toxicity. Thus, the use of alternative substrates with lower toxicity and more readily biodegraded than DX might be a promising option for the selective enrichment of DX-degrading bacteria in wastewater treatment.

Tetrahydrofuran (THF) is a favorable growth substrate for most DX-degrading bacteria (Mahendra & Alvarez-Cohen 2005; Sei et al. 2013a; Inoue et al. 2018; Yamamoto et al. 2018b), and is used as a primary substrate for co-metabolic DX degradation (Zhang et al. 2017). Recently, we found that the DX degradation ability of the microbial consortium in DX-contaminated industrial waste landfill leachate can be significantly enhanced by repeated cultivation on THF, and the resultant DX degradation ability did not rely on co-metabolism, but metabolism (Inoue et al. 2020). In addition to the inherent properties of THF, which is readily biodegradable and not as environmentally persistent as DX, and has low toxicity and bioaccumulation potential (Fowles et al. 2013), the results of our previous study suggested that THF can be a useful substrate for the selective enrichment of DX-metabolizing bacteria. However, there have been limited attempts of THF supplementation to boost the intrinsic DX degradation ability of environmental microbial consortia (Sei et al. 2010, 2013b; Xiong et al. 2019). In particular, its effectiveness remains unexplored on the microbial consortia that have not experienced historical exposure to significant DX and thus expectedly held only marginal abundance of DX-degrading bacteria.

This study aimed to determine the effectiveness of THF for enhancing the DX-degrading ability of activated sludge lacking historical exposure to DX. Activated sludge from a municipal wastewater treatment plant was cultivated on THF in a sequential batch mode. The DX degradation ability of the resultant enrichment culture was assessed without the co-presence of THF and compared with that of another enrichment culture cultivated on DX. The changes in the abundance and compositions of bacterial communities and soluble di-iron monooxygenase (SDIMO) genes were also tracked before and after the sequential cultivations. Some of the SDIMO families are known to be associated with the initial oxidation of the carbon atom adjacent to the oxygen atom in DX, which is a rate-limiting step in DX biodegradation and precedes the cleavage of the high-energy C-O bond of the cyclic ether structure (White et al. 1996). Thus, SDIMO genes have been applied as key functional genes for identifying potential DX-degrading populations in microbial consortia (He et al. 2018; Xiong et al. 2019; Inoue et al. 2020).

Activated sludge collected from a municipal wastewater treatment plant in Osaka, Japan, on October 17, 2017, was used as the seed for the enrichment experiments. This wastewater treatment plant mainly receives domestic wastewater, and has not experienced the inflow of detectable DX. The activated sludge sample was transported on ice to the laboratory, centrifuged (2,000×g, 4 °C, 5 min) and washed twice with basal mineral medium (BMM; pH 7.0) (Yamamoto et al. 2018b).

The sludge sample was inoculated into 300 mL-Erlenmeyer flasks containing 100 mL of BMM supplemented with 100 mg-C/L of DX or THF as the sole external carbon source and 5 mg/L of allylthiourea for nitrification inhibition. The enrichment cultures supplemented with DX and THF were designated as DX-enrichment and THF-enrichment, respectively. The initial sludge concentration was adjusted at 500 mg/L as total suspended solids (TSS). Abiotic control without sludge inoculation was prepared for both DX and THF to evaluate their volatilization. The flask for DX-enrichment was capped with a silicone plug, while that for THF-enrichment was tightly sealed with a rubber stopper to avoid extreme volatilization (vapor pressure: 3.9 and 18.9 kPa at 20 °C for DX and THF, respectively). All test systems were prepared singly. The cultures were incubated at 28 °C in the dark with rotary shaking (120 rpm). At regular intervals of 1 week, 40 mL of the culture was collected, washed once with BMM, and concentrated twice via resuspension to 20 mL BMM. Subsequently, 10 mL of the concentrated culture was transferred to fresh medium. The enrichment process consisted of a total of eight batch cycles. During the enrichment process, the concentrations of DX, THF, and dissolved organic carbon (DOC) in the medium were measured periodically.

The DX degradation abilities of the seed sludge and enrichment cultures after the 2nd, 4th, 6th, and 8th cycles were evaluated via single batch experiments. For enrichment cultures, the concentrated ones, which were prepared for transfer to the next enrichment cycles, were applied. One milliliter of the seed sludge or enrichment culture was inoculated in 50-mL vials containing 19 mL of BMM with DX at a final concentration of 100 mg-C/L. An abiotic control was also prepared. The cultures were incubated under the above cultivation conditions, and DX concentrations were determined periodically for 21 or 22 days. All experiments were conducted in duplicate.

TSS concentration was measured according to the method of dry cell weight concentration. Briefly, the sample was filtered through a Whatman GF/B filter (pore size, 1.0 μm; GE Healthcare Life Sciences, Buckinghamshire, UK) and dried at 110 °C for 2 h. The dry cell weight was determined as the difference in filter weight with and without dried cells. DOC concentration was analyzed using the total organic carbon analyzer, TOC-V_CSH (Shimadzu, Kyoto, Japan). DX and THF concentrations were determined using a gas chromatograph GC-2014 (Shimadzu) equipped with a flame-ionization detector (FID), as described previously (Inoue et al. 2020). The samples used for analyses of DOC, DX, and THF were prepared via centrifugation (1,500×g (for DOC) or 10,000×g (for DX and THF), 4 °C, 5 min), filtration through a 0.45-μm cellulose acetate filter (Advantec, Tokyo, Japan), and acidification with 2 M HCl.

The seed sludge and enrichment cultures after the 6th and 8th cycles (prior to batch DX degradation experiments), which were stored at −20 °C, were subjected to microbial community analyses. Microbial DNA in each sample was extracted with the FastDNA SPIN Kit for Soil (MP Biomedicals, Solon, OH, USA).

Illumina MiSeq sequencing was carried out by targeting 16S rRNA and SDIMO genes at Bioengineering Lab. Co., Ltd (Kanagawa, Japan), as described previously (Inoue et al. 2020). Briefly, primers 515F (5′-GTGCCAGCMGCCGCGGTAA-3′) and 806R (5′-GGACTACHVGGGTWTCTAAT-3′) were used to amplify the V4 region of bacterial 16S rRNA genes, while primers NVC57 (5′-CAGTCNGAYGARKCSCGNCAYAT-3′) and NVC66 (5′-CCANCCNGGRTAYTTRTTYTCRAACCA-3′) were used to amplify the SDIMO α-subunit genes (Coleman et al. 2006). The current grouping of SDIMO is supported by phylogenetic analysis of α-subunit proteins (Coleman et al. 2006). Sequencing was performed on an Illumina Miseq platform (Illumina, San Diego, CA, USA) with a 2×300 bp paired-end sequencing. The raw sequence reads for 16S rRNA and SDIMO genes were deposited in the DNA Data Bank of Japan (DDBJ) Sequence Archive database under the accession numbers, DRA010369 and DRA010370, respectively.

The raw sequence reads of the 16S rRNA genes were processed as described previously (Inoue et al. 2020), where clustering into operational taxonomic units (OTUs) was conducted with a 97% similarity threshold. The alpha diversity of the bacterial communities was evaluated with the Shannon index.

The raw sequence reads of the SDIMO genes were processed according to a previous study, with minor modifications (Inoue et al. 2020). Representative sequences from all OTUs of the obtained SDIMO genes were transformed to the deduced amino acid sequences, and the closest amino acid sequences were identified using a NCBI BLAST search (https://blast.ncbi.nlm.nih.gov/Blast.cgi). The OTUs whose deduced amino acid sequences did not fit SDIMO were excluded from further analysis. The remaining OTUs were classified into the six SDIMO groups (groups 1–6) based on the deduced amino acid sequence (He et al. 2017), whereas SDIMO group 5 was further divided into three subgroups 5A, 5B, and 5C, which were associated with the propane/methane monooxygenase (PRM/MMO) in Gram-positive bacteria, the PRM/MMO in Gram-negative bacteria, and the THF/DX monooxygenase (THM/DXM), respectively (Inoue et al. 2020).

The concentrations of total bacterial 16S rRNA genes and THM/DXM genes (SDIMO group 5C) were quantified by real-time PCR. Bacterial 16S rRNA genes were amplified using primers 331F and 797R (Nadkarni et al. 2002). Primers thm-F and thm-R designed based on the thmC gene in Pseudonocardia sp. D17 (Yamamoto et al. 2018a) were applied to specifically quantify the THM/DXM gene abundance. Real-time PCR was carried out on an ABI Prism 7000 sequence detection system (Applied Biosystems, Foster City, CA, USA). PowerSYBR Green PCR Master Mix (Applied Biosystems), and GeneAce SYBR qPCR Mix α (Nippon Gene, Tokyo, Japan) were respectively used as the SYBR Green real-time PCR reagent for quantification of 16S rRNA and THM/DXM genes, and PCR amplifications were performed with the thermal profiles described previously (Nadkarni et al. 2002; Inoue et al. 2020). Standard DNA for 16S rRNA and THM/DXM genes was respectively prepared from genomic DNA of Escherichia coli HB101 and Pseudonocardia sp. D17 (Sei et al. 2013a) using a DynaExpress TA PCR cloning kit (pTAC-1) with Jet Competent Cell (BioDynamics Laboratory, Tokyo, Japan). The detailed procedures were described in our previous report (Inoue et al. 2020). The PCR amplification efficiencies for 16S rRNA and THM/DXM genes were 83–96% and 79–90%, respectively.

16S rRNA gene amplicon sequencing of the seed sludge and enrichment cultures after the 6th and 8th cycles yielded 25,129–43,497 bacterial 16S rRNA gene reads per sample (157,081 reads in total), which were classified into 4,596 OTUs (Table 1). Variations in the OTU number and Shannon index revealed the reduction of bacterial community diversity owing to enrichment with DX or THF.

The OTU compositions of the SDIMO genes in seed sludge and enrichment cultures are shown in Figure 4(a) and 4(c), respectively. In seed sludge, 18 OTUs were detected. OTU 5 and OTU 10, both of which were allocated to group 5B (PRM/MMO in Gram-negative bacteria), were the two most dominant OTUs, with relative abundances of 38.2% and 21.6%, respectively. As a whole, group 5B was most dominant, accounting for 75.7% of the total SDIMO genes. After enrichment with DX or THF, almost all OTUs detected in seed sludge decreased, and alternatively limited numbers of OTUs (3–7 OTUs) appeared. Further, the dominant OTUs differed depending on the enrichment substrate. In the DX-enrichment culture, OTU 2 allocated to group 5A (PRM/MMO in Gram-positive bacteria) became most dominant, followed by OTU 4 (group 5A). Consequently, group 5A was most dominant, accounting for 94.2% and 87.1% at the 6th and 8th cycles, respectively. Besides, OTU 9 allocated to group 5B increased to 10.4% at the 8th cycle. However, in the THF enrichment culture, group 5C (THM/DXM), which was almost exclusively contributed by OTU 1, was most dominant at 91.8% and 81.3% at the 6th and 8th cycles, respectively. Group 5A (mainly OTU 6) also increased to 17.2% at the 8th cycle.

In our recent study, we demonstrated the DX degradation ability in industrial landfill leachate microbial consortium by repeated cultivation on THF (Inoue et al. 2020). The microbial consortium had continuously been exposed to DX for a long period and consequently had detectable DX degradation activity. In contrast, this study used activated sludge in a municipal WWTP, which was not previously exposed to DX and had a detectable DX degradation ability, as the seed microbial consortium (Figures 1(a) and 2). Nevertheless, DX degradation ability could be induced by four-times repeated cultivation on THF (Figure 2). Therefore, the proposed strategy is applicable to enhancing the DX degradation ability of microbial consortia, irrespective of the intrinsic ability.

In addition to the present study, recent studies succeeded in enriching DX-degrading bacteria from uncontaminated soil (Nam et al. 2016; He et al. 2018; Ramalingam & Cupples 2020) and activated sludge (Nam et al. 2016), although they applied DX as the enrichment substrate. Thus, although the environmental distribution of DX-degrading bacteria had been regarded to be limited (Sei et al. 2010), DX-degrading bacteria may be distributed relatively extensively in the environment, irrespective of the pollution history. The minor presence of DX-degrading bacteria in the environment might be caused by the shortage of their specific growth substrates, their low growth rate, etc., which should be clarified in a further study.

Surprisingly, THF could awaken the DX degradation ability after 4 weeks, which was the effectiveness for DX-enrichment culture, although the strength of DX degradation ability differed between the two enrichment (Figure 2). Such finding is inconsistent with that of our previous study where the DX degradation ability was enhanced earlier by cultivation on DX than THF (Inoue et al. 2020). As shown in Figure 1(b), THF could be biodegraded and mineralized from the initial cultivation cycle. Previous studies revealed higher cell yields on THF than DX by DX-metabolizing bacteria (Chen et al. 2016; Inoue et al. 2018). The results suggest that although THF is inferior to DX as the enrichment substrate of DX-degrading bacteria in terms of selectivity, its readily degradability and high cell yield allowed the enrichment of DX-degrading bacteria and consequent enhancement of the overall DX degradation ability with an equivalent effectiveness to DX itself. Thus, THF may be a promising biostimulation agent to enable the potential DX biodegradation activity of microbial consortia, where DX-degrading bacteria is present quite marginally.

As aforementioned, THF has been well-known as the primary substrate for co-metabolic DX degradation (Zhang et al. 2017). Therefore, previous studies attempted to induce the co-metabolic DX degradation by simultaneous supplementation of DX and THF (Sei et al. 2010; Xiong et al. 2019), and consequently the usability of THF to enrich DX-metabolizing bacteria remains unclarified. In this study, only DX was amended to the batch DX degradation experiments, and the enrichment culture was washed and thereafter used in batch DX degradation experiments. Owing to the experimental settings, the occurrence of co-metabolic DX degradation using THF as the primary substrate can be ignored or is minor in our batch DX degradation experiments, and the induction of DX biodegradation by THF-enrichment culture was judged to be attributed to the enrichment of DX-metabolizing bacteria capable of utilizing DX as the sole carbon source. Therefore, interestingly, our results indicate that DX-metabolizing bacteria that can use both THF and DX as the sole carbon source would be enriched even by cultivation with THF.

Although the DX degradation ability was enhanced by repeated cultivation on both THF and DX, their microbial communities in THF- and DX-enrichment cultures differed notably in terms of the overall phylogenetic composition, dominant populations (Figure 3), and SDIMO gene composition (Figure 4). The difference in the microbial community likely resulted in distinct strength of DX degradation ability of the enrichment cultures (Figure 2). As summarized in Table 2, the dominant microbial populations and SDIMO groups in DX-degrading enrichment cultures differed between this study and previous studies according to the enrichment conditions, including the microbial source, enrichment substrate and its concentration, and cultivation conditions. Regarding the SDIMO catalyzing the initial rate-limiting DX oxidation step, SDIMO group 5C (THM/DXM) was dominant in the THF-enrichment culture in this study (Figure 4(c); Table 2). Similarly, SDIMO group 5C was dominant in the THF- and DX-enrichment cultures obtained from microbial consortium of industrial landfill leachate in our previous study (Inoue et al. 2020). The THF-enrichment cultures in the two studies were also similar to each other based on the phylogenetic groups of dominant bacteria. These findings may suggest that THF can enrich similar DX-degrading microbial consortia.

In this study, the DX degradation ability in both DX- and THF-enrichment cultures declined between the 6th and the 8th cycle of enrichment (Figure 2). Concurrently, the abundance of the dominant SDIMO groups (groups 5A and 5C in DX- and THF-enrichment cultures, respectively) was slightly declined (Figures 4(c) and 5). In the enrichment culture amended with THF in our previous study (Inoue et al. 2020), the composition of SDIMO genes was unchanged, and the DX degradation ability was continuously enhanced. Therefore, it was suggested that the instability of the dominant SDIMO-carrying populations in the enrichment cultures led to the worsening of DX degradation ability in this study. In contrast, recent studies reported that DX biodegradation in mixed microbial consortia is not ascribed to a single strain, but occurs via the synergistic effects of different strains (Aoyagi et al. 2018; Chen et al. 2021). In particular, Chen et al. (2021) investigated the pathway and associated enzymes of DX degradation in a microbial consortium using integrated omics techniques, and proposed that DX biodegradation and mineralization occur through multiple pathways attributed to the collaboration of multiple strains, including those capable of DX degradation and those incapable of degrading DX but capable of transforming DX degradation intermediates. This study only focused on the microbial populations carrying SDIMOs, which are the most probable enzymes for the initial oxidation of DX to identify the potential contributors to DX degradation, and did not investigate the DX degradation intermediates and bacteria capable of degrading them. However, different populations coexisted and fluctuated during the enrichment process where limited substrates (THF or DX at 100 mg-C/L) were amended (Figure 3). Thus, similar cooperative DX degradation by multiple populations likely occurred and disturbance of the balanced synergy may cause the decline of DX degradation ability in the enrichment cultures of this study. Therefore, it is suggested that not only the enrichment of efficient DX-degrading populations, but also the balanced maintenance of overall populations that degrade DX and its intermediates and indirectly support the DX degradation are necessary to stably achieve a high DX biodegradation in a mixed microbial consortium. Further study is required to identify microbial populations responsible for the degradation of DX and its intermediates and their interactions that affect the overall DX degradation ability in a mixed microbial consortium, in addition to the in-depth understanding of DX biodegradation intermediates and pathways.

This study determined the effectiveness of THF at enriching DX-degrading bacteria and enhancing DX degradation ability in activated sludge lacking historical DX exposure. The results revealed that THF can induce the DX degradation ability as effectively as DX through enrichment of bacterial populations with DX metabolic ability, and indicated the utility of THF as a stimulant to awaken and enhance the DX degradation ability in microbial consortia, irrespective of the strength of the intrinsic ability. The enriched microbial populations differed according to the substrate, and SDIMO group 5C (THM/DXM) was the dominant contributor in the initial DX oxidation in the THF-amended enrichment culture. The weakening of the DX degradation ability in the enrichment culture after a noticeable increase suggested the importance of not only the stable maintenance of dominant SDIMO-carrying populations but also the balanced maintenance of dominant SDIMO-carrying populations and other populations to maintain high levels of DX biodegradation in a complex microbial consortium.

This study was partially supported by JSPS KAKENHI (grant numbers JP16K12624 and JP19H04301). We would like to thank Editage (www.editage.com) for English language editing.

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

The authors declare there is no conflict.